Download Review of information, policy and legislation on species translocation

yes no Was this document useful for you?
   Thank you for your participation!

* Your assessment is very important for improving the work of artificial intelligence, which forms the content of this project

Document related concepts

Extinction wikipedia, lookup

Conservation biology wikipedia, lookup

Occupancy–abundance relationship wikipedia, lookup

Biological Dynamics of Forest Fragments Project wikipedia, lookup

Latitudinal gradients in species diversity wikipedia, lookup

Ecological fitting wikipedia, lookup

Theoretical ecology wikipedia, lookup

Bifrenaria wikipedia, lookup

Introduced species wikipedia, lookup

Island restoration wikipedia, lookup

Molecular ecology wikipedia, lookup

Biodiversity action plan wikipedia, lookup

Habitat wikipedia, lookup

Reconciliation ecology wikipedia, lookup

Habitat conservation wikipedia, lookup

JNCC Report
No. 261
Review of information, policy and
legislation on species translocation
A report commissioned by the Joint Nature Conservation Committee
as a background for future policy formulation
J.M. Bullock, K.H. Hodder, S.J. Manchester & M.J. Stevenson
Institute of Terrestrial Ecology
Furzebrook Research Station
Wareham, Dorset
BH20 5AS
Further copies of this report can be obtained from:
Species Conservation Branch
Joint Nature Conservation Committee
Monkstone House
City Road
Peterborough PE1 1JY
ISSN 0963-8091
Review of information, policy and legislation on species translocations
SUMMARY ..............................................................................................................................5
Background ........................................................................................................7
Methodology and structure ................................................................................7
Development of workable definitions................................................................8
Development of a framework for assessing the environmental effects of
translocations ...................................................................................................15
1.4.1 Biodiversity..........................................................................................15
1.4.2 A refinement of definitions..................................................................17
1.4.3 Methods for quantifying changes in biodiversity ................................20
Scientific background ......................................................................................21
1.5.1 Genetics................................................................................................22
1.5.2 Ecology ................................................................................................24
Background to the main legislation concerning translocations………………26
TRANSLOCATIONS OF SPECIES NATIVE TO THE UK ................................27
2.1.1 Background ..........................................................................................27
2.1.2 Types of translocations of native species.............................................28
Impacts of translocations of native species and factors affecting their success
2.2.1 Ecological impacts ...............................................................................29
2.2.2 Factors relating to the spread of pathogens and parasites...................29
2.2.3 Genetic impacts....................................................................................31
2.2.4 Factors relating to successful translocation for conservation ..............31
2.2.5 Environmental/economic impacts........................................................32
2.2.6 Factors relating to management ...........................................................33
2.2.7 Factors related to animal welfare.........................................................33
Summary of guidelines ....................................................................................33
2.3.1 UK Guidelines .....................................................................................33
2.3.2 International Guidelines.......................................................................43
Summary of legislation ....................................................................................50
2.4.1 Reintroductions ....................................................................................50
2.4.2 Possessing wild animals.......................................................................51
2.4.3 Possessing wild plants..........................................................................53
2.4.4 Releases into or species removal from protected areas .......................53
2.4.5 Release of captive organisms...............................................................54
2.4.6 Import and release of non-native stock................................................54
2.4.7 Amenity restocking of native species ..................................................56
Translocations for conservation.......................................................................58
2.5.1 Overview of reintroductions ................................................................58
2.5.2 Reintroductions of species extinct in the UK ......................................59
2.5.3 Reintroductions of regionally or locally extinct species......................64
2.5.4 Re-enforcement of existing populations ..............................................70
2.5.5 Creation of new populations to conserve vulnerable wild populations73
2.5.6 Relocations to rescue individuals or small populations.......................74
2.5.7 Conservation seed mixes and plantings ...............................................76
Review of information, policy and legislation on species translocations
Translocations of native species for purposes other than conservation...........83
2.6.1 Fisheries and angling ...........................................................................83
2.6.2 Crustacea and molluscs in aquaculture ................................................87
2.6.3 Gamebirds ............................................................................................88
2.6.4 Falconry ...............................................................................................89
2.6.5 Bee-keeping .........................................................................................90
2.6.6. Translocations and releases for aesthetic purposes..............................91
2.6.7 Releases for animal welfare .................................................................93
2.6.8 Translocations for scientific research ..................................................94
General discussion ...........................................................................................95
2.7.1 General discussion of translocations for conservation.........................95
2.7.2 General discussion of translocations for purposes other than
conservation ...................................................................................................102
Summary conclusions concerning the translocation of species native to the UK
3.1.1 Background ........................................................................................106
3.1.2 Types of introduction.........................................................................106
Impacts of the introduction of non-native species .........................................107
3.2.1 Ecological impacts .............................................................................108
3.2.2 Impacts relating to the spread of disease ...........................................109
3.2.3 Genetic impacts..................................................................................109
Summary of guidelines ..................................................................................109
3.3.1 UK Guidelines ...............................................................................................110
3.3.2 International Guidelines.................................................................................112
Summary of legislation ..................................................................................117
3.4.2 Import, keeping, release and control of alien species - UK legislation ...
3.4.3 Releases into protected areas .............................................................123
Fish and shellfish stocking for aquaculture ...................................................123
3.5.1 Background ........................................................................................123
3.5.2 Impacts of introductions ....................................................................124
3.5.3 Discussion ..........................................................................................127
3.6 Biological control.................................................................................................129
3.6.1 Background ........................................................................................129
3.6.2 Impacts of introductions ....................................................................130
3.6.3 Discussion ..........................................................................................132
Wildfowl and game stocking .........................................................................133
3.7.1 Background ........................................................................................133
3.7.2 Impacts of introductions ....................................................................135
3.7.3 Discussion ..........................................................................................137
Amenity and ornamental planting, stocking or collections ...........................138
3.8.1 Background ........................................................................................138
3.8.2 Effects of introductions......................................................................141
Pets and domestic animals .............................................................................145
3.9.1 Background ........................................................................................145
3.9.2 Effects of introductions......................................................................145
Review of information, policy and legislation on species translocations
3.9.3 Discussion ..........................................................................................147
Forestry ..........................................................................................................148
3.10.1 Background ........................................................................................148
3.10.2 Effects of introductions......................................................................149
3.10.3 Discussion ..........................................................................................150
Crop species ..................................................................................................150
3.11.1 Background ........................................................................................150
3.11.2 Effects of introductions......................................................................151
3.11.3 Discussion ..........................................................................................152
3.12 Fur animals.....................................................................................................152
3.12.1 Background ........................................................................................152
3.12.2 Effects of introductions......................................................................153
3.12.3 Discussion ..........................................................................................155
3.13 Accidental introductions ................................................................................155
3.13.1 Background ........................................................................................155
3.13.2 Effects of introductions......................................................................156
3.13.3 Discussion .........................................................................................159
3.14 Discussion on introduction of non-native organisms.....................................160
3.14.1 Perceptions of non-native species in the UK .....................................160
3.14.2 Assessing the potential for invasion and spread of alien species.......161
3.14.3 Negative effects of introductions - present and future.......................164
3.14.4 Control of introduced species - techniques and problems .................165
3.14.5 Regulation and risk assessment of introductions ...............................168
3.14.6 Control of alien species - problems with legislation..........................170
3.14.7 Further research .................................................................................171
3.15 Summary conclusions concerning introduction of species not native to the UK
4.1.1. Background ........................................................................................175
4.1.2 Genetic modification..........................................................................176
4.1.3 Types and uses of GMOs......................................................................177
Summary of guidelines and policies ..............................................................178
4.2.1 UK Guidelines ...................................................................................179
4.2.2 International Guidelines.....................................................................179
Summary of legislation ..................................................................................181
4.3.1 International Conventions and European Legislation........................181
4.3.2 UK Legislation - Environmental Protection Act 1990 ......................181
Environmental impacts of GMO introduction ...............................................186
4.4.1 Types of release and precautions .......................................................186
4.4.2 Future changes in the use and release of GMOs................................188
4.4.3 Potential impacts of GMO release .....................................................189
Discussion ......................................................................................................198
4.5.1 Does genetic modification pose different risks to conventional
4.5.2 Risk assessment for GMO release .....................................................199
4.5.3 Involvement of nature conservation bodies in GMO assessment......200
Review of information, policy and legislation on species translocations
Summary conclusions concerning introduction of genetically modified
organisms .......................................................................................................201
TRANSLOCATION OF SPECIES ASSEMBLAGES .........................................203
Summary of guidelines and policies ..............................................................204
Summary of legislation ..................................................................................205
Case studies....................................................................................................206
Consequences of community translocation ...................................................215
5.5.1 Effects of species translocation..........................................................215
5.5.2 Effects particular to translocations of species assemblages..............216
Discussion ......................................................................................................226
5.6.1 Influences on the outcome of community translocation ....................227
5.6.2 Policy on the use of community translocation...................................230
5.6.3 Increasing the value of a translocation..............................................233
5.6.4 Surveying and monitoring..................................................................235
Summary conclusions concerning translocation of species assemblages......236
GENERAL CONCLUSIONS OF THE REVIEW................................................238
Translocations and their effects on biodiversity ............................................238
Costs of translocations ...................................................................................241
Monitoring and databases ..............................................................................242
Assessments of risks and of benefits - formats for guidelines.......................243
Co-ordination among organisations...............................................................244
A European framework of legislation............................................................246
Education and legislation...............................................................................247
Summary general conclusions of the review .................................................247
BIBLIOGRAPHY ....................................................................................................253
Review of information, policy and legislation on species translocations
This report reviews the available information concerning translocations, as
background for a future policy statement to be drawn up by the statutory UK
conservation agencies (the Joint Nature Conservation Committee, English Nature,
Scottish Natural Heritage, the Countryside Council for Wales and the Department of
the Environment, Northern Ireland).
Translocations of species for conservation reasons in the UK include reintroduction,
population supplementation, sowing and planting for habitat restoration, and
relocation of populations and of species assemblages.
Non-conservation translocations for commercial, amenity and aesthetic reasons are
carried out using native and non-native species, and genetically modified organisms.
Precise definitions of the different types of translocation are given, as well as other
important terms used in the review.
Four subject areas are covered in separate chapters: species native to the UK, species
not native to the UK, genetically modified organisms, and species assemblages. The
types of translocations carried out within the subject area are reviewed and
representative case studies are presented.
A set of definitions of genetic, species and ecosystem biodiversity are determined, and
are used as a framework with which to assess the environmental effects of each type
of translocation.
For all types of translocation, the adverse effects on biodiversity are assessed using
the case studies. For conservation translocations, the factors affecting the success and
benefits of the translocation are determined as well.
Existing policies and guidelines of UK and international organisations relating to each
type of translocation are summarised and assessed in the light of the reviews of case
UK and European legislation and international conventions relevant to translocations
in the UK are also summarised and possible improvements are suggested which
would allow better regulation and the amelioration of adverse effects on biodiversity.
While certain types of translocation are well regulated in the UK (e.g. GMO release,
release of non-resident alien animals), others poorly covered by legislative controls
are translocation of most native species, release of most alien plants or animals
resident in the UK and control of problem alien species.
Some areas require a revision or coordination of the approach to the regulatory
process: GMO releases, translocation of species assemblages, release of non-native
Review of information, policy and legislation on species translocations
Areas requiring further research are highlighted, and these illustrate a need for more
coordinated and structured monitoring and databasing of current and future
A major aspect of any translocation must be consideration of the maximum benefit to
biodiversity and/or the minimum risk of adverse impacts. It is recommended that
project planning and risk and 'benefits' assessment procedures should be carried out
before any translocation takes place.
It is recommended that the formulation of new policy and guidelines by the statutory
conservation agencies should involve other UK and international organisations
involved in carrying out, advising on, or licensing translocations.
Review of information, policy and legislation on species translocations
1.1 Background
Species translocation has developed in the UK in response to two separate needs. Firstly, the
coordinated reintroduction, relocation and population supplementation of UK native species
has become accepted as an important component of nature conservation in the UK. Secondly,
releases of both non-native and native species are carried out for commercial, amenity and
other reasons, especially in agriculture, aquaculture and forestry. Releases of genetically
modified organisms (GMOs) are also likely to increase and to be carried out at larger scales.
These areas are all related and give rise to environmental concerns in terms of their positive
or negative effects on genetic, species and ecosystem biodiversity. The development of
procedures for assessing and implementing translocations has, however, occurred in a series
of largely uncoordinated advances in knowledge. This is not surprising, as each area of
science and conservation has its own distinct set of specialists, each producing
recommendations or guidelines for their area of expertise.
There is a proliferation of information in all areas concerning specific translocations, but the
studies have not been coordinated to present a broad approach to the general problems and
concerns. This lack of an objective overview has lead to the large number of recent magazine,
newspaper articles and books focusing upon such issues. These have offered contradictory
opinions ranging from laissez-faire to predictions of massive declines in native populations
from any form of translocation. There is a growing need, therefore, to address this confusion
by reviewing current issues and understanding, and identifying areas where our knowledge is
One of the objectives of Biodiversity - The UK Action Plan, published by the UK
Government in 1994, is to 'update and publicise guidelines on translocations, reestablishments, introductions and restocking' of species. To fulfil this objective, the Joint
Nature Conservation Committee (JNCC) is co-ordinating the production of a policy
statement, to be drawn up, jointly by the JNCC, English Nature, Scottish Natural Heritage,
the Countryside Council for Wales and the Department of the Environment Northern Ireland.
This review aims to supply the background information needed to inform this policy
statement. It should be noted however, that the review is not exhaustive (that would produce
an enormous document), but rather uses selected case studies to assess the state of current
1.2 Methodology and structure
The purpose of the project is to produce a thorough review of current knowledge and expert
opinion in the area of species translocation and to evaluate existing legislation, policy and
case histories to inform the formulation of new policy.
The introductory section provides some background information necessary to the review.
This comprises: summaries of relevant legislation and scientific theory; precise definitions of
Review of information, policy and legislation on species translocations
terms to be used; and the development of a framework for assessing the impacts of
translocations on biodiversity.
The main chapters of the review reflect the four main types of translocation carried out in the
UK: translocations of species native to the UK, translocations of species not native to the UK,
introduction of GMOs, and translocations of species assemblages. The last subject area
strictly falls within the translocation of native species, but the importance of the subject
merited a separate chapter. Within each chapter a definition and overview of the subject area
is followed by a review of all recent policy and guideline statements made by UK, European
and international organisations with an interest in the subject. These were obtained by direct
contact with the organisations or by reviewing the relevant publications. All UK and
European legislation and international conventions relevant to the type of translocation are
also reviewed.
Case studies and scientific information on the type of translocation were obtained by direct
contact with individuals and organisations from the UK and abroad, by requests on
international email and by searches on the bibliographic database BIDS and using CD-ROM
Silver Platter searches. These are organised into categories dependent on the motives for the
translocation and are reviewed in the context of the framework of the consequences for
biodiversity. Most case studies are from the UK - reflecting the UK bias of the review - but
others are drawn from abroad to provide extra information and to aid the search for
generalities. The case studies and published assessments are used to draw conclusions on the
impacts of each type of translocation. The individual discussion sections consist of reviews of
general results, gaps in knowledge concerning impacts, the benefits and drawbacks of
policies and legislation in the light of the results, and recommendations for future actions.
The final chapter draws together some general conclusions and recommendations from this
review. A list of contacts is included. A bibliography containing our references and the
results of our literature search is listed in a separate volume.
1.3 Development of workable definitions
Before discussing the problems and consequences of translocations it is important to define
precisely the terms to be used. The existing terminology for translocations - introduction,
reintroduction, restocking, re-establishment, etc. - has a different set of definitions in each
document. The main documents we consider below use subtly different definitions, but other
documents (e.g. JCCBI 1986) are radically different. We base some of our definitions on
those of the IUCN and NCC guidelines and policy statements, although some modification
has been necessary. Therefore, although some may disagree with the exact terms used (e.g.
some prefer 're-establishment' or 'restocking' to reintroduction), it is the definitions which are
important. The existing definitions are listed and discussed in order to address their
shortcomings. Definitions of other terms (e.g. genetically modified organism, population,
genotype) follow the current consensus of opinion.
Review of information, policy and legislation on species translocations
IUCN guidelines 1987
The International Union for the Conservation of Nature and Natural Resources (IUCN) define
translocation, quite simply, as the movement of living organisms from one area with free
release in another (IUCN 1987). Within this framework, they distinguish three different
classes of translocation.
Introduction of an organism is the intentional or accidental dispersal by human agency of a
living organism outside its historically known native range.
Reintroduction of an organism is the intentional movement of an organism into part of its
native range from which it has disappeared or become extirpated in historic times as a result
of human activities or natural catastrophe.
Restocking is the movement of numbers of plants or animals of a species with the intention of
building up the number of individuals of that species in that habitat. (Later definitions rename
this process reinforcement or supplementation [IUCN 1995]).
These definitions give rise to the question of how one defines an organism. Organism is used
both to mean the individual plant, animal or micro-organism which is translocated, and as a
vague taxonomic unit, to encompass a greater variety of taxa than 'species'. Within this
review we find that much of the controversy surrounding translocations revolves around the
taxonomic unit considered. For example, the reintroduction of a native species may involve a
non-native sub-species, or the opposition to the introduction of a genetically modified crop
plant may not be because it is a non-native species but because of the genetic modification.
NCC guidelines 1990
The Nature Conservancy Council guidelines on species translocations (NCC 1990) provide
more detailed definitions of these terms in relation to issues peculiar to the UK and we use
these as the basis of our definitions. One significant difference from the IUCN definitions is
the extension of the term introduction to include the translocation of organisms from one
region of Britain to another. This provides the basis for assessing the potential impact of a
translocation not in terms of national boundaries (e.g. native vs non-native species), but
through an ecological approach. However, the term 'organism' is again used ambiguously,
being defined: 'includes species, sub-species, race, hybrid, strain, modified or genetically
manipulated fauna, flora, bacteria or viruses; all stages of life cycles'.
Reintroduction - the deliberate or accidental release of a living organism into the wild in
areas where that kind of organism was indigenous in historic times but is no longer present.
Introduction - the deliberate or accidental release of living organisms into the wild in areas
where that kind of organism does not occur naturally, and has not occurred since the last
glaciation (or during historic time). The term applies to introductions into Great Britain or
into a county or habitat within Great Britain.
Review of information, policy and legislation on species translocations
Restocking - the release of a living organism into the wild into an area where it is already
Relocation - the transfer of an organism away from its current site.
IUCN guidelines 1995
The IUCN guidelines produced by the Reintroduction Specialist Group in 1995 refer more
specifically to the translocation or reintroduction of the taxonomic unit species. More
importantly, they acknowledge that a lower denominator (e.g. sub-species or race) may be
referred to, so long as it may be unambiguously defined. This avoids the ambiguity of
'organism', although extra terms are needed to cover the range of meanings implied by
organism, especially for the genetic aspects of translocations.
Reintroduction: an attempt to establish a species (or lower taxonomic unit) in an area which
was once part of its historical range, but from which it has been extirpated or become extinct
('re-establishment' is a synonym, but implies that the reintroduction has been successful).
Translocation: deliberate and mediated movement of wild individuals or populations from
one part of their range to another.
Reinforcement/Supplementation: addition of individuals to an existing population of
Conservation/Benign Introduction: an attempt to establish a species, for the purpose of
conservation, outside its recorded distribution but within an appropriate habitat and
eco-geographical area.
Definitions of terms for this review
Taxonomic, ecological and genetic units
A single living individual of any of the five kingdoms - Animalia,
Plantae, Fungi, Protista or Monera - in any life-stage (i.e. including seeds,
spores, fertilised eggs, etc. and gametes). We will not use this term in the
ambiguous taxonomic sense used in previous definitions.
The genetic constitution of an organism. One can talk of genotypes for
specific characters - e.g. blood types in humans or Mendel's smooth and
wrinkled peas - and of the overall genotype for all of the characteristics of
an individual. Although different individuals of a species may have the
same genotype for a specific character, each individual produced as a
result of sexual reproduction (and therefore genetic recombination) has a
different overall genotype from all other individuals of that species. For
this reason it is inappropriate to talk of native and non-native or local and
non-local genotypes when discussing translocations. A group of organisms
does not have single genotypes that can be designated native or non-native.
It is more accurate to talk of populations and races in this context.
Review of information, policy and legislation on species translocations
A group of organisms, all of one species, within a particular
geographical area. Populations may occur naturally or as a result of
human activity, e.g. in domestication or captivity or as crops or
plantations. Populations of a species are distinguished by their genetic
isolation from
each other, i.e. there is little or no exchange of genes ('gene flow') among
populations and each population forms a separate 'gene pool'. Where
populations are widely separated (e.g. several kilometres or even several
hundred kilometres apart) gene flow can be taken to be insignificant. There
may be some gene flow between adjacent populations (e.g.
metapopulations), but this is much less than the gene flow within each
population. Where the species reproduces sexually the population is
therefore, simply a group of interbreeding organisms. Where reproduction
is asexual such gene flow does not occur. However, single populations of
both types are taken to consist of a single gene pool and, in practice,
populations are distinguished as being groups of individuals of a species
geographically or ecologically separated from other groups of that species.
Ecological separation covers those situations where populations co-occur
but gene flow is restricted, e.g. where the populations breed at different
times of the year.
The population therefore forms a genetic unit which one can describe by
its gene frequencies and even the mean genotype for particular characters,
and these can be used to determine genetic differences between
populations. It is scientifically correct to talk of native and non-native
populations (rather than genotypes) of a species when considering
One or more populations of a species showing genetic differentiation
from other populations of that species. Although populations are more
or less genetically isolated the genetic differences among them may not be
large. However, where the genetic differentiation is large such that there
are clear morphological, ecological or physiological differences the
populations may be called different races. Such differentiation usually
occurs through geographic separation and one race may consist of many
populations in a particular region. Race is a broad term and we use it to
encompass any taxonomic or genetic unit below a species; i.e. it is an
infraspecific taxon. Thus, it incorporates sub-species (a rather arbitrary
allocation of a Latin name to a race of a species, see Allaby 1994), ecotype
(where genetic adaptation to particular environmental conditions has
occurred; e.g. metal-tolerant plant ecotypes), variety (essentially a
synonym of race), cultivar (a race produced by horticultural or
agricultural techniques) and strain (a race with only small differences
from others).
Review of information, policy and legislation on species translocations
Genetically modified organism (GMO) An organism modified by an
artificial technique of genetic modification which is then capable of
either replication or transfer of the inserted genetic material to other
organisms (ACRE 1993). The modification may be either direct organisms modified by molecular techniques - or indirect - the offspring of
these directly modified organisms. Therefore, this definition excludes
organisms in which genetic material has been altered: 1. through natural
mating or natural recombination (thus including conventional breeding), 2.
by mutagenesis, or 3. by cell fusion (plants only).
Synonyms: Genetically engineered organism, Genetically manipulated
organism, Transgenic organism.
The status of a species or race
The CITES criterion for extinction over the whole world is that a species
is not definitely located in the wild during last 50 years. Within an area
(site, region, country, etc.) a species can be judged to be extinct after
extensive searches have found no living individuals after a defined period
of years. The IUCN Red List Categories document defines this as 'when
there is no reasonable doubt that the last individual has died' (IUCN
1994). A supplementary Red List term is 'extinct in the wild', meaning
the situation when a species is known to survive only in cultivation, in
captivity or in a naturalised population(s) well outside of the past
geographical range of the species.
A species or race which occurs naturally in an area. Technically this
includes any species or race whose dispersal into an area has occurred
independently of human activity and could have occurred at any time,
including the present. However it is usually construed as a species or race
that is thought to have occurred in an area since prehistoric times (NCC
1990); subsequent colonisation is assumed to have occurred through an
human agency. The time period is sometimes stated as since at least the
last glaciation, c. 14,000 BP, (NCC 1990) or since the Neolithic, c. 6,000
BP (Webb 1990). Often in this review the area considered will be the UK,
but it may also be at a smaller scale such as a region, county, vice-county
or site.
To separate these two scales we will use the terms:
Nationally native native to a particular country or state;
Locally native
native in a particular region within a country or state,
e.g. a county, vice-county, etc.
For migratory species that spend only part of the year in the UK (e.g. some
birds), it may be necessary to introduce a caveat to the term 'native'.
Species or races which do not breed in the UK could be called 'native 12
Review of information, policy and legislation on species translocations
non-breeding', as opposed to 'native - breeding' for other migratory
A native population is one which occurs within the particular area.
Synonym: Indigenous.
Not native. A species or race that does not occur naturally in an area;
i.e. it has never occurred there or its dispersal into the area has been
mediated by humans. As the converse to 'native', species colonising an
area in historic times, since the Neolithic or since the last glaciation are
commonly described as non-native. A non-native population is one which
occurs outside the particular area.
Synonyms: Non-indigenous, Alien, Exotic.
A non-native species or race which, after escape or release, has
become established in the wild in self-maintained populations.
An organism which has been kept in domestication, captivity
(animals) or cultivation (plants) but which, after escape or release,
now lives in the wild state. This also applies to descendants of such
released or escaped organisms. A feral population is one consisting of such
A species, race or other taxon that is restricted to a particular country
or region is endemic to that area.
Presumed natural area
The geographical range a species or race is thought to have
occupied in historical times. Although 'historical' is vague, this phrase
generally corresponds with the native range.
Concerning translocation
A general term for the transfer by human agency of any organism(s)
from one place to another (based on NCC 1990).
Donor site
Site from which translocated organism(s) originates. 'Donor population'
can be used in the same sense.
Recipient site
Site where translocated organism(s) is released.
Synonym: Release site, Receptor site.
The wild
Any conditions in which organisms can disperse to other sites or can
breed with individuals from other populations (e.g. by dispersal of
pollen, or visitation by these individuals from other populations) (based on
Review of information, policy and legislation on species translocations
NCC 1990). Thus, this can include natural conditions and semi-natural and
agricultural land, gardens, ponds and open glasshouses, but excludes
sealed laboratories and glasshouses. The precise definition depends on the
species involved. Fish in a garden pond with no water outlets (and not
liable to flooding) may not be in the wild, but winged insects or windpollinated plants in the same pond are in the wild.
The NCC (1990) definition is 'any circumstance in which organisms can
freely breed and disperse.' The added criterion that the organisms can
freely breed is not necessary - it is sufficient that the organisms can
A wild animal, plant or other organism is one occurring in the wild.
The deliberate or accidental release of an organism(s) into the wild in
areas (e.g. country, region, site, etc.) where the species or race is not
native. The term applies to translocations within the UK or into the UK
from other countries (development of NCC, 1990 and IUCN, 1995
guidelines). The term also applies to the release of GMOs into the wild.
Conservation introduction The deliberate release for conservation purposes of
individuals of a species or race outside of the native range (based on
IUCN 1995).
Reintroduction The deliberate or accidental release of a living organism(s) into the
wild to areas (e.g. country, region, site, etc.) where the species or race
was native but has become extinct (based on NCC, 1990) (GMOs are
excluded from this definition). Where a species is reintroduced the race
may be non-native. This case may therefore be described as the
introduction of a race as well as the reintroduction of a species.
Reintroduction is sometimes termed as 're-establishment' and some people
prefer the latter, but the former is the term used by the IUCN. We have a
different definition for re-establishment (see below).
Supplementation The translocation of individuals of a species or race into a site where
there is a pre-existing population(s) of that species or race. The
translocated organisms supplement the existing population(s) and may
interbreed with the resident organisms.
Re-enforcement A distinct form of supplementation that is undertaken for
conservation purposes, to increase the population size at a site. An aim of
re-enforcement is for the translocated and resident organisms to interbreed.
A distinct form of supplementation that is undertaken for amenity
purposes e.g. restocking of fish by angling groups or of wildfowl for
shooting. Interbreeding of the translocated and resident organisms does not
necessarily occur.
Review of information, policy and legislation on species translocations
A type of translocation where an organism(s) is transferred away from
the donor site because that site is under threat (based on NCC, 1990).
The species is generally absent from the recipient site. If the species does
already exist at the recipient site, the relocation also results in a
Synonym: Rescue translocation.
The formation of a self-sustaining population of the translocated
species, race or GMO, i.e. some of the organisms survive to produce
offspring. Another definition requires only that some organisms survive.
The former definition can be termed 'permanent establishment' and the
latter can be 'temporary establishment'.
Re-establishment Where a reintroduction results in establishment.
Captive breeding The managed breeding in captivity of animals of a species or race
which usually occurs in wild populations (i.e. non-domesticated species
or races). These animals have been captured from the wild or are the
descendants of captured individuals.
Artificial propagation
A broader term for captive breeding, which encompasses
all kingdoms as well as animals. For example, the propagation of plants
in cultivation by seed production or tissue culture.
Development of a framework for assessing the environmental
effects of translocations
In this section we shall develop a set of criteria which can be used in carrying out a
standardised and objective appraisal of the effects of a translocation. These have two
purposes: 1. to allow assessment of the aims and outcomes of the case studies which we shall
be examining; and 2. to provide a framework for judging the desirability, and the
environmental risks, of translocations in the future.
1.4.1 Biodiversity
The Convention on Biological Diversity (Anon 1994a), agreed at the 'Earth Summit' in Rio
de Janeiro in 1992 lead to the UK Action Plan (Anon 1994b), produced by the Government,
and Biodiversity Challenge (Anon 1994c), produced by the UK voluntary conservation
organisations. Both documents state that a major policy aim should be to increase or maintain
biodiversity in the UK. We therefore use biodiversity as a yardstick with which to measure
the effects of a translocation. One can assess a translocation: 1. carried out for conservation
Review of information, policy and legislation on species translocations
objectives (e.g. a reintroduction) in terms of whether it increases biodiversity or acts to halt or
retard a decline in biodiversity; 2. carried out for conservation or other motives (e.g. an
introduction of a genetically modified crop plant) in terms of whether it significantly
decreases biodiversity.
A definition of biodiversity is therefore necessary. In its broadest sense biodiversity is the full
variety of life on earth, or in our case, the UK. However, a more precise set of definitions is
required to allow direct measurements and comparisons of biodiversity; but the definitions
must also encompass the range of attributes of the broad definition - one must decide on
reasonable measures to describe the variety of life. We will develop a set of definitions to
include the full range of the attributes of biodiversity as specified in the numerous recent
publications on this subject (especially the UK Action Plan and the Biodiversity Challenge).
We shall not arrange these definitions in terms of priority - e.g. is maintaining the individual
genetic structure of a population more or less important than increasing population size? - but
rather we shall construct a list for use and subsequent discussion in this review.
Biodiversity is often divided into three categories: species, genetic and ecosystem (e.g. The
UK Action Plan). Species biodiversity is usually defined as species number ('richness') or by
a measure of the number of species and their relative abundances (i.e. synonymous with
'species diversity') . The latter measure is usually expressed in one of a variety of diversity
indices of which the Shannon and Simpson indices are the most popular (see Magurran
The IUCN (1980) define genetic biodiversity broadly as 'the range of genetic material found
in the world's organisms'. One measure of genetic biodiversity which is sometimes used is the
number of species. Species can be distinguished from each other by their genetic differences
(which are by definition greater than intraspecific differences) and thus a greater number of
species indicates a greater variety of genotypes. This is a trivial definition and it is more
useful to concentrate on the genetic biodiversity within single species. One suggested
measure is the number of gene copies in a group of individuals, i.e. the number of
individuals. This is also trivial and we should measure genetic biodiversity by assessing the
amount of heritable variation in a group of individuals.
Several techniques can be used to detect different genotypes and thus the variety of
genotypes. The most common techniques are: conventional analysis of phenotypic variation,
and using crossing experiments to analyse heritability of characteristics; isoenzyme analysis
(detection of different forms of an enzyme by electrophoresis) or analysis of other gene
products; and molecular techniques to detect DNA differences, including analysis of
Restriction Fragment Length Polymorphisms (RFLPs), Random Amplified Polymorphic
DNA (RAPDs) and microsatellites (see Gray 1995 for an introduction to these techniques).
While these techniques are complex and expensive, they are the only scientifically valid way
to measure genetic biodiversity.
Ecosystem biodiversity is a term created to take account of the variety in nature exhibited on
the scale of the landscape - different ecosystems, communities, species assemblages, etc.
However, it is a vague term and it has not been defined explicitly by parameters that are
accessible to measurement.
Review of information, policy and legislation on species translocations
1.4.2 A refinement of definitions
For the purposes of this review these definitions of biodiversity need to be refined. In doing
this we must reflect the priorities and objectives for conservation in the UK. Thus an increase
in biodiversity involves the attainment, or progress towards the attainment, of one of these
objectives. These objectives are discussed in the UK Action Plan and the Biodiversity
Challenge, and we use these as the basis for our definitions.
UK Action Plan objectives (p15)
To conserve and where practicable to enhance:
a) the overall populations and natural ranges of native species and the quality and range of
wildlife habitats and ecosystems;
b) internationally important and threatened species, habitats and ecosystems;
c) species, habitats and natural and managed ecosystems that are characteristic of local
d) the biodiversity of natural and semi-natural habitats where this has been diminished over
recent past decades.
Biodiversity Challenge objectives (p8)
To conserve internationally important species, habitats and ecosystems and to enhance
their conservation status where possible.
To conserve threatened species, habitats and ecosystems and to enhance their
conservation status where possible.
To conserve species, habitats and natural and managed ecosystems that are characteristic
of local areas and to enhance their conservation status where possible.
To restore degraded ecosystems to their natural status where practicable and to prevent
degradation in all ecosystems by maintaining their natural ecological processes.
To maintain genetic variation within ecosystems.
To contribute to the conservation of biodiversity on a European and global scale.
Some of these objectives are also expressed in the Bern Convention (Council of Europe
1979a) and the Habitats Directive (Council of Europe 1992). The former states a need to
conserve 'wild flora and fauna and their natural habitats', with particular emphasis on
'endangered and vulnerable species, especially endangered ones, and endangered habitats'.
The latter document gives as an essential objective the conservation of 'natural habitats and of
wild fauna and flora', especially 'endangered', 'vulnerable', 'rare', or endemic species and
'priority natural habitats'.
Review of information, policy and legislation on species translocations
Species biodiversity
Species number alone does not describe the species biodiversity of a site or region. The
invasion of Scot's pine (Pinus sylvestris) into a lowland heathland site or the colonisation by a
new non-native species into the UK both increase the regional species number, but neither
increase the species biodiversity of the region if biodiversity is defined in terms of
conservation objectives. For example, smaller heathland sites tend to contain a greater total
number of invertebrate and plant species than larger sites (Webb & Hopkins 1984, Webb &
Vermaat 1990). This is because the species number is elevated by edge effects such that a
large number of vagrant species (common elsewhere but not normally associated with
heathland) have spread onto the smaller sites from the surrounding non-heath vegetation. If
the conservation criterion is to maintain typical heathland species then the smaller sites have
a higher species diversity (species richness) but do not have a higher species biodiversity
(richness of typical heathland species).
Therefore, one definition of species biodiversity is the number of typical species. Within a
patch of a particular vegetation type or biotope1 certain species are typically found, such as
heather Calluna vulgaris and the sand lizard Lacerta agilis on dry heath or birdsfoot trefoil
Lotus corniculatus and the chalk-hill blue Lysandra corridon in chalk grassland. Typical
species need not be confined to a single biotope, but they could be found on a number of
different biotopes. These species are 'typical' to a number of biotopes.
On a larger scale, such as a county, a region of a country, or the whole of the UK, the typical
species can be those designated as native (of course, the typical species in a biotope need not
all be native, e.g. Aesculus hippocastanum in woodland). Thus, typical species are those
which conservationists desire to be maintained within the area under consideration.
As well as the number of these typical species, the number of individuals of these species can
be considered as a measure of biodiversity. Within a biotope patch the number of individuals
of each species can be measured, and at larger scales the total number of individuals can be
estimated or a count made of the number of extant populations. A further development of this
is to look for changes in the geographical range of each species - a decline in the number of
populations may result in a smaller geographical range for the species (see also Pienkowski
Within the set of typical species there is a smaller set of species of higher conservation
priority - the 'priority' species. These are generally those species which are seen as
particularly important element of the flora and fauna of the UK and/or which are particularly
at risk of extinction. Risk of extinction is categorised by the IUCN (1994), from greatest to
least, as 'critically endangered', 'endangered', 'vulnerable' and 'lower risk', and risk is assessed
according to the number of individuals, the rate of reduction in numbers, the species range
and the rate of reduction in range. The UK Action Plan states that priority species - species
an assemblage of species; what is sometimes called a community or a habitat - such as a dry heath, a chalk
grassland or a deciduous woodland.
Review of information, policy and legislation on species translocations
for which conservation targets should be set - qualify under the following categories in
priority order (see p63 in the UK Action Plan).
Globally threatened species
Threatened endemics in the UK
Species of international importance
Species that occur in a Red Data Book and are declining
Species that occur in a Red Data Book
The Biodiversity Challenge also lists over 1,000 UK species which are considered to require
special action for their conservation and these therefore may be seen as priority species. The
Wildlife and Countryside Act 1981 (see also Whitten 1990), the Wildlife Order (Northern
Ireland) 1985 and the European Union Habitats Directive (Council of Europe 1992 - given
effect in the UK in The Conservation [Natural Habitats, &c.] Regulations 1994) give shorter
lists of species which may be seen as requiring special priority. Palmer (1995) gives lists of
priority plant species in the UK which are derived from all these sources. These lists are
compiled for the UK as a whole, but species threatened at a more local scale, such as regions
or counties, can also be assigned priorities, for example using county red data books (e.g.
Mahon & Pearman 1993), or species on boundaries of their ranges.
A decline in numbers or the loss of one of these priority species can be considered to
comprise a greater reduction in species biodiversity compared to a decline in a species not on
the priority list, or one of lower priority. Therefore, in determining changes in biodiversity we
can give a weight to each species according to the conservation priority it has. This weighting
could be mathematical. Vane Wright et al. (1991) (see also Faith 1992, Crozier 1992)
calculated the 'taxonomic diversity' of a set of species by assigning a weight to each species
according to its taxonomic distance (calculated by phylogenetic analysis) from the other
species. However, a mathematical approach would be rather artificial and of little use for the
purposes of assessing translocation impacts. A species-by-species assessment of effects
would be more efficient and informative.
In this set of definitions of species biodiversity only the typical and priority species are
considered. The non-typical species - vagrant or invasive species not usually found on the
biotope or species not native to the region - are simply not included. The presence of a nontypical species should not be considered automatically to constitute a decrease in biodiversity
(i.e. non-typical species being given a negative weight). The effects of non-typical species on
biodiversity should only measured indirectly through their effects on the typical species.
Genetic biodiversity
Above we suggested that a non-trivial measure of genetic biodiversity is the genetic variation
within single species in the area of interest. The individual species looked at may be a typical
species or, more conservatively, certain of the priority species. Within single sites or over
larger areas (e.g. the UK) the genetic diversity of the species can be measured using the
techniques described above.
Another form of genetic biodiversity is the pattern of genetic differences among populations.
Each population of a species has certain frequencies of genes - its 'genetic structure' - which
Review of information, policy and legislation on species translocations
may be different from other populations. A decline in these differences among populations
constitutes a loss of genetic biodiversity.
Ecosystem biodiversity
This form of biodiversity might be better described as 'landscape biodiversity' - 'ecosystem' is
a precise term (Allaby 1994). Landscape biodiversity can be described by using some
'landscape unit' to separate land areas with distinguishable and different assemblages of
species. Such units may be the communities and sub-communities of the National Vegetation
Classification, biotopes, the 32 land classes of Britain in the Countryside Survey (Barr et al.
1993) or even biomes2. Each of these comprises a successively larger spatial unit of
The number of landscape units in a specific area would not be a good measure of landscape
biodiversity - more fragmented landscapes would have a higher landscape biodiversity. As
for species, each type of landscape unit could be assigned a value. For example, a valley mire
has a higher conservation value than an improved grassland. The Habitats Directive, the UK
Action Plan and the Biodiversity Challenge assign priority to certain types of landscape unit
such as lowland heath, limestone pavements, Caledonian pinewoods, etc. (see also the UK
Plant Conservation Strategy, Palmer 1995). These are types of habitat (more correctly,
biotopes) and they are generally types which are declining and contain priority species.
Individual sites within these types are also assigned priority status and can be protected as
SSSIs (ASSI in Northern Ireland), NNRs, Special Areas of Conservation (Habitats Directive),
Special Protection Areas (EC Birds Directive), Marine Nature Reserves and Local Nature
Biodiversity within a region decreases if the areas of particular priority species assemblages
decline. Increases in the area of these assemblages indicates an increase in biodiversity.
Increased fragmentation of priority assemblages may be seen as an additional decline in
1.4.3 Methods for quantifying changes in biodiversity
We can therefore construct a list of definitions of biodiversity. These are practical measures
which should be seen as relative - an increase or decrease in one of these measures compared
to a start point (e.g. before the translocation) constitutes an increase or decrease in
biodiversity. These measures fall into two categories: applying to single populations or
species assemblages; or applying on a larger, regional basis. Two of these measures are the
most often used to assess biodiversity: the number of typical species and the population sizes
of particular species. These are relatively easy to measure, but they are limited in giving only
a partial description of species biodiversity.
biomes are the largest geographical unit describing species assemblages associated with the world's major
climatic regions, e.g. desert, tropical rain forest, temperate forest, etc.
Review of information, policy and legislation on species translocations
Single population/species assemblage
Species biodiversity
1) Number of typical species
2) Number of priority3 species
3) Population size of each typical species
4) Population size of selected priority species
Genetic biodiversity
1) Amount of heritable genetic variation in each population of selected4 species
2) Genetic structure of each population of selected species5
Species biodiversity
1) Number of typical (native) species
2) Number of priority species
3) Total number of individuals of each selected species
4) Number of populations of each selected species
5) Geographical range of each selected species
Genetic biodiversity
1) Total amount of genetic variation in each selected species
2) Amount of genetic difference among the populations of each selected species
3) Pattern of genetic differences among the populations of each selected species5
Landscape biodiversity
1) Area of each priority landscape unit
2) The continuity of each priority landscape unit (i.e. the lack of fragmentation)
1.5 Scientific background
In this section we give a short summary of some ecological and scientific theories and
concepts relevant to a consideration of the benefits and problems associated with
translocations. This is to provide a background to the following chapters and is set out as a
series of definitions.
However defined.
i.e. a set of typical or priority species.
For this measure a change, rather than a decline, indicates a decrease in biodiversity.
Review of information, policy and legislation on species translocations
1.5.1 Genetics
General introductions to conservation genetics are given in the relevant chapters of Soulé
(1986), and Fiedler & Jain (1992) and by Falk & Holsinger (1991), Thornhill (1993) and
Gray (1995). This summary is drawn from these references.
Genetic diversity
The genetic diversity of a population is its amount of heritable genetic variation and there
are a number of ways of measuring this, including isoenzyme electrophoresis, molecular
techniques or morphometrics. Parameters used to describe this variation are: P, the proportion
of loci that are polymorphic; H, the average frequency of heterozygous individuals per locus;
and A the average number of alleles at each locus.
The ability of a population to adapt to a changed environment by selection is considered to be
dependent upon the level of genetic variation; more variation increases the probability of a
population having some individuals or traits pre-adapted to the new environment. This
selection would result in a change in gene frequencies of a population and thus its genetic
structure and may result in a loss of genes from the gene pool - the total amount of genetic
information within a population. However a lack of ability to undergo such genetic
adaptation to an environment may be unimportant if individuals show phenotypic plasticity
- the ability to show phenotypic changes in response to environmental change - which allows
improved performance in the new environment.
Genetic variation differs among populations of a species and among species for a number of
reasons. If a population has been subjected to strong selective pressure for a certain trait then
this may lead to low genetic variation for that trait, or even fixation - where all members of a
population are homozygous for a particular allele at the relevant locus. For traits for which
stability is important, e.g. those involved in breeding, there may be fixation throughout a
Fixation or low genetic variation may also arise through chance and could result in poor
adaptation to the environment or a harmful loss in ability to respond genetically to
environmental changes. It may occur through genetic drift - the random fluctuation in gene
frequencies between generations through chance (rather than selected) inequalities in
reproductive success - if all, or almost all, the successfully breeding individuals in one
generation show no genetic variation for a particular gene locus. Drift to fixation is most
likely in populations with a small effective population size (the average number of
individuals in a population that contribute genes to the next generation - it is usually less than
the number of individuals in the population) as a greater proportion of the population genetic
variation will reside, on average, in each individual.
Review of information, policy and legislation on species translocations
More isolated populations may also show low genetic variation. The tendency of drift and/or
selection (which occur in all populations) to reduce genetic variation can be counteracted by
the spread of new genes from other populations ('gene flow') which may have different
genetic structures through the effect of different selective pressures or a different direction of
genetic drift. Gene flow will be less between more distant populations.
Present low genetic variation can be the result of genetic bottlenecks in the past, when
populations became small through chance (e.g. demographic stochasticity) or environmental
changes, resulting in a loss of genetic variation through drift. Although the population may
have increased again, the genetic variation remains low. Bottlenecks may occur as a part of
founder effects, whereby the colonists of a new population represent only a part of the gene
pool of the original population(s) they have colonised from. The resulting, established,
population may differ quite strongly from the original populations purely because of the
founder effects.
Low genetic variation may also result from increased inbreeding between related individuals
(of similar genotype). Because this can increase homozygosity, it may expose deleterious or
even lethal recessive genes in the offspring, leading to inbreeding depression - a decreased
average vigour (performance, survival, etc.) of individuals in the population. Inbreeding is
likely to be more common in small populations.
Because of founder effects, bottlenecks, selection and drift, populations of a species can
differ widely in the amount of genetic variation. However, species may differ in the average
amount of genetic variation because of their evolutionary history (past changes in gene
pools), ecology (species that exist in small and/or isolated populations may show lower
variation for the reasons given above) or mating systems.
Because small population size can lead to a decline in genetic variation and cause inbreeding
depression it has been suggested that conservationists should utilise the concept of Minimum
Viable Populations (MVP - see Soulé 1987), whereby populations should be maintained at a
large enough size to avoid inbreeding depression and loss of genetic diversity.
However, it is difficult to know what the MVP of a species is, and whether one can generalise
at all given the differences between populations of a species. Certainly, there can be no
generalisations over species, and the often-quoted MVPs of 500 (effective population size) to
allow future adaptive change and 50 to avoid serious effects of inbreeding were not intended
to be taken as universally valid by the original author (see Soulé 1987). Indeed, the ideas that
low genetic diversity leads to poor persistence of populations and a high level of inbreeding
leads to inbreeding depression cannot be applied to all species. A number of species which
are known to have very low genetic diversity over a large number of populations, such as
cheetah Acinonyx jubatus, fallow deer Dama dama, and Furbish's lousewort Pedicularis
furbishiae, persist in nature, and some, such as the plants Spartina anglica, Avena barbata,
Bromus tectorum, and Chondrilla junceum (see Falk & Holsinger 1991, Gray 1995), are very
common and are expanding their ranges. While inbreeding depression is commonly found in
experimental inbreeding of normally outbreeding species, one would expect that inbred
populations which have survived in the wild would have already been purged of the
deleterious recessive alleles, by loss of individuals with these alleles. Thus, the population
would be tolerant of further inbreeding. It is clear that the genetic history and breeding
Review of information, policy and legislation on species translocations
biology of a population is an important predictor of its response to changes in population size
and inbreeding. For instance, if a number of individuals are removed from a population for a
translocation project, they would be more likely to exhibit subsequent inbreeding depression
and other consequences of reduced population size if they are from a large and/or outbred
population than if they are from a small and/or inbred population.
Gene flow
Selection often leads to individuals in a population of a species exhibiting sets of coadapted
traits which are coded for by coadapted gene complexes. These may be fixed or show low
variation and are inherited intact. However gene flow from other populations or interbreeding
with individuals within the population which do not exhibit these traits may lead to a
disruption of the gene complexes, and a decline in performance. This outbreeding
depression may sometimes lead to infertility, as many coadapted traits are concerned with
mating systems. As a general rule, if populations or individuals are more genetically different
the effects of outbreeding depression may be more severe. However, such effects are hard to
predict as outbreeding between distantly related or unrelated individuals may counteract
negative effects of inbreeding and result in increased vigour.
Hybridisation between individuals of different species (or sub-species) can be thought of as
exaggerated outbreeding, and usually involves congenerics. As for intraspecific outbreeding,
it could lead to increased vigour (hybrid vigour or heterosis) or have negative effects.
If there is a great amount of gene flow from one species to the other (e.g. from a crop plant to
an uncommon wild relative), there may be genetic swamping of the latter by the former and
this may result in introgression - the incorporation of the genes of the second species into the
gene pool of the first, with the (at least local) loss of the second species as a distinct entity.
Sympatric species may be less likely to hybridise, or at least to show extensive hybridisation,
because there has been a long period to allow mating barriers to evolve or for introgression to
become complete. However, secondary contact between allopatric species can lead to
extensive hybridisation. This is particularly apparent in birds (Cade 1983).
1.5.2 Ecology
We will only cover two ecological subject areas here of particular relevance to translocations.
General background is given in Soulé (1986, 1987), Gilpin & Hanski (1991), Fiedler & Jain
(1992) and Edwards et al. (1994).
Regional spatial dynamics
Metapopulation theory forms a theory for the large-scale dynamics of species which is
equivalent to island biogeography for communities. If a species forms populations on
discontinuous habitat patches within a region the patches may be linked by inter-dispersal of
individuals (or genes, see below) - forming a metapopulation. Metapopulation theory
provides a set of predictions for the persistence and abundance of a single species in a
landscape. The persistence and size of the metapopulation is determined by the spatial
Review of information, policy and legislation on species translocations
dynamics of individual species in response to the spatial configuration of the landscape. The
longevity and size of populations in biotope patches are affected by isolation and area
Metapopulations will be smaller (i.e. consist of fewer populations) and have a greater
probability of extinction as these patches become smaller and more isolated from each other.
In the classic or Levins-style metapopulation the populations either tend to go extinct
readily or depend upon immigration from other populations to remain extant. There is a
balance of colonisations and extinctions across the local populations, and the probability of
extinction is equal and independent among the populations. Thus, the metapopulation is
fragile and disruption of any of the habitat patches could cause extinction of the whole
metapopulation. Its seems unlikely that many metapopulations will persist in this way and
Harrison (1994) suggests that classic metapopulations are improbable in nature (she also
found few convincing examples) and suggests that three other types are more likely. In
source-sink, mainland-island or Boorman-Levitt metapopulations some populations (on
smaller patches) are prone to extinction and recolonisation, but one or more (on larger
patches) populations persist more or less indefinitely. These persistent 'source' populations
provide immigrants which maintain or recolonise the smaller 'sink' populations. These
metapopulations are less fragile and their survival is dependent only upon the survival of
source populations. Patchy populations are not metapopulations - the species forms a
continuous population over the patches studied and there is good dispersal among these
patches. Non-equilibrium metapopulations are again not real metapopulations - the
populations in a region are virtually isolated from each other and there is no recolonisation
after local extinctions; there are no metapopulation dynamics.
The patchy population concept deserves elaboration. There may be such a great amount of
dispersal among patches occupied by a plant or animal species that it forms a patchy
population (i.e. the patches do not fulfil the definition of population in 1.3). Some animals,
especially vertebrates, will behave differently and will forage between habitat patches, which
therefore provide patches of resource for a population.
As well as dispersal of individuals among populations, there may be genetic metapopulations
whereby there is gene flow among populations and this affects the genetic structure of
populations (see above). Dispersal of genes and individuals are not equivalent (obviously for
plants) - a single individual entering a population may have little effect on the population
dynamics, but may introduce many novel genes.
Minimum viable populations
The concept of Minimum Viable Populations (MVP) states that a species will have a
minimum population size below which the population will go extinct through ecological or
genetic factors. The genetic factors were described above, but the ecological factors related to
small population size may lead to extinction. Demographic stochasticity is the usual chance
variation over time in birth and death rates experienced by a population, and environmental
stochasticity is the random environmental variation over time which may affect a population.
In small populations there is a higher probability that such random fluctuations will affect all
individuals in the same way - e.g. all individuals will fail to reproduce (demographic
stochasticity), or a late frost will kill all individuals (environmental stochasticity) - and that
Review of information, policy and legislation on species translocations
this will lead to population extinction. A MVP has been defined as the smallest isolated
population having a 99% chance of remaining extant for 1000 years. General prescriptions on
population sizes are avoided now, and ecologists tend to use this idea in order to model and
assess extinction possibilities of certain types of population, or of particular target
populations (e.g. Goodman 1987, Guerrant 1992).
Background to the main legislation concerning translocations
The legislation concerning each type of translocation is dealt with in detail in the relevant
chapters. Here we summarise the background of the most important legislation.
Wildlife and Countryside Act 1981
This forms the main statutory provision for species and site conservation in Great Britain. Its
different Schedules list animal and plant species to which certain prohibitions apply. The
Wildlife (Northern Ireland) Order 1985 fulfils the same role for Northern Ireland and has a
virtually identical structure to the Wildlife and Countryside Act, but has different species
listed in its Schedules. One aim of this legislation was to implement the Bern Convention.
The Schedules can be revised and in Great Britain major revisions of the schedules of
protected animals (other than birds) and plants take place every 5 years (the 'quinquennial
Bern Convention on Conservation of European Wildlife and Natural Habitats
The Bern Convention (signed in 1979 and ratified in the UK in 1982) is a Council of Europe
Convention covering the protection of animals and plants and their habitats. This specifically
encourages reintroductions, providing they are well researched, and calls for strict control of
non-native species. The requirements of the Convention are mandatory on the contracting
EC Directive on the Conservation of Natural Habitats and of Wild Fauna and Flora, EC
Birds Directive
The 'Habitats' Directive (92/43) was approved in 1992 and implementation began in 1994. It
was developed to implement the Bern Convention and introduces procedures to conserve
certain threatened European habitats and species through the setting up of Special Areas of
Conservation (SAC). This Directive is implemented in Great Britain in the Conservation
(Natural Habitats & c.) Regulations 1994 and similar provisions are being developed for
Northern Ireland. SACs, along with Special Protection Areas set up under the Birds Directive
(EC Directive on the Conservation of Wild Birds) will form a network of protected sites in
Europe in the scheme NATURA 2000.
Review of information, policy and legislation on species translocations
UN Convention on Biological Diversity
This was signed in 1992 following the UN Conference on Environment and Development
and was ratified in 1994. It contracts the parties to develop programmes and strategies for the
conservation and sustainable use of biological diversity. This led to the development of a UK
Biodiversity Action Plan (Anon 1994a).
2.1 Introduction
2.1.1 Background
Translocations of species native to the UK fall into two types: conservation translocations
and translocation for purposes other than conservation. Translocations carried out for
conservation have become much more common in recent years as habitats have been altered
and fragmented, causing isolation of populations, preventing many species from maintaining
viable populations without human intervention and often rendering in situ conservation
measures inadequate. Consequently, many conservationists and conservation organisations
consider that translocation is likely to become an increasingly important tool in conservation
(Cade 1986, Griffith et al 1989, Maunder 1992, Akeroyd & Wyse Jackson 1995, Bright &
Morris 1994).
Translocations of native species for purposes other than conservation have been widely
practised throughout history in the UK for commercial, sporting and aesthetic purposes. The
popularity of such translocations has varied over the years but recent increases in
translocations of some taxa (e.g. salmonids and lepidoptera) have raised concern about
possible consequences.
The review does not treat cases of accidental escapes - e.g. of Atlantic salmon Salmo salar
from fish farms - separately. The problems associated with escapes are generally the same as
those of deliberate release and the individuals will often have been translocated to the point
of escape.
In some circumstances the motives or perceived objectives may differ from the category
allocated to the translocation. For instance, the unofficial release of butterflies is classified in
this review as an aesthetically motivated activity. However, those involved will frequently
consider that their releases constitute an act of conservation. In other cases positive
conservation may result from a translocation although this was not the primary motive. The
Review of information, policy and legislation on species translocations
release of capercaillie Tetrao urogallus in Scotland on hunting estates constituted a
reintroduction of an extinct species. The results of translocations conducted for scientific
purposes may be applicable to later work on the conservation of species.
In this chapter we consider all types of translocation involving species native to the UK.
These have been split into categories defined by their objectives. Case studies and other
examples are used to illustrate each category, but an exhaustive inventory of UK
translocations of native species, for conservation or other purposes, is beyond the scope of
this review.
2.1.2 Types of translocations of native species
These are split into the following sections.
Translocations for conservation:
Reintroductions (of species extinct in the UK or regionally or locally extinct species)
Re-enforcement of existing populations
Creation of new populations to conserve vulnerable wild populations
Relocations to rescue individuals or small populations
Seed mixtures and plantings (also used for amenity purposes)
Translocations for purposes other than conservation:
Fisheries and angling
Crustacea and molluscs in aquaculture
Translocations and releases for aesthetic purposes
Releases for animal welfare
Translocations for scientific research
2.2 Impacts of translocations of native species and factors
affecting their success
Prior to consideration of any translocation of native species, certain assessment criteria need
to be established. These include the known and potential impacts of translocations and the
factors which might influence success or lack of success. For clarity these criteria are
organised into sections. The contents of the sections are outlined below in order to introduce
the factors which will be considered for each class of translocation.
Review of information, policy and legislation on species translocations
2.2.1 Ecological impacts
Deleterious ecological impacts associated with translocations of native species include any
changes in the interactions between organisms or between organisms and their habitat which
lead to a reduction in biodiversity, specifically changes in species abundance or extinction of
It seems inevitable that translocations, particularly introductions or reintroductions, will have
some effect on ecosystem and community processes at the recipient site, although this effect
may be very small. Translocated individuals may compete with resident conspecifics or other
species. The release of a new predator could have a negative impact on resident prey or the
release of a prey species may attract more predators and lead to increased pressure on
resident individuals. Increased herbivore numbers may affect plant populations.
Removal of individuals for translocations can potentially have negative impacts on the donor
populations if large numbers are removed or if the donor population is too small to sustain the
The obvious benefits of successful re-establishment of species which were nationally,
regionally or locally extinct are the enrichment of local biodiversity and the improvement of
chances of survival of the species concerned.
Re-established species may also increase biodiversity through their impact on the recipient
area. For instance, by altering the habitat so that it can be colonised by other species or by
controlling population size of certain species by herbivory or predation and allowing
populations of other species to increase. The conservation value of an area may also be
generally enhanced by the improvement or creation of habitat for species reintroductions. The
creation of new populations may enhance survival of metapopulations by increasing the
metapopulation size and possibly, by providing source populations to replenish sink sites.
2.2.2 Factors relating to the spread of pathogens and parasites
This is strictly related to ecology but is treated separately because a considerable volume of
literature discusses the dangers of transmission of pathogens via translocations for
conservation and amenity purposes. Declines or extinctions of infected species may occur,
resulting in a loss of biodiversity. The results of some introductions of pathogens overseas
have been dramatic enough to cause considerable alarm that they might reach the UK via
translocations of infected individuals. The losses of Atlantic salmon Salmo salar in Baltic
fisheries due to a skin parasite translocated during aquaculture is an excellent example
(NASCO 1995, see below). However, it may be desirable in a translocation programme to
include the associated diseases and parasites of the species, because these can be considered
Review of information, policy and legislation on species translocations
part of biodiversity. Pathogens and parasites may also serve to regulate the numbers of a
translocated species.
Review of information, policy and legislation on species translocations
2.2.3 Genetic impacts
Supplementation or restocking and the spread of reintroduced individuals and their progeny
into resident populations may affect the genetic structure of the wild populations.
Translocations may also influence the amount of genetic variation among populations and the
pattern of genetic differences between them.
Release of animals from captivity, and transplantation of plants from artificial propagation,
may result in a number of deleterious genetic impacts depending on the genetic constitution
of the captive or cultivated stock.
Captive or cultivated populations of animals or plants are inevitably exposed to selective
factors which are different from those acting on their wild relatives and which may cause
them to become genetically adapted to life in captivity. This is probably particularly true of
species in commercial use where maintaining wild-type characteristics may be a
disadvantage. Individuals from a captive/cultivated population may therefore carry traits
deleterious in the wild environment. Outbreeding depression may put the extant population at
risk if interbreeding occurs between a resident population and individuals released from
captivity. This is particularly important when very large numbers of individuals are released
and cause genetic swamping of the original population.
The use of non-local wild stock may also lead to outbreeding depression when the species
concerned consists of genetically heterogenous locally adapted populations.
Other risks include: exposure of deleterious traits already present in the wild population as a
result of genetic mixing; and the breakdown of genetic barriers which normally prevent
hybridisation with congeners due to introduction of genetically distinct stock.
When resident populations exhibit inbreeding depression or genetic drift and fixation due to
small population size or isolation, introduction of genetic variation through translocations
may be beneficial. It may lead to increased genetic diversity and decreased inbreeding
depression. This may occur as a result of population supplementation or the creation of new
populations in metapopulations.
2.2.4 Factors relating to successful translocation for conservation
This section is used only in the consideration of conservation type translocations. Poor
success of a translocation for amenity purposes is classed as a management issue and treated
under the appropriate section. Ecological, genetic and human factors all apply to the outcome
of translocations.
Review of information, policy and legislation on species translocations
Ecological factors
Habitat requirements at the recipient site may be complex and difficult to identify for
reintroductions. For instance, it may be difficult to identify suitable recipient sites because the
recipient site and the donor, or other extant sites which could be used as models, occur in
different climatic regions. Even when there is a known cause of extinction, such as hunting, a
reintroduction may still fail because the habitat may have become unsuitable since extirpation
of the species.
Environmental or demographic stochastic processes are more likely to prevent the
establishment of populations if small numbers of individuals are released (Caughley 1994).
For instance, distance between individuals may prevent successful breeding or unusual
climatic conditions could cause the demise of a whole population.
For successful animal re-establishments it is essential that the individuals are able to behave
appropriately in the wild. As much of bird and mammal behaviour is learnt during
development, individuals released from captivity are unlikely to survive unless they are
suitably trained or familiarised with the environment of the receptor site.
Genetic factors
In the case of highly endangered species the donor stock may be derived from very few
individuals and this may result in low genetic variability. This might lead to inbreeding
depression and a reduced chance of survival of the population.
If individuals are released which are not locally adapted they may simply fail to survive or
breed. In other cases if the translocated stock is not locally adapted and there is genetic
mixing between these individuals and native populations this may lead to outbreeding
depression. This will not only make the translocation less likely to succeed but may put an
extant population at risk (see 'Genetic impacts').
Human factors
Public attitudes may hinder successful re-establishment of populations. This is particularly
true of animals regarded as pests, dangerous or otherwise undesirable, and also applies to
habitat management for reintroduced species.
2.2.5 Environmental/economic impacts
This category includes alterations to the landscape resulting from translocations that are
considered to be deleterious for financial or amenity reasons. Although this area falls outside
our definitions of biodiversity, it should be considered.
Review of information, policy and legislation on species translocations
2.2.6 Factors relating to management
This section refers to the effects of translocations for purposes other than conservation on
nature conservation and on management of stocks for amenity purposes. Research required
for conservation management may be seriously hindered by unrecorded releases (e.g. Stubbs
1995). Although this does not have a direct effect on biodiversity, the indirect effects on
conservation management may have consequences for biodiversity. Relocations to rescue
individuals or populations or establishment of new populations or communities may also
have indirect effects on conservation because they may promote the attitude that such
measures can replace in situ conservation. A similar effect related to attitudes is seen in
gamebird management where restocking may cause other management techniques required to
protect wild stocks to be neglected (Hudson & Rands 1988).
2.2.7 Factors related to animal welfare
Although welfare issues do not constitute a threat to biodiversity, they need to be taken into
consideration during translocations of vertebrates for any purpose. There is a danger that
animals might suffer during temporary captivity and transport. However, as the survival of
the subjects of a translocation is so important it seems unlikely that their welfare would be
neglected at this stage. A greater risk is posed by unsuitable releases. Animals released into
inappropriate habitat may suffer due to a lack of access to adequate food or shelter. These
considerations obviously do not generally apply to invertebrates although one example of an
insect welfare issue is included.
2.3 Summary of guidelines
This section summarises the policies (statements of position) and guidelines (recommended
procedures) of UK conservation and other organisations. These are not implemented by law,
but represent, at most, agreed codes of conduct.
In many cases the guidelines have been condensed although some sections are transcribed
verbatim. Some terms have been changed to conform with our definitions.
2.3.1 UK Guidelines
General Guidelines
1) Nature Conservancy Council (1990). Review of NCC Policy on Species Translocations in
Great Britain.
Previous guidelines by NCC (1983, 1987) simply reiterated the legal basis of the control of
reintroduction of species extinct in Great Britain and gave no policy statements or guidelines.
Review of information, policy and legislation on species translocations
A discussion document on NCC policy in 1988 (NCC, 1988) drafted some statements but
NCC (1990) was more comprehensive. The Review of NCC Policy on Species
Translocations in Great Britain (NCC, 1990) developed draft NCC policy on all aspects of
translocation. These guidelines are those currently endorsed by EN, CCW (L. Howe, pers.
comm.) and SNH (SNH 1994).
The basic principles of the document were that:
Reintroductions, relocations and supplementations, when carried out for reasons of
nature conservation, can have a valid role and should be encouraged where
Before implementation, each case should be carefully assessed and then monitored
after release.
Accidental translocations should be avoided and those carried out for reasons other
than conservation should be subject to legal controls or consultative procedures.
Assessment of a proposed reintroduction should consider the following.
Reintroductions should be restricted to areas in which the species is present or is
presumed to have formerly occupied in historical times - the 'presumed natural area'.
Will the donor population survive the removal of the individuals for the reintroduction
and will the donor site be adversely affected by this removal?
Does the donor race pose a threat to the genetic integrity of existing species in or near
the release site?
Is the donor race as genetically similar to the extinct population as possible?
Is the donor population healthy?
If the donor population is captive bred, have there been behavioural changes?
Have the factors causing the original extinction at the donor site been identified and
Does the release site have a conservation importance which could be undermined by
reintroduction of a long-absent component of a previous ecosystem?
Is there a programme for monitoring the species and its effects over the anticipated
range of its spread?
Formal consultative and advisory procedures are suggested which could involve the
Each proposal should be considered by a specialist panel.
Each proposal should have a management plan, to include the following information:
details of the species or race concerned, its distribution and ecology;
purpose of the reintroduction;
origin of the donor organisms;
effects on the donor site;
details of possible interactions with other species or races at the release site and
appraisal of consequences ;
Review of information, policy and legislation on species translocations
details of the management team and consultant specialists;
description and reason for choice of the release site;
the habitats surrounding the release site;
the past history of the species or race on the site and reasons for its extinction;
the present or past distribution within the presumed natural area;
assessment of the likelihood of natural recolonisation;
prediction of the rate of re-establishment of the reintroduced species or race into the
community at the release site;
potential food sources at the release site;
prediction of effects of this re-establishment on the community;
the anticipated spread of the species or race away from the release site;
appraisal of possible effects on genetic structure of existing populations near the
release site;
proposed site management;
details of the reintroduction procedure - number and life history stage of individuals
and methods and timing of release;
the monitoring programme;
proposed publicity.
A national register of reintroductions should be established.
Re-enforcement (named restocking in the NCC document)
In general, NCC (1990) stated that:
re-enforcement of declining populations by translocation of organisms may be of
conservation value;
such projects should be well researched and planned;
if the cause of the decline is not removed, re-enforcement is of little use;
They recommended:
consultation with NCC or other specialists before a re-enforcement is carried out;
preparation of a management plan, as for reintroductions;
monitoring of the outcome
In principle, relocation of single species is seen as undesirable and should only be
used as a last resort.
Conservation bodies should be consulted prior to relocation.
Procedure for introductions, reintroductions or re-enforcements should be followed,
as appropriate.
2) Stubbs (1988). Towards an introductions policy
This report was produced by Wildlife Link on behalf of 15 UK conservation organisations as
a contribution to the development of a UK policy on translocation. Guidelines were proposed
Review of information, policy and legislation on species translocations
for assessing introductions (see Chapter 3), reintroductions and re-enforcements (named
restocking). It was proposed that these guidelines would be implemented by a single UK
authority set up to licence those translocations that currently require official authorization,
and releases of biological control agents or GMOs (reiterating the UKINC 1979
recommendation, see below).
Each of the guidelines on reintroductions and single-species relocations was also seen in the
later NCC (1990) document, with one important exception.
All reintroduced species should have protected status, at least in the early stages of the
Extra guidelines were suggested for re-enforcements, which were not all stated explicitly in
the NCC (1990) document.
The cause of the current low population must be understood and eliminated.
The recipient site must be able to support the desired population size.
The re-enforcement should not cause an avoidable loss of genetic integrity.
There should be no risk of transmission of novel pathogens or parasites from the
released individuals to the original population
3) UK committee for International Nature Conservation (1979). Wildlife introductions to
Great Britain. (Linn report)
This review of introduction, reintroduction, re-enforcement (named restocking in the report)
in Great Britain was carried out by the independent 'Working Group on Introductions', with
strong input from the NCC. Many conservation and other organisations were consulted and
the report contained some general recommendations. These recommendations were repeated
with greater detail in the Wildlife Link (Stubbs, 1988) and NCC (1990) documents. A major
recommendation was
the setting up of an 'Introductions Authority' to assess and monitor introductions and
reintroductions and to draw up guidelines.
A difference from the later documents was the statement that reintroductions within Britain,
using species still extant in Britain will have less potential to result in pest problems than
reintroduction of extinct British species. Therefore
reintroductions within Britain (i.e. of extant British species) and supplementations
will generally require less careful assessment and monitoring and may not need to be
considered by the Introductions Authority.
Review of information, policy and legislation on species translocations
4) Society for the Promotion of Nature Reserves (1970). A policy on introductions to nature
The concerns of this document were the same as later guidelines (e.g. NCC, 1990), and were
given with less detail.
5) National Trust. Unpublished guidelines on reintroductions 1989
This internal document sets out a short list of guidelines to be followed in reintroductions of
species which have become locally or nationally extinct in the UK. Knowledge of the
following is required.
The former natural occurrence of the species.
Whether the species is known to have been lost from the site.
Why the species has been lost from the site.
That there is a small chance of natural recolonisation.
That the site contains sufficient habitat.
That there is a source of individuals of 'appropriate genetic form'.
That the donor population will not be put at risk.
The following are also necessary.
Approval by the Trust's Nature Conservation Panel.
Adequate provision for post-reintroduction monitoring.
Guidelines for plants
6) Birkinshaw (1991). Guidance notes for translocating plants as part of recovery plans
This report to the NCC gave a detailed protocol for carrying out reintroductions of plants as
part of species recovery programmes. Many of the recommended actions followed the NCC
(1990) guidelines, with emphasis on the following points.
A good understanding of the species' ecology is necessary - through literature
reviews, consultation with experts, field studies and experiments.
The donor population should have an appropriate genetic structure:
if possible, it should originate from a site close to and/or of a similar habitat to the
recipient site;
if possible, wild populations of rare plants should not be used;
inbred populations should not be used, or should be mixed with other populations;
where it is produced by artificial propagation, it should be checked for hybrids.
Review of information, policy and legislation on species translocations
7) Botanical Society of the British Isles (1991).Guidelines for the transfer of rare vascular
The Rare Plants Translocation Panel (now defunct) was a set up by the BSBI to provide
advice and consultation for proposed translocations of rare vascular plants, comprising those
listed in the British Red Data Book (Perring and Farrell 1983) and several others. Guidelines
were developed over time (e.g. BSBI, 1988), and BSBI (1991) was the most recent version.
Prior considerations
The panel should be consulted before rare plants are translocated - with the exception
of non-invasive, short-lived species translocated into artificial or urban habitats.
Translocations of local, rather than national, rarities should be reported to the
Biological Records Centre and consultation with the Panel is advisable.
Conditions under which consent may be given
Introductions (i.e. into areas or sites where the species is not locally native - in this
document, these are sites with a distance of over 1km from sites where the species has
been recorded) of rare plants are discouraged.
Reintroduction should only be carried out where the species either:
has recently become extinct in the proposed recipient area; or
is threatened in an area, and there is a site(s) within that area (within 1km of occupied
sites) that contains suitable habitat unoccupied by the species.
The recipient site must contain appropriate habitat.
The recipient site must have some form of long-term protection.
The translocation must not threaten other rare plants.
The donor population should be as near as possible, and within 1km, of the recipient
If material other than seed is used, to avoid accidental translocation of other species, it
is recommended to cultivate the plants prior to release.
Guidelines for insects
8) Joint Committee for the Conservation of British Insects (1986). Insect re-establishment - a
code of conservation practice
This committee, representing British entomological groups and other more general
conservation organisations, drew up guidelines for reintroduction (called re-establishment in
this document) of insects. The JCCBI called for a national policy on reintroductions and saw
their guidelines as a precursor to this.
The JCCBI emphasised that each reintroduction project should be considered on its own
merits and that all proposals should be discussed in full. However,
insect reintroduction should be particularly of nationally threatened species.
Review of information, policy and legislation on species translocations
The guidelines were similar, although less detailed, to those of NCC (1990). They stated the
need for:
clear objectives;
understanding of the species' ecology;
understanding and reversal of the cause(s) of extinction;
appropriate habitat at the recipient site;
avoidance of deleterious effects on the donor population - this may be avoided by
captive breeding;
use of a donor population of similar ecological type to the extinct population(s) being nearby and/or environmentally similar;
monitoring of the outcome using standard record sheets;
reporting of results to the Biological Records Centre (BRC) and the JCCBI.
An extra recommendation, not seen in other guidelines was,
inclusion of specific parasites in the reintroduction - to allow the conservation of these
species as well as their host.
9) Oates & Warren (1990). A review of butterfly introductions in Britain
This was a very detailed review of reintroductions, supplementations and introductions (NB,
Oates & Warren's terminology was quite different from ours) of butterflies, for the JCCBI
and WWF. Two major recommendations were that:
there is a need for a detailed national policy to regulate butterfly releases;
there is a need for a national strategy to identify species and regions where
reintroductions would benefit butterfly conservation.
10) The British Butterfly Conservation Society (BBCS) _
Previous BBCS guidelines on butterfly releases were included in Oates & Warren (1990).
These were seen as a further development of the JCCBI guidelines. Although more detail was
given, there were no substantial differences from the previous document and all aspects are
also covered in the NCC (1990) guidelines.
The guidelines were greatly modified in 1995 on the basis of the new IUCN (1995)
guidelines on reintroductions. The rationale for the new guidelines was as follows.
Restoration (i.e. reintroduction or supplementation) of Lepidoptera to their natural
habitats has considerable potential as a conservation measure, where long-term
species decline or extinction has occurred.
However, releases for other motives - to ensure a supply for livestock collections or
breeding or for aesthetic purposes - should be discouraged.
Review of information, policy and legislation on species translocations
This lead to a general policy.
There is a need for a national strategy which targets species and regions for
reintroduction programmes.
Reintroductions should be carried out within a Species Action Plan.
Reintroductions should not be carried out merely because captive stocks exist.
Therefore, the following criteria should be fulfilled before a reintroduction is carried out.
Priority measures of habitat and species protection, management and monitoring are
being carried out.
The species or race must be extinct or threatened with extinction at a national or
regional level or at an important site.
If a reintroduction is carried out, the following should be considered.
The habitat requirements of the species or race should be known.
The reason for decline at the receptor site should be understood and remedied.
Extinction should be confirmed at the receptor site (≤5 years of recorded absence).
The mobility and distribution of the species should be such that natural recolonisation
of the site is unlikely over the next 10-20 years.
The restoration or creation of networks of populations or metapopulation is preferable
to single site reintroductions.
Sufficient numbers of individuals should be used to increase the chances of
establishing a genetically diverse population.
The donor stock should be, as far as possible, the closest relatives of the original
population and genetic studies should be carried out when doubt exists.
The receptor site should be within the historical range of the species.
The donor population should not be harmed.
Other species at the receptor site should not be harmed.
Captive bred stock should be healthy and genetically diverse (i.e. not normally captive
bred for >2 generations).
At least five years of post release monitoring should take place and contingency plans
should exist for possible problems.
Approval should be obtained from the BBCS and other relevant bodies.
The entire process should be fully documented and recorded with BBCS and JCCBI.
11) Sheppard (1995). Guidance notes for invertebrate translocations and introductions English Nature Species Conservation Handbook
This document refers to reintroduction of invertebrates. The guidelines were produced 'in the
absence of any firm policy on such releases'. Translocations are seen, in the context of habitat
fragmentation, as a legitimate means of:
assisting natural dispersal;
spreading the risk of local extinctions.
Review of information, policy and legislation on species translocations
Criteria for project acceptability are very similar to the BBCS (1995) guidelines, requiring:
a justification of the reintroduction
little harm to the existing community of the donor site;
no genetic or other abnormalities in captive bred stock
the donor stock is as genetically similar as possible to the extinct population;
no threat to the 'genetic integrity' of animals at the receptor site;
the species is extinct at the receptor site;
the donor site be within the species' range, although releases into sites with no record
of the species are discouraged;
the receptor site is able to support the species and will be managed appropriately;
no adverse effects on other species at the receptor site;
a programme plan;
contingency plans against failure or effects on the donor population;
reporting of the project to JCCBI and EN.
Guidelines for herpetofauna
12) Conservation Committee of the British Herpetological Society (1983). Herpetofauna
Translocations in Britain - A Policy
13) English Nature (1994a). Translocations: rationale and objectives
14) Ginn (1983). The ecology and conservation of amphibian and reptile species endangered
in Britain
All three documents gave some guidelines on the reintroduction and supplementation of
native British herpetofauna. Ginn (1983) and EN (1994d) were both produced by the NCC or
EN and the latter was a discussion document for a species recovery programme for the sand
lizard Lacerta agilis. The BHS guidelines were produced because of a perceived need for
specific guidelines for herpetofauna. Apart from one aspect, the contents of all three differed
from each other only in the amount of detail, and contained less detail than NCC (1990).
The one difference was in the policy concerning where one should translocate species. Ginn
(1983) and EN (1994a) used the standard of NCC (1990) and other guidelines in stating that
native species should be translocated only into sites where they have recently become extinct
or within the 'presumed natural area' of the species - the geographical range it is thought to
have occupied in historical times. This requires knowledge of the current distribution of the
species. The BHS (1983) statement on this differs from all other guidelines in stating that
translocation of nationally native species can be used for
'extending distribution to unrecorded or poorly-recorded areas where this seems
reasonable', i.e. the introduction of nationally native species to areas where they are
presumed local aliens.
Review of information, policy and legislation on species translocations
This also accepts translocations without a comprehensive knowledge of the species'
Guidelines for birds
15) RSPB
The RSPB have no published guidelines on reintroductions, but G. Williams has given us a
policy statement. G. Williams states that this has been influenced by the IUCN (1995)
The RSPB see reintroductions as a potentially valuable tool for:
re-establishing important elements of national biodiversity;
contributing to the long-term survival of vulnerable and declining species by reextending their range;
to serve as examples of positive conservation to generate support for conservation and
encourage conservation-friendly land-use;
to encourage cooperation between collaborating organisations.
However, the RSPB will only support international reintroductions from overseas to the UK
if all of the following criteria are fulfilled.
The species became extinct largely through human activities.
Suitable habitats are available and there are not likely to be serious effects on other
Natural recolonisation is unlikely.
There is no risk to the 'genetic constitution' of other populations of the species by the
introduction of other, potentially interbreeding, sub-species.
The birds are obtained legally and without detriment to donor populations.
Necessary licences have been obtained.
The programme is fully documented.
The translocation is to an area formerly occupied by the species.
These also apply, where relevant, to translocations within the UK.
16) Wildfowl and Wetlands Trust. Black (1991). Reintroduction and restocking: guidelines
for bird recovery programmes
These guidelines, written by J. Black of the WWT, were the result of a meeting organised by
the WWT and the International Council for Bird Preservation in 1988 and we take them as
official policy of the WWT.
The main considerations in this policy were as follows.
Reintroduction and re-enforcement are useful only when a population has been lost or
reduced to critically low levels and no other measures will restore a viable population.
Review of information, policy and legislation on species translocations
Such programmes must be part of wider conservation efforts.
Supplementation is risky because of possibilities of introduction of disease,
introduction of 'disruptive genetic material' or causing competition.
A great deal of preparation should go into a reintroduction or re-enforcement programme and
there should be a subsequent long-term input. A number of stages are suggested.
An initial feasibility study should consider:
the species' ecology and habitat requirements;
effects of environmental changes in the receptor site on the species;
a cost-benefit analysis for the local human populations;
the number and taxonomic status of the donor stock (it must be as genetically and
ecologically close as possible to the original population).
The programme should only then go ahead if:
the receptor site can support a viable population and the original causes of the species
decline have been remedied; the original population is extinct or small;
disruption to humans is minimized and agreed.
Planning should include:
education of local human populations;
observation of regulations;
development of an optimal captive breeding programme,
including teaching of young birds;
health screening of birds;
development of an optimal release strategy;
identification of indicators of success.
Release and monitoring should involve:
releases of birds of appropriate age and in appropriate group sizes;
monitoring of the population and the community and ecosystem.
Finally, there should be periodic assessment of the success of the programme, with
adjustments of techniques where necessary, and documentation and dissemination of
both successes and failures.
2.3.2 International Guidelines
General Guidelines
17) IUCN (1987). The IUCN position statement on translocation of living organisms
The reintroduction section was superseded by IUCN (1995). The later document also
encompassed re-enforcement of native species (named restocking in IUCN 1987), but had
lost some more specific recommendations made in the earlier guidelines.
Re-enforcement may be useful where:
it is considered that a small population may be becoming dangerously inbred;
a population has dropped below critical levels and recovery by natural growth will be
dangerously slow;
Review of information, policy and legislation on species translocations
artificially high rates of immigration are required to maintain outbreeding between
isolated populations on biogeographical islands.
Re-enforcement should only be considered where:
the apparent non-viability of the population is due to genetic problems and not from
habitat degradation or over-utilisation of the population;
the causes of population reduction have been removed and natural increase still does
not occur;
the desired population size is sustainable.
Attention should be paid to the genetic constitution of stock for re-enforcement.
In general, genetic manipulation of wild stocks should be minimised to reduce effects
on the ability of the species to survive.
Genetically impoverished or cloned stocks should not be used.
Source of stock for re-enforcement.
The organisms to be released must be of the same race (race is not defined) as the
recipient population.
Where the recipient population is at the edge of the species range, the source stock
should be obtained from populations living under the same conditions - to maintain
any ecotypic adaptations.
If zoo stock is used, the breeding history and origin of the animals should be known.
If zoo stock is used, the dangers of introducing novel diseases into wild populations
must be avoided.
If supplementation is purely to release captive animals, rather than to re-enforce wild
populations, it is safer to carry these out as reintroductions into sites where there are
no pre-existing populations.
18) IUCN (1995). Guidelines for reintroductions
The Re-introduction Specialist Group of the IUCN's Species Survival Commission developed
these guidelines for reintroductions and re-enforcements, in response to the '... increasing
occurrence of re-introduction projects world-wide, and consequently, to the growing need for
specific policy guidelines to help ensure that the reintroductions achieve their intended
conservation benefit, and do not cause adverse side-effects of greater impact' (these
guidelines did not cover reintroductions or restockings for short-term, sporting or commercial
purposes - where there is no intention to establish a viable population). These expanded and
superseded the less comprehensive Position Statement on the Translocation of Living
Organisms (IUCN, 1987). The main messages were that
'each re-introduction proposal should be rigorously reviewed on its individual merits'
and that
're-introduction is always a very lengthy, complex and expensive process'.
Review of information, policy and legislation on species translocations
The guidelines were developed in the context of IUCN's broader policies on biodiversity
conservation and sustainable management of natural resources and were based on extensive
review of case-histories and wide consultation across a range of disciplines. These are
influencing domestic policy. Butterfly Conservation (M. Warren pers. comm.) and RSPB (G.
Williams pers. comm.) policies are strongly influenced by this document, and SNH also
refers to the IUCN guidelines (A. Kerr pers. comm.).
Aims and objectives of reintroduction
The principal aim should be to establish a viable, free-ranging population in the wild
of a species or race which has become globally or locally extinct in the wild. It should
be reintroduced within the species' former natural habitat and range and should
require minimal long-term management.
The objectives of a reintroduction may include:
to enhance the long-term survival of a species;
to re-establish a keystone species (in the ecological or cultural sense) in an ecosystem;
to maintain and/or restore natural biodiversity;
to provide long-term economic benefits to the local and/or national economy;
to promote conservation awareness; or a combination of these.
Multidisciplinary approach
A team with a full range of suitable expertise is required, possibly including persons
from governmental natural resource management agencies, non-governmental
organisations, funding bodies; universities; veterinary institutions, zoos (and private
animal breeders) and/or botanic gardens.
Pre-project activities
a. Biological
i. Feasibility study and background research
The individuals to be reintroduced should preferably be of the same subspecies or
race as those which were extirpated.
It can be helpful to make a study of genetic variation within and between populations
of the species.
Detailed studies should be made of the status and biology of wild populations (if they
exist) to determine the species' habitat requirements, population ecology and
behaviour, spatial dynamics, predators and diseases.
If any other species has taken the place of the species concerned at the release site(s),
the consequences should be assessed.
The build-up of the released population should be modelled under various sets of
conditions, in order to specify the optimal number and composition of individuals to
be released per year and the numbers of years necessary to promote establishment of a
viable population.
Review of information, policy and legislation on species translocations
ii. Previous reintroductions.
Previous reintroductions of the same or similar species should be used to inform the
reintroduction protocol.
iii. Choice of release site.
The release site should be within the natural habitat and historic range of species. For
a re-enforcement there should be few remnant wild individuals. For a reintroduction,
there should be no remnant population because the reintroduction may lead to disease
spread, social disruption and introduction of alien genes.
iv. Evaluation of reintroduction site
The site should satisfy the habitat and landscape requirements of the species for the
foreseeable future. The area should have the carrying capacity to support a viable
population in the long term.
The previous causes of the decline of the species at the site should be identified and
eliminated. This could involve a habitat restoration programme.
v. Availability of suitable release stock
It is desirable that source individuals should come from wild populations.
If there is a choice, the source population should ideally be closely-related genetically
and show similar ecological characteristics to the original native stock.
Artificially propagated stock must be from a population which has been soundly
managed both demographically and genetically, according to the current principles of
conservation biology.
Reintroductions should not be carried out merely because captive stocks exist, nor
solely as a means of disposing of surplus stock.
The donor population (wild or captive) should be assessed to ensure that it will not be
endangered by removal of individuals.
Animal stock must be subjected to a thorough veterinary screening. Any animals
having non-endemic or contagious pathogens must be removed.
vi. Release of captive stock
Most species of mammals and birds rely heavily on learning as juveniles for their
survival; they should be given the opportunity to acquire the necessary information to
enable survival in the wild, through training and their captive environment.
b. Socio-economic and legal requirements
Socio-economic studies should be made to assess impacts, costs and benefits of the
re-introduction programme to local human populations as well as an assessment of
attitudes of local people to the proposed project.
Adequate measures should be taken to minimise negative effects of human activities
in the re-introduction area.
Review of information, policy and legislation on species translocations
If the species poses potential risk to life or property, these risks should be minimised.
Relevant existing provincial, national and international legislation and regulations
should be checked.
Planning, preparation and release stages
Approval of relevant government agencies and land owners, and coordination with
national and international conservation organisations.
Construction of a multidisciplinary team for all phases of the programme.
Identification of short-and long-term success indicators and prediction of programme
Securing adequate funding for all programme phases.
Design of pre- and post- release monitoring programme.
Appropriate health and genetic screening of release stock.
Appropriate veterinary or horticultural measures as required to ensure health of
released stock throughout programme.
Development of transport plans for delivery of stock to the country and site of reintroduction.
Determination of release strategy (acclimatisation of release stock to release area,
behavioural training, group composition, number, release patterns and techniques,
Training of individuals involved in the long-term programme; public relations;
involvement where possible of local people in the programme.
The welfare of animals for release is of paramount concern through all these stages.
Post-release activities
These must comprise:
post release monitoring - direct (e.g. tagging, telemetry) or indirect (e.g. spoor,
demographic, ecological and behavioural studies of released stock;
study of processes of long-term adaptation;
investigation of mortalities;
interventions (e.g. supplemental feeding, veterinary aid, horticultural aid) when
decisions for revision or discontinuation of programme where necessary;
habitat protection or restoration to continue where necessary;
continuing public relations activities;
evaluation of cost-effectiveness and success of reintroduction techniques;
publication of results.
19) Boitani (1976). Reintroductions: techniques and ethics
This WWF sponsored conference produced a manifesto on species translocations which was
agreed by participants from six European countries. All of the concerns raised are also in the
later IUCN guidelines.
Review of information, policy and legislation on species translocations
Guidelines for plants
20) Akeroyd & Wyse Jackson (1995). Handbook for botanic gardens on the reintroduction of
plants to the wild
This report for Botanic Gardens Conservation International and the IUCN Species Survival
Commission set out in detail the procedures a botanic garden should go through in planning a
plant reintroduction project. The subjects covered and concerns raised were essentially the
same as in IUCN (1995), although there was more discussion. The most important aspect of
this extra discussion concerned genetic aspects.
For existing cultivated or wild-collected material.
Have there been studies of the genetic variation of existing wild populations and
cultivated material?
Is there evidence for inbreeding depression, genetic erosion or hybridisation?
When collecting wild material.
Samples should try to capture 95% of the genetic variation of the population.
Guidelines for aquatic organisms
21) International Council for the Exploration of the Sea (1995). ICES code of practice on the
introductions and transfers of marine organisms 1994
This code covered translocation of species within their native ranges (what ICES called
'transferred species') as well as introductions outside native ranges; where both are for
aquaculture. The same guidelines apply for both translocation types and are considered under
22) North-east Atlantic Commission (1995). Introductions and transfers including the
amendments proposed by the European Union
These proposed guidelines for translocation of salmon Salmo salar for farming, ranching or
restocking were produced by this Commission in consultation with the EU because:
translocation of Atlantic salmon poses genetic, ecological, disease and parasite risks
to the wild Atlantic salmon.
A number of recommendations were made.
Movements into the Commission area of Atlantic salmon or their eggs originating
from outside the Commission area should not be permitted.
Epidemiological zones (zones free of specific pathogens or parasites) should be
mapped within the Commission area and movements of live salmonids or their eggs
prohibited between these zones. A number of pathogens and parasites are named,
including Gyrodactylus salaris.
Review of information, policy and legislation on species translocations
Procedures for detecting outbreaks of previously unknown pathogens or parasites
should be strengthened, although procedures to reduce the spread of these by
movement of salmonids are not decided.
Movement of salmonids or eggs from hatcheries to areas containing Atlantic salmon
stocks should only take place from facilities regularly inspected for, and free from,
Guidelines for seed mixes and plantings for restoration
23) Anon. (1994d). Wild flower plants and seeds
This guidance leaflet produced by BSBI, JNCC, EN, Plantlife, RSNC and WWF
recommended that:
wild flower seeds should never be scattered in the countryside (i.e. outside of
for landscape use, it is preferable to use seed of native British origin.
24) Akeroyd (1994). Seeds of destruction? Non-native wildflower seed and British floral
This report, produced by Plantlife and endorsed by EN, produced no guidelines, but it
cessation of the use non-native variants and agricultural cultivars;
the use of seed of native British origin;
the use of seed of local provenance.
25) Ministry of Agriculture Fisheries and Food. Guidelines for Farmers on Environmentally
Sensitive Areas
These guidelines, covering ten ESAs, recommended that species-rich seed mixes sown when
reverting arable land to permanent grassland should be:
'indigenous seeds, where practical of British origin.'
The sowing of certain grass species is mandatory and
where wild flowers are sown, these should be typical of the grassland types of the
area, e.g. chalk grassland.
26) Countryside Commission (1995). Countryside stewardship: handbook and application
The countryside stewardship guidelines for recreating grassland on arable land stated that:
where possible this should be by natural regeneration;
Review of information, policy and legislation on species translocations
if sowing is necessary, this should be with a grass-mix appropriate to the area and
soils and 'where possible should be native and of British origin';
wild flowers may be added, by spreading hay from local grasslands, or by including
them in the seed mix.
27) Department of Transport, Scottish Office Industry Department, The Welsh Office,
Department of the Environment Northern Ireland (1993). The wildflower handbook
This is an extremely detailed advice note, produced by the Government departments
concerned with highways for the use of those involved in road building, on the use of
wildflower seed mixes for planting roadsides. Details are given on the ecology of a number of
important species. It also calls for:
Use of only native species.
Preferential use of British/UK strains.
28) English Nature (1992). Flowers in the grass
This sets out methods for sowing species rich grasslands or for using seed-mixes to introduce
species to grasslands. Planning and monitoring of the project are both advised. Seed-mixes
are recommended to contain:
slow-growing, fine leaved cultivars or native strains of grasses, not fast-growing
agricultural or amenity cultivars;
wildflower seed, although there is no recommendation as to the source of seed.
29) Nature Conservancy Council (1988c). Native trees and shrubs for wildlife in the United
This advises on species to be used in the MAFF Farm Woodland Scheme. It recommends
species should only be planted within their native range;
local stock should be used preferentially, especially for certain species.
2.4 Summary of legislation
This section summarises UK and EC legislation and international conventions which are
relevant to controlling translocations of native species.
2.4.1 Reintroductions
Several international conventions and European legislation directly address reintroductions.
Review of information, policy and legislation on species translocations
Bern Convention
The Bern Convention specifically encourages reintroductions. Article 11 requires the parties
to 'encourage the reintroduction of native species of wild flora and fauna when this would
contribute to the conservation of an endangered species'. However, the Convention also
requires that 'a study is first made in the light of the experiences of other Contracting Parties
to establish that such reintroduction would be effective and acceptable'.
Recommendation R(85)15
Recommendation R(85)15 of the Council of Europe on the Reintroduction of Wildlife
Species was developed from the Bern Convention. It laid out a number of points to consider
in planning a reintroduction and suggested that reintroduction projects should be preceded by
ecological and socioeconomic research and should proceed only after the causes of the
species' disappearance have been remedied.
Convention on Biological Diversity
Article 9 of the Convention on Biological Diversity requires the parties to adopt, as far as
possible and appropriate, measures for the recovery and rehabilitation of threatened species
and their reintroduction into their natural habitats under appropriate conditions.
EC Habitats Directive
This has a cautious approach to reintroductions of species requiring strict protection (i.e.
those listed in Annex IV of the Directive). Article 22a requires that member states study the
desirability of reintroducing native species for conservation, taking into account other
experience (in other member states or elsewhere) and the opinion of the public.
UK legislation
Despite the UK being part of the EC and a signatory to the above Conventions, there is no
legislation in place in the UK that specifically encourages or regulates reintroductions. Other
legislation must be used to regulate particular and limited aspects of a reintroduction project.
2.4.2 Possessing wild animals
Wildlife and Countryside Act 1981 and Wildlife (Northern Ireland) Order 1985
The Wildlife and Countryside Act, and the Wildlife (Northern Ireland) Order make it an
offence to take from the wild, control, possess, injure or sell certain animal species. This
legislation could therefore be used, to some extent, to regulate and prevent the translocation
of certain endangered native species from wild donor populations.
Review of information, policy and legislation on species translocations
Under Section I of the Wildlife and Countryside Act and Article 4 of the Wildlife (Northern
Ireland) Order it is an offence to take (i.e. possess alive or dead) any wild bird or egg of any
wild bird. A wild bird is defined as any species that is ordinarily resident in Great Britain or
is a regular visitor to Great Britain in the wild state. Poultry and game bird species are
excluded, as are any birds that have been bred in captivity. Schedule 1 in both pieces of
legislation lists bird species (nearly 100 in the Wildlife and Countryside Act) which are
afforded extra protection. The penalties for offences concerning these birds are greater than
for species not on the Schedule. It is also an offence to disturb nesting birds or the dependent
young of the species on the Schedule. A general exemption is that any bird, including those
species listed in Schedule 1, which is injured can be taken from the wild with the intention of
caring for it and releasing it at a later date. Licences allowing exemptions to these
prohibitions can be issued by the relevant Secretary of State (Environment, Agriculture,
Scottish Office, Welsh Office for different activities) or statutory conservation agency. These
licences can be used for conservation purposes. Some wild bird species, listed in the
Schedules 4, can be kept in captivity if they are registered and ringed. Under the Wildlife and
Countryside Act the registration and ringing must be in accordance with the Wildlife and
Countryside (Registration and Ringing of Certain Captive Birds) Regulations 1982 and the
Wildlife and Countryside (Registration and Ringing of Certain Captive Birds (Amendment))
Regulations 1991.
Schedule 5 of the Wildlife and Countryside Act (nearly 100 species) and of the Wildlife
(Northern Ireland) Order lists certain other animal species, which include mammals, reptiles,
amphibians, insects and other invertebrates, which are given similar protection against
possession as the Schedule 1 bird species. As for birds, injured Schedule 5 animals can be
taken from the wild to be cared for and later released. The Secretary of State and the statutory
conservation agencies can issue licences allowing exemption from these restrictions for
various reasons, including conservation.
Birds and Habitats Directives
The EC Birds Directive lists a number of bird species and the EC Habitats Directive lists a
number of animal and plant species which are afforded special protection. However, very few
UK species are named in these Directives (see Schedules 2 and 4 in The Conservation
(Natural Habitats & c.) Regulations 1994, which implement the EC Habitats Directive in
Great Britain).
Protection of Badgers Act 1992
Under this Act it is an offence to take badgers Meles meles from the wild without a licence. If
the capture of badgers is required for conservation measures, the statutory conservation
agencies administer this licensing.
Conservation of Seals Act 1970
This makes it an offence to take grey seals Halichoerus grypus or common seals Phoca
vitulina during specified close seasons. If thought necessary the Secretary of State can extend
this restriction to other times of the year. Licences for taking seals may be issued, but the
consent of the statutory conservation agencies is necessary.
Review of information, policy and legislation on species translocations
Whaling Industry (Regulation) Act 1934
The catching of all species of cetacean within UK coastal waters is prohibited under this Act
(as amended by Part V of the Fisheries Act 1981).
2.4.3 Possessing wild plants
Wildlife and Countryside Act 1981 and Wildlife (Northern Ireland) Order 1985
These make it an offence to uproot any wild plant unless authorised by the owner or occupier
of the land or a statutory authority. A wild plant is defined as one growing wild and of a kind
which ordinarily grows in Britain in a wild state. (This definition is confusing in that it can be
interpreted to include non-native species which have become widely established in the wild,
e.g. Rhododendron ponticum). Offences could include the digging up of plants for relocation
or for artificial propagation to provide stock for a reintroduction, supplementation or
introduction. Schedule 8 in both pieces of legislation lists a number of species of plant that it
is an offence for anyone to pick, destroy, uproot or sell. Sale provisions cover material which
is alive or dead, also any derivative (including seeds) from the plant. Licences allowing these
activities for several reasons, including conservation, can be granted by the statutory
conservation agencies or the Department of the Environment. This legislation does not
restrict the collection of seed from wild plants for sowing elsewhere unless it is seed of a
Schedule 8 species. Native plants may be grown in cultivation if they have been collected
legally from the wild.
2.4.4 Releases into or species removal from protected areas
Natural Parks and Countryside Act 1949, Wildlife and Countryside Act 1981, Natural
Heritage (Scotland) Act 1991
Sites protected under these UK Acts include National Nature Reserves, Sites of Special
Scientific Interest, Natural Heritage Areas, National Parks and Areas of Outstanding Natural
Beauty. It can be argued that release of organisms into such sites or the removal of resident
organisms may damage such sites. Potentially damaging operations on an SSSI would
generally include taking or releasing species, so the statutory conservation agencies must be
consulted before either of these actions is undertaken
European legislation
Sites protected under European legislation, such as Special Protection Areas (EC Directive on
the Conservation of Wild Birds, 92/43) (and, in the future, Special Areas of Conservation
under the EC Habitats Directive) or international Conventions, such as Ramsar sites
(Convention on Wetlands on International Importance) are protected against damaging
Review of information, policy and legislation on species translocations
actions. Particularly, the Habitats Directive, and the resulting UK Regulations (The
Conservation (Natural Habitats & c.) Regulations, 1994) protect certain sites containing
animal and plant species listed in Annex II of the Directive.
2.4.5 Release of captive organisms
Wildlife and Countryside Act 1981 and Wildlife (Northern Ireland) Order 1985
Individuals of native species of animals may legally be held in captivity. Even species listed
in Schedules 1 and 5 of Wildlife and Countryside Act and of Wildlife (Northern Ireland)
Order may be held either under licence, as a result of tending of injured animals or as a result
of captive breeding. Similarly, native plants may be grown in cultivation if they have been
legally collected from the wild or grown from wild-collected or artificially-propagated
propagules. There are legal constraints against the release to the wild of only very few captive
or cultivated individuals of native species.
Three native bird species are on Schedule 9 of the Wildlife and Countryside Act; the
capercaillie Tetrao urogallus, the white-tailed eagle Haliaeetus albicilla and the barn owl
Tyto alba. It is an offence to release or to allow to escape any individuals of these species.
The barn owl was added to the Schedule in 1992, explicitly to constrain widespread releases
of captive barn owls. This was both to safeguard the welfare of such birds, and for
conservation motives (G. Williams, RSPB and W. Parish, DOE pers. comm.). It is not clear
why the capercaillie or the white-tailed eagle (both species are reintroduced to the UK) were
placed on the first version of the Schedule in 1981. G. Williams of RSPB (pers. comm.)
suggests that the only reason was a general wish to have a measure of control of releases of
these species. G. Williams also points out that the ruddy duck, mandarin duck and Carolina
wood duck were not problem species when they were placed on the Schedule in 1981, but
there was a perception that the releases of these species should be controlled. Altogether, this
suggests a degree of confusion as to the precise purposes of Schedule 9 at the time of its
instigation. However, there is a Schedule 9 Working Group within the statutory conservation
agencies which is attempting to develop a clearer approach to the use of Schedule 9 (M.
Palmer pers. comm.).
The licensing procedure for such releases is the same as that for releases of alien species
under Section 14 of the Wildlife and Countryside Act, although the different purposes of such
releases are considered. Chapter 3 describes the licensing procedure. The Wildlife (Northern
Ireland) Order Schedule 9 contains no native species.
2.4.6 Import and release of non-native stock
Wildlife and Countryside Act 1981 and Wildlife (Northern Ireland) Order 1985
Section 14 of the Wildlife and Countryside Act and Article 15 of the Wildlife (Northern
Ireland) Order (for a background see Chapter 3) prohibits releases of animals 'of a kind' not
ordinarily resident in or regular visitors to the UK without a licence issued by the relevant
Review of information, policy and legislation on species translocations
Secretary of State. DOE have been advised that this can refer to non-native races or subspecies of native animals as well as non-native species (W. Parish pers. comm.). A
reintroduction project involving an alien race of the chequered skipper Carterocephalus
palaemon (M. Warren pers. comm.) required a licence in relation to Section 14 of the
Wildlife and Countryside Act. Therefore, Section 14 or Article 15 will in most circumstances
be used to control reintroductions using donor populations from outside of Great Britain or
Northern Ireland. This is, however, not a possibility for plant species, there being no general
prohibition against the release of non-native plants in Section 14 or Article 15.
The licensing procedure for such releases is the same as that for releases of alien species
under Section 14 of the Wildlife and Countryside Act and is described in Chapter 3.
However, the particular purposes of such releases, and the usual aim to establish selfsustaining populations are taken account of in the risk assessment by the Advisory
Committee on Releases to the Environment (ACRE) and the statutory conservation agencies.
Animal Health Act 1981 and Plant Health Act 1967
These Acts (see Chapter 3 for a full description) could be used to control the imports of
individuals of native species which are carrying pathogens.
Convention on International Trade in Endangered Species of Wild Fauna and Flora
CITES, concluded in 1973, established a list of endangered species for which international
trade is controlled by a permit system. Trade is prohibited, unless in exceptional
circumstances, for species in Appendix 1 and is regulated for species in Appendices II and
III. These Appendices are regularly revised. Export permits are required for international
trade in the species in Appendices I and II, and Appendix I species also require an import
permit. CITES was implemented in the EC in 1984 by Regulation 3626/82, and in the UK by
the Endangered Species (Import and Export) Act 1976, the Control of Trade in Endangered
Species Regulations (Enforcement) 1985 and the Control of Trade in Endangered Species
Regulations (Designation of Ports of Entry) 1985 (the latter two enforced the EC regulations).
In the UK permits are issued by the Secretary of State in consultation with a scientific
authority (currently the statutory conservation agencies). The only relevance to species
translocation is that the import for reintroduction projects of species listed in the CITES
Appendices will require the relevant permits. However, few UK species are in these
EC Balai Directive
Directive 92/65, which came into effect in 1994, is concerned with disease transfer and
regulates the import and export of animals within Europe and between Europe and the rest
the world. Zoos and other animal collections can seek approved status under the Directive by
complying with a strict veterinary protocol. Approved centres can then only receive animals
from other approved centres. In a few cases, this Directive could restrict the import of
animals into the UK for captive breeding programmes or reintroduction projects.
Review of information, policy and legislation on species translocations
2.4.7 Amenity restocking of native species
Salmon and Freshwater Fisheries Act 1975
Section 30 of this Act prohibits the release of any fish or shellfish (Mollusca or Crustacea),
native or alien, into English or Welsh inland waters without a licence from the relevant Water
Authority (now the relevant region of the NRA - see Water Act 1989). Therefore, this
legislation applies to this chapter and to Chapter 3. Two forms of exemption from licensing
are allowed (R. Crawshaw pers. comm.). Under Salmon Act 1986, release into MAFF
registered sites needs no licence under Section 30. NRA are not consulted about registering
of sites. However, MAFF require records of movement of live fish on or off registered sites.
This exemption is aimed at commercial trout and fish farms. The translocation of ornamental
species into ornamental sites for non-angling purposes and into contained sites is also
exempted. This is aimed at fish enthusiasts stocking small ponds (these usually must be <0.5
acre) or aquaria. It is not a statutory exemption, but is a pragmatic attempt to reduce
Licensing of fish and shellfish translocations in inland waters under Section 30 of the Salmon
and Freshwater Fisheries Act 1975
Because the Salmon and Freshwater Fisheries Act gives no guidance as to when to refuse
consent and each NRA region acts autonomously in giving licences, each region has its own
policy. However, these policies are similar and there have been recent moves towards a
national NRA policy (R. Crawshaw pers. comm.). At the moment all the regions tend not to
licence translocations where there are disease problems in the stock or if exotics are involved,
unless the release is into contained waters. Otherwise the regions take a case-by-case
Under the Act and as general NRA policy: consent must be written for each translocation;
consent is never given retrospectively; consent can be revoked at any time; the applicant must
be the person introducing the fish and must accept full liability; and the NRA reserves the
right to require health checks. The application form currently used by Southern Region NRA
requires details on
the position and a description of the site of release;
the conservation status of the release site;
the supplier and source water (donor site) of the fish;
the proposed date of release;
the species, size and number of fish to be released;
health checks carried out.
If fish are to be released into an SSSI the relevant statutory conservation agency must be
consulted. EN require the following information to assess the possible impacts of such a
translocation (M. Gibson pers. comm.).
Review of information, policy and legislation on species translocations
Will new species be introduced?
Will the species have a deleterious effect on existing flora or fauna?
Translocation of certain species is particularly discouraged: e.g. carp or
bream in still waters; or the alien species grass carp, zander, wels catfish,
landlocked salmon or rainbow trout in any waters.
There may be adverse effects on other fish species or wild stock, e.g. by
competition, predation or interbreeding (the latter may be especially
important when captive-bred salmonids are released).
The replacement of a coarse fishery by a game fishery is likely to have
effects on the whole ecosystem.
Will fish populations be significantly increased?
Will the proposed increase exceed the carrying capacity of the water body,
and will fertilisation be needed to sustain their increase? Neither situation is
likely to be acceptable on an SSSI.
Is there likely to be increased predation on invertebrates?
Is there likely to be competition for food with other species?
Will the increase affect fish species of conservation importance?
Will there be changes in water quality?
Will there be increased human pressure on the site, resulting in disturbance to animals
or damage to riparian features?
Salmon Act 1986
Under the Salmon Act the same rules apply for salmon Salmo salar translocations in Scotland
as for freshwater fish or shellfish in the Salmon and Freshwater Fisheries Act 1975. The Act
allows proprietors of salmon fisheries within a district in Scotland to form a district salmon
fishery board which must grant permission before salmon or sea trout Salmo trutta can be
released into rivers or lakes.
Fish Health Regulations 1992 and Fish Health Regulations (Northern Ireland) 1993
These Regulations are described in greater detail in Chapter 3. They prohibit the import of
live or dead fish or shellfish (Mollusca or Crustacea),or their eggs or gametes, from zones not
approved as free from certain diseases. include imports of native species for restocking, e.g.
salmon Salmo salar. Imports must be licensed by MAFF or its equivalents.
Sea Fisheries (Shellfish) Act 1967
This requires a permit to be obtained from MAFF for the release of molluscs or crustaceans
into any coastal waters or adjacent land areas.
Agriculture Act 1986
This provided for the creation of Environmentally Sensitive Areas. Farmers can enter into
agreements with MAFF on the management of areas within ESAs. The guidelines for
reverting arable land to permanent grassland within ten ESAs, recommend that the speciesrich seed mixes sown should be 'indigenous seeds, where practical of British origin.' The
Review of information, policy and legislation on species translocations
sowing of certain grass species is mandatory and where wild flowers are sown, these should
be typical of the grassland types of the area, e.g. chalk grassland.
Fodder Plant Seeds Regulations 1993
These implement the EC Directive 66/401 and control the marketing of fodder plant seeds.
The relevance to conservation seed mixes for creation or enhancement of plant communities
is that the marketing of certain fodder species, listed in Schedule 1, is restricted. The variety
of each of these species sold in a seed mixture must be entered on the National List of Plant
Varieties and the seed must have been grown in accordance with the regulations and have
met all the requirements for its certification. This includes the seed crop having been grown a
minimum isolation distance from other conspecific crops to avoid cross-pollination.
Therefore, if these requirements are not met, e.g. if the seed is grown and harvested in wild
conditions, it cannot be marketed.
2.5 Translocations for conservation
The primary purpose of any translocation for conservation must be to enhance the long term
survival of a species, sub-species or race. This section is split according to the immediate or
specific objectives involved and these are illustrated using case studies. The impacts
associated with each type of translocation and the factors influencing the outcome are then
described and assessed using the case studies and other examples.
2.5.1 Overview of reintroductions
A small number of reintroductions have taken place in the UK in previous centuries, for
instance red squirrels Sciurus vulgaris were re-established in Scotland and Ireland in the 18th
and 19th centuries following releases (Tittensor 1977). More recently, the use of
reintroductions has become increasingly common in conservation projects (Pinder 1979,
Cade 1986, Griffith et al 1989, Maunder 1992, Bright & Morris 1994, Akeroyd and Wyse
Jackson 1995, IUCN 1995). Concerns about the methods used, dissemination of results and
coordination of action between organisations are reflected in the number of guidelines that
have been produced (see section 2.3). Reviews of the subject (e.g. Morris 1986, Yalden 1986,
Maunder 1992) indicate the debate concerning which species are appropriate subjects for
reintroductions. Techniques used in reintroduction projects are discussed in these and other
reviews. They include innovations, such as radio-telemetry and micropropagation, which are
becoming more widely used. Detailed description of techniques is beyond the scope of this
review but their influence on the results of species translocations is discussed where relevant.
Effort invested in reintroduction projects throughout the world has been largely concentrated
on vertebrates. Nevertheless, it is probable that extinction rates of invertebrates in the UK
have matched or exceeded those of vertebrates and vascular plants (Thomas & Morris 1994).
Recently more attention has been paid to other groups but this is often restricted to the more
charismatic examples of those groups. For instance, until recently nearly all insect
Review of information, policy and legislation on species translocations
conservation efforts have involved butterflies (New et al. 1995), although a range of other
invertebrate groups is being considered under the EN Species Recovery Programme and the
SNH Species Action Programme.
It is only in the present century that conservation motives have become important in plant
translocations (Maunder & Ramsey 1994). As a result of the short history of plant
reintroduction work, long generation times of woody plants and the episodic nature of
regeneration, most plant reintroductions can still be regarded as experimental. Evidence has
yet to be presented for their long term success (McMahan 1990, Maunder 1992). The need
for an integrated approach to plant conservation was indicated by Whitten (1990) and this
includes the use of translocations and ex situ collections. Since the 1970s botanic gardens
have become increasingly involved with plant conservation by playing an important role in
reintroduction projects (McMahan 1990, Bramwell 1991, Maunder 1992). Similarly, zoos
have become more involved in producing animals for reintroduction, this function is given as
one of the three main reasons for keeping animals in zoos (Pinder 1977).
Consideration of reintroductions of nationally extinct species and those of regionally or
locally extinct species are treated separately in this review. However, impacts associated with
translocations of species and factors affecting their success, may often be the same or similar
in these and other sections, such as species re-enforcements. Attention is drawn to this where
2.5.2 Reintroductions of species extinct in the UK Background and case studies
Reintroductions in this category may receive considerable media attention but are much less
frequent than translocations of species still extant in the UK. Examples of reintroductions of
species extirpated from the UK are given in tables 2.1 and 2.2.
Well known examples of reintroductions of species extinct in the UK include translocations
from northern Europe of two typically spectacular species, the large blue butterfly Maculinea
arion and the white-tailed sea eagle Haliaeetus albicilla. In addition to their aesthetic appeal,
both species have international conservation status. All five species of the genus Maculinea
are classed as 'endangered' or 'vulnerable' by the IUCN (Thomas 1995a), and the white-tailed
eagle is recognised by the International Council for Bird Preservation (ICBP) as among the
28 most threatened birds which breed in Europe. Consequently, they are priority species for
the conservation of biodiversity in the UK. A less famous example is stinking hawk's-beard
Crepis foetida. This species and M. arion are on the northern margins of their geographic
ranges, and the implications of this for their conservation are discussed below.
The potential impact of some reintroductions are perceived to be so great that proposals for
releases of these species provoke considerable controversy. This is particularly true of the
possible reintroduction of mammals such as the European beaver Castor fiber or the wolf
Review of information, policy and legislation on species translocations
Canis lupus to the UK. The proposed reintroduction of the European beaver to the UK is
treated as a case study.
Large blue butterfly Maculinea arion
M. arion was previously found in the midlands and south west of England. It became extinct
in the UK in 1979 (Thomas 1995a). Re-establishment work using larvae from Sweden has
been carried out by a group of organisations collectively known as the Joint Committee for
the Conservation of the Large Blue (JCCLB). More recently, in 1992, English Nature
initiated a five year recovery programme which involved support and collaboration from
JCCLB and Butterfly Conservation. Early results of reintroduction into habitat restored by
changed management are very promising (English Nature 1994b, Thomas 1995a).
White-tailed sea eagle Haliaeetus albicilla
The last known nesting attempt by white-tailed sea eagles in the UK was in 1916, but prior to
their decline it is thought that there were about 200 pairs on the north and west coasts of
Scotland and Ireland. The species was probably even more widespread earlier in history.
Natural recolonisation from Europe seems extremely unlikely as the species has also suffered
serious declines in Europe and only 19 winter visitors have been recorded from the continent
since 1959 (Elliot et al. 1991, Evans et al. 1994). The first two attempts to re-establish this
species in 1959 and 1968 were unsuccessful. More recent releases of Norwegian eagles on
the Isle of Rhum were initiated by the NCC (Elliott et al. 1991). These resulted in records of
a breeding population of up to 10 pairs. In 1992 releases recommenced with supplementary
releases of 10 juvenile eagles each year (I. Evans in prep.). Evans et al. (1994) concluded that
the breeding population of sea eagles was too small to be viable and that these further
releases were necessary to try to establish a sustainable population.
Stinking hawk's-beard Crepis foetida
Stinking hawk's-beard was last recorded in Kent in 1980. Plants from this site were
fortunately established in cultivation and are now being used to provide source stock for
experimental translocations as part of the English Nature Species Recovery Programme
(English Nature 1994b). The experiments were intended to help understand the ecological
requirements of the species and to achieve a long-term self-sustaining population of the
species at the recipient site (Ferry 1994).
Table 2.1: Examples of species extinct in the UK currently or recently subject to
reintroduction attempts
White-tailed sea eagle Haliaeetus albicilla: SNH, RSPB.
Large blue butterfly Maculinea arion: English Nature-Species Recovery
Programme (EN-SRP), Joint Committee for the Conservation of the Large Blue,
Butterfly Conservation.
Large copper butterfly Lycaena dispar: EN-SRP; University of Keele.
Review of information, policy and legislation on species translocations
Stinking hawk's-beard Crepis foetida: EN-SRP; Royal Holloway College.
Table 2.2: Examples of species extinct in the UK currently subject to feasibility studies or
proposals for reintroduction.
European beaver Castor fiber: SNH; Wildlife Conservation Research Unit
(WILDCRU), University of Oxford.
Wolf Canis lupus: Highland Wolf Fund
European beaver Castor fiber
The European beaver was widely distributed in the UK in historic times (Clutton-Brock
1991) but had become extinct by the 16th century (MacDonald et al. 1995). The desire to reestablish the beaver in the UK is not new. Two 19th century authors (cited in Pinder 1977)
report three releases of Canadian beavers Castor canadensis in this country. Despite the
successful reintroductions of C. fiber in 13 European countries (MacDonald et al. 1995)
beavers are still classed as vulnerable by the IUCN/SSC Rodent Specialist Group (Amori &
Zima 1994). Recently several authors have debated the possibility of reintroducing the beaver
to the UK (Lever 1980, Morris 1986, Yalden 1986, Lever 1994, Halley 1995, MacDonald et
al. 1995) and SNH are currently supporting a feasibility study for reintroduction of this
mammal in Scotland (V. Fleming pers. comm.). Halley (1995) and MacDonald et al. (1995)
suggest that the proposal to reintroduce this species to the UK is justified for a number of
reasons including: the cause of its extirpation is known to no longer be operating; and it
seems that adequate habitat is available. Impacts of reintroductions of species extinct in the UK and factors affecting their
Ecological impacts
Re-establishment of species has potential to cause undesirable ecological impacts and if the
given species is nationally extinct this will often make the impacts hard to predict.
Possible ecological threats include damage to resident populations of prey species through the
reintroduction of a predator such as the white-tailed sea eagle. This is hard to predict but
studies of the effect of goshawks on wood pigeon Columba palumbus populations in Europe
showed that the return of this predator to the UK would be unlikely to have a marked effect
on wood pigeons (Kenward 1979).
Alterations to habitat by reintroduced species may be harmful to other species. Lever (1994)
suggests that reintroduction of the beaver to the UK might impede migration of Atlantic
salmon Salmo salar. Experience from Europe however, suggests that the effects of beavers
on salmon are likely to be slight as dams are rarely sited on the fast flowing areas used by
salmon for spawning (Halley 1995). According to Halley (1995) Norwegian salmon fisheries
are not adversely affected by beaver populations that have increased dramatically since 1900.
Review of information, policy and legislation on species translocations
Reintroductions of keystone species may have major effects on entire ecological
communities, and these may be beneficial for conservation. Proponents of the reintroduction
for the European beaver to the UK point out that it could benefit conservation in more ways
than by conservation of a single species. Beavers in Europe are considered to increase
breeding habitat for waterbirds by stabilising wetlands through dry periods and creating pools
used by spawning trout. Beavers also increase the total biomass of invertebrates by replacing
running water taxa with pond taxa with resulting knock-on positive effects for aquatic
invertebrate feeders (Balodis 1995, Halley 1995, MacDonald et al. 1995). Vegetation around
beaver ponds can benefit native herbivores such as water voles Arvicola terrestris. This
species also uses beaver lodges for shelter as do otters Lutra lutra (MacDonald et al. 1995).
Habitat management required for the reintroduction of this species may benefit the wildlife
value of an area (MacDonald et al. 1995).
Genetic impacts
The reintroduction of a nationally extinct species poses little threat of genetic impact at the
recipient site in comparison with re-enforcements of populations or releases of regionally or
locally extinct species. However, there is some risk that non-native genotypes might not have
genetic barriers to hybridisation with resident congeners.
Factors relating to successful reintroduction
Ecological factors
When a species is extinct in the wild, identification of the recipient site for a translocation
may be difficult and Maunder (1992), dealing with plants, writes that knowledge of the
habitat where original stock was collected is not necessarily helpful because the early stages
in development may have different requirements from the mature plant. In addition,
knowledge of the ecology of an extant population may not be applicable when the population
is located in a different geographic region to the site of the reintroduction. This point is
particularly pertinent as many subjects of reintroduction efforts in the UK are near the
margins of their geographic range, for example, stinking hawk's-beard and the large blue
butterfly. In such cases there may be important differences in ecology between sites in the
UK and locations situated nearer to the centre of the range. Existing populations may be in
decline as well and, therefore, the habitat they are occupying may not be optimal.
Changes in the ecology of the recipient site since the extinction of a species may influence
the outcome of a reintroduction. Territorial competition with fulmars Fulmarus glacialis and
other seabirds is thought to be responsible for the loss of white-tailed sea eagles Haliaeetus
albicilla which were released on Fair-Isle in Scotland. There were no breeding fulmars on the
island when the white-tailed eagle was last resident and their potential effect on the
reintroduced birds was not anticipated (Elliott et al. 1991).
The ecological factors contributing to the extirpation of a species may not be immediately
apparent. Considerable research effort may be required to discover what the ecological
Review of information, policy and legislation on species translocations
requirements of a species are. Studies showed that the large blue butterfly Maculinea arion
was much more ecologically specialised than had previously been suspected. The larvae feed
on wild thyme Thymus praecox, and then the later instars live as predators in myrmicine red
ant nests. Although the larvae are found in the nests of several species of red ant they can
only survive in the nests of Myrmica sabuleti. The UK distribution of this ant species is
heavily influenced by microclimate; it is only found on south facing slopes of heavily grazed
grassland. The gradual abandonment of such slopes by agriculture, followed by the massive
decline of rabbit Oryctollagus cuniculus populations in the 1950s, resulted in a loss of this
habitat type and the extinction of the large blue butterfly (Thomas 1989a, 1995a).
Genetic factors
The main genetic problems associated with failures to re-establish populations are likely to be
due to the selection of genotypes that are not suitably adapted or are from inbred captive
The failure of the reintroduction of the Essex emerald Thetidia smaragdaria which became
extinct in the wild in the 1990s is thought to be due to inbreeding and the poor performance
of captive stock which was obtained from the original wild population (English Nature
Where stock are obtained for reintroduction from captive breeding or cultivation there is a
risk that ex situ breeding programs have selected genotypes adapted to artificial conditions.
There is also the danger of accidental selection of poorly adapted genotypes from ex situ
collections (Kay 1993) or donor sites. This is discussed further in section 2.4.3.
Human factors
The possible effects of future changes in land use should also be taken into account when
planning reintroductions because they may influence the chances of successful establishment.
For instance, it is probable that the survival of the white-tailed eagle in Scotland is linked to
traditional land use so that it is possible that their success could be jeopardised by future
increases in afforestation which could lead to reductions in the availability of sheep carrion
(Elliot et al. 1991).
The sympathy of local organisations, landowners and other members of the public may be an
important influence on the success of a reintroduction. This aspect is discussed further for
reintroductions of regionally or locally extirpated species (section 2.5.3).
Environmental/economic impacts - European beaver Castor fiber
Negative impacts
Lever (1994) writes that differences between the UK and continental Europe make the
possible impacts of the reintroduction of the beaver unpredictable and possibly undesirable.
He suggests that beaver dams might cause flooding to farmland, cause direct damage to crops
including conifers, or interfere with generation of hydroelectricity. In a review by Morris
(1986) it is suggested that if beavers were reintroduced to the UK the resulting impacts might
Review of information, policy and legislation on species translocations
be comparable to those seen after the introductions of the alien rodents coypu Myocastor
coypus and musk rat Ondatra zibethica to the UK (see chapter 3).
These potential impacts are discussed in MacDonald et al. (1995) and Halley (1995) in the
context of impacts caused by beaver in continental Europe. Southern Sweden is used as an
example because it is intensively farmed and probably more densely populated than rural
Britain. Few problems are associated with the coexistence of beaver and human populations
in this region. It is shown that direct damage to crops in most areas of Europe is slight,
especially when compared to damage caused by other animals. Unlike coypu, beavers stay
close to the river bank so that damage caused to crops is localised. Damage to conifers is
unlikely as in Europe beavers avoid streams exclusively bordered by conifers. MacDonald et
al. (1995) indicate that there is no evidence for damage to hydroelectric schemes by beavers.
They refer to studies which show that beavers can colonise upstream of hydroelectric
barrages without any negative impact and that beavers are unlikely to survive downstream of
these installations due to the sudden fluctuations in water levels.
When they do occur, effects on dams or direct damage may be remedied quite easily and
cheaply (Halley 1995, MacDonald et al. 1995).
Positive impacts
MacDonald et al. (1995) cite studies from Poland, France and Latvia which indicate the
positive environmental and economic role that beavers can play. By their influence on stream
flow beavers improve water quality, stabilise ground water levels, aid flood control, and aid
water conservation. In several European countries beavers generate income through their
appeal as a tourist attraction and game animals.
2.5.3 Reintroductions of regionally or locally extinct species Background and case studies
The successful re-establishment of regionally or locally extinct species benefits the
conservation of biodiversity by widening the geographic range of the species concerned with
the intention of making its future in the UK and worldwide more secure. Tables 2.3 and 2.4
list a number of species for which reintroductions are being carried out or are planned within
the UK.
Red kite Milvus milvus
The effects of persecution reduced the UK distribution of the red kite to a small area of
central Wales. The justification for attempting to reintroduce this species to the UK is
strengthened by the fact that its global population (reviewed by Evans & Pienkowski 1991) is
rather fragmented and is reduced to a relatively small proportion of its former range (JNCC
1994a). Two attempts to establish red kites of Spanish origin in Wales in the first half of this
century were not successful (I. Evans in prep.). In 1989 the RSPB and the NCC (now the
JNCC) began a programme for reintroduction of this species. This involved the experimental
Review of information, policy and legislation on species translocations
release of red kites from various UK and overseas origins into two sites in the UK. By 1995
nucleus populations had been established of about 30 pairs in southern England and 13 pairs
in northern Scotland which show breeding success comparable with that in the donor sites.
This suggests that the habitat is capable of supporting this species (Evans et al. 1994,
McGrady et al. 1994). These populations are now considered to be self-sustaining and the
programme now hopes to establish more populations which will eventually link with the
populations already established (I. Evans in prep.).
Plymouth pear Pyrus cordata
Contraction in the UK range of many species is the result of changes in land use.
Reintroductions may be considered when sites providing appropriate habitat management are
available. The Plymouth pear Pyrus cordata is considered worthy of conservation as it is one
of Britain's rarest plants and, being a wild relative of the cultivated pear, is a potentially
valuable genetic resource. There have been attempts to re-establish it in several parts of its
historic range in the UK. A recovery plan was initiated in the 1990s and suckers from the
wild trees were used to reintroduce the plant to two sites. Interestingly, although the decision
has been made to reintroduce this plant, there is some doubt regarding its native status in the
UK (Jackson 1995).
Chequered skipper Carterocephalus palaemon
This butterfly is thought to have become extinct in England due to the rapid changes in
woodland management since 1945 which were probably responsible for the decline or loss of
many species of butterfly. Recent changes in habitat management on a number of sites have
lead to the possibility of reintroduction, so a programme has been started which hopes to
establish a network of self-sustaining colonies (Warren 1995, M. Warren pers. comm.). This
involves the introduction of individuals taken from the Ardennes in France. These were used
in preference to Scottish populations of the species (see below).
Table 2.3: Examples of species regionally or locally extinct currently or recently subject to
reintroduction attempts in the UK.
Plymouth pear Pyrus cordata: Royal Botanic Gardens Kew.
Ribbon-leaved water plantain Alisma gramineum: EN-SRP, Institute of Terrestrial
Fen orchid Liparis loeselli: EN-SRP, Sainsbury Orchid Project (Kew), Norfolk
Naturalists Trust.
Starfruit Damasonium alisma: DCMP Plantlife.
Strapwort Corrigiola litoralis: EN-SRP.
Fen ragwort Senecio paludosus: EN-SRP.
Review of information, policy and legislation on species translocations
Wart-biter cricket Decticus verrucivorus: EN-SRP, International Institute of
Entomology, Invertebrate Conservation Centre (London Zoo).
Field cricket Gryllus campestris: EN-SRP, Invertebrate Conservation Unit (London
Reddish buff moth Acosmetia caliginosa: EN-SRP, Federation of Zoological
Gardens of Britain and Ireland.
Chequered skipper Carterocephalus palaemon: Butterfly Conservation.
Natterjack toad Bufo calamita: EN-SRP, CCW, University of Sussex.
Sand Lizard Lacerta agilis: EN-SRP, CCW.
Red kite Milvus milvus: JNCC, RSPB.
Dormouse Muscardinus avellanarius: EN-SRP, Royal Holloway College.
Red squirrel Sciurus vulgaris: SNH, (EN-SRP previous translocations and currently
a feasibility study)
Table 2.4: Examples of species regionally or locally extinct which are currently subject to
feasibility studies or proposals for reintroduction in the UK.
Schleicher's thread-moss Bryum schleicheri: SNH, RBGE
Alpine sow thistle Cicerbita alpina: SNH, Plantlife.
Scottish primrose Primula scotica: SNH, SWT.
Twin flower Linnaea borealis: SNH, Plantlife.
Sticky catchfly Lychnis viscaria: SNH, RBGE, Plantlife.
Shore dock Rumex rupestris: EN-SRP, Institute of Terrestrial Ecology, Plantlife.
Perennial knawel Scleranthus perennis prostratus: EN-SRP, Suffolk Wildlife Trust.
Fen violet, Viola persicifola: EN-SRP, Institute of Terrestrial Ecology.
Oblong woodsia Woodsia ilvensis: SNH.
Medicinal leech Hirudo medicinalis: SNH.
Lagoon sandworm Armandai cirrhosa: EN-SRP, Southampton University, Nature
Conservation Bureau.
Review of information, policy and legislation on species translocations
Northern brown argus Aricia artaxerxes: SNH.
Ladybird spider Eresus cinnaberinus: EN-SRP.
Vendace Coregonus albula: SNH
Giant goby Gobius cobitis: EN, Marine Biological Association, Plymouth.
Smooth snake Coronella austriaca: EN-SRP.
Chough Pyrrhocorax pyrrhocorax: Paradise Park, Cornwall.
Pine marten Matres martes: Peoples Trust for Endangered species, Bristol
University. Impacts of reintroductions of species regionally or locally extinct in the UK and
factors affecting their success
Ecological impacts
The deleterious ecological impacts that may result from translocations in this category are
similar to those described for reintroductions of nationally extinct species.
The possibility of damage to donor sites is mentioned in a discussion of the proposal to
reintroduce the chough Pyrrhocorax pyrrhocorax to Cornwall. It appeared that a magazine
article regarding the Cornish project unintentionally incited illegal collecting of chough (see
NCC 1989).
Translocation of locally or regionally extirpated species to additional sites can be an
important element of the conservation of biodiversity at a regional or national level by
helping to secure the future of the species. For instance, the intention to create networks of
populations or metapopulations of species such as the red kite Milvus milvus and the
chequered skipper Carterocephalus palaemon.
Factors relating to transmission of pathogens and parasites
The possibility of transmission of pathogens is exacerbated when plants are cultivated ex situ
prior to reintroduction because of the risk of exposure to alien pathogens (Kay 1993). For
example, cucumber mosaic virus was found in cultivated specimens of the Chatham Island
Forget-me-not Myosotidium hortensia (Thomson 1981). Reintroductions using infected
individuals may have serious consequences because the pathogens could potentially spread
from recipient sites to established populations elsewhere.
Review of information, policy and legislation on species translocations
Genetic impacts
If reintroduced populations spread into the range of native populations which are genetically
distinct there is a potential for loss of local adaptation. The problems associated with
outbreeding depression are discussed more fully under considerations of population reenforcement in section 2.5.4.
Factors relating to successful reintroductions
Ecological factors
As with the reintroduction of nationally extinct species, release or transplantation of
individuals into appropriate habitat is of paramount importance. For animal species this
includes appropriate food sources for each stage of development, suitable climate throughout
the year, and lack of excessive competition or predation. Proposed reintroductions may not
receive official approval when research indicates that the available habitat is unsuitable. For
instance, this was one factor in the official disapproval of the proposed reintroduction of the
chough Pyrrhocorax pyrrhocorax to Cornwall (NCC 1989).
Habitat requirements have been considered in great detail in the investigation of the
feasibility of reintroducing the chequered skipper Carterocephalus palaemon to eastern
England (Warren 1995). The proposal recommends that continental stock would be more
suitable for reintroduction than stock from Scotland. The argument was based primarily on
ecological factors. The report found that Continental populations used similar habitat (damp
woodland) and food-plants to former populations in eastern England but that Scottish
populations were found on open moorland and used a different food-plant. It was also noted
that the climate in the continental sites resembled the recipient sites more closely than the
Scottish climate. Ecological preference for the continental stock is supported by genetic
considerations (see below) in the report.
Genetic factors
The low productivity of the native Welsh population of red kites Milvus milvus precluded the
use of entirely local individuals for reintroductions. Most of the birds were obtained from
European sources although this could lead to the introduction of individuals poorly adapted to
the local conditions. Evans and Pienkowski (1991) however, state that it is unlikely that there
are any 'relevant genetic differences' between continental and British stock of red kites as
they are similarly sized and look the same and because red kites occurred continuously
throughout Europe until recently. However, this is a tenuous argument without proper genetic
studies. To attempt to reduce the possibility of deleterious genetic mixing the released birds
were obtained from populations as close as possible to the native population (JNCC 1994a).
Conservation of the Plymouth pear Pyrus cordata has encountered several genetic problems.
Hybridisation experiments have shown that viable seed can be readily obtained from crosses
of the Plymouth pear with cultivated varieties of Pyrus communis. The danger of
hybridisation with cultivated pears is a potential threat to both original and reintroduced
populations. Attempts were made to select sites sufficiently distant from cultivated pears to
Review of information, policy and legislation on species translocations
be outside the range of insect pollinators. Unfortunately foraging distances were not available
for bumblebees Bombus spp. which may be a major pollinator. Additional threats of
introduction of non-local genetic races may arise if European individuals of the species are
imported into the UK for planting in gardens.
The example of the Plymouth pear also illustrates some of the genetic problems encountered
when establishing cultivated stocks which are intended to be used for reintroduction. RAPD
DNA analysis showed that there were no major genetic differences within the two remaining
populations of this species in the UK. This accounted for the lack of seed set due to selfincompatibility. Following discovery of some genetic differences between the two UK
populations controlled crosses resulted in seed production (Jackson 1995).
Warren (1995) states that there is no strong genetic argument in favour of using either the
continental or Scottish stock for reintroduction of the chequered skipper Carterocephalus
palaemon to eastern England. Genetic considerations however, support the ecological
arguments in favour of using the continental stock. The report considers that continental stock
are more likely to be able to adapt to the recipient sites in England than Scottish stock for a
number of reasons. Continental populations are likely to have been isolated from the former
English sites for a shorter period than those in Scotland. It is therefore possible that the
former are more genetically similar to the extinct English butterflies. The long isolation of the
butterfly in Scotland and its restricted distribution suggest that it may have become highly
adapted to local habitats. In contrast, the continental populations breed in a variety of
different habitats which are often situated close together. The stock from the continent may
therefore show greater ability to adapt to novel environments than Scottish stock. Early
results of genetic analysis show that the continental stocks have greater heterozygosity.
Human factors
Many predators in the UK have suffered severe range contraction due to persecution.
Changes in attitudes and legislation have made it possible to consider re-establishment of
species once regarded as vermin over their former range. The success of the reintroductions
of the red kite Milvus milvus in the UK show that re-establishments of animals once targeted
by gamekeepers may now be possible. Other species in this category which are now being
considered for reintroduction include the pine marten Martes martes in England (Bright &
Harris 1993).
In species once subject to persecution the possibility that released individuals might be killed
by humans must be considered. In the feasibility study for reintroduction of the pine marten
to the UK Bright & Harris (1993) surveyed attitudes of the British public by questionnaire.
Their results indicated that 58% of organisations which were questioned supported the
proposed reintroduction but most of these attached caveats such as the protection of gamerearing interests. More importantly they indicate that experience from overseas has shown
that the attitudes of a minority opposed to reintroduction can be crucial. This consideration
applies equally to potentially unpopular reintroductions of nationally extinct species in the
UK such as the wolf.
Review of information, policy and legislation on species translocations
2.5.4 Re-enforcement of existing populations Background and case studies
Release of individuals into existing populations may be justified when that population is at
very low numbers and is therefore vulnerable to extinction by chance factors. As in
considerations of reintroductions, it should also be possible to demonstrate that the factors
responsible for the original decline are no longer operating and that harmful impacts will not
ensue. Examples of species subject to current, recent or planned re-enforcement programmes
in the UK are given on Table 2.5 and 2.6.
Lady's Slipper Orchid Cypripedium calceolus
This is an extreme case of a vulnerable small population as only one wild plant remained in
Northern England. The colony size was boosted by the introduction of material from ex situ
propagation (English Nature 1994b).
Eurasian otter Lutra lutra
Otter numbers have declined substantially in the UK, and especially England, since the 1950s
and the surviving populations are thought to be so fragmented that they may not be viable
(Morris 1993). However, otter populations in the English midlands are now recovering and
expanding eastwards (T. Tew et al. unpublished report).
Pesticide pollution is likely to have been the major cause of decline. Exposure to
polychlorinated biphenyls (PCBs) is thought to be important in the decline of otter
populations in most European habitats. At a seminar on the Eurasian otter in 1994 the
importance of ensuring that otters would not be at risk from this contaminant at recipient sites
was stressed (de Jongh 1995). Waterway pollution with pesticides has decreased in the UK
but areas considered to be suitable for recolonisation by otters appear to have remained
vacant (Jefferies et al. 1986). Jessop (1991) and Jefferies et al. (1986) consider that these
areas have remained vacant because the fragmented populations may be too small to allow
dispersal and colonisation.
To try to increase otter numbers, vacant areas in East Anglian rivers were identified using
spraint surveys (Jessop 1991) and groups of captive bred otters referred to as 'breeding units'
were released into those areas (Jefferies et al. 1986). Other releases have taken place since
this time in the eastern areas which have fragmented populations (T. Tew et al. unpublished
The recent UK framework for conservation of the otter produced by the JNCC (T. Tew et al.
unpublished report) states that translocation should be used as a complimentary tool to the
promotion of natural recolonisation and that reintroduction should be used only as a last
resort. There is also a requirement that releases should only occur where natural
recolonisation is unlikely in the foreseeable future. Given the further research and population
Review of information, policy and legislation on species translocations
modelling recommended in the same report, it may in future be possible to be more precise
about what is meant by 'the foreseeable future'. The implications of this are important when
considering the recommendations of the codes of practice for release of otters produced by
the JNCC in response to the framework document (T. Tew et al. unpublished report). This
proposes that otters may be released in the paths of expanding populations to aid
colonisation. Other suggested reasons for release are to boost the size of a small (presumably
non-viable) populations and as a test to discover whether the cause of decline has been
correctly determined and ameliorated. The latter suggestion may be at variance with the
IUCN guidelines adopted by the framework which state that prior to releases, the factors
leading to the extirpation should be removed.
Table 2.5: Examples of species currently or recently subject to re-enforcement in the UK.
Lady's Slipper orchid Cypripedium calceolus: EN-SRP, Sainsbury Orchid Trust
Potentilla rupestris: Humphries Rowell Associates
Strapwort Corrigola litoralis: EN-SRP, Field Studies Council (Slapton Ley).
Sticky catchfly Lychnis viscaria: SNH, RBGE.
Saxifraga caespitosa: EN, CCW, University of Liverpool.
European otter Lutra lutra: The Otter Trust, Norfolk.
Table 2.6: Examples of species currently subject to proposals for re-enforcement in the UK.
Oblong woodsia Woodsia ilvensis: SNH. Impacts of population re-enforcements and factors affecting their success
Genetic impacts
Population re-enforcements can lead to genetic mixing between genetically differentiated
populations. This has the potential to introduce traits which are locally maladaptive.
When populations are as vulnerable as the lady's slipper orchid Cypripedium calceolus
threats of outbreeding depression are minor in comparison to the threat of imminent
extinction. Nevertheless, in this case the transplanted material was derived from the original
stock so outbreeding depression is unlikely unless the stock has been exposed to
inappropriate selective pressures during propagation.
Poorly adapted traits in plants may be introduced during re-enforcement by chance or
unconscious selection of individuals for transplantation. Kay (1993) lists types of plant
genotypes which may be favoured in unconscious selection. These include plants producing
small but numerous seed, reduced seed dormancy, low allocation to defence, precocious
growth and flowering and increased self-compatibility. Although these plants might be
Review of information, policy and legislation on species translocations
unable to establish, the spread of their deleterious genes into a vulnerable population could
have serious consequences. The spread of maladaptive genes appears counter-intuitive but is
possible when a native population is swamped by released individuals during a population reenforcement.
A frequently cited example from overseas of the problems associated with outbreeding
depression resulting from genetic mixing is the re-enforcement of a herd of reintroduced ibex
in Czechoslovakia. Bezoars Capra ibex aegagrus and Nubian ibex C. ibex nubina were added
to the reintroduced herd of Tatra mountain ibex C. ibex ibex. The offspring rutted earlier in
the year than the Tatra mountain ibex and this resulted in production of kids at the coldest
time of year and the subsequent extinction of the whole population (Templeton 1986).
If very low genetic variation can be demonstrated between populations it may be appropriate
to recommend that there should be minimal constraints on moving animals around the UK for
restocking. For example research has shown that British natterjack toads Bufo calamita are
genetically depauperate in comparison with conspecifics in Europe. Empirical evidence
supports the suggestion that British populations can be transferred freely between sites. For
instance, toads from a tidal site showed no more tolerance of salinity than those from an
inland heath (Denton et al. 1995).
In the case of re-enforcements of otter Lutra lutra, populations in the UK genetic aspects
have been considered. Locally derived stock might be preferred to maintain genetic integrity
of otter populations (T. Tew et al. unpublished report). Jessop (1991) on the other hand
suggests that released and native otter populations in East Anglia might benefit from an
'interflow of genes'. This makes the point that outbreeding may be beneficial and the
introduction of a new genetic stock may counteract such problems as inbreeding depression
and genetic drift towards maladaptive traits.
In the recent report on a framework for conservation of the otter in the UK (T. Tew et al.
unpublished report), the need for research on the genetic diversity of wild otter populations is
indicated. Its results may clarify these issues.
Factors related to successful restocking - Eurasian otter Lutra lutra
Ecological factors
The composition of the release groups of otters were based on field observations of otters in
Britain. It was hoped that releasing animals of 18 months old would eliminate high mortality
experienced by young otters. Groups of two females and one male were reared and released
together. Groups of this type are observed in the wild and it was hoped that they would
provide a focus for population establishment (Jefferies et al. 1986). The possible effects of
captive breeding on behaviour are not discussed except that the contact with humans was
Sites chosen for release of otters in East Anglia were considered to be free of otters but
adjacent to occupied sites (Jefferies et al. 1986) on the basis of spraint surveys (Jessop 1991).
This should prevent competition for resources with native conspecifics. Kruuk (1989)
however, considers that the captive bred otters may well have been released into an existing
Review of information, policy and legislation on species translocations
but declining population. This is because Kruuk & Conroy (1987) showed the presence or
absence of faeces is an unreliable method for estimates of population parameters and that
otters may range over stretches of river in excess of 70 km whereas the vacant sites were
identified on the assumption that otters would use less than 30 km.
Released otters could also experience competition from the non-native mink Mustela vison. A
recipient site was therefore chosen which was free from this species (Jefferies et al. 1986).
One of the main criteria considered in re-enforcements and reintroductions is that the original
cause of the decline is known to no longer be operating. In the case of the otter the pesticide
contamination held to be responsible for the population decline was thought to be adequately
reduced (Jefferies et al. 1986). The water was apparently unpolluted at the recipient site and
coarse fish were abundant. Nevertheless, when a sample of chub Leuciscus cephalus from the
receptor site were analysed 'low concentrations' of pesticides and mercury were found. This
was thought to 'not outweigh the advantages of the site' (Jefferies et al. 1986). Otters in north
Norfolk have however shown high PCB contamination and the otter release in the area
appeared to fail rapidly. Fish from the site in north Norfolk had PCB concentrations similar to
those shown by Dutch studies to cause rapid reproductive failure in mink (Mason 1991). The
pesticides in East Anglia are therefore a serious cause for concern and it seems likely that the
primary cause of population decline is still operating in this area. Mason (1991) concludes
that otters will be unable to colonise former sites in the lowlands in the UK due to pesticide
contamination and that apparent range expansion into lowland areas (e.g. the West Midlands)
is only sustained by continued emigration from adjacent upland areas. Currently, PCB's and
heavy metals are still considered to pose a threat to the otter and monitoring and research into
causal links between pollutants and population decline are recommended (T. Tew et al.
unpublished report).
Genetic factors
The possibility of inbreeding in the captive bred otters used in the East Anglian release
project appeared to be addressed as the 'breeding units' which were released consisted of
unrelated individuals (Jefferies et al. 1986).
Human factors
Accidental deaths on the roads and in traps set for other animals are a common cause of otter
mortality and could threaten the success of a release programme (T. Tew et al. unpublished
report). In East Anglia the cooperation of riparian landowners was considered prior to the
release of otters (Jeffries 1986).
2.5.5 Creation of new populations to conserve vulnerable wild populations
This includes sites which may have been previously occupied by the species, and sites which
have no records of the species but appear to be suitable receptor sites. The same possible
impacts apply as for reintroductions.
Review of information, policy and legislation on species translocations
Maunder (1992) stresses the prudency of establishing additional populations of species which
are reduced to one population and hence vulnerable to extinction. If a species is extremely
rare and appears to be impossible to conserve ex situ the establishment of new populations
may be the only method of conserving them (McMahan 1994). Plugs of three rare Breckland
lichens, Buellia astrella, Fulgensia fulgens and Squamaria lentigera were transplanted
because populations at the original sites were declining probably due to pollution and
changes in land use (English Nature 1994b).
New populations may also be created to deliberately divert attention from vulnerable wild
populations. For instance, orchids which are at risk due to public pressure could be protected
in the same way by creating 'orchid garden' nature reserves (Farrell & Fitzgerald 1989).
The use of new habitats such as motorway verges for creation of new populations of
threatened plants is recommended by Green (1981).
2.5.6 Relocations to rescue individuals or small populations Background and case studies
This type of translocation is often necessitated by destruction of habitat by development. In
the 1950s one of the rarest plants in the UK, the brown bog rush Schoenus ferrugineus, was
transplanted due to imminent flooding of its site by hydro-electric development. The plant
eventually disappeared from this site and was considered extinct but has subsequently been
located elsewhere in the UK (Brookes 1981). Relocations may also involve 'habitat
translocation' and this is discussed in Chapter 5.
As the crested newt Triturus cristatus is classed as widespread but uncommon, there are
frequent conflicts between conservationists and land developers when sites are threatened.
Translocations of this species are reviewed by Oldham et al. (1991). Using the information
from the NCC licensing section, the National Amphibian Survey Database and personal
contact, they identified eighty-six translocations between 1970 and 1990. Due to the lack of
adequate monitoring and availability of reliable information the review was unable to
produce conclusive evidence for the success or failure of translocations of crested newts.
In the recent past destruction of sites has lead to numerous rescue translocations of
populations of sand lizards Lacerta agilis. The Recovery Plan produced for this species by
English Nature in 1994 hopes to rationalise translocations of the sand lizard so that the work
can contribute to the restoration of a representative range for this species in the UK (Corbett
Relocations to rescue badgers are considered in section 2.6.6 'Releases for animal welfare'. Impacts of rescue translocations and factors affecting their success
Review of information, policy and legislation on species translocations
Ecological impacts
Relocated individuals may compete with resident conspecifics or interact with other species
of conservation value.
Environmental impacts
The perception that relocation techniques are equivalent to conservation will make it harder
to argue for development planning which avoids conservation areas (Allen 1994, Maunder
1992). Public perception is illustrated by an article in the Guardian ('Protestors dig in as
butterfly starts High Court Flap' 3r June 1995) which states that 'planning permission was
granted on condition that the firm moved an SSSI to a nature reserve about a mile away' (see
also Chapter 5).
Factors related to success of the relocation
Ecological factors
Rescue translocations often transfer organisms to carefully managed habitat. Success will
depend on the ability to maintain appropriate management and to ensure that the animals
remain at the site. The tendency for crested newts to escape from their enclosed 'conservation
area' was identified as the main problem in a pilot project designed to test proposals to
translocate populations of this species to protect them from mining operations (Horton 1994).
The failure of the transplantation of the brown bog rush to nearby sites which appeared to be
similar has been cited to illustrate how a lack of ecological understanding can render
transplantation unreliable (Morton 1982). Some aspect of the habitat at the receptor site may
have been inappropriate or the founder population may have been too small.
Oldham et al. (1991) note that probable reasons for failures of crested newt translocations
are: a founding population that was too small, unsuitable habitat, and presence of predators.
Human factors
During land development there is an increasing trend to expect that populations and even
habitats can be moved (Cleave 1995). Ensuing rescue projects tend to be of a reactive nature
precluding adequate planning and hence reducing the chances of success. Rescue
translocations have a greater chance of success if ecological knowledge regarding
translocations is efficiently disseminated. According to Oldham et al. (1991), for at least 33%
of known failures the result could have been predicted from current knowledge of crested
newt ecology. This emphasises the need for better communication and dissemination of
Review of information, policy and legislation on species translocations
2.5.7 Conservation seed mixes and plantings Background and case studies
The other sections in this review on translocations of native species consider programmes
concerned with the conservation of single species. However, a major form of translocation in
the UK involves the sowing or planting of many plant species at one site in projects aimed at
the conservation of plant communities. Such translocations might be described by the general
term 'habitat restoration' as they have the general aim of creating or supplementing fully
functioning communities of plants, animals, microorganisms, etc. Although this may be the
(sometimes unstated) ultimate aim, there are different types of project, with different
rationales, which use conservation seed mixes and/or plantings. Some of these reasons
involve amenity purposes, but, because all show some commonalty of approach, all are
discussed here.
Seed mixes for habitat creation - conservation reasons
These are used for two reasons: in projects to mitigate for the loss of communities through
changes in land use, e.g. road building, pipeline laying, quarrying, etc.; or in projects to
restore rarer community types, usually on land previously used for agriculture or forestry.
Thus, such projects are carried out on unvegetated areas, or with the aim of replacing the
existing vegetation (e.g. of arable weeds and volunteer crops). A large number of projects are
being carried out under a number of schemes, such as the MAFF Environmentally Sensitive
Areas and Farm Woodland Scheme, and the Countryside Commission Countryside
Stewardship Scheme. Such restoration can provide new habitat for species in recovery
programmes, and is being used in some of EN's Species Recovery projects. Some of the more
advanced work in this field is considering how one can restore a whole community of plants,
animals and soil flora and fauna (e.g. Davis 1989, Morris et al. 1994).
Plantings for habitat creation - conservation reasons
Planting out of 'pot-grown' plants is often used to supplement the use of seed mixes in habitat
creation projects.
Seed mixes and plantings for habitat creation - amenity reasons
Habitat creation for amenity purposes may have a number of motivations, although nature
conservation is usually a consideration. These include: creating pleasant surroundings in an
urban environment, sometimes for education; stabilising and/or landscaping of derelict land,
e.g. waste tips, quarry and mine workings etc.; and vegetating roadside verges. Seed mixes
and plantings were probably first used for restoration of derelict land (e.g. Bradshaw &
Chadwick 1980, Bradshaw 1983, EAU 1988, Wali 1992). Education and creation of pleasant
surroundings in urban areas is a very popular reason for restoration (EN 1992). Roadside
verges are commonly sown or planted with a variety of species, and, although the main
reason is to create an attractive landscape, conservation is also considered. For instance, the
Review of information, policy and legislation on species translocations
'Wildflower Handbook' issued by Department of Transport, Scottish Office Industry
Department, The Welsh Office, Department of the Environment Northern Ireland (1993)
promotes this technique because of the 'attractive natural appearance', the interest to road
users, landscaping of the road, and provision of conservation islands or corridors.
Seed mixes and plantings for enhancement/diversification
These are used in existing communities in an attempt to increase the plant species richness
and, usually, with the aim of directing vegetation development towards a more valued
community type. Species are sown into the vegetation (e.g. using slot-seeding) or planted.
As described by Parker (1995) and the Wildflower Handbook (DTp et al. 1993), sowing and
planting for habitat creation and enhancement has been, and is being, carried out for a variety
of vegetation types: mesotrophic, acid and calcareous grasslands, saltmarshes, woodlands,
hedgerows, heathlands, moorlands and peatlands.
Procedures for successful habitat creation and enhancement are extremely well researched
and are the subject of a recent comprehensive review for EN by Parker (1995). T. Wells of
ITE has carried out extensive research into the techniques and ecology of creation and
enhancement of species-rich grasslands (Wells et al. 1981, 1986, 1989, Wells 1989, 1990),
both for conservation and amenity aims. Other reviews or descriptions of techniques in the
UK include: Buckley (1989), for a variety of vegetation types; EAU (1988) and Pywell et al.
(1994), for heathlands; and the Wildflower Handbook (DTp et al. 1993), for roadside verges.
The situation abroad is much the same; there are many reviews and publications on
restoration techniques on a variety of sites (Wilson 1986, Jordan et al. 1987, Berger 1990,
Wali 1992). We do not aim to reiterate these reviews, but instead will consider the wider
impact of such translocations. Unfortunately, these have received little practical attention.
Although some concerns, such as use of local races, have been discussed extensively, the
implications have not been studied in the field and are not represented by any case studies.
We will aim to highlight some of these gaps in our knowledge. Impacts of the use of conservation seed mixes and plantings and factors affecting its
Ecological impacts
Deleterious - Damage to the donor site
Collecting seed or plants may damage plant populations at the donor site. SSSIs and nature
reserves are popular sources of seed and repeated collections may affect the plant
populations. Stevenson et al. (1994) were careful to visit a number of sites and to harvest
each site only lightly when suction harvesting seed for a chalk downland restoration project.
Drake (1994) reports a study of suction harvesting of seed where many insects were injured
or killed by the harvester. As well as avoiding sites with strong invertebrate interest, it is
suggested that techniques are used to avoid injuring the invertebrates. These are: use of brush
harvesters in preference to suction harvesters, harvesting at low intensity, not harvesting
repeatedly in the same places, and allowing any intact insects to escape from the harvest.
Review of information, policy and legislation on species translocations
Deleterious - Use of non-native species
It is not clear how common this is in the UK (Parker 1995), although it does occur in some
tree-planting schemes (Soutar & Peterken 1988) and some grassland projects (Akeroyd
1994). Non-native species are used more commonly in other countries (Berger 1993). As
discussed in Chapter 3, there may be problems with invasion of other communities (see also
Berger 1993).
Deleterious - Use of non-native sources of native species
The use of UK native species which are not native to a region could cause similar problems
to the use of non-native species (Soutar & Peterken 1988). Akeroyd (1994) presented
evidence that some species used in grassland restoration projects are being planted outside
their native
ranges and suggested that these might out-compete native races (see below). EN (1993)
suggested that specialist herbivores may perform worse on such non-native types, but there is
no evidence to support this.
The beneficial aspects of restoration are well rehearsed (Bradshaw 1983, Gilpin et 1987,
Lamb 1993, Berger 1990, Parker 1995). As well as creating new populations and new habitat,
and having landscaping and amenity value, restored semi-natural communities may affect the
landscape ecology of an area. The dynamics and survival of metapopulations may be
enhanced and the use of local native species may improve continuity of the local landscape
and can therefore reduce fragmentation.
Transmission of pathogens and herbivores
As with any translocation of native species, pathogens and herbivores may be translocated as
well. These may form important parts of a functioning community and so may be beneficial
rather than deleterious.
Genetic impacts
Seed may be obtained from hay cuts, hand gathering or mechanical harvesting from
appropriate donor sites, topsoil, litter or seed merchants (Gillham & Putwain 1977, Wells et
al. 1981, 1986, 1989, EAU 1988, Parker 1995), and plants can be grown on from this seed or
transplanted as turves (Rawes & Welch 1972, Wathern & Gilbert 1978, Pywell et al. 1994;
these are used as inocula for species rather than as a method of community translocation).
The source of the material used in a restoration project will affect the genetic constitution and
structure of the resulting populations.
Deleterious - Use of non-native races or agricultural varieties
The guidelines reviewed in section 2.3 all suggest that seed or plants of UK or British origin
be used. Akeroyd (1994) provides the most extensive discussion of this issue and suggests
Review of information, policy and legislation on species translocations
that use of seed that is not native in origin (i.e. alien races of native species or agricultural
varieties) could lead to (in his words):
confusion as to the natural distribution of the plants in Britain;
confusion of complex and ancient patterns in the landscape and creation of a facsimile
of countryside;
competition between native and perhaps more vigorous introduced plants of the same
crossing between native and introduced plants, leading to erosion of native genetic
While the first two points are management problems (and are discussed below), the latter two
points concern disruption of genetic structure of local populations. Akeroyd (1994) states that
many suppliers of seed use non-native stock. D. McIntyre (pers. comm.) of Emorsgate Seeds
points out that many seed merchants provide 'British produced seed' which is often an
agricultural variety, of perhaps foreign origin, which is planted as a seed crop in Britain; the
seeds produced are labelled 'of British origin'. Akeroyd (1994) names a number of forb
species commonly planted in restoration projects, but which are often of foreign origin or are
agricultural varieties: Trifolium pratense, T. repens, Lotus corniculatus, Medicago lupulina,
Anthyllis vulneraria, Onobrychis viciifolia, Sanguisorba minor, Leucanthemum vulgare,
Achillea millefolium and Chrysanthemum segetum. He also reports two high-profile
restoration projects where non-native seed of native species was used. If the stock is an
agricultural variety it can be very different in growth form and ecology from the native races.
Therefore, as for crop varieties (see Chapter 3) or GMOs (see Chapter 4), the non-native
races could spread to other communities and replace local races, or hybridise with local races.
This will change the genetic structure of local populations. Despite these concerns, there is
little actual research into the consequences of using non-native races and there is no
information on
whether there is a serious conservation problem. It is highly probable that non-native races or
agricultural varieties will not establish well in semi-natural conditions because they are not
locally adapted. In particular, fodder varieties are developed as crops rather than to survive in
perennial communities. Thus, non-native races are more likely to die out or undergo severe
selection pressures, perhaps involving spread of local genes into the introduced population. In
fact, because they are locally adapted, ecological or genetic spread of local races into the
restoration site is more likely than such spread of the introduced race out of the site.
Therefore, the only likely problem with the use of non-native races or agricultural varieties is
that the restoration could be unsuccessful (see below). However, it does seem wasteful and
unnecessary to use such seed or plant sources in restoration projects.
It is important to note that many of the agricultural varieties are widely sown in agricultural
systems in the UK. Therefore, semi-natural communities are probably exposed to gene-flow
from these varieties even if they are not used in conservation seed mixes.
One reason such stock is used is that native seed is expensive (R. Snazell pers. comm.). D.
McIntyre (pers. comm.) suggests two factors which could bring down the cost: use of lower
seeding rates (e.g. Stevenson et al. 1995), and a greater demand for such seed. D. McIntyre
Review of information, policy and legislation on species translocations
(pers. comm.) suggests that greater demand could be brought about by tighter specifications
for ESAs and other schemes and would like to see a future where seed for use in an ESA is
actually produced within the ESA from local stock.
Another reason may be that the Fodder Plant Seeds Regulations 1993 (see section 2.4.7)
actually make it illegal to market seed of listed species which is wild-collected or produced in
conditions not meeting the requirements of the Regulations. Those species listed in Schedule
1 that might be used in restoration projects are Agrostis canina, A. stolonifera, A. capillaris,
Festuca ovina, F. rubra, F. arundinacea, F. pratensis, Poa nemoralis, P. pratensis, P.
trivialis, Arrhenatherum elatius, Dactylis glomerata, Lolium perenne, Phleum pratense,
Lotus corniculatus, Medicago lupulina, Onobrychis viciifolia, Trifolium pratense, T. repens,
and Vicia sativa. Some of the forbs are species mentioned by Akeroyd (1994) as being
commonly sown using non-native seed.
MAFF is aware of this conflict between the legislation and conservation interests and
therefore applies the Regulations 'as lightly as possible' with regard to conservation seed
mixtures where it is satisfied that there is no risk of the seed being used in agriculture or
horticulture (D. Hallam pers. comm.). The UK is also lobbying for an amendment to the EC
Directive to relax marketing restrictions on seed mixtures for restoration.
Deleterious - Seeds of non-local provenance
The guidelines also suggest that seed or plants of local provenance be used. Although 'local'
is rarely defined, this stipulation is more strict than the use of native UK races, but similar
reasons are given: success of the restoration (see below) and maintenance of genetic
biodiversity. Parker (1995) suggests that this will maintain the local genetic integrity of each
species. The Wildflower Handbook (DTp 1993) suggests that if the restoration is near areas
of high conservation value, local provenance should be used to 'avoid problems of unsuitable
gene types spreading in'. Handel et al. (1994) suggest that the genetic diversity of a species
should be understood before it is used in restoration programmes, and suggest that addition of
new plants could cause an 'invasion of new genotypes that could swamp the local gene pool'.
Millar & Libby (1994) also express concern over 'genetic contamination' from the use of nonlocal stock. Guinon & Allen (1990), in a coastal strand restoration in the USA, collected seed
from within a radius of a few km from site - to maintain the genetic integrity of region's flora.
Stevenson et al. (1994), to maintain the local genetic integrity (Exton et al. 1991) of the
populations, collected seed using a suction harvester and hand collection from a variety of
chalk grassland sites within several kilometres of the restoration site. However, there are, to
our knowledge, no studies of the consequence of using plants or seed of local vs non-local
These arguments seem sensible and, in the absence of any data are an application of the
precautionary principle. However, the wish to preserve genetic integrity is a vague one. There
is great variation in the genetic structure of many plant populations and the pattern of
differences among populations (Gray 1995). Large genetic changes can occur between
adjacent populations (e.g. Billington et al. 1988) and therefore, locality is no guarantee of
similarity of genetic constitution. As with non-native races, any major differences between
the introduced and local populations are likely to represent better adaptedness of the local
Review of information, policy and legislation on species translocations
populations and result in changes to the introduced population. Even if the introduced genes
spread into the local populations, the implications are unclear.
Allowing natural colonisation may seem an attractive proposition to ensure local provenance,
but this will only produce the desired community if there is an appropriate seedbank or there
are nearby sources of an appropriate seed rain. For example, Pywell et al. (1994) found good
regeneration of heathland from the seedbank of an agricultural grassland and Gibson et al.
(1987) found rapid development of chalk grassland on an ex-arable site because there were
ancient chalk grasslands nearby. However, studies on other grasslands (Bullock et al. 1994)
or ex-arable sites (R. Pywell pers. comm., Stevenson et al. 1995) have shown poor
colonisation of desirable species despite the proximity of seed sources. Bradshaw (1983) also
points out that the process of natural colonisation will be very slow. As found by Wells et al.
(1976), some species may take many decades to colonise, even if they occur locally.
There are no studies of possible beneficial effects on the genetics of local species but
enhancement of metapopulation dynamics and the introduction of new genes into the gene
pool of a region may act to counter possible inbreeding depression and genetic bottlenecks in
existing populations.
Factors relating to successful habitat creation or enhancement
The procedures necessary for satisfactory establishment of a diverse plant community or
enhancement of an existing plant community are discussed extensively in Wells et al. (1981,
1983, 1986, 1989), Wells (1989), the 'Wildflower Handbook' (DTp 1993), EAU (1988) and
in a recent EN review (Parker 1995). These have different foci - e.g. the EN review considers
restoration of semi-natural communities, whereas the Wildflower Handbook is much more
concerned with creation of an attractive sward (although emphasis is placed on creating local
community types) - but provide very similar advice. We will not repeat these discussions, but
the points to consider are:
clear planning and setting of objectives;
choice of a site and environment appropriate to the required community type - using
detailed surveys and consideration of previous land use;
preparation of the site - e.g. topsoil stripping to reduce fertility;
choice of appropriate species and vegetation type for the site and its geographical
location - to avoid unsuccessful restoration, or creation of a geographically,
ecologically or visually inappropriate community;
use of a range of species that sufficiently represents the vegetation type;
choice of appropriate source of seed and technique for gathering seed - use of UK
races and plants of local provenance;
treatment of harvested seeds - cleaning, storage, after-ripening, etc.;
testing of seed viability;
choice of species to plant - those with low germinability or ability to establish;
identification of species which may colonise naturally;
techniques of seedbed preparation and seeding or planting;
appropriate design (species ratio) of the seed mix;
Review of information, policy and legislation on species translocations
use of nurse crops;
appropriate sowing and planting dates;
techniques for enhancing germination and/or establishment;
methods of attracting colonisation and use of the site by animals, including
appropriate initial and long-term aftercare and management;
monitoring procedures.
Most of this is ecological common sense. The only points we wish to discuss further are the
use of UK races and the use of plants of local provenance. As we discussed above, some
projects use native species but from a non-native source. Another possibility is that
agricultural varieties of native lineage may be used. Because these may be poorly adapted to
the environment of the restored community (e.g. fodder Lotus corniculatus used in a chalk
grassland) this may lead to failure of the species to establish.
Similar arguments are used to promote the use of seeds or plants of local provenance. Most
authors (DTp et al. 1993, Parker 1995) suggest that these are more likely to thrive than plants
brought in from another area. Handel et al. 1994 suggest that the genetic diversity of a
species should be understood before it is used in restoration programmes and suggest that
inattention to natural genetic variation may lead to use of inappropriate genotypes unable to
tolerate local environments. Millar & Libby (1994) report the case of forest planting in the
USA using local species but of non-local stock, and they showed that these trees showed
higher mortality and poorer growth than local trees.
Factors relating to management and planning
The use of habitat restoration as a form of mitigation for the loss of communities to changes
in land use - such as development - raises the concern that it will be used in the place of in
situ conservation and will be viewed by developers as adequate compensation. This will form
an indirect impact on existing communities. It is commonly stated by conservationists that
restoration can never amount to full mitigation or replacement of lost communities (e.g.
Hopkins 1988, Parker 1995). We discuss this issue in detail in Chapter 4. It is sufficient to
make the extra point here that creation of a new community with seeds and plantings will be
even less effective than turf or soil transfer at replicating the lost community, because of the
unpredictability of outcome and because it is impossible to re-create the complete species
complement (including invertebrates and microorganisms) and the genetic structure of the
threatened community.
Regionally or locally non-native species
Planting of UK natives outside of their native ranges occurs (Akeroyd 1994) and this may
cause confusion as to the distribution of the species, its status in the wild and what
conservation management of the species is needed.
Review of information, policy and legislation on species translocations
2.6 Translocations of native species for purposes other than
2.6.1 Fisheries and angling Background and case studies
Commercial fisheries and angling interests are having an increasing impact on freshwater and
marine ecology in the UK and overseas. Fish faunas in the UK are changing from natural and
stable stocks of many species to unstable (i.e. a population not at dynamic equilibrium)
artificially-maintained populations with just a few commercially valuable species (Maitland
1987). Fish may be moved for stock enhancement, to provide specimen fish or bait fish for
anglers and as accidental releases from farms. In the 1980s an accidental translocation
resulted in the establishment, and subsequent rapid growth in numbers of ruffe
Gymnocephalus cernua in Loch Lomond, which probably escaped while being used as live
bait by anglers (Maitland 1987, 1995). Other translocations carried out for angling interests
include Barbel Barbus barbus (Maitland 1987), pike Esox lucius and grayling Thymallus
thymallus. The latter is removed annually from miles of chalk streams where it competes with
the more valuable trout stocks (A. Ibbotson pers. comm.).
For at least 100 years Atlantic salmon Salmo salar stocks have been enhanced by rearing in
hatcheries and release of smolt. Where young salmon were raised from eggs taken from the
wild and usually released into their parental stream the genetics of salmon stocks were
probably little affected by supplementation (Elliot & Mills 1989). Today natural recruitment
is augmented by young or eggs from hatcheries that may not use local stock. The procedures
are generally carried out through the District Salmon Fisheries Boards in Scotland. Despite
the wide practice of, and heavy commitment to, enhancement the benefits are often unclear
and there is no apparent scientific basis for the work (Harris 1978, Maitland 1985). Land
locked salmon Salmo salar sebago is a non-native race which is released to increase the
appeal of stillwater recreational fisheries.
Recently salmon farming in the west of Scotland has become extremely successful and
widespread. Most suitable sites were already utilised by the late 1980s, and according to
Maitland (1989), the biomass of farmed salmon in Scotland was 14 times that of native
salmon. Inevitably fish from these establishments find their way into local populations either
through accidental escapes or deliberate release of excess broodstock for enhancement. Large
numbers of fish are known to escape from salmon farms. Ninety thousand escaped after a
ship collided with cages, 185,000 escaped from one farm after a storm and 1.6 million fry
escaped from a hatchery in Scotland (Maitland 1989). High proportions of captive reared
salmon have been reported among spawning populations in Norway and Iceland and salmon
of captive origin have also been reported in commercial catches in the UK and throughout the
north east Atlantic and there is evidence that they breed successfully (Crozier 1993).
There is no large scale fish ranching in the UK although it practised elsewhere in Europe, e.g.
Iceland (Elliot & Mills 1989), and in the USA (Hindar et al. 1991). A small pilot ranch has
Review of information, policy and legislation on species translocations
opened at the DAFS Freshwater Fisheries laboratory, Pitlochry to investigate the biological
and commercial viability of salmon ranching in the UK (Maitland 1985).
2.6.I.2 Impacts of translocations and factors relating to their success
Factors relating to the spread of pathogens and parasites
An introduced pathogen may either alter selection pressures by favouring resistant
individuals, or it may threaten or even wipe out local populations. The impacts of pathogens
are particularly well documented for salmonids. In a report by the North East Atlantic
Commission working group on introductions and transfers (NASCO 1995) considerable
concern was expressed about the possible adverse effects of salmonid translocations on wild
stocks (this included genetic and ecological effects as well as pathogens- see below). The
working group demonstrated serious currents threats using case studies on a skin parasite
Gyrodactylis salaris and a bacterium Aeromonas salmonicida which causes furunculosis.
G. salaris is thought to have been introduced to Norway by stocking with resistant Swedish
stock in the mid-1970s. The parasite can cause high mortality in salmonids and the only
means of eliminating it is to poison the whole river system and restock. By 1984 the
Norwegian salmon fisheries had sustained losses of between 250-500 tonnes. More recently
the parasite has been detected in rivers in Finland, Russia and West coast of Sweden. It has
been reported on hatchery fish in Spain, Germany and Denmark. There are no records in the
UK but experimental work has shown that two Scottish salmon stocks are susceptible to the
Aeromonas salmonicida was probably introduced to Britain in the 1920s in imported trout.
Subsequent outbreaks of furunculosis in wild salmon populations led to legislation against
introduction of diseases in fish (Diseases of Fish Act 1937 updated in the Fish Health
Regulation 1992). Recent outbreaks of furunculosis in 74 Norwegian rivers are considered to
have resulted from the escape of farmed salmon.
Genetic impacts
Many species of fish exist as genetically isolated populations both within and between
catchments. These may show considerable local adaptation. This has been particularly well
studied in salmonids and Crozier (1993), Hindar et al. (1991) and Elliot & Mills (1989) cite
several studies which show that local populations of salmonids are typically genetically
distinct and that the divergence among populations represents local adaptation. Consequently,
translocations between catchments may result in loss of adaptation (Maitland 1985,
Wellcome 1988) as a result of outbreeding depression (Templeton 1986). Serious adverse
effects of indiscriminate stocking which can be related to genetic changes have been
illustrated by Altukov (1981). Over a two year period more than 350 million fertilised eggs of
chum salmon Onchorhynchus keta were transferred between two rivers in the USSR. In the
fourth year the genetic characteristics of the stock at the recipient site had shifted towards the
donor stock and subsequently the population size declined dramatically.
Review of information, policy and legislation on species translocations
In Crozier (1993) reference is made to three recent studies in Scotland and Ireland which
indicate that salmon culture can cause genetic changes relative to the wild population.
Hutchings (1991) lists ten studies providing experimental evidence that there are heritable
differences in behavioural traits between wild and cultured salmonids and between wild
populations. Concern about the recent increase in fish culture in temperate waters and its
potential genetic effects on wild populations therefore appears to be well founded. Elliot &
Mills (1989) cite examples of hybridisation between Atlantic salmon and trout Salmo trutta
after release of captive bred fish in Newfoundland and Spain and also clear evidence of
introgression between native and non-native strains of trout in Ireland. Although some
aquaculturists and authors suggest that translocations might enhance natural salmonid
populations by increasing genetic variation (Kapuscinski & Cannan 1984), Hindar et al.
(1991) showed in their review that where genetic effects of releases of cultured fish on
salmonids have been documented, they always appear to be negative in comparison with the
unaffected populations. The North East Atlantic Commission Working Group on
Introductions and Transfers conclude that the evidence for genetic risks to salmonids through
translocations is less certain than that for pathogens, but the potential risk is serious (NASCO
Ecological impacts
Threats to native stocks from farmed salmon might include reduction of the productivity of
wild stocks by ecological interference, for instance, if they spawn later they could destroy the
nests of wild-type females (Webb et al. 1991). There is however, evidence that released fish,
which have been reared in captivity or taken from another river, have reduced fitness and
survival rates. Fleming et al. (1995) showed that cultured salmonids have lower reproductive
ability than their wild counterparts and cites five studies supporting this view. In Elliot &
Mills (1989) three studies are listed which indicate that released salmonids have relatively
low survival rates. Wild stocks may also be affected by the attraction of predators to culture
facilities. Hindar et al. (1991) mention a ranch near Oregon where wild stocks of salmon
appeared to be depleted for this reason.
Other possible impacts include changes to invertebrate, amphibian and plant populations,
especially if fish are introduced to previously fishless waters. Discussion and solutions
Some conservation problems are similar for translocations of non-native and native fish.
Section 3.5 should be consulted for further discussion of these problems.
The potential damage to wild fish through translocations by commercial and angling interests
are well illustrated by the literature on salmonids. The problems and possible solutions are
specific to fisheries and angling so they are discussed here. Threats include the introduction
of pathogens, ecological interference and adverse genetic changes. Guidelines and measures
to protect wild salmonid stocks have been suggested in several reviews including Hindar et
al. (1991), Elliot & Mills (1989) and Maitland (1985). The most recent proposals have been
produced by the North East Atlantic Commission Working Group on Introductions and
Transfers (NASCO 1995) and are intended to be complementary to recommendations
adopted by the North Atlantic Commission in 1994 (CNL(94)53). All documents agree that
Review of information, policy and legislation on species translocations
more control of translocations is necessary to protect wild salmonid stocks, the need to keep
farmed and wild stocks separate was considered particularly important. Alarm caused
recently by the impact of the parasite Gyrodactylus salaris has raised the profile of the
problems associated with translocations but awareness of genetic risks is also high and many
proposals are intended to protect wild stocks from both pathogenic and genetic risks.
Proposals include the maintenance of pristine stocks. The NASCO (1995) proposal to classify
Atlantic salmon rivers is intended to provide a basis for management measures appropriate
for each class of river. The report of the working group states that this is intended principally
to limit genetic impact on wild stocks. This measure could be used to offer protection to
stocks of 'high status'. Protection of 'pure native stock from genetic mixing' is also advocated
by the National Rivers Authority (NRA 1993), this will presumably be carried out by
enforcing Section 30 of the Salmon and Freshwater Fisheries Act 1975. The results of the
recent review 'The Identification of cost-effective stocking strategies for migratory salmonids'
(NRA R&D notes, not yet released) may be relevant to developing policy for salmonid
Bearing in mind the frequent lack of demonstrable need for supplementation of salmon
fisheries, Elliot and Mills (1989) suggest that a requirement to show the necessity for stock
enhancement might be imposed. Hindar et al. (1991) propose that genetic differences
between captive and wild stocks should be kept to a minimum. This might be achieved
through a requirement to obtain broodstock locally and to keep it in captivity for no more
than one generation.
NASCO (1995) propose to ban certain types of translocations which pose a threat to the
genetic integrity of local stocks. These include salmon originating from North America and
transgenic salmonids. The former is because of major genetic differences have been
described between European and North American salmon stocks. North American protocols
ban the import of reproductively active European salmon. Another proposal of the NASCO
working group is to map epidemiological zones and control the movements of fish between
Contamination of wild stocks might also be reduced by 'localisation' (Hindar et al. 1991), i.e.
keeping the culture facilities or points of release of ranched stock as far as possible from
other rivers or to restrict aquaculture to specific areas. For example, it is hoped that the
ranched fish from the pilot project in Pitlochry will home to the point of release, effectively
separating them from wild stocks (Maitland 1985).
Other management options include improving access to spawning sites, the encouragement of
rearing and stocking sterile fish and reducing fishing pressure as alternatives to enhancement
(Elliot & Mills 1989). Stocking with sterile fish should perhaps be approached with caution.
Hindar et al. (1991) hypothesise that sterile but sexually active males might cause problems
by competing in courtship bearing in mind that similar techniques have been used in
biological control of mosquito's. Closed culture, i.e. operating from inland sites from which
escape is impossible, has been implemented on a farm in the USA (Hindar et al. 1991) but
may often be impractical. Selective harvesting could ameliorate problems caused by release
of cultured fish. This has been demonstrated in a mixed inland fishery. Gene frequencies in
naturally reproduced young indicated that the majority of parents were wild fish and not the
Review of information, policy and legislation on species translocations
more abundant hatchery fish. Wild fish therefore needed to be protected to maintain the stock
size. Consequently, the harvest of wild fish has been restricted by requiring that they are
released after capture. Selective harvesting in the marine environment would be aided by
understanding the differences in the migration patterns of wild and released fish (Hindar et al.
2.6.2 Crustacea and molluscs in aquaculture Background and case studies
Flat oysters Oystrea edulis and mussels Mytilus edulis have traditionally been moved
between sites in the UK to establish and enhance fisheries. In the Exe estuary the majority of
mussel beds were created artificially (McGrorty et al. 1993). Today artificial mussel beds are
enhanced by moving larvae from other sites which may be quite distant (S. McGrorty pers.
comm.). Shellfish are also regularly brought into the UK from overseas.
MAFF are currently nearing completion of study into the possibilities of ranching European
lobsters Homarus gammarus (a native species) in the UK (MAFF 1995). Several thousand
young lobsters have been released at inshore and offshore sites over a 5 year period. Results
at this stage indicate a 50% survival rate but cannot show whether the released animals are
augmenting or displacing the wild stock. Impacts of translocations and factors relating to management
Factors relating to pathogens and parasites
There are two notifiable pathogens of bivalves (Marteilia and Bonamia) which infect mussels
and oysters, and an important disease of lobsters known as Gaffkaemia, could be spread
through translocations (MAFF 1994). Guidelines and legislation help to control the spread of
parasites and pathogens. Persons intending to move shellfish between sites must follow the
guidelines provided by MAFF, WOAD and SOAFD and they must obtain licenses. The
commercial importance of shell fisheries ensures that effort is put into implementing the
controls set out by these guidelines.
Genetic impacts
As with other taxa the potential for outbreeding depression exists when individuals from nonlocal populations are introduced. In the case of lobsters little appears to be known about the
genetics of native populations. Early results of research by MAFF indicate that lobster
populations around the UK appear to be fairly genetically similar so it is possible that
translocations between populations might not be deleterious. MAFF however, are aware of
the possible risks of loss of local adaptation through genetic mixing and therefore intend to
recommend the precautionary principal (C. Bannister pers. comm.).
Review of information, policy and legislation on species translocations
Factors relating to management
There is a risk that the release of invertebrates for commercial interests might displace rather
than enhance wild stock.
2.6.3 Gamebirds Background and case studies
In the UK the main lowland game-bird species are exotic, i.e. the pheasant Phasianus
colchicus and red-legged partridges Alectoris spp. (Hudson & Rands 1988). However, use of
native species for game is common in some other countries. Of translocations reviewed by
Griffith et al. (1989) in Australia, Canada, Hawaii and the USA (1973-1986), 90% involved
native game species. For instance the wild turkey Melleagris gallopavo is extensively
released for game in the USA (Kleiman 1989). In the UK a limited amount restocking is
carried out with the native grey partridge Perdix perdix with birds from Denmark and other
European counties (I. Evans in prep.), and experimental restocking trials have been
conducted with red grouse Lagopus lagopus scoticus. In the past the red grouse has been
released in large numbers in the UK but the releases and their outcome have apparently not
been controlled or monitored (Picozzi 1989).
The capercaillie Tetrao urogallus became extinct in the UK in the late 18th century. Swedish
stock was successfully introduced in the 1830s by a number of translocations onto Scottish
estates (Lever 1977). They became abundant and were even regarded as pests in the 1950s.
More recently numbers of capercaillie have declined and a management plan has been
published, but the option of restocking is not discussed (Moss & Picozzi 1994). In their
policy statement (G. Williams pers. comm.) the RSPB report limited restocking of
capercaillie and black grouse but state that this is not recorded and fails to meet some of their
criteria for approval of translocations. In addition, as capercaillie is listed on Schedule 9 of
the Wildlife and Countryside Act 1981 restocking should be licensed. Impacts of translocations and factors relating to their success
Ecological and economic impacts
Restocking with the grey partridge in the UK has a record of poor success, probably because
brooder reared chicks are likely to have poorly developed predator avoidance behaviour
(Potts 1986). Potts (1986) concludes that, due to the large financial investments required and
the poor results obtained, restocking partridges using current techniques is not beneficial to
partridge conservation. In fact low survival rates and reproductive success of released grey
partridge may lead to damage to wild populations if the released birds compete for territories
or pair with their wild relatives (I. Evans in prep.). Similar conclusions were drawn by Price
(1994) during restocking trials with red grouse in Scotland which also showed that
supplementation of this sort was not commercially viable. Price (1994) found that the
breeding density of wild grouse declined during her release experiment although there was no
direct evidence that the decline was caused by the release. Hudson & Rands (1988) state that
release of hand-reared birds for game is associated with decreased management effort and
heavier shooting of wild stocks. If native game birds are to be restocked care would have to
Review of information, policy and legislation on species translocations
be taken to avoid these problems. As with translocations of other taxa there is the risk that
released gamebirds might carry novel pathogens which could infect local populations.
Genetic impacts
According to Picozzi (1989), there is a high probability of hybridisation when large numbers
of gamebirds are released into a wild population. He cites evidence that there are sixteen
types of natural interspecific hybridisation in the Tetronidae. Hybrids with pheasant
Phasianus colchicus and black grouse Tetrao tetrix were produced during the dispersal of
released capercaillie in Scotland. However, these are thought to pose no threat as they are
rarely recorded in the wild today (Picozzi 1989).
Undesirable genetic traits may be introduced into a wild population as a result of releasing
captive-reared stock. This is particularly true when the source animals derive from farms
where breeding regimes have deliberately selected for domestically desirable traits. For
instance, farms provide grey partridge for release in the UK. These birds are bred for
maximum egg production and not for an ability to survive and breed in the wild (Game
Conservancy Advisory Service 1992).
Presumably, the import of grey partridge stocks from Europe may cause a threat of
outbreeding depression through
genetic mixing if UK populations are locally adapted. There does not appear to be any
mention of this in the literature.
2.6.4 Falconry Background and case studies
Translocation of native species occurs through falconry due to accidental and deliberate
releases. A large number of native birds of prey are kept by falconers in the UK and a high
proportion are lost or released (Kenward 1974). For instance, I. Evans (in prep.) reports DOE
records of 31 goshawks Accipiter gentilis officially recorded as lost between January 1993
and March 1994.
As birds kept for falconry are trained to hunt they might be expected to survive in the wild
and it has been shown that a proportion of lost falconry goshawks can enter the wild
population. Falconry equipment may still be attached to birds breeding in the wild and many
of those birds are morphologically different (larger) than those that would be expected to
colonise naturally from the continent (Marquiss 1981). Impacts of translocations and factors relating to management
Review of information, policy and legislation on species translocations
Ecological and genetic impacts
Damage to wild raptor populations through the influx of birds from falconry is not reported in
the literature. However, potential threats due to uncontrolled releases are present as with
other taxa. These include undesirable genetic and ecological changes and transmission of
pathogens. Research is advisable to determine when releases of falconry birds might be
harmful to wild populations.
Unrecorded releases may mislead ecological research. Marquiss (1981) reports that rapid
increases in goshawk populations in the UK in the 1970s may have been affected more by
numbers of birds imported than by intrinsic population growth.
2.6.5 Bee-keeping Background and case studies
The possible impacts of bee-keeping on native honey-bees (Apis mellifera mellifera lehzeni)
in the UK are reviewed by Elmes (1989). In the 1920s honey-bee stocks fell to a very low
level and the Ministry of Agriculture encouraged the import of bees, especially the Italian
subspecies A. m. ligustica. Subsequently, the native honey-bee was thought to have become
extinct. However Elmes (1989) writes that most studies have indicated that some regions of
Britain still support colonies of honey-bees genetically very similar to the ancestral
Currently, most commercial apiarists stock A. m. ligustica or hybrids of imported stock. In
the past the development of distinct domestic strains has been hampered by the reproductive
behaviour of the species which precludes control of the paternal line. However, more
breeders are now using artificial insemination techniques to overcome this problem.
Techniques may also be developed to produce genetically modified strains for pollinating
specific crops.
There is also a growing trend to rear and translocate species other than honey-bees to
pollinate crops whereas traditionally only honey-bees have been transported for this purpose
(Williams 1995).
Current legislation restricts imports and translocations of honey-bees in the UK for
commercial purposes. Nevertheless, the potential impacts on wild honey-bee colonies of
pathogens and genetic changes related to past translocations need to be understood to allow
effective conservation management. There is therefore a clear need for research on the
ecology and genetics of wild honey-bees in the UK. Impacts of translocations and factors relating to management
Genetic impacts
Genetic material from commercial strains may enter wild populations through the escape of
drones, which are not husbanded as carefully as queens, and the sale of bees to cottage
Review of information, policy and legislation on species translocations
bee-keepers effectively releases them to the wild. Accidental translocations of this sort may
pose a threat to the native honey-bee. It is not possible to predict the effects of genetic
exchange between domestic and wild stocks at the moment because there is virtually nothing
known about the abundance, distribution or genetic characteristics of wild honey-bees in the
UK. It is also unknown whether the various races and hybrids vary in their ability to pollinate
plants native to the UK.
Factors relating to pathogens and parasites
The introduction of a pathogenic mite Varroa jacobsoni into European honey-bee colonies
via the import of honey-bees from stocks originally introduced from Asia to Europe and its
subsequent spread is a good example of the serious and unpredictable effects of species
2.6.6. Translocations and releases for aesthetic purposes Background and case studies
Animals and plants are frequently released simply because the individuals or organisations
involved find them aesthetically pleasing, but motives for translocations can often be more
complicated, involving well meaning but misguided attempts at species conservation. It could
be argued that a substantial part of the motivation for officially approved reintroduction
projects is aesthetic. Conservation of endangered species has tended to focus on birds and
mammals (Lyles & May 1987), and charismatic species have been much more likely to be
chosen for reintroduction projects than species with less emotional appeal (Whalley 1989).
Approaches are now changing; for example EN's Species Recovery Programme involves 62
species of which only 11 are birds or mammals.
Native butterflies are commonly released in schools and gardens for aesthetic reasons
(Thomas 1989b). This is easily achieved because larvae and pupae of butterflies are readily
and cheaply available from institutions such as the British Rare Breeds Society (Lepidoptera
Section). It is possible that between a third and a half of the existing colonies of Strymonidia
pruni originate from this type of release (Thomas 1989b). Some releases may have a simple
amenity motive such as the release of aberrant lepidopterans by collectors hoping to increase
the proportions of these 'valuable' individuals in wild populations (Thomas 1989b). However,
many unofficial releases of Lepidoptera carried out in the UK might be considered to have
conservation value by the persons involved (e.g. Smith 1992). In their 'Review of Butterfly
Introductions in Britain and Ireland' Oates & Warren (1990) studied 323 attempts at
establishments. They found that 17% were released for amenity purposes and another 29%
consisted of indiscriminate dumping of excess broodstock. Moreover, 47% of the releases
had 'conservation objectives' but 45% of the releases for 'conservation' post-1960 were
introductions to sites with no previous records. This gives some indication of the scale of
releases but of course can only include releases that have been notified.
The involvement of hobbyists in butterfly breeding and releases dates back to the 1920s when
the Royal Entomological Society established its 'Lepidoptera Protection Committee', which
had a policy of introducing threatened species into new districts. Enthusiasm for amateur
Review of information, policy and legislation on species translocations
butterfly releases has continued but has been officially disapproved of since the 1940s (e.g.
JCCBI guidelines, see section 2.3.1). This has effectively driven the practice underground, so
that although there was a dramatic increase in such releases in the 1980s this has been
impossible to monitor (Thomas 1989b, Oates & Warren 1990, M. Warren pers. comm.).
Other taxa are also translocated by members of the public and amateur groups because of
their aesthetic appeal and in some cases a belief that the action will benefit the species. Prior
to the inclusion of the species in Schedule 9 of the Wildlife and Countryside Act, release of
barn owls Tyto alba had become common enough to arouse concern of the conservation
bodies (Hanna 1992, JNCC 1994b). Native plants are occasionally reported outside their
usual geographic range after being moved for aesthetic reasons. Ellis (1987) reported the
discovery of Helianthemum apennium in Gower, prior to learning that the species had been
transplanted from Somerset. Several other examples are reported in the BSBI News but it is
difficult to determine the scale of the problem. In many instances it may be difficult to
determine source of unusual records of species. The geographic range of the Dorset heath
Erica ciliaris has been intensively studied and appeared to be restricted to Dorset but recently
there have been records of this plant in the New Forest. It is unknown whether this is the
result of a deliberate transplantation, seed dispersal from a nursery or natural dispersal of
seeds e.g. by birds (Chapman & Rose 1994). Impacts of translocations and factors relating to their success
Factors relating to conservation management
The largely uncontrolled and undocumented releases of butterflies are recognised as a serious
problem in conservation management (Thomas 1989b, Smith 1992, Stubbs 1995). The
butterfly populations of reserves and other sites are often monitored to assess their suitability
and carrying capacity in order to plan habitat management regimes. (Stubbs 1995) reports
that unofficial releases in such areas during research have disrupted some projects and have
undermined the effort invested in butterfly recording schemes. Likewise if the Dorset heath
arrived in Hampshire through human intervention its presence misleads ecological research.
The release of barn owls without regard to the distribution of wild conspecifics caused
concern that they might compete with resident owls for food territory and mates (Hanna
1992, JNCC 1994b).
Factors relating to welfare and the spread of pathogens
Captive-bred barn owls were known to suffer unusually high mortality following release by
members of the public (Cayford & Percival 1992). This was largely due to unsuitable release
techniques and habitats. Ad hoc releases are also thought to increase the risk that released
individuals might carry disease into local areas (JNCC 1994b).
Genetic impacts
Indiscriminate releases may also arouse concern from a genetic viewpoint (Stubbs 1995).
Thomas (1989a, b) notes that there is probably no pure native population of Marsh Fritillary
Euphydryas aurinia left in Hampshire after regular releases of captive individuals of mixed
Review of information, policy and legislation on species translocations
genetic stock over the last 20 years. However, he also points out that morphological rather
than ecological characteristics are generally used to distinguish races and subspecies of
lepidoptera and that much research is required on the ecological genetics of this taxon to
discover whether mixing of local races can harm the species. In conclusion, Stubbs (1995)
recommends the precautionary principle established in the JCCBI guidelines. Discussion
Where haphazard and potentially problematic releases attract publicity it is sometimes
possible for conservation organisations to effectively argue against them. For instance in
1984 the Guernsey Tourist Board attempted to launch 'Project Papillon' which aimed to breed
vast numbers of butterflies for release in Britain. The NCC argued that the scheme was
inappropriate and that the butterflies would generally be unable to find suitable habitat and
would therefore die. This argument relates to welfare rather than conservation and is unusual
in that insect welfare considerations tend not to be high profile (Whalley 1989). As a result of
the arguments against it the project was scaled down (Oates & Warren, 1990). Problems such
as unforeseen genetic changes or transmission of disease could have resulted from such a
large scale release and a possible lesson from this example is that welfare issues tend to carry
more emotional impact and hence can be more effective than arguments for preservation of
The majority of cases of translocations in this category are not high profile. The fact that
aesthetic and amenity releases, including those with a well intentioned but misguided
conservation motive, are generally conducted by amateurs who are acting, in some cases,
without official consent, makes this area of translocation activity very difficult to monitor. In
an attempt to rectify the problems associated with indiscriminate release of butterflies the
JCCBI set out a code of practice, 'Insect Re-establishment- A Code of Conservation Practice'.
This aimed to encourage people to document and notify releases but has met with virtually no
cooperation from those involved in private releases. It would appear the voluntary principle
has failed. Consequently, the British Butterfly Conservation Society has proposed that all
British Red Data Book and notable butterflies be added to Schedule 9 of the Wildlife and
Countryside Act (Stubbs 1995). This legislation has been used to help prevent irresponsible
releases of barn owls (JNCC 1994b). The proposal has engendered a hostile reaction from
some amateur butterfly release enthusiasts, which illustrates the lack of understanding
between elements of the amateur groups and the conservation and scientific bodies. One
individual claimed to have taken part in or supplied stock for just under a thousand butterfly
releases in the UK. He was delighted to think that the 'experts' would be fooled into thinking
that his surreptitiously released stock were a natural population (White 1995). Due to the rift
between parties interested in butterflies even if some species are added to Schedule 9 the
licensing of release activities would be extremely difficult to police. Concentration on liaison
between groups to encourage cooperation and education is clearly important in this case.
2.6.7 Releases for animal welfare
Charismatic animals such as birds of prey, hedgehogs Erinaceus europaeus and badgers
Meles meles found as accident victims or orphans may be rehabilitated and released in the
Review of information, policy and legislation on species translocations
interests of animal welfare. Relocations of individual badgers or social units are often
conducted to avoid development operations or destruction for pest control. Translocations of
badgers are controlled by the Protection of Badgers Act 1992 which requires that capture of
badgers must be licensed by the statutory conservation agencies.
Although this type of translocation is primarily motivated by animal welfare objectives it
may actually result in welfare problems. In some cases animals such as badgers may suffer as
a result of the translocation and humane destruction may be preferable (Roper 1988).
Inappropriate releases may lead to animals being unable to find adequate food, or nest sites
and suffering to territorial competition with wild resident conspecifics (Morris et al. 1993).
Other problems relating to the effects of inappropriate release techniques or sites on the
chances of success of the translocation are the same as those discussed under 'Translocations
for Conservation'.
Prior to recent work by Morris & Warwick (1994) and Morris et al. (1993) there had been
little attempt to determine the fate of the hundreds of rehabilitated hedgehogs that are
released each year in the UK. Visual markers and radio-telemetry were used to study released
adult hedgehogs and juveniles which had virtually no experience of life in the wild. The
results indicated that mortality was surprisingly low, even for naive juveniles and that the
released individuals were able to feed adequately, build nests and integrate with their wild
Another interesting finding was that the source of adult animals appeared to be important.
Animals from urban origins moved surprisingly large distances to find what was presumably
familiar territory after release in a rural area. Nevertheless, they warn that release in urban
areas may still be inadvisable due to heavy traffic. They conclude that rehabilitating
hedgehogs is 'possible and worthwhile' and offer recommendations for good practice.
2.6.8 Translocations for scientific research Background and case studies
This category includes translocations intended to test a scientific hypothesis. By virtue of
their nature they should be better documented than other types of translocations and hence are
probably over-represented in the literature.
In the review by Oates & Warren (1990), 7% of 323 butterfly releases reviewed were for
scientific purposes. Thomas (1989c) cites genetic research on the scarlet tiger moth Panaxia
dominula by Sheppard (1951, 1961) and Ford (1971) which involved releases of individuals
containing known frequencies of a gene in new and existing populations.
Mammals have also been translocated in the UK for genetic research. For instance the
experimental release of house mice Mus musculus onto islands by Berry et al. (1991). This
provided valuable insights into the interactions between social behaviour and gene flow.
Review of information, policy and legislation on species translocations
Other experimental releases are intended to test ecological hypotheses. In 1954 HopeSimpson et al. transplanted rare and local plants in sites near Bristol to study the reasons for
the species range restriction and factors important in establishment. The authors took great
care to notify the community of the planned work to avoid causing misleading records.
Current experimental releases of the common buzzard Buteo buteo in areas where it is rare in
the UK, are intended to help explain the distribution and dispersal of this species in the UK
(S. Walls and R. Kenward unpublished report). The results may be applicable to species
conservation by the development of techniques for animal translocations which combine
study of a wild population and intensive monitoring of a small number of released
individuals. Red squirrels Sciurus vulgaris were released in southern UK to study interactions
between this species and grey squirrels Sciurus carolinensis (Kenward et al. 1995). The
primary motive for this work was scientific research but the results may be applicable to
planning future management of this species including possible reintroductions. Other recent
translocations of the red squirrel in the UK have been primarily intended as attempts at reestablishment so these are included in the relevant section.
2.7 General discussion
2.7.1 General discussion of translocations for conservation Discussion of legislation
The outlines of legislation and guidelines given in sections 2.3 and 2.4 show that
translocations for conservation are covered by international and national guidelines relating
to all taxa or to groups of taxa but that there is no legislation specifically relating to this type
of translocation.
Several animal species rare enough to merit translocations for conservation are listed on
Schedules in the Wildlife and Countryside Act 1981 which make it illegal to possess them or
to remove them from the wild unless a license is granted by the nature conservancy agencies.
This means that it is possible to regulate to some extent reintroductions, re-enforcements and
relocations of these species through licensing. However, the constraints on translocations of
plants are less strict. It is possible to collect seed of any wild plant (providing that seed of
Schedule 8 plants are not sold). This makes it feasible to carry out translocations of any plant
species. Therefore, there is no control on the release of native species to the wild.
There is a feeling among many conservation organisations that current legislation concerning
the translocation of native species is weak and requires a 'significant overhaul' (e.g. G.
Williams pers. comm.). A number of interested parties have proposed that many other native
species should be placed on Schedule 9 as a means of controlling and regulating releases,
partly because of a perception that the capercaillie Tetrao urogallus, the white-tailed eagle
Haliaeetus albicilla and the barn owl Tyto alba were listed for such a reason (see above).
Butterfly Conservation has proposed that all British Red Data Book and Notable butterflies
Review of information, policy and legislation on species translocations
should be added (Stubbs 1995). It has been suggested that the Schedule might be split - one
part listing the non-native species and the second part listing the native species. This would
clarify the intentions behind the listings. These suggestions are still under debate and need to
be explored with DoE before any firm proposals are made by the statutory conservation
agencies (M. Palmer pers. comm.).
W. Parish, of the DOE (pers. comm.), does not encourage such a use of Schedule 9, arguing
that such extensive additions would change the intentions of the Act; the barn owl was a
special case, and the reasons for the listings of the other two native birds are obscure. Finally,
it is relevant that, while the Wildlife and Countryside Act does not explicitly prohibit,
regulate or even mention reintroductions (apart from, by default, those involving animals of a
kind not ordinarily resident in Great Britain, e.g. the case of the chequered skipper), the
source document, the Bern Convention, explicitly encouraged reintroductions and stated a
need for some form of control. It is not clear whether this omission from the Wildlife and
Countryside Act was deliberate or not. The common statement that the Wildlife and
Countryside Act does not differentiate between introductions and reintroductions (e.g. NCC
1983) is only true if one considers reintroductions of species extinct throughout Great Britain.
Reintroductions and supplementations of species extant in Great Britain are simply not
considered. This issue clearly requires more debate. Discussion of guidelines
The elements of guidelines pertaining to translocations for conservation in the UK are listed
in detail in section 2.3 and will not be reiterated in this discussion. It is evident that detailed
guidelines covering all parts of the translocation process are available at the national and
international level. These include requirements for feasibility studies, management plans,
monitoring and documentation. Some guidelines are quite specific such as the suggestion by
the Botanical Society of the British Isles that translocations of rare vascular plants over 1km
from their donor sites should be discouraged. Often however, they are more generalised
because they are intended for a number of taxa or for all taxa and must interpreted for a
variety of circumstances.
Evans (1995) makes the point that reintroduction guidelines for birds (e.g. Black 1991) are so
generalised that their criteria are open to a variety of interpretations. Guidelines such as those
of the IUCN (1995) may be difficult to implement directly due to their general nature.
However, the generalised nature of guidelines is addressed in the IUCN guidelines, which
explain that generalisation is inevitable given the need to include all taxa. However, the need
for rigorous review of each proposal is stressed and the intention to provide handbooks for
individual taxonomic groups is mentioned. Guidelines for translocations of some individual
taxa have already been proposed. For instance, those suggested for rescue translocations of
crested newts in the UK (Herpetofauna International Ltd. 1991). Other taxa already have
guidelines produced by the IUCN such as those produced by the IUCN Otter Specialist
Group (de Jongh 1995). Bramwell (1991) suggests that interpretation of guidelines should
occur during discussion of reintroduction and re-enforcement proposals on a case by case
basis so that interpretation takes place in fora of interested parties as opposed to within
individual organisations.
Review of information, policy and legislation on species translocations
The IUCN (1995) document contains the most modern and detailed guidelines and it would
seem sensible to use these as a basis for any new UK guidelines. Monitoring of translocations
The need for adequate monitoring and documentation is also stressed in the guidelines. The
literature reflects the need for more rigorous attention to this aspect and draws attention to the
fact that a major problem concerned with translocation policy is implementation of the
Maunder (1992) indicates the lack of adequate documentation for plant reintroductions in the
literature worldwide. Less than 25% of studies in a literature search covering over 200
references indicated what techniques were used or whether monitoring was carried out after
the translocations. Similar inadequacies in monitoring of released animals is indicated by
Bright & Morris (1994) Bright & Harris (1993) and Oldham et al. (1991).
This lack of adequate documentation may be especially acute for reintroductions that are
regarded as failures (Griffith et al. 1989). Conservation and wildlife management projects
have been criticised for not conducting or recording reintroductions scientifically. A change
in attitude is required so that the rejection of the hypothesis that a viable population of a
species may be established by a certain method, is viewed not as a disaster but as an
advancement of knowledge (MacNab 1983).
It may be appropriate for agencies providing permits and/or funding to ensure that data are
recorded rigorously and to provide facilities to make it accessible by databasing (Griffith et
al. 1989). Only through effective communication and application of accumulated knowledge
are the chances of success of reintroductions likely to be improved. Nevertheless,
requirements for recording and databasing should operate to focus and aid reintroduction
attempts and not increase what is perceived as bureaucratic encumbrance which saps energy
from wildlife management (Cade 1986). The insect establishment recording scheme set up by
JCCBI and the BSBI plant translocation database are largely unused now (M. Warren, V.
Morgan pers. comm.) but provide examples of the necessary approach. Some directions for future research on translocations for conservation
The importance of adequate research cannot be overemphasised. This should cover genetics,
ecology and other factors associated with translocations such as pathogen transmission.
Ecological research can help to indicate which taxa might respond well to reintroduction
attempts. For instance Quinn et al. (1994) have shown that the distribution of a sample of
scarce plants in the UK is correlated with their dispersal ability. Those which dispersed
poorly had more aggregated distributions. They suggest that this may imply that
translocations may be useful in the conservation of such species. The research should also
attempt to be predictive.
The formal advisory procedure in the NCC (1990) guidelines suggests that the rate of reestablishment and the possible effect of the establishment on the community should be
predicted. In feasibility studies this is often approached with comparative methods using
Review of information, policy and legislation on species translocations
extant populations or translocations at other sites. More predictive results are likely to be
gained if this approach is combined with modelling or by using trial or experimental releases
where feasible. Although the requirement for adequate research prior to translocations is
emphasised in the guidelines there is no specific reference to the benefits of trial releases,
transplantations or of modelling.
The role of modelling in application of the theories of population viability and vulnerability
to conservation translocations is discussed in Guerrant (1992) who shows models that
incorporate genetic and demographic parameters and based on the concepts of Minimum
Viable Population (MVP) and its derivative, Population Viability Analysis (PVA).
Simulation modelling with VORTEX (a stochastic population simulation model (Lacy 1994))
was used in the recent study of the feasibility of reintroducing the beaver Castor fiber to the
UK (Macdonald et al. 1995). The results demonstrated the utility of this technique in
formulating strategy for reintroductions. They showed that different results might be obtained
with sequential and unsupplemented releases and that the possibility of creating
metapopulations should be explored.
Obviously the limitations of models must be considered and they should be paramaterised as
carefully as possible.
Experimental translocation
The advantages of using experimental translocations to increase the chances of success of
reintroductions is stated in Mistretta (1994), and Evans (1995), in a review of avian
reintroductions, emphasises the importance of experimental releases. In the case of avian and
other vertebrate taxa the monitoring of individuals may require the use of visual markers
and/or radio-telemetry techniques.
The combination of ecological study of the species with intensive post-release monitoring can
indicate whether the habitat of the recipient site is suitable by interpretation of the behaviour
of released individuals. This is referred to as the 'Noahs Dove' approach in Kenward and
Hodder (in prep.).
The use of experimental translocations is equally applicable to more sedentary species and
certainly more easily applied. Many translocations could be included in the experimental
category merely by more rigorous attention to planning, monitoring and documentation.
Negative effects on other species
Conservation translocations may result in population declines of resident species at the
release site and other site that the species colonises subsequently. This may be especially
noticeable when top predators, such as birds of prey, are released and reduce the numbers of
the prey species. Effects on the release site are inevitable and the possibilities should be
assessed, as much as possible, before the translocation. Experimental releases may help such
Review of information, policy and legislation on species translocations
an assessment. It should the be possible to determine whether the effects are acceptable, or
even desirable (e.g. if a species causing conservation problems is decreased)
Comparative ecological research
Ecological requirements of a species subject to translocation may be assessed at extant sites
in the UK. However, where species are extinct in the UK or reduced to extremely small
populations ecological research overseas may be necessary. The Plymouth pear Pyrus
communis is only found in habitat created by human activity in the UK so studies of its
ecology in woodland in Europe were necessary to choose a site for reintroduction (A. Jackson
pers. comm.).
Research into tests for pathogens
Griffith et al. (1993) found, in a survey of translocations of terrestrial mammals, only 24%
included professional examination of individuals for disease or parasites prior to release.
Potential threats caused by the transmission of pathogens as a result of species translocations
could be reduced by the development of diagnostic tests, improvement of quarantine systems
and better research into the incidence, distribution and risk of disease in wild and captive
populations (Wolff & Seal 1993). Appropriate application of resources for disease
investigation should be aided by the development of models for prioritising infectious
diseases (Munson & Cook 1993). These would include tests for viral, bacterial or
mycoplasmic infection (Maunder & Ramsey 1994).
Genetic research
Recently there have been great advances in the understanding of the genetics of populations
and species. By examination of DNA sequence variation it is possible to make inferences
about many aspects of populations including genetic diversity and gene flow between
populations and species.
Many reports or critiques of translocations mention genetic considerations. Concerns are
expressed regarding such processes as 'contamination of local genetic stock', 'genetic erosion',
or 'spread of unsuitable genotypes'. Thus, the use of non-native or even non-local, stock, is
often seen as negative in itself, with little regard for any actual positive or negative impact on
biodiversity. In this report we have suggested that problems may arise due to difficulties in
establishment of poorly adapted individuals or swamping of resident populations with
maladaptive traits. However, careful choice of individuals for translocation could be
beneficial due to increases in genetic diversity of a species within an area; particularly if
adverse effects of inbreeding or genetic bottlenecks can be reduced. Such cases may merit the
use of individuals from large populations such as might be found in SSSIs or nature reserves
for translocations.
The use of non-local donor populations may, in some cases, have only minor impacts. Rapid
selection or introgression with local populations could result in a well-adapted introduced
population exhibiting similar morphological and ecological characteristics to local types.
Review of information, policy and legislation on species translocations
In Mistretta (1994) the following questions are listed in relation to management decisions
regarding rare plants. These indicate directions that genetic research on any taxa might take
in consideration of reintroduction or re-enforcement programmes.
What is the optimal level of genetic relatedness of potential mates? i.e. How
important is the risk of inbreeding depression versus the maintenance of local
adaptation or co-adapted gene complexes (outbreeding depression)?
What sort of populations should be chosen for captive breeding or reintroduction
Should effort be concentrated on species that although rare appear to possess
sufficient genetic diversity to allow future adaptation or on those that already are
genetically depauperate and may be suffering from inbreeding depression?
What is the value of a rare versus a common allele?
Mistretta (1994) concludes by warning that both the advantages and limits of genetic analysis
should be recognised. The danger is that awareness of the need to preserve genetic diversity
can induce a paralysis preventing important conservation activities from going ahead until a
full range of genetic analyses are made. The review suggests that where genetic analyses are
not possible theoretical predictions should be used to guide strategy. Environmental and social factors
The success of translocations is highly dependent on the availability of appropriate habitat. In
the UK it is important to recognise that conservation of biodiversity takes place in the context
of a landscape altered by human activity (Thomas & Morris 1994). The extinction of several
species of mammal (Yalden 1986) and many other taxa may be attributed to anthropogenic
causes, but other species may have survived to the present or recent past because of human
activity. Thomas (1989a, 1995b) suggests that most butterfly species currently restricted to
warm microclimates in the UK would probably have become extinct between 4,000 and
2,500 years ago were it not for artificially created refugia. These species may be seen
popularly as part of a 'natural' landscape, but in fact, were maintained by traditional
management practices and suffered population declines or losses largely due to changes in
agricultural practices. The importance of establishing appropriate habitat management
concurrently with translocations for conservation is emphasised by this point.
Reintroduction projects are unlikely to succeed without adequate research into habitat
requirements and availability and better integration with habitat management (Maunder
1992). This may require practical site management at a local level or less direct measures for
larger areas. The approach to habitat management for reintroduced species will depend on the
species concerned. An important factor is the geographic range used by individuals and
populations. The designation of SSSIs and appropriate management within these areas is
suitable for sedentary species (e.g. many butterflies) or those with small home ranges (e.g.
dormice Muscardinius avellanarius). Elliott et al. (1991) suggest that in the case of a wide
ranging species such as the sea eagle Haliaeetus albicans which appears to be dependent on
traditional land use, support systems such as 'Environmentally Sensitive Areas' are more
Review of information, policy and legislation on species translocations
It is equally important to consider that, given the influence of human activity on the
environment in the UK, the loss of a species to a region, even if within historical times, is not
sufficient to justify reintroduction. Some species, particularly those which are wide-ranging,
may simply not be compatible with the present environment. This argument is used by
opponents of attempts to reinstate certain species, but may not be appropriate where the
conditions responsible for the species decline or loss have changed. The reduction or
cessation of persecution or hunting of some avian and mammal groups in the UK has made it
possible to consider or attempt several reintroductions. Examples include the white-tailed sea
eagle Haliaeetus albicans, red kite Milvus milvus, beaver Castor fiber and pine marten
Martes martes. Dispersal
Reintroductions of regionally or locally extirpated species are unnecessary if the species can
recolonise without human intervention. Butterflies and birds are mobile and it might be
assumed that they should be relatively capable of recolonising naturally within a reasonable
time-scale. However, many apparently mobile taxa exist as populations which are sedentary
or have rates of dispersal restricted by their ecology or behaviour. For instance, Thomas
(1984, 1995b) shows that many species of butterfly have low dispersal rates in the modern
European landscape; about 85% of butterfly populations in the UK are closed. This means
that they are unlikely to recolonise naturally and that some sites may be under-utilised. This
last suggestion is supported by the success of introductions of several sedentary species of
The rate of spread of many animal species may be slowed by philopatric behaviour. The
effects of philopatry have been illustrated for the red kite by Newton et al. (1994). That is,
birds wandered widely in their first year but returned to breed close to their natal area, this
resulted in a gradual 'rolling front' of expansion in the breeding range, rather than rapid
colonisation of distant sites.
Evans (1995) lists a number of birds including the goshawk Accipter gentilis which were
subjects of reintroduction attempts when they were nationally or regionally extinct in the UK
and in his view unlikely to naturally recolonise the area of reintroduction. In contrast, the
RSPB consider that the goshawk is unsuitable for reintroduction because it could naturally
recolonise (Everett 1978, G. Williams pers. comm.). The argument over the potential for
natural recolonisation might be resolved by considering what the acceptable rates of
recolonisation are for a given species. The BBCS guidelines, for instance, give an acceptable
rate of colonisation for butterflies. Communication and cooperation
A major obstacle to effective management of translocations in the UK is the lack of
dissemination of knowledge and paucity of communication and coordination between
interested parties (Oldham et al. 1991, Hanna 1992).
Improved communication, cooperation and coordination between governments, nongovernmental organisations, the academic community and the general public is necessary to
Review of information, policy and legislation on species translocations
maximise the chances of success of translocations for conservation (Miller et al. 1994). Plant
conservation for instance, can benefit from active partnerships between ecologists, geneticists
and horticulturists at botanic gardens (Bramwell 1991). Specialist knowledge contributed by
groups such as horticulturists and falconers has been vital in many reintroductions. For
instance the use of a falconry technique, 'hacking out' when releasing birds of prey such as
the red kite (Evans et al. 1994).
The coordination of interested parties in translocations for conservation is an important role
for the statutory conservation bodies. The contribution of translocations to conservation
Do translocations for conservation divert funds from less glamorous conservation work or do
they act as a flagship to attract political and public support? Is it really expensive? The latter
depends on perspectives and almost all reviews on the subject mention the high cost of
reintroduction projects (Kleiman 1989, Maunder 1992, Lamb 1992, IUCN 1995). Cade
(1986) however, estimates that it would cost $5 million a year to restore the world's
threatened birds of prey, equivalent to the cost of building one armoured car. In addition,
reintroductions of charismatic species may generate funds for conservation either by
attracting funding directly or by providing a focus for ecotourism. Thomas (1995a) points out
that the reintroduction of butterflies in the UK and Holland has generated rather than
competed for conservation resources. Public awareness of conservation issues is raised by
work on threatened species and single species can be used to secure or raise the status of
important sites (Maunder 1992). Jones et al. (1995) point out that species conservation can
provide a major incentive for habitat protection and ecological restoration work.
General discussion of translocations for purposes other than conservation
UK guidelines developed by the NCC in 1990 and currently endorsed by the statutory bodies
state that this class of translocation should be subject to legal control or consultative
procedures and that accidental releases should be avoided. Legislation and guidelines which
apply to translocations for purposes other than conservation are outlined in section 2.3 and
2.4, and much of the discussion on legislative control in the previous section applies to this
type of translocation as well. In each part of section 2.6 the extent of research, management
options and the development and implementation of guidelines and legislation are discussed.
Some general points can be drawn from this discussion.
Concern about this type of translocation is focused in two main areas.
1. Commercial and amenity interests concerned with economic threats posed by
translocations e.g. damage to native stocks through the transmission of pathogens.
A major area of commercial interest is that of salmon Salmo salar fisheries. In the case of
fisheries licensing provides legal control. Concern about the economic threats associated with
the possibility of damage to native stocks which may result from species translocations (e.g.
transmission of pathogens) has lead to research and development of guidelines. Due to the
Review of information, policy and legislation on species translocations
migratory nature of the fish the guidelines are necessarily international in their development
and implementation. Proposals include the mapping of epidemiological zones and control of
fish movements between them. Other strategies address the ways in which deleterious
impacts of accidental or deliberate releases might be ameliorated (further details in section
The need for consultative involvement with the nature conservation bodies in the
development of guidelines for translocations of species by commercial interests should be
stressed. Working groups may only be concerned with conservation of one species and may
not be aware of or interested in conservation of biodiversity in the ecological community as a
2. Conservation interests concerned with the possible impacts of translocations on
This includes possible deleterious ecological and genetic impacts of translocations for
commercial or amenity purposes. Movements of fish by anglers for bait or specimen fish for
instance may cause serious alterations to the ecological community. Again consultation with
the conservation bodies is important when policies on licensing this type of activity are drawn
Another important area of concern is the release and transplantation of species for amenity
purposes. The discussion in section 2.6.6 indicates that this can be an area of considerable
debate. This is particularly relevant to the release of native lepidoptera by hobbyists. Large
numbers of undocumented releases may cause undesirable environmental impacts and
seriously mislead the ecological research essential for effective conservation. The difficulty
of implementing guidelines to control this practice, the suggestions for legislative control and
the response to this are described. In this case the failure of the acceptance of the guidelines is
unfortunate as many of the individuals involved are well meaning towards conservation.
Cooperation and partnership between hobbyists, conservation bodies and the academic
community can be a great asset to butterfly conservation (New et al. 1995). Skilful liaison
and dissemination of information is required to promote this cooperation and help to
discourage the development of distrust between interested parties.
Two important points are illustrated by this example.
1. That the implementation of guidelines may be extremely difficult and that this should be a
prime consideration when they are drawn up.
2. That public attitudes are extremely influential and much of the difficulty encountered in
implementation of policy and guidelines may be due to inadequate education, communication
and dissemination of information.
Review of information, policy and legislation on species translocations
Summary conclusions concerning the translocation of species native
to the UK
Types of translocation of native species for conservation are reintroductions, reenforcements, benign introductions, relocations and seed-sowings and plantings for
habitat restoration.
Non-conservation motivated translocations of native species occur for fisheries,
shellfish aquaculture, gamebirds, falconry, bee-keeping, animal welfare, research and
aesthetic reasons (releases by hobbyists and other parties).
Any translocation of native species can affect the fauna and flora at the release site
and the surrounding landscape through: competition, predation, herbivory and other
interactions, spread of pathogens and parasites, alteration of the habitat, or gene flow
between translocated and resident populations of conspecifics.
This review has found good evidence for a range of adverse effects of nonconservation translocations on the species and habitats at the release sites. These
effects are particularly worrying in fisheries.
Because conservation translocations are carried out to enhance biodiversity, adverse
effects on the release sites are fewer, although concern has been expressed over
genetic effects in general (see below), and the possibility of large ecological effects
from certain translocations such as that of the beaver Castor fiber.
The success of a conservation translocation may be dependent upon: the amount of
available habitat at the release site, the source of individuals for translocation, the
number of translocated individuals, the genetic structure of the translocated
population, and attitudes of the public to the translocation.
Provision of sufficient and appropriate habitat is usually well-studied in conservation
translocations and is achieved by restoration and/or management.
Methods for assessment of demographic (appropriate numbers of translocated
individuals) and genetic criteria for the success of a conservation translocation are
generally poorly developed, although advances are being made.
Genetic criteria for conservation translocations are particularly poorly understood.
There are perceived needs: to use large numbers of individuals to achieve high genetic
variation and avoid inbreeding depression in the translocated population; and to use
local donor populations to ensure that the translocated population is well-adapted to
the environment of the release site and to avoid spread of non-local genes to other
populations ('genetic contamination'). Genetic risks of non-conservation
translocations are perceived to concern loss of local adaptations and of genetic
variation of local resident populations. Many of these concerns are poorly researched.
Review of information, policy and legislation on species translocations
Other areas which could be researched to improve the value of conservation
translocations are: autoecology and population ecology of the species; ecological
modelling; development of experimental programmes for pre-translocation testing of
factors affecting success and effects on other species; development of procedures for
testing for pathogens and parasites; and research into large-scale and metapopulation
dynamics of the species and the possibility of positioning translocations over a
landscape such that these dynamics are created or enhanced.
There is also a need for wider reporting and dissemination of results in order to
advance the state of knowledge concerning conservation translocations. Monitoring of
translocations and databasing of results would be essential parts of such a programme.
There is a need for legal control of conservation and non-conservation translocations
in order to reduce the possible adverse impacts on biodiversity, and for the latter, to
allow a coordinated and expert approach to translocation of species of conservation
value. There is no legislation that specifically regulates translocations of native
species. Amenity stocking of fish and shellfish are strictly regulated, but other forms
of translocation are only partially covered by legal controls.
Schedule 9 of the Wildlife and Countryside Act forms a possible basis for legal
regulation of translocations of native species; each release of any of the named
species requires a licence and the statutory conservation agencies are involved in the
licensing process. However, there are very few native species listed on the Schedule
and such a use of the Act would require a large number of additions to the Schedule.
There are many guidelines by UK and international conservation organisations which
cover both conservation and non-conservation translocations of native species.
Guidelines for the former list requirements for feasibility studies, management plans,
monitoring and documentation. The recent IUCN (1995) guidelines provide the most
detailed requirements and should be used as the basis for UK guidelines. A general set
of guidelines could be agreed which would then be amenable to precise modification
for particular species groups (e.g. birds of prey or vascular plants) or particular
situations (e.g. assessment of licence applications for Schedule 9 species).
Some non-conservation translocations can be covered by the guidelines for
conservation translocations, e.g. aesthetic and animal welfare releases, but amenity
translocations (especially fisheries and shellfish) require different guidelines. The
guidelines should be concerned with the risks to biodiversity of such releases and
would inform the licensing processes that regulate such releases. The development of
guidelines should involve the relevant industries.
Review of information, policy and legislation on species translocations
Exotic species have been introduced into Britain by humans, deliberately or inadvertently,
since the Neolithic. Indeed, many introductions are of such long standing that they are
generally accepted as part of the British countryside, e.g. sycamore Acer pseudoplatanus and
rabbits Oryctollagus cuniculus.
The period of time approaching 1500AD was characterised by increasing perturbations to the
UK environment by humans, some migration of human populations and improving marine
transportation (di Castri 1990). From 1500AD, trade routes across biogeographic regions
were opened and mass migration of human populations occurred across continents. The rate
of species introduction into Britain increased rapidly once global travel was possible and the
reasons for deliberate introduction became much more diverse.
However, the impacts of alien species upon the British flora and fauna have not been as great
as found on other land masses. On remote islands, the impacts of exotic species have often
been severe because the native flora and fauna are depauperate and susceptible to invasion. In
the case of the British Isles, a land bridge connected the island to Europe until approximately
7500BP, and as a result the biota is essentially continental in character, although it consists of
fewer species than mainland Europe (Pennington 1969). By comparison with Britain and
continental Europe, the terrestrial and freshwater fauna of Ireland is depauperate reflecting its
isolation by sea since the last glaciation (Costello 1993).
Types of introduction
The introduction of non-native species has been either deliberate or accidental, and can be
split into categories according to the reasons for introduction. As for the previous chapter, we
investigate the impacts of these introductions using case studies. Lists of alien species which
have established in the wild in the UK. are given in several publications, including Lever
(1977), Brown (1986), Perry & Ellis (1994) and Eno (1995). We shall not repeat these lists,
but rather consider examples of the effects of such established aliens. A number of case
studies from other countries are investigated, to provide a more general overview of impacts
of aliens. The main categories we investigate are:
Fish and shellfish stocking
Biological control
Review of information, policy and legislation on species translocations
Wildfowl and game stocking for sport
Amenity and ornamental planting, stocking or collections
Pets and domestic animals
Fur animals
Accidental introductions
Impacts of the introduction of non-native species
This section supplies an outline of the effects non-native species can have on native
biodiversity. These effects are investigated in detail for each type of introduction.
It has been widely accepted that the addition of exotic species to assemblages of native
species can result in a decrease in native species richness. Invasion by introduced species is
generally thought to be associated with losses of native species because the latter may be
directly affected by the exotic invader (e.g. competition, predation, herbivory, etc.) or may be
adversely affected by habitat changes caused by the invader (McIntyre & Lavorel 1994).
Introduced species are often translocated without their own population regulators (predators,
parasites, pathogens), and populations may be able to increase within the limits of the new
environment. Some introduced species that establish will have native herbivores, predators,
etc. that regulate population growth, but other species will not and it is these species that may
affect native species and ecosystems severely.
For example, a population of the New Zealand oyster, Tiostrea lutaria, was introduced into
MAFF experimental shellfish beds in the Menai Straits. While the crabs Carcinus maenus
and Cancer pagurus (resident predators) were observed feeding on Pacific oysters
Crassostrea gigas, both species were reluctant to feed on T. lutaria, possibly due to
mechanical difficulties associated with prey handling. This study illustrates that the presence
of native predators would not necessarily regulate a population of introduced species.
Some introduced species are also thought to inhibit recruitment of native species. A
continuing reduction in recruitment to native populations, together with the usual mortality of
native adults may eventually lead to extinction and thus a reduction in the number of native
It is also widely believed that exotic plant species will not support such rich species
assemblages of associated invertebrates or epiphytes as native species. Native plant species,
having been present for longer periods of time than introduced species, have provided more
opportunities for exploitation by insects through evolution and co-evolution. Introduced tree
species are thought to support less diverse floral and faunal communities than most native
tree species (Southwood 1961, Kennedy & Southwood 1964). However, Boyd (1992) points
out that this may well be due to the unequal effort targeted at monitoring non-native species.
He states that sycamores Acer pseudoplatanus support as many species of mosses and
liverworts as ash Fraxinus excelsior and elm Ulmus procera, and better than oak Quercus
Review of information, policy and legislation on species translocations
robur, beech Fagus sylvaticus and other native species. It is also as rich in lichen species as
elm and beech, richer than birch Betula spp and alder Alnus glutinosa, and has a similar
number of species of fungus as oak, whilst having more than ash or alder.
The introduction of exotic species poses several problems that may be grouped into general
ecological effects, spread of disease and genetic effects. Ecological impacts and introduction
of disease may also lead to effects on genetic biodiversity.
3.2.1 Ecological impacts
Introduced species may utilise resources in common with native species. An introduced,
competitively superior species may exclude native species from vital resources. This may
affect both plants and animals.
Introduced herbivores may have direct impacts on plant populations (by grazing, trampling,
etc.), and have indirect effects on other species in the community (by altering habitat). This
may be a particular problem on more isolated islands or ecosystems, if species have evolved
without grazing pressure.
The introduction of exotic predators can have serious consequences for native biota. Islands
and other isolated systems, where species may have evolved without adaptations specifically
for avoidance of predators, are particularly susceptible to predation. There are several
alternative consequences of predation (Ebenhard 1988): 1) there is little effect on the prey
population; 2) the predator population is regulated by factors other than prey numbers and
cannot maintain a population large enough to have a significant effect on prey populations; 3)
the predator reduces the prey population until the prey population goes extinct; 4) the
predator and prey population dynamics are interdependent, leading to, e.g. a dynamic
equilibrium or regular cycles in population size
Survival of species following introduction of predators may depend upon the rapid
development of avoidance behaviour. For example, Dickman (1992) investigated the
responses of house mice Mus domesticus to two species of introduced mammalian predators
in Western Australia, the red fox Vulpes vulpes and cat Felis catus. Mice in areas containing
introduced predators generally selected relatively dense vegetation compared to mice in
predator-free areas. They also used sites of greater vegetation density on moonlit compared to
dark nights. Survival rates of predator-experienced mice a month after release into an area
containing predators were 2.5 times higher than those of their predator-naive counterparts.
Review of information, policy and legislation on species translocations
Introduced species may disrupt ecosystems by themselves becoming prey to a native
predator. Addition of prey species may lead to an increase in predator population numbers,
with the result that native prey species may be further affected.
Habitat alteration
Introduced species may not necessarily affect native species directly, but indirect effects may
be severe. The alteration of habitat form or function by, for example, alteration of the water
table or the fire regime, alteration of soil properties or through alteration of the structure of
the native vegetative community, may make the habitat unsuitable for native species. Again,
isolated populations are most at risk.
3.2.2 Impacts relating to the spread of disease
Introduced species may carry parasites and disease. This presents a problem to susceptible
species that have not previously been exposed to the parasite or pathogen, and thus have no
natural resistance. Again, isolated or island floras and faunas are most at risk.
Genetic impacts
Introduced species may hybridise with related native taxa. Hybridisation, when successful,
can be thought of as an exaggerated form of gene flow (Grant & Grant 1994). The change in
genetic constitution and changes in phenotype can be considered a loss in biodiversity.
However, hybridisation may also affect the adaptedness of native species to the local
environment if it affects genetic correlations (i.e. the inheritance on co-adapted traits). The
outcome of hybridisation on genetic correlation depends on the allometries of the co-adapted
traits of the hybridising species. When the allometries are similar, the effect of hybridization
is to strengthen genetic correlations. Only with species of different allometries are genetic
correlations likely to be weakened or eliminated by hybridization.
Introduced species may fragment and reduce the abundance of native populations. This may
lead to a loss of genetic variation due to genetic drift and inbreeding.
Summary of guidelines
This section summarises the policies (statements of position) and guidelines (recommended
procedures) of UK conservation and other organisations relating to introduction of non-native
species. These are not implemented by law, but represent, at most, agreed codes of conduct.
In many cases the guidelines have been condensed although some sections are transcribed
verbatim. Some terms have been changed to conform with our definitions.
Review of information, policy and legislation on species translocations
3.3.1 UK Guidelines
General Guidelines
1) Nature Conservancy Council (1990). Review of NCC Policy on Species Translocations in
Great Britain
No guidelines beyond a statement of UK legislation on introductions were given in NCC
(1983) or NCC (1987). NCC (1988a) began the drafting of a policy, which was more fully
developed in the NCC draft guidelines of 1990.
One element discussed in NCC (1988a), but missing from the later draft guidelines, was
the need to consider the likelihood of hybridisation, or other forms of gene transfer,
with native species or races that could alter the characteristics of the native species or
The NCC (1990) guidelines considered only new proposals for intentional introductions in
the context of NCC advice on these. These guidelines are those currently endorsed by EN,
CCW (L. Howe pers. comm.) and SNH (SNH 1994) and their basis was that
'a few [introduced non-native species] have posed ecological problems as pests,
carriers of disease or competitors with native species.'
The following should be considered in assessing a proposed introduction.
Can the species or race establish in the wild?
Will the species or race be restricted to a defined area or habitat?
Will the establishment be detrimental to the area or habitat concerned?
Are there methods to control establishment or spread?
What are the implications of spread from the area of establishment?
Would the native habitat of the species or race be damaged by removal of the
Will the release be monitored?
It was suggested that certain types of introduction should be subject to a code of
conduct and would require approval. These were:
animals other than game species;
plants other than commercial crops, or crops likely to establish in the wild beyond a
distance of five miles from the release site;
any release onto or adjacent to uncultivated (i.e. semi-natural or natural) land;
any release onto or adjacent to sites of recognised conservation value.
Review of information, policy and legislation on species translocations
Each proposal should include a management plan, similar to that proposed for
reintroductions, to contain information on:
details of the species or race concerned, its distribution and ecology;
origin of the organisms;
effects of removal of the translocated organisms on the donor site, if applicable;
purpose of the introduction;
other species that might be introduced along with the intended species, e.g. pathogens
or parasites;
details of the project team and consultant specialists;
description and reason for choice of the release site(s);
the habitats surrounding release site;
the anticipated ecological relationships of the introduced species with the ecosystem
of the release site(s);
mechanisms ultimately likely to limit numbers and distribution of the species;
the anticipated spread away from the release site;
appraisal of possible effects on genetic structure of existing populations near the
release site;
proposed site management;
details of the introduction procedure - number and life history stage of individuals and
methods and timing of release;
methods available to control the species after release;
the monitoring programme;
A national register of introductions should be established.
2) Stubbs (1988). Towards an introductions policy. Wildlife Link
The suggested guidelines of the Wildlife Link report (see Chapter 2 for a background) were
also seen in the later NCC (1990) guidelines, but with an exception concerning secondary
procedures to follow the initial approval of the introduction.
There should be a minimum period of 6 months consultation before final approval.
Field trials should be implemented.
3) UK committee for International Nature Conservation (1979): Wildlife introductions to
Great Britain. (Linn report)
The recommendations of this report (see Chapter 2 for a background) were contained, with
greater detail, in the later Wildlife Link (Stubbs 1988) and NCC (1990) reports.
Guidelines for herpetofauna
4) Conservation Committee of the British Herpetological Society (1983). Herpetofauna
Translocations in Britain - A Policy
The BHS guidelines on introduction of herpetofauna were very different from the other UK
guidelines described above.
Review of information, policy and legislation on species translocations
Translocations of British natives
Translocation of British natives within Britain, but to sites outside the presumed
natural area (see NCC 1990) of the species are acceptable 'where this seems
Translocations of species not native to Britain
The reasons for introductions are:
for general interest as an experimental activity;
to enrich the British fauna, which is poor in herpetofauna;
for commercial purposes.
Types of introduction allowed.
Non-European species should not be introduced - they may compete with native
Some European species and races should not be introduced - they may hybridise with
native species or races.
Introductions of some European species and races are acceptable. Such species and
races coexist on mainland Europe with species native in Britain and may not occur in
Britain, either because they failed to colonise before the loss of the land bridge to
Europe, or because the climate is sub-optimal such that they cannot colonise Britain
rapidly by natural means.
Guidelines for birds
The RSPB simply opposes all artificial introductions of non-native bird species (G. Williams
pers. comm.).
International Guidelines
6) IUCN. (1987). The IUCN position statement on translocation of living organisms
The basis of these guidelines was that:
'the introduction of alien species ... has often been directly harmful to the native plants
and animals of many parts of the world and to the welfare of mankind.'
They advise that governments should implement the following approach to reduce the
damaging impact of introductions.
Review of information, policy and legislation on species translocations
Intentional introduction of non-native species
Such introductions:
should only be considered if clear benefits to humans or natural communities can be
should only be considered if no native species is considered suitable for the purpose
for which the introduction is being made;
should not be carried out into any natural habitat (a habitat not perceptibly altered by
humans), island, lake, sea, ocean or centre of endemism. Such areas should, where
possible, be surrounded by a buffer zone sufficiently large to prevent unaided spread
of alien species from nearby areas;
should not be carried out into any semi-natural habitat (one changed by human actions
but still resembles a natural habitat in its species diversity and complexity of species
interactions) unless there are exceptional reasons for doing so, and only when the
operation has been comprehensively investigated and carefully planned in advance;
can be made into habitat created by humans (artificial, arable, ley, forest plantations
or other predominantly monocultural systems) following assessment of effects on
surrounding semi-natural or natural habitats.
Planning an introduction: assessment
An assessment should be carried out to decide the desirability of the introduction. The
following should be considered:
the probability of an increase in numbers after introduction that damages the
the probability of invasion into other habitats;
how the introduction proceeds during all phases of the biological and climatic cycles
of the area of release;
the capacity of the alien species to affect native species by breeding with them;
the probability that interbreeding of the alien and a native species will produce a new
invasive polyploid species;
whether the alien is host to diseases or parasites that can spread to other species in the
area of release;
the probability of a negative effect on the continued existence or stability of native
populations through predation, competition, or other means.
The methods for control of the introduced species should be investigated and
subjected to risk analysis. No introduction should be made for which no acceptable
control is possible.
The environmental, aesthetic or economic benefits of the introduction should be
compared to the possible disadvantages.
Planning an introduction: experimental, controlled trial
A controlled experimental introduction should be made, under the following conditions.
The same stock should be used as that to be introduced extensively.
Review of information, policy and legislation on species translocations
The organisms should be free of diseases or parasites that could spread to other
The performance of the introduced species should be measured using the parameters
of the assessment phase and the suitability of the introduction be reassessed in the
light of the results.
Planning an introduction: the extensive introduction
The extensive introduction should be closely monitored and arrangements made to
restrict, control, or eradicate the species if necessary.
Accidental introductions
Accidental introductions should be discouraged where possible, with emphasis on the
On island and other isolated habitats, special care must be taken to avoid accidental
introduction of seeds of alien plants on clothing or of alien animals associated with
Measures should be taken to discourage the escape of farmed and captive bred alien
animals, which could breed with native species.
Measures should be taken to control the contamination of imported agricultural seed
with seeds of weeds and invasive plants.
Where an accidentally introduced species establishes successfully, the economic and
environmental effects should be investigated. If the effects are negative, measures
should be taken to restrict its spread.
Where alien species are already present
In general, introductions of no apparent benefit to humans and with negative
environmental effects, should be eradicated.
Priority areas for eradication of such species are:
islands with a high percentage of endemics;
centres of endemism;
areas with high species diversity or other ecological diversity;
areas in which a threatened endemic is negatively affected by the alien.
Special attention should be paid to feral animals, which can be the most aggressive
and damaging of alien species. Eradication should be considered. Where the feral
population(s) has value in its own right, but is damaging native flora and fauna, the
latter should take precedence. In this case, removal to captivity of the feral species
should be considered.
Biological control
As this may involve introduction of alien species, the same care should be taken as for
other intentional introductions.
Review of information, policy and legislation on species translocations
Guidelines for aquatic species
7) International Council for the Exploration of the Sea (1995). ICES code of practice on the
introductions and transfers of marine organisms 1994
ICES, of which the UK is a member, developed these guidelines in response to the
translocation of fish, molluscs, crustaceans and plants for marine aquaculture. This involves
both introductions and translocations within a species native range (called by ICES
'transferred species'). The '1994 Code' was drawn up to update and expand previous ICES
guidelines (e.g. ICES 1988). They perceived three problems.
Accidental coincident translocation of other species.
Harmful ecological effects of the translocated species.
Genetic impacts of translocated species.
The ICES will consider the advisability of a new translocation (i.e. not part of an ongoing
practice) based on a prospectus containing the following information.
The purposes of the translocation.
The translocation procedure.
The ecology of the species.
The area of origin.
The proposed area of release.
The ecological, genetic and disease impacts of the species in its native range and
The potential ecological, genetic and disease impacts of the species in the proposed
area of release, to include:
potential habitat breadth;
diet and potential changes in diet;
potential to hybridise with native species;
other genetic impacts;
effects of associated species, including pathogens;
potential to spread outside the release site and effects of this.
An assessment of risks and benefits, possibly including quantitative risk assessment.
Following approval the translocation procedure should involve the following.
Establishment of a stock for artificial propagation (i.e. a brood stock) in quarantine.
Sterilisation of effluents from the quarantine premises.
Evaluation of the health status of the stock - the translocation should proceed only
with a healthy stock.
A limited release into open waters of the first generation progeny to assess ecological
interactions with native species.
Continued study and monitoring of the outcome.
Submission of progress reports to ICES.
Review of information, policy and legislation on species translocations
Procedures for ongoing translocations that are part of current commercial practice.
Periodic inspection of material prior to translocation - if any pathogens or pests are
discovered the translocation must be discontinued.
Quarantining, inspection and control, where possible and appropriate.
Consideration and/or monitoring of genetic impacts on native species.
8) International Council for the Exploration of the Sea (1988): Codes of practice and manual
of procedures for consideration of introductions and transfers of marine and freshwater
This contains the most recent code of practice of the European Inland Fisheries Advisory
Committee (EIFAC), which updates that in EIFAC (1983). EIFAC advises on introductions
and 'transfers' (translocations within a species native range) for aquaculture within European
inland waters. The code of practice is virtually identical to the then code of practice of ICES
(similar to ICES 1995, but with no consideration of genetic problems), but with EIFAC as the
advisory body and recommending in addition,
that after translocation, every effort should be made to contain the species within the
water bodies or water courses into which it was released.
9) International Maritime Organisation (1993). Guidelines for preventing the introduction of
unwanted aquatic organisms and pathogens from ships' ballast water and sediment discharges
Resolution A.774(18) of the IMO is based on concerns about:
the translocation, by the discharge of ballast water from ships, of unwanted aquatic
plants, animals, disease bacteria and viruses ('harmful aquatic organisms').
The guidelines were drafted to advise governments of the appropriate measures to take
against such discharges, but these measures have not yet been implemented by the UK
A number of procedures to prevent translocation are discussed in detail, which involve the
following objectives.
Ensuring, if possible, that only clean ballast water and a minimal amount of sediment
is taken on to a ship.
Contaminated ballast water is not released (by non-release, discharge in accepted
areas or discharge into disposal or treatment facilities).
Ballast water is sampled before discharge into sensitive areas.
Education of ships crews and other relevant persons.
The document also discusses the implementation of these procedures.
Review of information, policy and legislation on species translocations
10) North-east Atlantic Commission (1995). Introductions and transfers including the
amendments proposed by the European Union
Although generally aimed at salmon Salmo salar restocking and translocation, these
guidelines also recommended that
no non-indigenous fish should be introduced into a river containing Atlantic salmon
without an evaluation which indicates no risks of adverse effects on the salmon.
ICES and EIFAC codes of practice should be followed if the introduction proceeds.
Summary of legislation
This section summarises UK and EC legislation and international conventions which are
relevant to controlling introductions of non-native species
3.4.1 Import, keeping, release and control of alien species - international conventions and
EC legislation
A requirement to prevent introductions of, or to control established or feral populations of,
non-native species is expressed in a number of international conventions and EC legislation.
EC Directives
Of EC legislation, the Directive 79/409 on the Conservation of Wild Birds requires member
states to see that any introduction of non-native bird species does not prejudice the local flora
and fauna. Article 22b of the EC Habitats Directive requires measures to regulate the
deliberate introduction of non-native species.
The EC Fish Health Directive 91/67 is described below under the Fish Health Regulations
EC Directive 77/93 (amended by 91/683 in response to the removal of frontiers in the EU)
concerns the passage of organisms harmful to plants and plant products and calls for bans on
introductions of certain organisms (listed in Annex A). This is primarily concerned with agroeconomic implications of pest introductions.
International Conventions
The Bern Convention requires that the introduction of non-native species be controlled.
Recommendation R(84)14 of the Council of Europe Concerning the Introduction of NonNative Species was based on the Bern Convention. It calls on member states to prohibit the
introduction of non-native species into the natural environment, with possible exceptions only
if an expert study of the consequences has been carried out. Accidental introductions should
be prevented as far as possible.
Review of information, policy and legislation on species translocations
The Convention on Biological Diversity requires that the introduction of non-native species
which threaten ecosystems, habitats or species should be prevented or that such species
should be controlled or eradicated. The Bonn Convention on the Conservation of Migratory
Species of Wild Animals encourages similar measures against non-native species which
threaten endangered migratory species. The parties to the Bonn Convention are currently
negotiating an Agreement on African/Eurasian waterfowl (see de Klemm 1995, Holmes &
Simons 1995) which contains a provision which would require the parties to prohibit the
deliberate introduction of exotic species, to take steps to prevent their accidental introduction
and to prevent species already introduced from endangering native species.
The 1982 United Nations Convention on the Law of the Sea, enforced in 1994, requires that
the member states take all measures necessary to prevent and control the intentional or
accidental introduction of alien species (and 'new species', i.e. GMOs) which could cause
harm to the marine environment.
3.4.2 Import, keeping, release and control of alien species - UK legislation
A number of pieces of UK legislation directly concern the keeping, release and control of
alien species, which reflects many of the concerns of EC legislation and international
agreements and provides a much more comprehensive legal framework than exists for
regulating projects involving native species (see Chapter 2). However, contrary to many of
these international statements, under UK legislation introductions are regulated rather than
Wildlife and Countryside Act 1981 and Wildlife (Northern Ireland) Order 1985
Section 14 of the Wildlife and Countryside Act and Article 15 of the Wildlife (Northern
Ireland) Order make it an offence to release or to allow to escape into the wild any animal 'of
a kind' which is not normally resident in or is not a regular visitor to Great Britain or
Northern Ireland in a wild state (see also Holmes & Simon 1995). In the Wildlife and
Countryside Act Schedule 9 Part I lists a number of feral non-native species of animal which
have become widely established in Great Britain (i.e. which are normally resident in Great
Britain), and which, again, it is an offence to release or to allow to escape. This is to prevent
increased numbers of these species in the wild. These are 11 mammals, 13 birds, three
reptiles, eight amphibians, six fish and four invertebrates. Another listed bird, the Canada
goose Branta canadensis occurs naturally as well as through introductions. Three native
species, the capercaillie Tetrao urogallus, the white-tailed eagle Haliaeetus albicilla and the
barn owl Tyto alba, are included in Schedule 9 for different reasons (see Chapter 2).
Schedule 9 Part II of the Wildlife and Countryside Act also lists two non-native vascular
plant species and 11 non-native marine algae (10 named species and all non-native laver
seaweed species, Porphyra spp) which it is an offence to plant or cause to grow in the wild.
These are species which have established in the wild in Great Britain or which may become
established in the future. Schedule 9 Part II of the Wildlife (Northern Ireland) Order is a
similar list of plant species whose introduction to the wild is prohibited.
Review of information, policy and legislation on species translocations
As for many aspects of the Wildlife and Countryside Act and Wildlife (Northern Ireland)
Order, licences can be issued to allow release of any non-native animal or species listed on
Schedule 9. The Department of the Environment issues such licences (except for fish and
shellfish) in consultation with the statutory conservation agencies. In England all
introductions (except fish and shellfish) are dealt with by the Toxic Substances Division of
DOE (although the European Wildlife Division of DOE has an administrative role with bird
introductions). Fish and shellfish (Crustacea and Mollusca) releases are licensed separately
by MAFF, SOAFD or WOAD, again in consultation with the statutory conservation agencies.
The general policy of the statutory conservation agencies is that unrestricted release of alien
species should definitely not be allowed (P. Clement, A. Kerr pers. comm.). Conditional
releases, such as of biological control agents into greenhouses (especially of species unable to
survive outside greenhouses) or of aquatic species into contained conditions (e.g. isolated
ponds, as long as the species cannot survive out of water) are thought of as less of a problem.
Licensing of animal (except fish and shellfish) and plant releases under the Wildlife and
Countryside Act
The Advisory Committee on Releases to the Environment (ACRE) assesses applications to
the DOE for the release of alien species (ACRE also advises on alien races and GMOs and
more detail on ACRE is given in Chapter 4). ACRE appraises the risks associated with the
release based on a simple risk assessment of the likelihood of certain harmful incidents
occurring as a result of the release (negligible, low, moderate or high) and the potential
amount of harm occurring as a consequence of the incident (negligible, low, medium or
severe). Such assessment is based on information of the biology of the species involved and
the circumstances of the release. Assessment will examine such factors as: the persistence of
the species in the wild, competitive advantages over native species, effects on non-target prey
or hosts, potential to invade other communities, ability to show rapid population increase, etc.
This is a similar procedure to that used by ACRE for assessing risks of GMO releases. The
application is also assessed by the relevant statutory conservation agency.
A new guidance note for applicants under Section 16 of the Wildlife and Countryside Act is
being prepared (W. Parish pers. comm.). This will give a background on Section 14 of the
Act and provide guidance on the information needed on applications. The information
requirements are very detailed and will cover the following:
the name and a description of the species and, where relevant, strain, cultivar, etc.;
the site of origin;
the site and conditions of pre-release rearing facilities (including disease outbreaks);
the natural distribution and habitat of the species or race;
the lifecycle, ecology and biotic interactions of the species or race;
information relevant to assessing environmental impacts - the ecology, physiology
and pathology of the species or race;
the purpose of the release;
the conditions of the release - numbers and lifecycle stage to be released and
frequency and duration of releases;
the position and attributes of the release site(s), including conservation status;
post-release monitoring methods;
methods to minimise or avoid spread outside the release site;
Review of information, policy and legislation on species translocations
Standard licences for release are issued with conditions and allow the release of a single
species or race into one or more named sites only for a specific, named purpose. Other usual
conditions are: freedom from disease; the Secretary of State be notified of spread outside of
the release area; and only the donor sites named in the application to be used. The licences
are always for limited periods of time (usually several months) after which releases cannot
continue without a further licence and, unless explicitly allowed in the licence (e.g. when a
release is to establish a self-sustaining population), the released organism(s) must be
Most applications are for the release of invertebrate biological control agents (insects and
mites). These present a more complicated chain of events: a company supplies the control
agent to a number of users who release the agent (mostly into glasshouses). DOE consider
glasshouses to be 'the wild' as organisms may be able to escape from them, although the
potential for the species to survive outside the glasshouse environment enters into the risk
assessment (W. Parish pers. comm.). It would be very difficult to assess each release so DOE
issue two types of licence: the Suppliers licence and the Growers licence. The first type is
issued to the company supplying the control agent and they must make a standard application
for release of a Section 14 animal. The risk assessment takes account of the risks involved in
this form of release. The licence usually lifts the Section 14 prohibitions on the supplier for
research and development and for supply of the agent to customers, but imposes conditions
on the supplier. Only disease-free specimens of the named species, taken from the specified
sources may be supplied. The Secretary of State must be notified of establishment of the
agent outside the research and development premises. The supplier may supply to a customer
only the number of individuals required for a specified rate of application and for a specified
maximum area of application. Again the licence is only valid for a limited period of time and
must be subsequently renewed.
The supplier must also provide the customer with a copy of a Growers licence, which
imposes restrictions on the user of the control agent. Again, the licence applies to only the
named species, supplied by the named company and only for the specified use. The
application rate of the agent must not exceed that recommended by the supplier and any
escapes must be reported to the Secretary of State. The licence is valid for a limited period.
The Scottish and Welsh Offices and DOE(NI) follow very similar procedures (I. Holt pers.
comm., L. Howe pers. comm., R. Weyl pers. comm.)
Licensing of fish and shellfish releases under the Wildlife and Countryside Act 1981
These licences are issued by MAFF, SOAFD or WOAD. A general licence has been granted
for any release of rainbow trout Onchorhynchus mykiss, Pacific oysters Crassostrea gigas
and Portuguese oysters Crassostrea angulata but all other alien species require a specific
licence for single releases of single species at a single site and for a specific time period,
usually two months. To allow escapes or secondary translocations is an offence. Licences are
usually issued subject to conditions which attempt to ensure that the species cannot escape
from the release site (D. Linskey pers. comm.). The application form for such releases asks
for detailed information on:
Review of information, policy and legislation on species translocations
the species to be introduced; the life stage to be released (eggs or fish/shellfish);
the size and number of the fish or shellfish to be introduced;
the place of origin of the fish or shellfish;
the location and a detailed description of the receptor site;
whether the site is legally protected in any way (e.g. an SSSI);
the risks of escape;
the precautions to be taken against escape;
the purpose of the introduction;
the date of the introduction.
The statutory conservation agencies advise MAFF on the implications of any introduction on
the native flora and fauna and their habitats. A major concern is whether a proposed
introduction will be in or near an SSSI (M. Gibson pers. comm.)
Salmon and Freshwater Fisheries Act 1975
Under Section 30 of this Act the appropriate region of the NRA must license any release of
freshwater fish or shellfish (Mollusca and Crustacea) or their eggs into inland waters in
England or Wales. This Act and the licensing procedure, including NRA and EN guidelines
are described in detail in Chapter 2. For non-native species, a licence under the Wildlife and
Countryside Act 1981 is also required.
Fish Health Regulations 1992, Fish Health (Amendment) Regulations 1993-1994 and Fish
Health Regulations (Northern Ireland) 1993
These Regulations, resulting from the EC Fish Health Directive 91/67, prohibit the import of
live or dead fish or shellfish (Mollusca or Crustacea), their eggs or gametes, from zones
within the EC not approved as free of certain diseases. Imports must be licensed and licensing
conditions by MAFF, WOAD, SOAFD (1994) lay down rules aimed at preventing the
transfer of diseases of fish, molluscs and crustaceans in aquaculture.
The Shellfish and Specified Fish (Third Country Imports) Order 1992
This has much the same function as the Fish Health Regulations, but for imports from outside
the EC.
Pesticide Regulations 1986
The Wildlife and Countryside Act 1981 and Wildlife (Northern Ireland) Order 1985 cover the
release of alien animals and plants. The Pesticide Regulations (under the Food and
Environment Protection Act 1985) cover releases of alien fungi, viruses, bacteria, protozoa
and other microorganisms as biological control agents. MAFF and its equivalents license the
releases of such organisms. Recently MAFF have arranged that applications for such releases
will be seen by ACRE who will advise MAFF in the same way as they advise on releases of
alien animals and plants (see above).
Destructive Imported Animals Act 1932
Review of information, policy and legislation on species translocations
This Act (amended by the Destructive Imported Animals Act 1932 (Amendment)
Regulations 1992) imposes further restrictions on the import (rather than the release) and
keeping (sometimes imposing precautions for caging) of certain mammals; e.g. musk rats
Ondatra zibithica, coypus Myocastor coypus, grey squirrels Sciurus carolinensis, minks
Mustela vision, Arctic foxes Alopex lagopus and 'non-indigenous rabbits' - i.e. rabbits other
than Oryctolagus cuniculus. MAFF may license imports for research or exhibition.
Zoo Licensing Act 1981 and Dangerous Wild Animals Act 1976
These both require precautions against the escape of captive non-native species considered to
be dangerous to humans.
Animal Health Act 1981
Under this Act MAFF, or its equivalents, can make orders to prevent the introduction of
disease through the import of animals, carcases, eggs or any other animate or inanimate thing
by which disease can be transmitted. Although aimed at domesticated animals, this legislation
could be used for controlling imports of wild animals.
The Import of Live Fish (Scotland) Act 1978 and Import of Live Fish Act 1980
The Import of Live Fish (Scotland) Act (Scotland) and the Import of Live Fish Act (England
and Wales) provide more specific regulations concerning fish imports. The Secretary of State
can make orders prohibiting, or requiring licences for the import, keeping or release of live
fish or fish eggs of alien species which might harm the habitat of, compete with or prey on
any freshwater fish, shellfish or salmon Salmo salar. The statutory conservation agencies are
consulted before such orders are made.
Fisheries Act (Northern Ireland) 1966
This empowers the Department of Agriculture of Northern Ireland to introduce legislation to
prohibit the introduction, unless under permit, into certain waters of fish species which would
be detrimental to the fishery.
Fish Health Regulations 1992 and Fish Health Regulations (Northern Ireland) 1993
These Regulations (the former applying to Great Britain) seek to prevent the introduction of
certain disease species to the UK and prohibit the import of live or dead fish from zones not
approved as free of these diseases. These apply to native and non-native fish species (see
section 2.4).
Plant Health Act 1967
This Act and the Orders made under it (including the Plant Health (Great Britain) Order 1993
and the Plant Health (Forestry) (Great Britain) Order 1993) were designed to control pests
and diseases of agricultural, horticultural and forestry plants, but the legislation is broad
enough to cover wild plants as well. MAFF in England, the Secretaries of State in Scotland
and Wales and, for matters relating to forestry throughout Great Britain, the Forestry
Review of information, policy and legislation on species translocations
Commission are all given powers to prevent or control plant pests. Pests are defined as
harmful insects, bacteria, fungi, plants, animals and all other agents causative of transmissible
disease. These powers include the removal, treatment or destruction of any infected plants or
seeds, prohibition of keeping live individuals of the pest, prohibition of import of pests or
infected items and powers of entry and inspection and to take direct action if official
requirements are not fulfilled.
Releases into protected areas
Conservation areas designated under UK or EC legislation or international Conventions
(Natural Parks and Countryside Act 1949, Wildlife and Countryside Act, Natural Heritage
(Scotland) Act 1991, EC Directive on the Conservation of Wild Birds, EC Habitats Directive,
Convention on Wetlands of International Importance; see section 2.4 for a fuller discussion)
should have extra protection against releases of non-native species, either directly onto the
site or onto nearby sites from which the species can colonise.
Fish and shellfish stocking for aquaculture
In recent years the international transport of live fish, or eggs, for the purposes of aquaculture
or fisheries stocking has increased. In the past, smaller numbers, of fewer species, were
introduced for recreational fishing only, and the natural stocks of fish and shellfish were
exploited to provide food. However, concerns about this exploitation of natural resources,
and an opportunity to profit, have meant that although natural stocks of fish are still being
exploited, large-scale, and widespread, farming of fish and shellfish for food production now
Not all areas contain species suitable for large-scale farming, and so exotic species (or UK
natives which are locally non-native, see section 2.6.1) have been introduced in many cases
for this purpose. In general, fish used in aquaculture have been chosen for certain
characteristics, such as the ability to mature rapidly, high reproductive output, phenotypic
plasticity and wide habitat preferences and feeding habits. Unfortunately, such characteristics
mean that these species are often highly invasive and able to establish.
A high proportion of the freshwater fish species in the UK are non-native. Of the 50 fish
species forming wild populations in England, twelve species are aliens (Grice 1994). Most
requests for section 14 licences in the UK are for grass carp Ctenopharygdon idella, Wels
catfish Silurus glanis and zander Stiztedion lucioperca releases (Gibson 1995). Grass carp
releases are mostly for control of aquatic plants, but are also used to increase the diversity of
fisheries. Wels catfish are also introduced to increase the interest in recreational fisheries (see
zander in section 3.7). Britain has also received shipments of oysters Crassostrea gigas and
crayfish Astacus leptodactylus and Pacifastacus lenuisculus for aquacultural purposes.
Review of information, policy and legislation on species translocations
It is estimated that more than 30% of introduced fish species in inland waters in Europe have
come from aquaculture (Holcik 1991). Carp Cyprinus carpio, Tilapia spp. and rainbow trout
Onchorhynchus mykiss have been the most frequently introduced alien species in Europe.
Impacts of introductions
Individual introduced species can have a variety of impacts. For example, effects of salmonid
introductions include competition, predation on native salmonids and other fish and
environmental alteration by digging of redds in stream bottom substrates during spawning
(Krueger & May 1991). Other, more specific examples are as follows. Ecological impacts
Strong competitive interactions are thought to be rare in undisturbed fish communities. Niche
shifts and character displacement aid in avoidance of competition between native fish
species. However, when an exotic species, that has not co-evolved with the native species, is
introduced into a native ecosystem, the mechanisms for the avoidance of competition
between the native species and the exotic have not evolved, and competition may result in the
loss of native species (Crivelli 1995).
Prior to the 1970s the British Isles had only a single species of freshwater crayfish,
Austropotamobius pallipes - the white-clawed or Atlantic stream crayfish (Holdich & Reeve
1991). It appears to have had a widespread distribution with extensive large populations.
Until recently the British Isles has not had a recognised crayfish industry, but since the 1970s
large shipments of the long-clawed or Turkish crayfish Astacus leptodactylus and the North
American signal crayfish Pacifastacus leniusculus have been imported into Britain for
culinary and aquacultural purposes (Richards 1983, Lowery & Holdich 1988, Palmer 1994).
Both these species have been introduced, or have escaped, into the wild where they are now
expanding their ranges, often into waters previously occupied by the native species (Holdich
& Reeve 1987, 1989, Palmer 1994). In areas where invasive alien species of crayfish have
been introduced, the native species are often eliminated once the crayfish carrying capacity of
the water body is reached (Holdich 1988, Hobbs et al. 1989).
Review of information, policy and legislation on species translocations
Predation on native species by introduced fish is common. Predation may result in
displacement, or extinction, of native fish species (Crivelli 1995). Fish communities lacking
in native predators are the most vulnerable to predation by exotics, since they have generally
not evolved the necessary avoidance mechanisms (Townsend & Crowl 1991).
Attraction of predators
High densities of farmed fish and food attract predators and scavengers which may in turn
displace local species. For example, tern Sterna spp. breeding colonies in Scotland have been
displaced from islands in sea lochs due to influxes of gulls attracted by the fish (Beveridge et
al. 1994). A study conducted in Argyll, Scotland (Carss 1993) found that shags
Phalacrocorax aristotelis attracted to fish farms attacked fish through the netting, taking only
the smallest stock. An added impact was that most fish eaten near the fish farms were wild
fish that congregated around the fish cages. Thus, wild stocks may be affected by predators
attracted to the fish farm. This is not a particular problem with non-native fish; high stocking
densities of native species in fisheries may also attract predators (see section 2.6).
Habitat alteration
This can occur in a variety of ways. Certain fish species may cause an increase in turbidity
which can affect other fish species and may change the use of the waterbody by fish-eating
birds. For example, bottom-feeding fish, e.g. Common carp Cyprinus carpio and goldfish
Carassius auratus, increase water turbidity by churning of sediment while feeding (e.g.
Richardson & Whoriskey 1992). By predation on zooplankton, grazing pressure on
phytoplankton is reduced, and this results in blooms.
Habitat alteration may also occur through intense grazing of aquatic plants, making the
habitat less suitable for the resident native species. Grass carp Ctenopharyngodon idella and
crayfish Procambarus clarkii can cause extensive damage to macrophytes, completely
removing submerged vegetation and redbuds in some places (Stott 1974, Crivelli 1995).
Complex effects: the example of crayfish
Ecological effects of introduced species may be more complicated, as in two cases involving
introduced crayfish. Populations of the only native crayfish in Sweden, Astacus astacus, have
been significantly reduced due to a disease known as crayfish plague (Soderback 1995). The
North American crayfish Pacifastacus leniusculus was introduced into Sweden and has
spread. Where these two species occur together the introduced species is replacing the native
species. Pacifastacus has a higher rate of population increase, and appears to be dominant
over Astacus in interference interactions, suggesting competitive exclusion. Recruitment of
young to Astacus populations has been much reduced, probably by interspecific competition
and predation, resulting in increased mortality of young Astacus, and by reproductive
interference suppressing the less common species. The two species do not segregate in habitat
use, implying that there is no refuge for the native species.
Review of information, policy and legislation on species translocations
In northern Wisconsin lakes, the introduced crayfish Orconectes rusticus is replacing O.
propinquus, a previous invader, and O. virilis, a native crayfish (Garvey et al. 1994). Fish
predation and crayfish-crayfish competition interact to influence crayfish replacements.
Largemouth bass Micropterus salmoides predation modifies the outcome of interference
competition among the three crayfishes. Competitive interactions among the crayfish
influence susceptibility to fish predation. The less aggressive native species is excluded from
shelters by the invaders, and interaction with congeners and attacks by bass both increased
behaviour likely to result in predation (increased activity and swimming). In areas of
sympatry where predators are selective and prey species compete, predation and competition
interact to determine community structure. Impacts relating to the spread of disease
The impacts of the release of microorganisms or parasites from aquaculture operations has
been poorly studied. There is much concern that the escape of pathogenic organisms from
fish farms may contribute to the spread of disease from farmed to wild stock. When disease
organisms are introduced unintentionally with non-native species, the non-native species is
often a vector for the disease, being relatively immune to adverse effects. However, the
native assemblage receiving the introduction, and thus the pathogen, may have no resistance
to the introduced disease.
When intensive fish husbandry began, infectious diseases were responsible for much fish
mortality (Bullock & Wolf 1986). Many diseases will have been introduced with live fish, for
example, Furunculosis, a systemic bacterial infection, which is responsible for thousands of
fish deaths each year, was probably introduced into the United States with introduced brown
trout Salmo trutta from Europe.
Of particular concern to Britain is the threat posed by crayfish plague to the only native
British species of crayfish Austropotomobius pallipes (Holdich & Reeve 1991, Palmer 1994).
Crayfish plague is caused by the pathogenic fungus Aphanomyces astaci. This fungus is
endemic in North American signal crayfish (the plague vector) which are relatively immune
to its effects, whereas all European, Asiatic or Australasian crayfish tested are susceptible
(Unestam 1975). Outbreak of the disease in a population may lead to total mortality in a short
time (Soderback 1995). Several countries, including Britain and Ireland, had escaped the
disease until recently. Genetic impacts
Farmed species are not as genetically variable as wild species due to the unnatural selection
of fish, and inbreeding. Hybridisation with wild fish may result in the introduction of
maladaptive genes to wild populations (Beveridge et al. 1994) and lead to introgression of
gene pools.
However, hybridisation and introgression between native and introduced fish faunas has been
difficult to detect because traditional methods of differentiation between taxonomic groups
rely upon morphological criteria. Individuals with intermediate morphologies have been
designated as hybrids. However, this assumption is not universally valid. Ferguson et al.
Review of information, policy and legislation on species translocations
(1988) reported that first generation hybrids between westslope cutthroat trout Salmo clarki
lewisi and Yellowstone cutthroat trout Salmo clarki bouvieri have meristic counts higher than
or as high as the parent with the higher count. Furthermore, introgressed fish with less than
10% genes from other taxa are morphologically indistinguishable from pure species (Leary et
al. 1984). It is therefore possible that many more hybrid fish exist than are currently
identified by their morphology.
3.5.3 Discussion
Stocking of fisheries with alien or native (see section 2.6.1) has been widespread in the UK.
Many fisheries have developed into highly managed systems, with very high stocking rates
and use of feeding and fertilisation to maintain high fish densities. Because of this history,
there are few 'natural' (i.e. largely unchanged and unmanaged by humans) fisheries remaining
in the UK. The traditional aim of fisheries management has been to increase productivity and
to provide recreational angling, with little regard for the conservation of native species or
communities. EN, CCW and the NRA are currently producing a joint publication with the
aim of changing this approach to fisheries, to produce more diverse fisheries (in terms of
biodiversity) which are more sustainable and valuable for conservation. The NRA is also
working on a revision of policy in order to achieve these conservation aims (M. Gibson pers.
Negative impacts from the introduction of non-native species for aquaculture include
competition, predation, habitat alteration and the spread of disease. The only possible way to
protect native species from novel diseases is by preventing entry of the disease into
assemblages of native species. As described in section 3.4, the EC Fish Health Directive
prohibits all live fish and shellfish imports without a MAFF licence to prevent the transfer of
diseases in aquaculture. The Fish Health Regulations 1992 go further in attempting to prevent
diseased stock from entering Britain by prohibiting the import of live or dead fish from zones
not approved as disease-free.
In order to preserve the genetic integrity of locally adapted populations, the challenge is to
characterise these units relative to geographical references so that informed decisions on
introductions can be made (Ferguson 1990). To preserve genetic integrity of native species,
non-native species that seem likely to hybridise should not be introduced into systems
containing native species at risk.
The decline of native British crayfish stocks caused by crayfish plague and the spread of alien
crayfish species has resulted largely from entrepreneurial interests (Holdich & Reeve 1991).
Fisheries authorities at the time did not question these subsequently harmful introductions,
although many conservationists had reservations (Marren 1986). Austropotamobius pallipes
has since been protected by the Wildlife and Countryside Act 1981. Action needs to be taken
to protect the native crayfish and its habitat, since the introduction of disease severely affects
the native species, and even disease-free non-native crayfish displace the native. A JNCC
report (Palmer 1994) on conservation of the native crayfish recommended that the NRA,
MAFF, SOAFD, WOAD and DANI should regulate the keeping of all alien crayfish species
using the Import of Live Fish Act 1980, the Import of Live Fish (Scotland) Act 1978 and the
Review of information, policy and legislation on species translocations
Fisheries Act 1966 (see section 3.4). This would involve the creation of certain 'no-go' areas
for the keeping of crayfish: the whole of Scotland, Northern Ireland, the NRA regions of
Northwest, Northumbria and Yorkshire, Severn Trent, and Welsh, and a number of
catchments in the NRA regions of Anglia, Southern, Wessex and Southwest. Other
recommendations are to revise containment procedures, control the use of alien crayfish as
live bait, and develop and implement measures for eradication of alien crayfish in particular
areas. Holdich & Reeve (1987, 1989) have suggested similar measures. MAFF and SOAFD
have been convinced of the necessity of creating 'no-go' areas and legislation will be laid
before Parliament early in 1996 to implement this.
It would appear that, given the obvious dangers of introducing exotic fish and shellfish into
native waters, even for aquaculture, any proposed introduction should be thoroughly assessed
prior to consent. A thorough knowledge of the species' biology and ecology, together with
case studies from other countries involving the same species, is essential for the decisionmaking process. The characteristics of successful aquaculture species lend these species to
ready establishment in the event of escape into the wild, and thus each proposal must be
judged separately. In the interests of preventing further damaging introductions, various
authors have agreed that pre-introduction assessments of proposed exotic fish should follow
the guidelines for biological control suggested by entomologists (Arthington 1991, Li &
Moyle 1981).
Li & Moyle (1981) suggested a modification of biological control screening programmes for
use in fisheries management.
No introductions should be made into the few aquatic systems left that show little
evidence of human disturbance.
Introductions should be considered mainly for systems that have been so altered by
human activity that it is necessary to create a new community.
Introductions should be considered mainly for bodies of water that are sufficiently
isolated that uncontrolled spread of the introduced species is unlikely.
Any system being considered for an introduction should be inventoried thoroughly.
From the inventory, a species list should be developed that organises the species into
functional groups by habitat and trophic position.
Estimates should be made from the information available for the functional groups
with which the proposed introduction is likely to interact.
Oligotrophic systems or nutrient-poor systems are not good sites for species
The following criteria are suggested for a species that is a candidate for an introduction.
The species should be part of a co-adapted trophic assemblage, members of the
assemblage already being present within the system.
It should have a narrow niche breadth.
It should have low vagility.
It should be free of contagious diseases and parasites exotic to the system.
Review of information, policy and legislation on species translocations
Careful examination of a proposed introduction in the light of such criteria should result in an
evaluation of its long-term impact on the aquatic community into which it may be inserted, an
evaluation of possible impact on communities to which it might spread, and a reconsideration
of improving conditions for native species already present, rather than adding another
Whilst Li & Moyle's (1981) suggestions appear sensible and do concur with the guidelines
and regulatory criteria we reviewed above (sections 3.3 and 3.4), some appear unrealistic for
aquaculture. That the introduced species should be part of a co-adapted trophic assemblage,
species of which are present at receptor site, would surely preclude most introductions,
except those into extremely modified waters. Also, the selection of species with narrow niche
breadths would make most suitable aquaculture candidates unsuitable.
A thorough investigation of each proposed introduction should at least minimise the chances
of negative environmental impacts occurring. An interesting case study from the literature
demonstrate the use of pre-introduction assessments, and the subsequent refusal, based on
biological and ecological criteria, of the application to introduce a non-native species. When
the introduction of channel catfish Ictalurus punctatus into New Zealand for aquaculture was
proposed in 1987, a thorough environmental impact assessment was carried out (Townsend &
Winterbourn 1991). The EIA detailed the fish's biology and ecology, and on the strength of
that document, supervised trials were considered necessary to evaluate impacts on flora,
fauna and the environment. It was accepted that channel catfish would escape from
aquaculture, and that breeding populations would become established. The fish can tolerate a
wide range of physical and chemical conditions, and is thus able to occupy a wide range of
freshwater environments. It is omnivorous, and feeds on the bottom, in the water column and
at the surface, and large individuals may be piscivorous. Thus, a wide range of species may
be vulnerable to predation. Although unable to accurately predict the outcome of the
introduction, the available evidence suggested that adverse affects were likely. The danger of
local or global extinction of any of New Zealand's native species was considered too great,
and as a result the application to introduce channel catfish was rejected.
3.6 Biological control
Biological control usually involves the regulation (but not extermination) of a pest species by
introduced natural enemies. The pest species is frequently, although not always, an
introduced alien species that has not only established itself, but has achieved pest status due
to uncontrolled population expansion. Pests are usually defined in economic terms, but
conservation criteria can be applied. Van Driesche (1994) considers an introduced species to
be a pest when the species attacks or threatens the continued existence of native species or
alters ecosystems such that the existence of whole biological communities is at risk.
Indigenous species generally have natural enemies that regulate the population at a lower
than the maximum possible density (carrying capacity), including predators, pathogens,
parasites or herbivores. A problem occurs when a species is moved from its native range to
Review of information, policy and legislation on species translocations
an area where these regulatory organisms are not present. In theory, biological control offers
the ideal
solution. The control agent will be host-specific and should not attack native species. One of
the characteristics of a biologically controlled system of populations is that prey and predator,
or host and parasite populations are reciprocally density-dependent (Caltagirone & Huffaker
1980). Also, for all practical purposes these systems are self-sustaining and permanent.
However, problems with biological control have been detected.
Most examples of biological control programmes are drawn from outside the United
Kingdom. Classical biological control has not been truly tested in Britain (myxomatosis was
not introduced to the UK in a considered biological control programme, see section 3.13),
although it was considered for bracken control (Fowler 1993). This is partly because Britain
has not been exposed to as many, or as damaging, alien pests as other parts of the world.
Most current introductions to the UK licensed under the Wildlife and Countryside Act 1981
are of glasshouse biological control agents (W. Parish pers. comm.).
Pest species subjected to biological control programmes throughout the world include:
aquatic plants, terrestrial plants, herbivorous arthropods, predatory and parasitic arthropods,
non-arthropod invertebrates and vertebrates.
Biological control agents include: parasites, predators, herbivores, and pathogens.
Impacts of introductions Ecological impacts - Risk to non-target organisms
Biological control agents, although chosen to be host-specific, may switch to feed on other
species and thus affect non-target organisms. Howarth (1991) listed factors affecting the
degree of risk to non-target organisms.
Permanency of the agent in the environment. The chances of a non-target organism
being affected negatively increases with the length of time the control agent is in the
environment. The more generations for which an agent persists, the greater is its
potential to spread, and the greater is the possibility of host and habitat shift
Host range. Polyphagous agents have the potential to affect non-target organisms.
Agents with narrow host ranges are less likely to affect other species.
Habitat range. Species with a greater habitat range can invade a greater number and
variety of communities.
Genetic adaptability. The generation time of invertebrates is much shorter than that of
higher organisms, and consequently they have a greater tendency for genetic change.
The shift from ecological specialisation to generalisation in some insects may have a
relatively simple genetic basis.
Behaviour of the control agent. Dispersal ability, host-searching and host-handling
abilities can enhance a control agent's chances of increasing habitat range and
attacking non-target organisms.
Review of information, policy and legislation on species translocations
Vulnerability of the target region. Most extinctions caused by biological control
agents have occurred on islands or in freshwater habitats. In part this may be due to
the greater use of biological control on islands, and to superior documentation of
extinctions in these habitats.
Hopper et al. (1993) discussed the possibility of host switching. Parasitoid species that attack
more than one host species may have subpopulations that are adapted to different hosts. All
the cases of intraspecific variation in host specificity found in the literature involve
endoparasitoids. Endoparasitoids must overcome specific host defences, and parasitoid and
host metabolisms are often intimately connected. Thus it is not surprising that switching host
species by endoparasitoids would involve some cost. Whether ectoparasitoids suffer a similar
handicap is unknown. Hopper et al. (1993) conclude that, given the lack of knowledge about
which parasitoids are likely to show strong host specificity, one should attempt to collect
from, and rear on, the target pest where possible. One should also collect from alternative
hosts in cases where use of alternative hosts is important, e.g. for survival during periods
when the target pest is not present.
The island of Guam, in the Marianas chain, has 20 species of butterfly. In the last 40 years,
several butterfly species have declined dramatically in numbers, leading to extinction in some
cases. Nafus (1993) suggested that biological control organisms may have been a
contributory factor to this decline - over 100 exotic species have been introduced for
biological control since 1911. Whilst most of the insect biocontrol agents were specialists, the
other generalist species with wide host ranges have affected non-target species. Non-target
butterflies have apparently been parasitized by some of the insect species released as control
Acacia longifolia, a shrub indigenous to south-eastern Australia, was introduced into South
Africa for dune binding during the early nineteenth century (Dennill et al. 1993). It has
subsequently invaded river systems and catchments in a larger area. The wasp Trichilogaster
acaciaelongifoliae, which galls the reproductive buds of A. longifolia, was introduced from
Australia into South Africa in 1982 and 1983 to reduce the rate of invasion. The wasp
populations have reduced the reproductive potential of A. longifolia by more than 90%
throughout the coastal region of the Cape Province. An important reason for using the wasp
against the weed was the alleged host specificity of gall-forming insects which would prevent
the wasp attacking those other Australian acacias grown in South Africa, namely A.
melanoxylon and A. mearnsii. However, the wasp has spread onto the other Acacia species. Ecological impacts - Enhancement of target species
In 1870, the Indian mongoose Herpestes auropunctatus was introduced to the islands of the
West Indies to control rats on sugar plantations (Pimental 1980). Unfortunately, two species
of rat were present - the black rat Rattus rattus and the brown rat R. norvegicus. The
mongoose was able to control the population of the ground-nesting brown rat, but not the
tree-nesting black rat. When the two rats occur together, the brown rat is the dominant
competitor. In controlling the brown rat population, the mongoose enabled the black rat to
proliferate. Ground-nesting lizard populations were also affected, leading to an increase in
another sugar-cane pest - the sugarcane beetle Eutheola rugiceps. Today, the mongoose
would have been found to be unsuitable before introduction through investigation of its
Review of information, policy and legislation on species translocations
ecology and also the island ecosystem. Indeed, Pimental (1980) stated that there had been no
recent reports of vertebrate control agents causing environmental problems. Ecological impacts - Habitat alteration
The grass carp Ctenopharygdon idella is used sometimes in the UK to control aquatic weeds
but can bring about excessive reduction of aquatic macrophytes (Stott 1974, Crivelli 1995).
This may affect native species that use areas of aquatic plants, for example, bream Abramis
brama require water plants to deposit their spawn, whilst other species may need underwater
cover as refuges. Impacts relating to the spread of disease
In the Indian mongoose project discussed above, the mongoose was found to be a vector for
rabies and leptospirosis.
It would appear from the literature that biological control introductions have not, in general,
had such negative impacts on the environment as other types of introduction (Crawley 1986,
1989, van Driesche 1994, Center et al. 1995). This can be attributed to the failure of many
control organisms to establish in the new habitat and/or their host specificity. Most of the
problems cited in the literature have arisen from the introduction of control organisms some
time ago.
There are risks associated with the introduction of any exotic species, but a thorough
screening process, and detailed ecological studies, involved in biological control programmes
should ensure that the potentially harmful organisms are screened out. In particular,
polyphagous or generalist biological control agents should not be introduced.
Screening of biological control agents
Before introduction of control agents, the risks of damage to non-target organisms must be
determined and certain information is necessary. CAB International Institute of Biological
Control have refined the pre-introduction survey by holding a series of international symposia
on the biological control of weeds. Their procedure (1986) now complies with the demands
of plant quarantine authorities of Australia, Canada, New Zealand and the United States.
The target species, and its origin, should be correctly identified.
All records of phytophagous/predatory and pathogenic organisms associated with the
target species, and closely related species, should be reviewed.
Information on the hosts of organisms closely related to the potential control agent
should be collected. If the potential agent belongs to a group of species that are
restricted to hosts that belong to a single systematic unit, then the potential control
agent may show desirable specialisation.
Review of information, policy and legislation on species translocations
Field surveys should follow to determine which of the organisms play a major role in
controlling the abundance and distribution of the target weed.
Finally, host specificity must be determined. This is achieved using a selection of species
which are potential hosts to the organism in question. Species to be considered are:
those related to the target species, and other recorded hosts of the candidate agent;
hosts of species closely related to the candidate agent;
morphologically and biochemically similar plants to the target species.
The intentional introduction of biological control agents is subject to the legislation covering
any intentional release of a non-native species. Provided that the pre-introduction screening
process is thorough, the risks associated with this type of introduction can be minimised
Wildfowl and game stocking
The introduction of species to Britain for sport is not new. Many of the species traditionally
associated with the British countryside are not native, but were introduced for this reason.
These include rabbits (also introduced for other reasons), and some species of pheasant, deer,
fish, and raptor.
Game fish
Zander or pike-perch Stizostedion lucioperca were originally limited to central and eastern
Europe, including Russia. Many attempts were made to introduce fish into Britain between
1860 and 1880 (Lever 1977), but it was really from the early 1960s that the distribution of
fish in England began to expand. Initial introductions were cautiously into ponds
unconnected to river systems due to the savage reputation of this fish (Maitland & Campbell
1992). Following a second introduction, redistribution of these stocks has occurred and the
range is still expanding in eastern England.
Rainbow trout Oncorhynchus mykiss is native to western North America, but has been widely
distributed throughout much of the world due to its fine sporting qualities and food value. In
1882, the first consignment of ova to Europe (Germany) all died (Lever 1977). In 1886,
25000 eggs were sent to Germany. The eggs hatched, and the resulting fish formed the basis
for much of the subsequent stock of European rainbow trout. The first shipment to England
occurred in 1884. Resulting fish were despatched to fish breeders in England, Scotland and
Ireland. Rainbow trout are notorious for escaping from waters into which they have been
introduced (Maitland & Campbell 1992), but of the thousands of British waters in which
rainbow trout have been stocked, only in a very few instances have self-sustaining
populations arisen. The main prey appear to be shrimps, water lice and water boatmen.
Review of information, policy and legislation on species translocations
Pike Esox lucius is indigenous to south-east England, but was introduced into Ireland
possibly in the fifteenth century (Maitland & Campbell 1992). It is now widespread in
Ireland, except in the north-west and south-east coastal areas. Its distribution in Britain and
Ireland, has been extended by coarse anglers.
The largemouth bass Micropterus salmoides is native to North America, but has been widely
introduced because of its fine sporting and eating qualities (Maitland & Campbell, 1992). In
the British Isles, despite a series of introductions, it has only established in the extreme south
of England at two sites.
Game birds
The Red-legged partridge Alectoris rufa is native to France. It was first introduced to the UK
about 300 years ago by gamekeepers to country estates. Lever (1977) reported that since
1830 there had been at least 40 different attempts to introduce it into the wild in 26 counties.
The releasing of chukar Alectoris chuka and hybrid partridges in Britain originated with
chukars imported into Italy and bred for their docile temperament. A small batch of eggs was
imported into Britain in 1966-7, and several thousand chukars were bred in captivity (Potts
1988). In 1970, several thousand chukars were released at North Farm on the Sussex Downs,
and an average of 2700 were then released per year until the early 1980s.
The pheasant Phasianus colchicus is a southern Palaearctic and north eastern Oriental
species. Highly prized for sporting purposes and decorative qualities, the pheasant is
extremely catholic in its choice of habitat, favouring particularly partly cultivated, partly
wooded country, with areas of thick undergrowth and dense plantations well supplied with
water. The earliest evidence of pheasants in Britain was around 1177 in Essex. Pheasants
were certainly known in Britain, if only in captivity, before the time of the Norman Conquest.
Exactly when the pheasant became fully naturalised in the English countryside as a feral
breeding bird remains uncertain (Hill & Robertson 1988).
Falconry species
The American red-tailed hawk Buteo jamaicensis was introduced for use by falconers and has
escaped in the UK (Murray 1970).
Mammals introduced for hunting
Fallow deer Dama dama are native to southern Europe, and their original appearance in
Britain is often ascribed to the Romans or Bronze/Iron Age Phoenicians. The deer were
certainly well established in Roman times. Fallow deer were introduced for hunting and to
many deer parks.
The rabbit Oryctolagus cuniculus may have been introduced to Britain during Roman times,
but the present British population appears to be derived from stock introduced during the
twelfth and thirteenth centuries. The animals were conserved in enclosed warrens for meat,
fur and sporting reasons. The rabbit was never legally regarded as game (Sheail 1984), but
was treated as game by sportsmen. Sharing the same natural predators as hares, pheasants and
Review of information, policy and legislation on species translocations
partridges, the rabbit benefited from the increase in game keepers from the eighteenth century
onwards. Commercial warrens were abandoned in the late eighteenth and nineteenth
centuries, but the survival of the species was already ensured by the changes in land use and
management of the time. Impacts of rabbits have generally been viewed from the commercial
point of view, in terms of their impacts upon crops.
There are also associated groups of introduced species: those introduced as game cover, for
example, snowberry Gaultheria spp., and species of fish introduced as fish bait, e.g. the
release of live bait has been responsible for the establishment of populations of roach Rutilus
rutilus and dace Leuciscus leuciscus in Ireland (Maitland & Campbell, 1992).
Impacts of introductions Ecological impacts
Fallow deer Dama dama damage in coppice regrowth within woodlands by browsing was
reported by Kay (1993). Levels of damage were highest on sweet chestnut Castanea sativa,
ash Fraxinus excelsior and lime Tilia cordata.
The zander Stizostedion lucioperca is mainly piscivorous as an adult (Maitland & Campbell
1992), whilst young fish feed largely on invertebrates, mostly zooplankton. The impact of
this fish on British stocks of prey species may be substantial, but systems generally stabilise
at lower levels of native species' abundances as a balance is established. However, adult
zander prey on their own young, which is probably an important population regulatory factor.
The effects of introduced zander on native fish communities has been studied in central
European waters, but remain unpredictable (Hickley 1986), with the abundance and diversity
of prey species and the presence or absence of other predators as contributory factors to the
outcome of zander introductions. It appears that zander may have a greater impact when there
is a reduced diversity of prey species, which is unfortunate since all British waters containing
zander have low species diversity.
There is a strong negative association between the distributions of introduced brown trout
Salmo trutta and native Galaxias vulgaris in New Zealand (Townsend & Crowl 1991). In
most cases, G. vulgaris is only found above waterfalls that are large enough to inhibit trout
migration. Predation is the most likely mechanism for the observed distributions. The effect
of the predation is so pronounced because of New Zealand's paucity of piscivorous fishes native fish fauna have evolved no defences.
It has been suggested that population declines in a number of rare woodland butterflies in
Britain have been caused by increased pheasant numbers which may eat caterpillars or larvae.
Although pheasants may predate upon insects, in order to lead to nationwide declines in
certain species, high rates of predation would need to be maintained over large areas (Clarke
et al. 1990). However, high densities of pheasants within release pens could affect insects,
Review of information, policy and legislation on species translocations
e.g. where release pens contain violets Viola riviniana, high pheasant densities can lead to
loss of violets, which are an important host plant for many fritillaries.
Habitat alteration
The great impact of the rabbit on the landscape of the UK has been well documented (see
Sheail 1991), but it is usually viewed as beneficial for UK biodiversity (see also section
3.13). The development and continued persistence of many important species assemblages
(e.g. chalk downland) has been strongly influenced by rabbit grazing. It is also an important
prey item for many UK predators.
Domestic and wild-type European rabbits Oryctolagus cuniculus have been liberated on
islands around the world (Flux & Fullagar 1992). Where large populations persist, islands
may be denuded of vegetation and soil. European rabbits are also thought to be responsible
for inhibiting recruitment of perennial herbs and shrubs in La Campana National Park, Chile
(McDonald et al. 1988). Once again, effects are felt most strongly on islands. Sheail (1984)
relates how the arrival of a mother and young on Porto Santa in the Madeira Islands in 1418
led to a population explosion that consumed native vegetation and crops, forcing the human
inhabitants of the island to leave.
Positive impacts
In high-elevation Hawaiian shrubland of Haleakala National Park, the alien game birds ringnecked pheasant Phasianus colchicus and chukar Alectoris chukar are the dominant avifauna
(Cole et al. 1995). The impact on native invertebrate populations is minimal. These birds
occupy, at least partially, an ecological niche once held by now-extinct or rare birds, and they
do not appear to be significant competitors with the endangered nene goose Branta
sandvichensis. Their role in facilitating seed dispersal and germination of native plant species
is beneficial in restoring degraded ecosystems. Genetic impacts
Hybridisation between introduced game bird species is common in the UK. These effects,
described below, are important for the game bird industry, but any effects on biodiversity will
be indirect; through changes in stocking rate or management of the alien birds.
There are several introduced species of pheasant in Britain, e.g. Phasianus colchicus, a
southern Palaearctic and north-eastern Oriental species, which was introduced for food and
game purposes. This species hybridises readily with other species of introduced pheasant
(Lever, 1977), both sporting and ornamental.
Hybridisation between chukar and red-legged partridge has occurred freely in captivity, and
from 1971 onwards increasing numbers of hybrids were released onto the South Downs and
other localities within Britain. Early studies suggested that in the wild, amongst the three
types of partridge (red-legged, chukar, hybrid), like tended to pair with like. However, a
significant proportion of red-legged partridges pair with chukars or hybrids. The chukar and
hybrids have been released for shooting purposes, but the breeding success in the wild of the
chukar and hybrid partridges is poorer than that of the red-legged partridge, with the result
Review of information, policy and legislation on species translocations
that the populations of chukar and hybrid can not support the shooting rate required to justify
the cost of rearing and numbers fall. While stocking of chukar and hybrid continue, red-legs
are affected more severely by shooting at a rate beyond their sustainable yield. To release,
shoot and restock with chukar and hybrids, instead of red-legged partridges, is catastrophic
for the latter species. Consequently, red-legged partridge numbers are declining, and may
have been caused in part by a higher shooting rate associated with the releasing of chukar and
hybrids. The red-legged partridge is more expensive to rear due to a lower fecundity on
game-farms, but in the wild only half as many birds would be needed to produce the same
sustained return.
However, other alien birds may hybridise with native species. American red-tailed hawks
have been found to be nesting with native buzzards Buteo buteo (Murray 1970), and this may
lead to hybridisation.
Impacts associated with the introduction of wildfowl and game species include herbivory,
habitat alteration or damage, predation, competition and genetic impacts. These impacts
could only be avoided by segregation of alien species from the native species at risk.
Again, game species must be subjected to the same pre-introduction assessment as other
intentional releases. The negative impacts associated with non-native game species can only
be avoided by prohibition of introductions into vulnerable areas (defined in terms of either
the biotope itself, or the resident species).
Ecological impacts of game species have been amplified by anthropogenic disturbances.
Native habitat is often altered by management to benefit the alien game species. In addition,
native habitat has been fragmented, with the result that native species may be unable to find
refuges from non-native invaders. Moyle & Williamson (1990), by example of the native fish
fauna of California, report that modification of habitat by water diversion to benefit
introduced species of fish is the principle cause of decline of native fish species.
Native species currently experiencing genetic alteration due to hybridization with non-native
species need protection from further genetic erosion. In the case of species other than birds,
'safe' areas could be established, where non-native invader and native species do not come
into contact with one another. Further introductions into such areas could be prohibited.
The question of introduced species hybridising with introduced species is an interesting one.
The prime objective of conservationists is to protect native diversity. Species such as
pheasants, even though they have been present in Britain for many centuries, are non-native
and have hybridised with other introduced species of pheasant. Protection of genetic diversity
in such mobile species would not be simple, and determination of 'pristine' stocks of pheasant
species would be difficult, if at all possible.
Review of information, policy and legislation on species translocations
Amenity and ornamental planting, stocking or collections
There is a large category of introductions comprised of species introduced into Britain solely
for their decorative qualities. Species of animal and plant have been introduced for this
purpose for several centuries. Certainly, ornamental gardening began in the Medieval period
(Nelson 1994), with specimens coming from Europe and West Asia, then later Africa, the Far
East and the Americas as trade routes opened across the world. Also included in this section
are collections of exotic specimens of plant and animal.
Examples of introductions to the UK in this category are as follows.
Garden plants
There are many ornamental plant species established in the wild in the UK. For example,
Hadleigh Great Wood, now Belfairs Nature Reserve, in Essex, contains a total of 374 plant
species, 54 of which are non-native. The high proportion of non-native species is accounted
for by several factors, including accidental escapes from an adjoining nursery, establishment
of species dumped in the wood as garden rubbish, and some garden species which were
deliberately planted by locals to 'brighten up' the wood.
Rhododendron ponticum was first introduced into Britain in 1763 (Elton 1958), and was
widely planted for game cover and ornamental purposes. In its native mountains of northern
Turkey it grows in mixed deciduous forests. In the UK it invades a range of habitats, and is
now a major forest weed, forming a dense understorey.
Giant hogweed Heracleum mantegazzianum was sold widely through the 19th century as an
attractive garden plant and is now widespread throughout the UK. It occurs in a variety of
vegetation types, including riparian habitats, agricultural land, road and railway
embankments, urban habitats and waste land (Nelson 1991, 1994).
Buddleja davidii forms dense stands on a variety of open habitats, from derelict city plots to
coastal landslips. It was introduced from China as a garden plant in the late 19th century
(Nelson 1994).
Japanese knotweed Fallopia japonica was introduced from South East Asia as a garden plant
(and, to a lesser extent, as a fodder plant) in the late 19th and early 20th centuries. It has
increased its range from a very few records between 1900 and 1940 to become very common
and widespread by 1970. It is found throughout the UK but is most common in the south and
west. It is a primary coloniser of unmanaged derelict or disturbed land, and is particularly
effective at colonising banks of downstream reaches of urban rivers (Beerling et al. 1994,
Palmer, 1994).
Review of information, policy and legislation on species translocations
Hypericum is a genus of about 420 species from the temperate zone (Robson 1994) that has
been cultivated in British gardens since medieval times. H. calycinum, a member of the
ground flora of broad-leaved woodland of south Bulgaria was introduced into Britain in 1676.
It became widespread in gardens, escaping to establish in semi-natural habitats. H.
xylosteifolium, of the south-eastern Black Sea area, arrived in Lancashire and Yorkshire, and
forms dense thickets by suckering. More recently there has been an influx of shrubby
Hypericum species from China and the Himalayas, which have had little real effect on the
natural flora as yet.
Other well-known garden escapes are Impatiens glandulifera and Gunnera spp. A commonly
seen, but less disliked, garden alien is Fuchsia magellanica which is found in hedgerows in
the southwest of the UK.
Aquarium plants
Pondweeds, such as Elodea canadensis, are sold widely in the UK and are often released into
water bodies deliberately or unintentionally. Swamp stonecrop Crassula helmsii, which
originates from Australasia, has been sold as a oxygenating plant for ornamental pools since
the early part of this century. Since then it has become widespread in England and Wales and
is still spreading (Dawson & Warman 1987). It can swamp other vegetation in the ponds and
lakes that it colonises.
Ornamental fish
Goldfish Carassius auratus is native from eastern Europe to Asia and China. The first record
of its presence in Britain was in 1742. Increasingly large numbers were introduced, and they
are also kept as pets.. They are still being introduced, and have formed self-perpetuating
naturalised colonies.
The bitterling Rhodeus amarus is native to continental Europe. It has been imported for many
years by fish culturists for sale to coldwater aquarists. Lever (1977) reported that the current
status in the wild was uncertain.
Ornamental introductions of sturgeon Acipenser sturio have sometimes led to establishment
on a larger scale through accidental or illegal, intentional releases.
Cage birds
The ring-necked parakeet Psittacula krameri, a native of Africa, Mauritius, Seychelles,
Burma, India, Sri Lanka, Malaysia and south western China, was a popular bird with
aviculturists and was known to nest in the wild and bring young to maturity from 1855 in
Britain (Lever 1977). It is reported in areas to the south, south-west and south-east of
London. Feral populations may have originated from free-flying homing colonies, escapes
from pet shops or exotic bird farms, or individuals may have been deliberately introduced.
They cause damage to crops, commandeer nesting sites of native species and can transmit
Review of information, policy and legislation on species translocations
The budgerigar Melopsittacus undulatus originates from Australia. Those in the wild in the
UK have escaped or have been deliberately liberated. There are currently breeding
populations on the Scilly Isles.
Ornamental wildfowl
Mandarin duck Aix galericulata originates from eastern Asia, in China and Japan, and was
first introduced to Britain in the 18th century. In 1830 two pairs were purchased by London
Zoo, and they bred for the first time in England four years later. The first recorded escape
was shot in Berkshire in 1866. The 11th Duke of Bedford introduced some to Woburn Park at
the start of this century, where they flourished.
Canada geese Branta canadensis are native to North America. The earliest mention of this
bird in Britain refers to birds in the collection of King Charles II in St James Park (in
approximately 1665). Estates all over Britain contained Canada geese among their collections
of wildfowl.
The golden pheasant Chrysolophus pictus is native to mountainous districts of central China.
It was kept in captivity in Britain as early as the 18th century, but it was not until 1845 that
one appeared in the wild in Norfolk. Individuals have been introduced onto estates in East
Anglia this century.
Lady Amherst's pheasant Chrysolophus amherstiae is native to the mountains of southwestern China and Upper Burma. It was introduced first in 1828, but did not breed until
1871. It's main centre is in the East Midlands. Reeve's pheasant Syrmaticus reevesi, native to
the mountains of central and north China, is also found in the UK.
Ruddy duck Oxyura jamaicensis is native to North America. In 1948, the Wildfowl and
Wetlands Trust obtained three pairs from the United States, which bred at Slimbridge reserve
in Gloucestershire. The ruddy ducks now feral in England are descendants of escapees from
the reserve. Earliest reports of ducks in the wild date from 1954. There was at least one
deliberate release of three or four juvenile females from Slimbridge to Chew as potential
mates for drakes already there. The increase in population and range in England has been
rapid. Principal reasons are probably the availability of suitable habitats; the existence of a
vacant niche for a freshwater bottom-feeding duck; lack of shooting pressure, and mild
climatic conditions. The population of ruddy ducks in Britain numbers approximately 3500
and the population is still increasing.
Deer introduced to deer parks
Reeve's muntjac deer Muntiacus reevesi is native to south China and Taiwan. The 11th Duke
of Bedford introduced these to Woburn, Bedfordshire at around the turn of the century, with
the first release from Woburn being in 1901. Numbers remained low, until further deliberate
introductions in the 1930s and 1940s, and also in the second half of this century, meant that
the subsequent spread was from several main foci. Spread has been faster than it would have
been naturally (natural spread occurs at about 1km per year). They are now established in
many areas of central England (Chapman et al. 1994).
Review of information, policy and legislation on species translocations
Japanese sika deer Cervus nippon were first introduced in 1860, when a pair were presented
to the Zoological Society of London for a collection in Regent's Park. Also in 1860, a stag
and three hinds were acquired by Viscount Powerscourt in County Wicklow, Ireland. These
deer were the basis for the original stocking of a number of parks in England and Scotland. A
number escaped and formed feral herds.
The Chinese water deer Hydropotes inermis originates in China and Korea. The 11th Duke of
Bedford established a breeding herd in Woburn park. Escapes became naturalised in areas of
Bedfordshire, Buckinghamshire, Hampshire, Northamptonshire, Cambridgeshire, the Norfolk
Broads and Shropshire.
Grey squirrel Sciurus carolinensis is native to north-eastern United States and south-eastern
Canada. The earliest recorded introduction into Britain was in Cheshire in 1876. From 1880
until 1929 there were many releases in all parts of the country. In 1938, it was declared illegal
to import them into Britain or keep them in captivity. It is now common in many parts of the
The number of species which have escaped from collections is low. In most cases impacts
have not been great either, the species mainly remaining restricted to the locality of escape.
Effects of introductions Ecological impacts
Feral Canada and greylag Anser anser geese are thought to have exacerbated damage caused
to reedswamps in the Norfolk Broads caused by coypu by their grazing activities (Boorman
& Fuller 1981).
Muntjac have caused extensive damage by grazing woodland plantings, although these is less
evidence of damage to areas of conservation value (Chapman et al. 1994).
Introduced ornamental fishes prey on aquatic insects, and it is thought that aquatic insect
diversity is threatened by alien aquarium fishes (Polhemus 1993).
Competition and habitat alteration
Crassula helmsii is a fast growing plant, with good dispersal and colonisation abilities, that
produces a dense sward which can smother all but the largest and established members of the
aquatic flora (Dawson & Warman 1987).
Rhododendron ponticum is capable of invading a variety of habitats, including woodland,
grassland and heath, and is particularly a problem in Ireland, western Scotland, southern
England and Wales (Thomson et al. 1993). A survey of the Snowdonia National Park in
Review of information, policy and legislation on species translocations
1985-6 estimated that 28% of the park contained rhododendron, although only approximately
1.6% of the area of the park was severely affected. Broadly speaking, rhododendron appears
to favour a maritime climate with high rainfall and acid soils available for seed regeneration.
It is particularly vigorous in deciduous and mixed woodland, e.g. invasions into oak Quercus
petraea and holly Ilex aquifolium woodland in South West Ireland may be responsible for
inhibition of native woodland regeneration (Usher 1987). The dense shade cast by the
rhododendron, together with an impenetrable litter layer, appear to be adversely affecting the
native species. Any soil disturbance near to existing seed sources exacerbates invasion.
Heracleum mantegazzianum is a problem because of its phytotoxicity, prolific vegetative
growth (enabling it to shade out and replace native flora) and the large areas of bare ground
exposed to erosion during winter following dieback of the plant (Dodd et al. 1994). Several
other alien plants in the UK, such as Fallopia japonica (Beerling et al. 1994) and Impatiens
glandulifera (Beerling & Perrins 1993) form dense stands which tend to swamp other
On the Galapagos island of Floreana, the dense thickets formed by Lantana camara are
thought to be endangering the threatened darkrumped petrel Pterodroma phaeopygia. The
petrel nests in ground burrows, and the spread of the Lantana thickets towards the largest
petrel colony is causing concern as unchecked the Lantana thickets may prevent the petrels
gaining access to, and leaving, their nest burrows (Cruz et al. 1986).
The invasiveness of non-native Acer platanoides in an isolated New Jersey Fagus
grandifolia-Quercus spp.-Acer saccharum forest preserve was investigated by Webb &
Kaunzinger (1993). Acer platanoides was present by 1915, and is now the second most
abundant canopy species. It seems likely to increase in importance in the future, because its
abundant shade-tolerant seedlings and saplings comprise 59% of all small woody stems, over
twice the amount of Fagus grandifolia reproduction and five times the Acer saccharum
There is growing concern over the impact that expanding deer populations may have on the
conservation interest of woodlands (Pollard & Cooke 1994). Main impacts are directly upon
vegetation, but indirect effects may be felt by invertebrates. In Monks Wood National Nature
Reserve, Cambridgeshire, an increase in the number of muntjac deer has affected vegetation,
e.g. increased grazing on flowers of lady's smock Cardamine pratense on which the larvae of
the orange tip butterfly Anthocaris cardamines, feed. Some eggs and larvae are lost to
grazing each year, but there is little impact on abundance of flowers or of larvae in the wood.
In mature woodlands, honeysuckle Lonicera periclymenum grows up trunks, producing long
trailing stems which the caterpillars of the white admiral butterfly Ladoges camilla feed
upon. Muntjac browse upon the honeysuckle, removing the lower leaves, significantly
changing the pattern of egg laying. It is not clear whether the population has been affected by
browsing, but a continuing increase in abundance of deer could have a significant indirect
effect on the behaviour of the invertebrate (Pollard & Cooke 1994).
Review of information, policy and legislation on species translocations
Benign species
The European bitterling Rhodeus sericeus was introduced into natural waters of New York
State at the turn of the century (Schmidt & McGurk 1982). It has been a popular aquarium
fish in Europe since the late 1800s. The bitterling's small size and plant diet preclude the
species from being a serious predator on other fishes. It is unlikely that it would compete with
native species for food since no other species utilise the same food source (Scott & Crossman
1979). Its small population size would indicate that the bitterling could not possibly reduce
the standing crop of diatoms in the river. Impact upon the ecosystem appears minimal, and it
is unlikely that the species would cause appreciable negative impact on any habitat in the
north-eastern United States.
Positive impacts
In a study by Owen (1986), introduced ornamental plants in gardens were found to support
many plant-feeding insects. Many conservationists believe that species of native plant support
a richer assemblage of herbivorous insects than do alien plants. The idea seems to originate
from counts of the numbers of species of insects found on trees growing in Britain
(Southwood 1961). The reasoning is that native plants have been present for a long time and
so there has been an opportunity for an array of insects to evolve to exploit them. Owen
(1986) studied a mature and well-established garden, which consisted of a mosaic of lawns,
herbaceous borders, vegetable patches, trees and shrubs. Between 1972 and 1985, 375 species
of flowering plant were listed. Also during this period, a total of 358 species of Lepidoptera
belonging to 28 families were recorded. Results indicated that substantial numbers of
Lepidoptera fed on alien as well as native species of plants. Owen (1986) concluded by
stating that it could be argued that the lepidopteran fauna of a garden is composed largely of
generalist rather than specialist feeders, and that in 'natural' areas there are more specialists
restricted to native species of plants. Impacts relating to the spread of disease
The importation of exotic aquarium species has been responsible for the introduction of
exotic disease and parasites. Introduced Japanese goldfish Carassius auratus infected with
Aeromonas monicida were responsible for goldfish ulcer disease in cultured and wild
Australian goldfish and carp populations (Langdon 1990). Impacts of competition and disease
The grey squirrel Sciurus carolinensis is a well known example of an introduced species
which has had adverse impacts on wildlife the UK. The native red squirrel Sciurus vulgaris
has been replaced by its congener throughout much of its former range in the UK (Usher et
al. 1992) and the exotic species is continuing to expand its range in the north (Lurz & Garson
1992 Skelcher 1993).
The reasons for this replacement are not fully understood. Local epidemics caused by a
parapox virus were associated with the decline of red squirrels in the 1940s (Keymer 1983),
but there is no evidence that grey squirrels can carry or transmit the disease (Gurnell 1987).
Epidemics and local extinctions of red squirrels often preceded the arrival of grey squirrels
Review of information, policy and legislation on species translocations
(Reynolds 1985), and red and grey squirrels have been able to coexist in other areas for some
years (Gurnell 1987). Grey squirrels appear to have a feeding advantage in deciduous woods
(Gurnell 1983, 1989) which may be caused by a better ability to tolerate phytotoxins in
acorns (Kenward & Holm 1993). This may explain the replacement of the native species in
this habitat.
The remaining red squirrel populations in mainland UK are in woods with a high conifer
content (Lurz & Garson 1992) and it may be possible to create refuges in conifer woodland
(Kenward & Holm 1989, Gurnell & Pepper 1993). However, grey squirrels may still have a
feeding advantage in conifer woodland which contains, or is near to, large seeded
broadleaved trees (Kenward et al. 1992). In addition, the possibility that some kind of
interference competition might benefit grey squirrels cannot be excluded (Gurnell & Pepper
1993, Kenward et al. 1995).
The possibility that oak woodland might need to be reduced to conserve red squirrels raises
an interesting dilemma for conservation, given the considerable value of this species for other
wildlife (Kenward & Holm, 1993). The need to acquire better understanding of the
mechanisms of replacement are important for this reason and because grey squirrels are now
established in Italy (Currado et al, 1987). Knowledge gained through research in the UK
might be applied in Italy to help prevent the loss of the native squirrel throughout Europe. Genetic impacts
The Japanese sika deer was introduced into Scotland approximately 80 years ago (Abernathy
1994). It is expanding its range, and hybridization between sika deer and the native red deer
Cervus elaphus is occurring. The genetic integrity of the Scottish mainland red deer is being
threatened by the sika deer invasion.
Ruddy duck is currently hybridising with Western Europe's most important breeding
population of white-headed ducks Oxyura leucocephala in Spain. The white-headed ducks
are not as vigorous, or as numerous, as ruddy ducks (conservation in Spain has increased the
O. leucocephala population from 22 individuals in 1977 to 786 in 1992), and are threatened
with the genetic effects of breeding between populations which may lead to genetic
introgression and the effective extinction of the white-headed ducks. Concern over the spread
of ruddy ducks from Britain into Europe led to the convening of International Ruddy Duck
Workshops (Anon 1993, 1994e) to discuss the problem of interbreeding and the fertile
hybrids produced. It is planned to cull ruddy ducks to avoid loss of the white-headed duck.
There is a great deal of evidence for hybridisation among a number of goose species in
Britain. The most numerous hybrids are those of the two most numerous species, the Canada
and greylag geese (Delaney 1993). It would appear that escaped and introduced geese will
breed freely with many other goose species.
Some plants which have escaped from garden cultivation also have the ability to hybridise
with species in the wild. For example, cultivated Hypericum spp. hybridise with native
species (Robson 1994). This leads to the possibility of a loss of genetic identity of the native
species, possibly with changes to locally-adapted gene pools.
3.8.3 Discussion
Review of information, policy and legislation on species translocations
There are many more ornamental species of plant than of animal. However, both types are
capable of affecting native biota negatively, as demonstrated by the case studies. The animals
introduced are mainly herbivorous, so may compete with other animals or feed upon plants,
whilst several of the plant species appear to inhibit recruitment of native species, both
animals and plants, by their dense growth habits.
Both ornamental plants and animals may have the ability to hybridize with wild relatives, and
this may only be avoided by limiting introductions to areas without native 'compatible'
relatives. Potential problem species may also need to be limited in their non-native
distributions to preserve ecological, as well as genetic, integrity. Webb and Kaunzinger
(1993) have suggested that in order to maintain or restore examples of relatively natural
forest communities, restrictions on plantings and the removal of individuals of invasive,
problem species near nature reserves may be necessary.
This suggests that again that imports of aliens, even of garden plants, should be subject to
rigorous assessment to determine the potential impacts on the environment should the species
escape from the area of the introduction. Species likely to escape and be detrimental to native
biota should be restricted.
Pets and domestic animals
Many animals imported as pets and domestic animals find their way into the wild, either by
deliberate release or accidental escape. Many of these species have bred in the wild. In the
UK his category includes feral populations of cats Felis catus, dogs Canis familiaris, goldfish
Carassius auratus (see section 3.8), sheep Ovis aries, goats Capra hircus, cattle Bos taurus,
rabbits Oryctolagus cuniculus (see section 3.7) and terrapins, e.g. European pond terrapin
Emys orbicularis and American red-eared terrapin Trachaemys scripta elegans.
Effects of introductions Ecological impacts
Impacts of introduced domestic animals are felt most strongly on endemic island flora and
fauna. Overgrazing causes degradation of vegetation, and feral and domestic predator species
cause declines in native fauna.
Review of information, policy and legislation on species translocations
Introduced sheep have had adverse effects on the vegetation and avifauna of Santa Cruz
island, California. Their overbrowsing is believed to have caused the extinction of Easter
Island's only native tree Sophora toromiro (Lever 1985).
Feral pig Sus domesticus impacts have been studies on the islands of Sao Tome and Principe.
Amongst others, they are known to cause degradation to natural plant communities, and their
rooting activity can assist the spread of exotic plants (Dutton 1994).
Impacts of feral cats on native species in the UK are often cited, but as Churcher & Lawton
(1987) point out, feral cats form only about one fifth of the British cat population. The
remaining 80% of the British cat population are house cats. In a study of approximately 70
domestic cats in a Bedfordshire village, Churcher and Lawton (1987) found that prey items
brought home consisted of 22 bird species and 15 species of mammal. In one year, the 70 cats
brought home 1090 prey items, but this may be as little as 50% of their prey. Most important
prey were woodmice (17%), house sparrows (16%) and bank voles (14%).
The impacts of cats have been severe in other countries, especially on islands, where they
have been blamed for the decline of many island species (Lever 1985). In New Zealand and
the surrounding islands the impact of cats has been so detrimental that control and eradication
programmes have been implemented (Fitzgerald & Veitch 1985).
The endemic lizard Urosaurus auriculatus of Socorro Island, Mexico, is an important prey
item for the feral cat (Arnaud et al. 1993). The impact of sheep on the natural vegetation,
together with the impact of the cats means that active measures need to be employed to
remove exotics and restore natural vegetation.
Feral dogs have caused reductions in numbers of native birds on Hawaii, and have led to
extinction of two subspecies of snakes in the West Indies (Lever 1985).
Another interesting case of predation concerns sheep. Domestic sheep are descendants of the
Asiatic mouflon Ovis orientalis, and so can be considered as non-native. On the island of
Foula, Shetland, sheep were seen to ingest bone-rich parts of unfledged Arctic terns Sterna
paradisea (Furness 1988). The sheep often severed legs or wings, and wounds were not
always fatal (unlike the decapitation of manx shearwater Puffinus puffinus chicks by red deer
Cervus elaphus on Rhum). This habit has not been widely documented, so may only occur in
rare situations where ruminants are grazing mineral-deficient vegetation on which there are
dense colonies of ground-nesting birds. Another possibility is that the type of injuries seen
may have been overlooked elsewhere or attributed to other animals, such as otters or mink. Genetic impacts
British populations of the European wildcat Felis silvestris had suffered a reduction in
numbers due to deforestation and persecution (Hubbard et al. 1992). Recovery this century
Review of information, policy and legislation on species translocations
has been due to reduced shooting and trapping, and to increases in the area of suitable habitat
(forestry plantations in Scotland). The increase in domestic cats, Felis catus, and hence feral
domestic cats in the wildcat habitat is complicating this recovery. Much hybridisation may
have occurred between Felis silvestris and Felis catus after the First World War, when the
wildcats recolonised former habitats which had become occupied by domestic cats. Hybrids
are frequent, but genetically distinct wildcats do remain, particularly in western Scotland.
Reasons for this endurance may be the remoteness of the populations, and little movement
between isolated feral cat populations in Scotland. In addition, local control programmes of
feral domestic cats may have further reduced the threat. Also, many encounters between
wildcats and domestic cats are antagonistic, with domestic cats sometimes being killed, and
this may inhibit establishment of feral cats in wildcat habitats. It is not clear whether the
genetic integrity of wildcat populations in Scotland is seriously threatened at present by feral
Felis catus. However, long-term conservation of wildcats may require control of feral
domestic cats, prevention of domestic cats entering wildcat habitats, and preservation of
wildcat habitats.
The impacts of domestic animals on flora and fauna of areas of introduction include habitat
damage by herbivorous mammals such as sheep and pigs, and predation by introduced
species such as cats and dogs.
These impacts can be devastating in some areas. Once again, islands have been most severely
affected, both by herbivores (e.g. sheep, rabbit) and by predators (e.g. cats). Cats,
particularly, have caused declines in abundance of many island species, including groundnesting birds and lizards.
In response to Proulx's (1988) call for greater control of domestic cats to reduce predation on
urban wildlife, Fitzgerald (1990) pointed out that it is not known which small mammals and
birds, if any, would maintain significantly higher populations if cats were removed. Indeed,
removal of cats may be deleterious in some circumstances. When the feral cat population in
one New Zealand forest was reduced, the population of ship rats (Rattus rattus) increased
fourfold (Fitzgerald, 1988). Ship rats in New Zealand forests are important predators of birds
and nests. Increased numbers of rats may be more detrimental to many bird species than cats
Impacts of domestic animals do not appear to be great in the UK at the present, possibly due
to the fact that most have not been allowed to form feral colonies. Further research may need
to be done to ascertain exactly what damage is being done by domestic animals, especially
the impact of cats.
Review of information, policy and legislation on species translocations
3.10 Forestry
3.10.1 Background
Many of the native UK woodland trees are slow growing species, generally unsuitable for
forestry. The lesser demand for hardwoods and their slower rate of growth has resulted in a
low economic justification for the maintenance of native species of hardwood in areas
primarily devoted to commercial forestry (Jeffers 1972). The faster growing, non-native
species of conifers have been widely planted. Introduced forest trees in the UK include the
European larch Larix decidua has been important timber tree since about 1780, especially for
boat building.
Norway spruce Picea abies is the traditional Christmas tree.
Sitka spruce Picea sitchensis originates from coastal parts of Alaska, and is the chief forest
tree in western Britain.
Corsican pine Pinus nigra var. maritima.
Austrian pine Pinus nigra var. nigra.
Weymouth pine Pinus strobius, an American species, is good timber for e.g. mouldings and
musical instruments.
Douglas fir Pseudotsuga menziesii is an American forest tree which is extremely widely
distributed in the UK due to the excellent qualities of its timber.
Pedunculate oak Quercus robur.
Sweet chestnut Castanea sativa is a very early introduction from southern Europe, probably
introduced into Britain by the Romans.
There is currently about 845,482 ha of land in Britain under plantation forestry, with 508850
ha in Scotland alone. In 1993, new planting and restocking took place on 10,830 ha of land
(9,657 ha of coniferous species, 1,173 ha of broadleaved species) (Forestry Commission
Clearcutting, intensive silviculture and single species forest management reduces habitat
diversity and decreases the density and diversity of breeding birds (Parker et al. 1994), The
introduction of these non-native tree species themselves poses several threats to native flora
and fauna.
Review of information, policy and legislation on species translocations
3.10.2 Effects of introductions Ecological impacts
Habitat alteration
The destruction of important biotopes by forest planting is a well-known negative impact, but
this is an issue of land use rather than of alien introductions. We shall not address these land
use issues, but shall concentrate on the impacts of non-native trees on native biota.
The planting of conifers on poorly-buffered acid and acid-sensitive soils is partly responsible
for the increased acidification of the British upland environment (Neal et al. 1992). The effect
of this, and more importantly of the felling of plantations, is the generation of more acidic
stream water with more available aluminium, and this is detrimental to some aquatic biota.
The effects of coniferous afforestation on rove beetles (Coleoptera, Staphylinidae) was
investigated in Kielder Forest in plantations of various age and in unplanted sites (Buse &
Good 1993). Tree planting decreased habitat availability for most beetles, but provided new
habitat for forest species. The greatest abundance, species richness and diversity occurred in
non-afforested sites.
In Natal, South Africa, the impacts of plantations of two exotic tree species on native
grasshoppers has been studied (Samways et al. 1991). The grasshoppers are considered to be
good indicator species for general grassland insect communities, and there is a strong positive
correlation between grasshoppers and grass species richness. Orthopteran abundance and
species richness in adjacent grassland was increased by patches of exotic cypress, whilst
exotic pine patches proved detrimental for the grasshoppers, far into the adjacent grassland.
A study of the ecology of Pinus plantations in New South Wales (Curry 1991) discovered a
reduction in diversity of native species compared to native plantations. Insectivorous bird
diversity was reduced by the reduced availability of food. It is probable that few native
invertebrates would be adapted to exploit the non-native pines.
Beneficial effects on natives
The pine beauty moth Panolis flammea is native to Great Britain. Before 1973, it was only
known to occur at low levels on Scots pine Pinus sylvestris. However, in the mid 1970s,
outbreaks were discovered in plantations of the exotic lodgepole pine, Pinus contorta. An
integrated pest management programme is being developed. The pine beauty moth had been a
common, but unimportant herbivore of native Scots pine prior to outbreak (Watt et al. 1991,
Evans et al. 1991). Impacts relating to the spread of pests and disease
Monocultures have often led to an increase in insects and disease problems (Larsen 1995).
Several factors may dispose single species stands to insect attack including lack of natural
enemies, high concentration of host plants, absence of alternative hosts and development of
Review of information, policy and legislation on species translocations
closer coincidence between insect and plant phenologies. This may cause economic problems
for foresters, but it is not a conservation problem.
The accidental introduction of disease with imported timber, e.g. Dutch elm disease is
discussed under Accidental introductions.
3.10.3 Discussion
Impacts associated with the introduction of non-native tree species for forestry include
acidification of soils and streams, contamination of streams by aluminium after felling,
reduced diversity of the native biota due to even-aged monocultures of exotic trees, and
diseases imported with foreign timber.
The problem of reduced native floral and faunal diversity associated with exotic tree
plantations could be dealt with by increasing the structural and floristic diversity of the
plantations (Curry 1991). By staggered planting, thus creating a range of successional stages,
and the retention of native vegetation within and near plantations, opportunities should be
presented to a wider range of native organisms. Forest managers could aim to increase
staphylinid diversity 'by design', particularly by varying tree species and age class so as to
develop greater biological and structural diversity (Buse & Good 1993). Habitat diversity
could be enhanced further by conserving representative areas of former land use, such as
farm fields, river banks and open moorland. Staphylinid species are favoured by forest edge
habitats, so would gain from the integration of small habitat units within plantations, resulting
in a beneficial 'knock-on' effect by providing prey for birds and small mammals.
Crop species
3.11.1 Background
Since humans first moved from one region to another, crop species have been moved with
them. Therefore, many long-established species were originally introduced for their utility
value, e.g. most agricultural crops like maize, wheat, tomatoes, herbs, medicinal plants, and
also species introduced for pasture improvement. Despite their widespread use, there has
been surprisingly little research into the impact of introduced crop plants in the UK.
However, the advent of genetically modified crop plants has stimulated research (see Chapter
Until the latter part of this century and the advent of GMOs, the only method for introducing
genetic variation into crops was by plant breeding. Two parental types, with desirable traits,
would be crossed and the progeny either selected for use or for further breeding. As a result
of such breeding programmes, many crop species exist as a range of cultivars.
Review of information, policy and legislation on species translocations
3.11.2 Effects of introductions Ecological impacts
Crop species might 'escape' from agriculture and have detrimental impacts upon the
environment (Raybould & Gray 1994). Volunteers populations may persist after harvest to
become weeds of cultivation, or feral populations may establish on non-agricultural land.
However this is rare, and a list by Raybould & Gray (1993a) of British crops forming
persistent feral populations has just one non-native species - oilseed rape Brassica napus
oleifera. Genetic impacts
Crop plants have always been subject to genetic modification. Historically, this occurred
through selection for individuals with increased production, naturally disease-resistant
strains, etc. but this century has seen the advent of genetic modification and transgenic crops
(see Chapter 4). Gene transfer could occur between the crop and a wild species by pollen
transfer, producing a hybrid.
Raybould & Gray (1993a,b) have identified three groups of crops on the basis of crosscompatibility with wild relatives (see Chapter 4 for a fuller discussion). The first group
consists of species with a 'minimal' probability of gene flow into wild relatives, and includes
potato, maize, wheat, tomato, some grain legumes, cucumber and sunflower. With these
species, the concern is that the crop itself may escape from cultivation, and that breeding may
have altered the persistence, weediness or invasiveness of the plant.
Group two species have a 'low' probability of gene flow, and are generally species with no
wild conspecifics but instead close, usually congeneric, wild relatives with which there is
limited sexual compatibility. This group includes oilseed rape, flax, raspberry, lettuce and
Group three species have a high probability of gene flow. Species in this group include sugar
beet, carrots, some cabbage cultivars, forage species ryegrass, clover and lucerne.
Overlapping geographical distributions and reproductive compatibility with wild relatives
indicate that the escape of transgenes via hybridization is possible and probable.
However, hybrids of crop plants and wild relatives are usually rare, often sterile and
relatively few populations persist, except where the parents remain in contact or where they
are able to spread vegetatively. This is because crop plants may bear several dominant traits
of domestication that are maladaptive in hybrids (Ellstrand & Hoffman 1990).
Review of information, policy and legislation on species translocations
3.11.3 Discussion
Crops that form feral populations and that have the ability to hybridize freely with wild
relatives can be identified (Raybould & Gray 1993a). Regions where crops and wild relatives
co-occur should be identified. In many cases this will reveal that the possibility of
introgression is confined to a certain region. For example, lettuce Lactuca sativa is cultivated
commercially in England in Kent and Lancashire. The wild species of lettuce, with at least
some degree of cross-compatibility, are common in Kent, but rare or absent in Lancashire.
The threats in this case are thus confined to Kent.
Klinger et al. (1992) reported that crop-weed hybridization rates between cultivated radish
and nearby wild radish (both Raphanus sativus) declined with increasing distance between
crop and weeds and that the effect of population size on the rate of hybridization was
Raybould & Gray (1993a) suggest several issues that are in need of further investigation.
Which natural habitats are most invasible.
Which modifications would be likely to improve crop species' performance in these
Regions where crops and wild relatives co-occur.
The method of gene transfer through populations of the crop's wild relatives.
The only possible method of ensuring that crop genes do not enter wild populations would be
to isolate crops from their wild relatives, or other species they might be compatible with.
However, in practice, isolation of outcrossing crops from their wild relatives may not be
practical or possible.
3.12 Fur animals
3.12.1 Background
The use of species for their pelts is another long-standing reason for introduction of nonnative species. The rabbit was kept for its fur in the past (see Section 3.7). More recently,
exotic mammals have been farmed for their fur in the UK, and these include the following.
The American mink Mustela vision was first introduced into Ireland in the early 1950s for fur
farming, and soon established feral populations. Population densities of American mink in
Ireland were highest on the River Glore in the Irish Midlands (Smal 1991). The populations
are believed to be at, or near, the carrying capacity of the environment and in all cases are self
regulating. Availability of crayfish Austropotamobius pallipes appears to be a major factor in
determining mink numbers and stability within populations in Ireland. Both otters Lutra lutra
Review of information, policy and legislation on species translocations
and foxes Vulpes vulpes kill mink, but in insufficient numbers to act as a regulatory
Coypu Myocastor coypus, a South American rodent of wetland areas, with webbed hind feet
for swimming and diving, were first imported for fur farming in about 1929 (Gosling &
Baker 1989). Fifty farms were established, mainly in the south and east of England near
natural water supplies, but all were closed before 1940. However, persistent feral populations
became established in Norfolk. By the mid 1940s a small population was well established in
the River Yare valley. By the late 1950s, despite sporadic control operations, coypus were
found over most of East Anglia. Between 1962 and 1965 MAFF started a trapping campaign
to reduce numbers because of increasing damage to crops, natural plant communities and
drainage systems, and the coypu has now been eradicated (Gosling & Baker 1989).
Muskrat Ondatra zibethicus was introduced from North America in the early years of this
century to allow farming of its fur, musquash. It became extremely widespread in the wild in
Great Britain, but was eradicated in the 1930s following a major programme of control
(Sheail 1988, Gosling & Baker 1989).
3.12.2 Effects of introductions Ecological impacts
The Norfolk Broads consist of a series of rivers and artificial lakes, fringed naturally with
reedswamp vegetation. In the past reedswamp tended to colonise large areas of the open
water, but during the 1970s much of the original reed was reported to have died (Boorman &
Fuller 1981). Marginal reedswamp is an important habitat for local bird species and protects
banks against erosion by wave action. It is also the precursor of the often species-rich fen
vegetation. Coypus caused major declines in reedswamp vegetation between 1950 and 1963.
Eutrophication led to the virtual elimination of aquatic macrophytes in many areas, with the
result that the wildfowl and coypu that formerly fed on the macrophytes increasingly turned
reedswamp vegetation as an alternative food source. The eutrophication also increased the
rate of sedimentation in the Broads, and the consequent deposits of soft mud raised the
susceptibility of reeds to grazing (the young rhizomes and shoots could be more easily
reached by herbivores). Thus, although the coypu appears responsible for much damage to
vegetation and consequent habitat alteration, this was only possible in combination with
Australian brushtail possums Trichosurus vulpecula were introduced into New Zealand to
establish a fur industry (Owen & Norton 1995). Escaped individuals bred, and the
populations expanded rapidly. The possum diet is dominated by a small number of food
types. Some species, especially short-lived 'sera' tree species are preferred. Most studies of
possum impacts have been in floristically diverse forests, where the preferred food species
Review of information, policy and legislation on species translocations
form an important structural component. High possum numbers have caused widespread and
progressive mortality, resulting in major changes in species composition and forest structure.
In Nothofagus forests, the palatable species form a small percentage of biomass, but
contribute much to species diversity. The selective browsing of the possums reduces
diversity, accentuating the already strong bias towards unpalatable biomass in these forests.
The impact of possums may be more severe in these depauperate forests due to heavier
browsing of the smaller number of preferred species. Key food resources for other species,
such as forest birds, may be removed by the possums.
The impact of feral mink on populations of native birds, mammals and fish along waterways
is currently being debated, although this predator has an impact on some species of
waterbirds, the magnitude of effects on other species are not known (Woodroffe et al. 1990).
One vulnerable species is the water vole Arvicola terrestris. There is a significant inverse
correlation between mink activity and vole activity. Therefore, it is probable that, in the long
run, mink will depress water vole numbers. Mink hunt on land and in water, so the voles
avoidance action of swimming does not deter the mink. Polyphagous predators, sustained by
a large variety of alternative prey, may wreak havoc with particularly vulnerable species,
eliminating them entirely from certain localities or ecosystems. By reducing population size
and fragmenting water vole colonies, mink pose serious long-term threat to survival of water
voles on British rivers.
When the otter Lutra lutra and American mink were studied in Sweden (Erlinge 1972) they
were found to be ecologically distinct, with the otter being more adapted to life in water.
During the warmer months, otter and mink differed in food and habitat preference, mainly
due to their differing ecological adaptations. However, in colder months, otter and mink
preferred the same habitat and food overlap is great.
The otter population in Sweden decreased in number at about the same time as the rapid
increase of the mink population, but probably because of water pollution and hunting rather
than directly due to the mink. The previous reduction of otter populations in the 1950s may
have made it easier for the mink to colonise. The mink appears to cause restriction of otters to
optimal habitat, whilst high densities of otters locally limits the population size of mink.
Introduction of the North American mink has also affected the European mink Mustela
lutreola (Maran & Henttonen 1995). Populations of the European mink have declined
throughout Europe this century, and in general the decline appears to be correlated with
environmental change. However, in addition to habitat destruction, introduced mink now
compete with the more specialised native mink and may have been disease vectors.
Habitat alteration
Both coypu and muskrat caused extensive damage to watercourses and riverbanks by
burrowing (Sheail 1988, Gosling & Baker 1989).
Review of information, policy and legislation on species translocations
3.12.3 Discussion
The impacts of mammals which have escaped from fur farm facilities can be various,
affecting vegetation through herbivory, predating upon native animals, or reducing floristic
diversity. Coypu and muskrat caused such problems that a programme to eradicate them in
the wild was put in place, and these have been successful. Mink, on the other hand, are now
fairly widespread in Britain and it would be very difficult to eradicate them completely. They
have been blamed for affecting many native species, although their true impact has yet to be
The restrictions on the import and keeping of animals such as mink and coypu (Destructive
Imported Animals Act 1932) should ensure that no further individuals of these species gain
entry into the wild.
3.13 Accidental introductions
3.13.1 Background
This is a very large category, covering species that have been imported accidentally (rather
than escapes of deliberately imported species), involving many varied forms of introduction,
and many species. The following are some examples.
Shipping has transported many marine organisms around the world, for example, in ballast,
cargo, or on the hull. For instance, the algae Solieria chordalis, Pikea californica and
Antithamnionella spirographidis, the crustacean Elminius modestus and the North American
annelid Hydroides dianthus have all been transported to British shores on ships' hulls (Eno
Aquatic species have also been introduced with commercial shipments of shellfish, e.g. the
Pacific cnidarian Haliplanella lineata, the molluscan gastropod Rapana venosa, and the algae
Polysiphonia harveyi and Sargassum muticum (Eno 1995).
Imports of grain have been responsible for introduction of species of beetle and arable weed
seed, and many plant species have been introduced with birdseed or wool shoddy.
Spartina alterniflora, a species native to the United States, was introduced into Britain
accidentally with shipping ballast. This species hybridized with Spartina maritima (native,
but at the northern limit of its natural distribution on the south coast) to produce a hybrid
whose allopolyploid is S. anglica (Thompson 1990). During the first half of this century S.
anglica was widely introduced to salt marshes to enhance sediment accretion for reclamation
of land, from where it established vigorous populations.
Invertebrates have also invaded after being stowaways during transport of, e.g. fruit or exotic
rooted plants. The Australian amphipod, Arcitalitrus domini, is found in south-western
Review of information, policy and legislation on species translocations
Britain, the Inner Hebrides and Ireland (Nelson 1994). In Ireland it is generally in or near
gardens, supporting the theory that it arrived in batches of rooted, growing plants.
The New Zealand flatworm Artioposthia triangiulata is a terrestrial planarian originally from
New Zealand. Its initial introduction into north-western Europe and subsequent dispersal was
probably facilitated by movement of containerised plants (Stewart & Blackshaw 1993). Irish
populations widely distant from one another are closely related genetically, which supports
the theory that dispersal is not natural. The first record in Scotland was from Royal Botanic
Gardens in Edinburgh in 1965 (Boag et al. 1994). From there it spread to nurseries and
garden centres, and from there to domestic gardens. In the last 30 years A. triangulata has
become well-established throughout most of Northern Ireland, northern England and
Scotland and is still spreading. The Australian flatworm Australoplana sanguinea is
spreading northwards from south-west England.
Unintentional vertebrate introductions have been rarer due to the more conspicuous nature of
vertebrates due to their size. However, there are exceptions, particularly among rodents. Rats
have been transported around the world aboard ships, the black rat Rattus rattus is thought to
have entered Britain in the luggage of returning Crusaders. Cytotaxonomic studies have
shown that R. rattus found throughout the world have common ancestry in southern India
(Armitage 1993).
3.13.2 Effects of introductions Ecological impacts
The New Zealand flatworm can grow to 15cm and weigh over 2g (Blackshaw 1990). It preys
upon earthworms and may be capable of inflicting severe losses on indigenous earthworm
populations to the point of elimination. There is no evidence so far of any natural regulation
of A. triangulata numbers other than shortage of food. They are subject, however, to a
relatively low upper lethal temperature, with low survival in temperatures above 20oC.
Temperature may therefore play a key role in determining distribution of this species, with
cooler habitats favoured, which helps explain the current distribution of the species in UK.
The effects of a reduction of earthworm populations on soil structure, nutrient cycling, crop
yield and wildlife are unknown, but beneficial effects of earthworms have been documented
(Boag et al. 1994). By denuding areas of earthworms, the flatworm is likely to have indirect
detrimental effects on other species reliant upon the earthworms, including species of birds
and mammals. The Australian flatworm may have similar effects.
Endemic terrestrial tree snails of the Hawaiian Islands are extremely sensitive to disturbance
because of their low population numbers and small geographic ranges (Hadfield et al. 1993).
Like many other plants and animals of oceanic islands, they have evolved no defences against
introduced predators and competitors. Many snails exhibit slow growth and late maturity.
Population growth typically depends on considerable longevity. Predation by alien predators,
rats and a North American predatory snail (both introduced accidentally) threatens the
Review of information, policy and legislation on species translocations
endemic snails. The predatory snail eats all sizes of snail and can drive populations rapidly to
extinction. Rats select larger snails and may leave an area before destroying all snails, thus
reproductive output is temporarily reduced but the population may survive. This illustrates
that introduced predators may impact the same native species in different ways, and that a
combination of impacts from different predators may cause a much greater impact.
The endemic New Zealand avifauna evolved in the absence of mammalian predators (Moors
1983). Three species of European rodent (Rattus rattus, R. norvegicus and Mus domesticus)
have since been introduced unintentionally together with three species of mustelid introduced
deliberately for rabbit control. All species kill indigenous and introduced birds. The mustelids
were found to destroy more nests with chicks in, whereas the rodents generally destroy nests
with eggs in.
The alga Polysiphonia harveyi, originally from the Pacific ocean, was introduced to the south
coast of England with shipments of oysters prior to 1908 (Eno 1995). It is an opportunistic
gap coloniser and its rapid growth rate means it may displace native species due to
abundance. It is now distributed along coasts of south and east England, western coasts to
Scotland, Ireland, and from Norway to the Mediterranean in Europe.
The alga Sargassum muticum occurs naturally in Japanese and Chinese waters. It too was an
unintentional introduction with commercial introductions of oysters, coming from either
northern France, Canada or the United States (Eno 1995). It has spread quickly along the
south coast of England, at a rate of approximately 30 km per year, and is now distributed
along the Channel coast and the east coast to Norfolk. It is expected that higher temperatures
in the future will encourage further spread of this species. It is possible that it will displace
native species, as it is already known to displace native species on the French Atlantic coast,
e.g. Laminaria saccharina and Zostera marina.
The Australasian crustacean Elminius modestus was imported to Britain on ships' hulls (Eno
1995). It is able to grow rapidly, and withstand reduced salinity, turbidity and lower
temperatures than the native barnacles Chthamalus spp. and higher temperatures than native
Balanus spp. In the north, Elminius competes with Semibalanus balanoides, while in
southern Europe Elminius competes with Chthamalus spp.
A New Zealand reptile, the tuatara (Sphenodon punctatus), is threatened by introduced
Pacific rats (Rattus exulans). The rats can cause extinction through competition for food and/
or predation on eggs and juveniles (Cree et al. 1995). Pacific rats inhibit recruitment of the
tuatara and have already caused extinction on four islands in the Mercury Group and the Hen
and Chicken Group of New Zealand.
Habitat alteration
There is only circumstantial evidence that Spartina anglica has a detrimental effect on other
Spartina species, the succession of plant communities and the abundance of invertebrates and
waders (Thompson 1990). Indeed, the hybrid may be beneficial. With the predicted rise in
Review of information, policy and legislation on species translocations
sea level caused by global climate change, the enhanced development of salt marshes by the
hybrid may lessen the expected losses of this habitat.
Beneficial effects on native species
The giant kangaroo rat Dipodomys ingens is endemic to the arid grasslands of California.
Widespread conversion of grassland to agriculture (habitat destruction), and the use of
rodenticides has made this an endangered species. Invasion of the remaining grasslands by
plant species of Mediterranean origin has been responsible for displacement of native species
(Schiffman 1994). The kangaroo rats (mainly granivorous) may have contributed to the shift
from native to exotic species by their continual modification of the soil and vegetation near
burrows. The rats are now dependent upon the exotic species for food, as the exotic species
depend on the rats to disturb the ground continually. This poses an interesting problem for
conservationists. Eradication of exotic species would probably have a significant negative
impact on the rat populations, whilst enhancement of rat populations would probably result in
an increase in the cover and diversity of exotic plant species.
Ecological studies
The systematics of most marine taxa are far from complete, and the discovery of previously
unrecognised species in regions affected by ballast water release (most coastal zones of the
world) must be viewed as potential invasions (Carlton & Geller 1993). For easily identifiable
species, unrecognised historical transport may have led to false conclusions of natural
cosmopolitanism. This confounds our understanding of historical patterns of dispersal, gene
flow, and speciation since geographic barriers to dispersal and gene flow are readily breached
by ballast water transport. Impacts relating to the spread of disease
Of considerable concern has been the discovery of cysts of the toxic dinoflagellates
Alexandrium catenella, A. tamarense and Gymnodinium catenatum in the ballast water of
ships entering Australian ports (Hallegraeff & Bolch 1992). These species can contaminate
shellfish with paralytic shellfish poisons and pose a serious threat to human health and the
aquaculture industry. Mid-ocean exchange of ballast water is only partially effective in
removing dinoflagellate cysts which have settled to the bottom of ballast tanks. When ballast
water and sediments from bulk cargo carriers involved in the export of woodchips from
Washington State to Japan were examined, the results indicated that the threat of introduction
of harmful algae, pathogens, predators and resource competitors is genuine (Kelpy 1993).
The nematode Anguillicola crassus was imported into Britain with infected Japanese eels
Anguilla japonica, from Taiwan via West Germany (Eno 1995). A. crassus spreads within
aquatic systems through intermediate hosts, and spreads between localities through transport
of infected eels. At high levels of infestation, affected native eels can show adverse effects.
The rabbit myxoma virus Myxomatosis cuniculi was officially introduced into other countries
to control rabbits, but not to Britain. The first reported outbreak of the virus in 1953 may
have been accidental, but it seems more likely that it was deliberately, although not officially,
introduced. Armour & Thompson (1955) estimated that the original outbreak in 1953 caused
nearly 100% mortality. The huge mortality caused further concerns, this time from
Review of information, policy and legislation on species translocations
environmentalists, that without rabbit grazing many habitats maintained by grazing might
have been lost (Sheail 1991), for example, steep chalk pastures traditionally grazed by sheep
were being maintained by rabbit grazing. With much reduced grazing pressure from the
decimated rabbit populations, scrubby vegetation developed on many sites. The extinction of
the British population of the large blue butterfly, Maculina arion, can be blamed partly on the
decline in rabbits. When the rabbit population crashed, the early successional grazed habitat
was replaced with taller grassland which no longer supported the butterfly (see section 2.5.2).
The loss of an important prey species may have also affected predators such as buzzards
Buteo buteo. However, the effects of myxomatosis are decreasing for two reasons. Firstly, the
virus strains currently found in Britain are less virulent than the strains in the original
outbreak in 1953. Secondly, genetic resistance to myxomatosis was detected in wild rabbit
populations in 1970, and resistance has been increasing (Trout et al. 1992).
The Dutch elm disease fungus Ceratocystis ulmi was introduced into Britain with imported
timber. The disease is spread by beetles of the genus Scotylus, or through the roots of
adjacent trees. The disease has greatly reduced the number of elm trees Ulmus spp. over
much of England and Wales. Through its impact on elm trees, the disease also affected
invertebrates dependent on elm and, indirectly, many farmland bird species (Osborne 1985).
The elm deaths alone were probably not of great significance to most bird species, but the
felling of dead elms caused a reduction in numbers and diversity of bird species. The dead
elms would have provided nest sites and alternative food sources, but it is unlikely that they
would have catered for as many species as live elms. Woodpeckers Dendrocopos spp and
nuthatch Sitta europaea appear to have benefited from the abundance of beetle larvae on dead
elms, and nest in elms more often than formerly, but the introduction of Dutch elm disease
may have had adverse effects on nesting kestrels Falco tinnunculus, stock doves Columba
oenas, barn owls Tyto alba and tawny owls Strix aluco (Osborne 1982).
3.13.3 Discussion
There have been accidental introductions to the UK of all types of plant, animal and
pathogen, and therefore the impacts are diverse, including competition, predation, herbivory,
habitat alteration and the transfer of disease.
Recent guidelines specifically mention accidental introductions. For example, IUCN (1987)
expresses concern over accidental introductions, especially to islands and isolated habitats.
Most regulatory bodies agree that steps should be taken to avoid unintentional release into the
wild of non-native species.
Many accidental introductions are avoidable, for instance, better seed cleaning has reduced
the incidence of 'weed' seeds in imported grain, and increased vigilance during packing and
transport of cargo might lead to early removal of 'stowaways'.
However, not all introductions are avoidable, e.g. seeds transported by vehicles, particularly
in tyres, are not easy to prevent. Instead of preventing the initial introduction, in order to
protect important, natural areas such as nature reserves, the creation of buffer zones between
Review of information, policy and legislation on species translocations
the reserve and the road together with monitoring and action to remove aliens where invasion
is occurring would help protect natural areas (Tyser & Worley 1992).
In the case of marine organisms, it is probably already too late to prevent major introductions.
The global transport of organisms has already blurred the international patterns of occurrence
of most species (Eno 1995). Detriment to the marine environment has gone unnoticed until
relatively recently, mainly due to the inaccessible nature of the environment in question.
However, species distributions are not the only issue associated with ballast water, since,
perhaps more seriously at the moment, disease organisms are also being spread by this
method. The most effective measure to prevent the spreading of toxic dinoflagellate cysts via
ballast water would be to avoid taking on ballast water during dinoflagellate blooms in the
water column. Enforcement of the IMO (1993) guidelines, aimed at preventing the release
into the wild of contaminated ballast water or sediment, could prevent further introductions of
disease and pathogens in this way.
3.14 Discussion on introduction of non-native organisms
3.14.1 Perceptions of non-native species in the UK
Resident aliens
Alien species are often defined as those not resident in the UK before a particular time in
prehistory, usually the Neolithic, c. 6,000 BP (Webb 1990) or the last glaciation, c. 14,000
BP. For species resident in the UK, the question of the native status of a species is more of
academic interest than an imperative for conservation. In some cases the distinction between
native and non-native may not straightforward, and often is the result of educated estimates
of the length of time a species has been resident in the UK. The pool frog Rana lessonae
provides a cautionary example of this problem. The frog was traditionally thought of as an
alien, or at least of questionable status, but its status has been changed to native following
recent findings that the frog has been resident in England for many centuries, perhaps even
since the last glaciation.
Even if a species is clearly non-native, but it is not known to cause conservation problems
(i.e. a decrease in biodiversity, however defined), and is not expected to increase its numbers
or range (i.e. it is long-established and is biogeographically stable), there is little need for
conservationists to concern themselves with that species. It would be extremely purist to want
to eliminate or control a species just because it is an alien. However, this response to our
alien species is commonplace, with a few exceptions made for charismatic or accepted
species, such as the little owl Athene noctua or the horse chestnut Aesculus hippocastanum.
Some aliens are seen as beneficial to the UK flora and fauna, although this is usually the
result of many hundreds of years of effects of the alien, which has come to be seen as part of
the natural biota of the UK. A famous example is the rabbit. A few species introduced to the
UK are actually subject to conservation measures in their native ranges, e.g. Mandarin duck
Aix galericulata and Roman snail Helix pomatia.
Review of information, policy and legislation on species translocations
If a resident species is causing a conservation problem then measures to control the problem
should be considered whether or not the species is an exotic. In fact, several native species
cause major problems - e.g. bracken Pteridium aquilinum, birch Betula spp and tor grass
Brachypodium pinnatum are all invasive native species of valued semi-natural communities although native vertebrates are rarely perceived in this way. The only difference may be that
complete eradication of a problematic alien species may be acceptable to some people,
whereas such a target for control of native species will rarely be acceptable.
This assessment of alien species based on their effects on biodiversity may not apply to alien
species which have formed wild populations in the UK only recently (however 'recently' is
defined). Species that do not seem to be causing problems may still be expanding their range
and/or building up populations, and may become a problem at a later date. It may be difficult
to assess whether species - other than those resident for many hundreds of years - have
achieved stability or not.
New introductions of aliens
The introduction of UK non-native species to the wild is a different question.
Supplementation of alien species which are already resident in the wild may increase
population sizes and geographic ranges, and this could cause a formerly benign species (i.e.
one not causing a problem) to have negative effects on biodiversity. As is the case for
introductions of UK natives into regions outside of their natural range (see Chapter 2), the
introduction of species alien to the UK into new areas may have unpredictable, and possibly
severe, consequences.
If an alien species is not found in the wild in the UK then it could have drastic effects when
introduced, especially if it is introduced without its associated, and possibly co-evolved,
natural enemies. One must be very circumspect about introducing it to the UK, and
assessment of potential problems should be carried out.
3.14.2 Assessing the potential for invasion and spread of alien species
The case studies have shown a wide range of impacts of introduced aliens on the ecology and
genetics of communities and species. These effects include: population decline or loss of
native species through, e.g. increased herbivory, predation, competition and disease load;
change in habitat characteristics through, e.g. grazing, tall growth (of plants), sediment
disturbance or soil disturbance; and hybridisation and introgression with native species.
However, many species have few perceivable effects, and others do not seem able to establish
in the wild in the UK. One only has to look at the vast number of alien plants grown in
gardens and greenhouses to realise how few are found outside these controlled habitats (see
Perry & Ellis 1994). Of the alien bird species forming feral populations in Great Britain (see
Vinicombe et al. 1993), only a few cause any perceivable conservation problems.
Williamson (1992, 1993a, Williamson & Fitter 1995a) has proposed the 'tens rule' which
states that, for a taxon (however defined), 10% of aliens imported into Great Britain appear in
Review of information, policy and legislation on species translocations
the wild, 10% of these establish, and 10% of the establishing species - i.e. 0.1% of imported
species - become pests. Taking 10% to represent 5-20%, this rule has been shown to hold for
a number of groups in Great Britain, such as angiosperms (Williamson 1993a) and pines
(Williamson & Fitter 1995b), and in other parts of the world (see Lonsdale 1994, Williamson
& Fitter 1995a); but it does not hold for all cases (Williamson & Fitter 1995a). It is not clear
what proportion of these cause conservation problems; 'pest' status does not necessarily
indicate an impact on biodiversity.
This rule is of ecological interest, but all it tells conservationists is that a small proportion of
introduced species can establish in the wild. The next question is, which species will establish
in the wild.
Characteristics of invading species
The species which will cause conservation problems are those that will invade natural and
semi-natural communities, especially if they build up large populations.
Many authors have attempted to define the characteristics of the perfect invader, often in an
attempt to predict which are the species that will invade and establish into ecological
communities. Baker (1965) listed characteristics of the perfect weed. In summary, it is a
plastic perennial which will germinate in a wide range of physical conditions, grows quickly,
flowers early, is self-compatible, produces many seeds that disperse widely, reproduces
vegetatively and is a good competitor. However, a species may exhibit all of these traits and
still not establish itself when introduced. Conversely, a species does not need any these
characteristics to successfully invade (Drake et al. 1989). Thus, although of interest, the list
of characteristics is of little predictive value.
Scott & Panetta (1993) investigated weed species in Australia, and concluded that many of
the exotic species of South African origin which are weeds in Australia demonstrate weedy
behaviour in their native ranges. Occurrence in a wide range of climates and the existence of
congeneric species that are weedy also help to predict 'weediness'. The length of time an
exotic species has been present may also be important. Scott & Panetta (1993) suggest that
many introduced species that are not displaying weedy characteristics have not been present
long enough to do so.
These analyses concern weediness, and thus the ability to establish in disturbed habitats.
These form only part of the UK landscape, and any assessment of invasiveness must take into
account all habitats. Williamson & Fitter (1995b) examined the characteristics of successful
alien invaders in the British flora and found few significant differences from native species.
Gray (1986) concluded that invaders of the British flora were not characterised by particular
genetic characteristics. The only common attribute of twenty plant species listed by Crawley
(1987) as his 'top twenty' most abundant and widespread British aliens, was the propensity to
form dense thickets. This result probably reflects a circularity, in that the most noticeable
aliens are those forming large monospecific stands.
Ehrlich (1986) attempted to define the invasive potential of vertebrate species, and concluded
that successful invaders: may have a large native range where they are abundant; are likely to
be vagile species; are generalist in their feeding habits; have short generation times; have
Review of information, policy and legislation on species translocations
high population genetic variation, and have the ability to function in a wide range of physical
conditions. Conversely, unsuccessful vertebrate invaders are more likely to have a small
native range where they are relatively rare, a sedentary lifestyle, a limited diet, long
generation times, little genetic variability and a narrow preference for a range of physical
conditions. O'Connor (1986) found that successful bird invaders in Britain tended be species
with a greater clutch size and lesser propensity for long-distance migration than unsuccessful
For insects, Lawton & Brown (1986 ) report that the size of an insect is related to the
probability of successful invasion, but that the relationship is sufficiently weak that insect
size is of no predictive value. Simberloff (1989) could make no generalisations about the
invasive potential of insect species.
Characteristics of invaded habitats
Invasibility is a measure of a community's susceptibility to colonisation by exotic species
(Smallwood 1994). It is widely accepted that the more disturbed habitats are more readily
invaded - thus, anthropogenic disturbance increases community susceptibility.
Plant communities may be ranked in terms of their invasibility, based upon the proportion of
bare ground and on the frequency and intensity of soil disturbance (Crawley 1987). More
disturbed communities include urban wasteland, arable fields and river banks. The length of
time that has elapsed since the last major disturbance, i.e. the successional age of the site, will
also influence the alien flora (Crawley 1987). Thus, urban wasteland will have a higher
proportion of exotic species than unmanaged, native woodland. Drawing upon the evidence
of pine invasions, Richardson et al. (1994) found that the most widespread invaders were
those with attributes allowing populations to persist in habitats subjected to disturbance at
frequent intervals.
The rate at which plant communities are invaded is also determined by biogeographic factors
such as the size of the available pool of exotic species and the rate of population immigration,
which will itself depend upon the isolation of the site, and the area of the target plant
community (Crawley 1987).
Oceanic islands are especially susceptible to establishment and spread of invaders because
biotic resistance to invasion is low (Elton 1958, Brown 1989). The biota of these islands,
having evolved in isolation from the mainland, is particularly susceptible to the disruptive
influence of exotic species. There are often very few species of predator or parasite on such
islands, thus there are few enemies of invading species (Lake & O'Dowd 1991). The often
highly restricted range and small population size of insular species, together with their limited
diversity of defences, makes island biotas particularly vulnerable to extinction, largely
through habitat loss or interactions with introduced species (Paulay 1994).
Inability to predict invasions
Different authors have found different attributes that can be correlated with invasive ability of
a species, and some have found no correlations. Therefore, no general conclusions can be
drawn on this matter. Although habitats can be ranked according to their vulnerability to
Review of information, policy and legislation on species translocations
invasion, again, as for invading species, predictions as to precisely which habitats will be
invaded, and which of those will be most affected by such invasions, cannot be made with
any great degree of certainty at present. Thus, various authors conclude that it is impossible
to have general descriptors of a good invader, or of a particularly invasible community and
that only a detailed study of the ecology of a species and its potential habitats can allow one
to make any sort of prediction as to its potential success or failure (Crawley 1987, van
Broembsen 1989, Simberloff 1989, Ruesink et al. 1995, Williamson & Fitter 1995b). This
was the general conclusion of an analysis by Brown & Willamson (1986) of the risks of
introducing novel species to the UK and of the SCOPE programme on biological invasions
(Kornberg & Williamson 1986, Drake et al. 1989). This is not surprising; the determinants of
invasiveness of species and invasibility of habitats will interact in a complex way.
3.14.3 Negative effects of introductions - present and future
The UK appears to have been fortunate so far, in that most of the recently introduced exotic
species have not caused major adverse ecological impacts (Brown & Williamson 1986,
Kornberg & Williamson 1986) (effects of ancient introductions are difficult to assess). For
this reason, many of the case studies used in this chapter were drawn from abroad. However,
Ireland might be expected to be affected more severely by the introduction of non-native
species, since the fauna and flora are not as rich as that of mainland Britain or Europe, and
the major detrimental impacts of introduced species are experienced by biota of smaller
A second point to make is that most new introductions (i.e. introductions of alien species not
previously introduced to the UK) at the moment are biological control agents - which, as a
type, seem to have an inherently lower risk than other types of introduction (see above) - and
these are mostly of species that are most unlikely to be able to establish outside glasshouses
(W. Parish pers. comm.).
However, as this review has shown, certain species introduced to the UK have caused
problems, and certain of these conservation problems are great and/or high profile. These
include the ruddy duck, the grey squirrel, coypus, North American signal crayfish, zander,
Rhododendron ponticum, Fallopia japonica and Crassula helmsii, and the developing
problem of the New Zealand flatworm.
National environmental change & introductions
The general decline in UK biodiversity (declines in species numbers and species abundances,
disruption and loss of semi-natural communities, etc.) may cause ecosystems to become more
susceptible to adverse effects of introductions of alien species in the future (e.g. small
fragmented populations will be more affected by invasions). Climate change may bring about
more specific increases in the risks from alien introductions, as certain introduced species
which are not currently affecting biodiversity to a great degree may come to do so in the
future (Hill et al. 1994). With the rise in temperature predicted by climatic modelling, there
will be scope for many species (particularly plants) to increase their ranges. The climate is
Review of information, policy and legislation on species translocations
expected to warm by 2oC by the year 2050, with winters warmer by 2oC and summers by
1oC. As a result, species ranges may shift northwards. Alien plants currently rare and
restricted in the south may become common; sporadic species may establish, and yet more
species might invade (Hill et al. 1994). It is uncontrolled increases in abundance and range of
non-native species that cause negative impacts on native biota (see above).
Brown and Williamson (1986) stress the importance of climatic matching for invasion
success. The greatest success rate for species invading Britain is of those originating in the
same climatic region. Many of the most problematic introductions around the world appear to
be from warmer climes, and most introduced species do not establish persistent populations
because they are not suited to the local conditions. Therefore, as the climate changes it is
possible that the UK will experience a higher impact from invading species (Hill et al. 1994).
3.14.4 Control of introduced species - techniques and problems
Once an introduced species has been perceived as having adverse ecological or
environmental effects, measures to remove the organism may be required. An invasive
species generally attains problem status following population and/or range expansion. There
appear to be relatively few successful control/ eradication programmes against such problem
species. In addition, control measures are usually retrospective in nature; control or
eradication is not thought necessary until the species becomes a problem. Unless total
eradication of the introduced species occurs, control measures will have to be supplemented
by further control and monitoring of the species.
As the examples given below illustrate, control programmes are generally very expensive and
require extensive research into the ecology, economics and politics of the problem. For these
reasons, it is clear that measures to prevent the initial increase and spread of a species - i.e.
preventing the problem in the first place - is by far the best approach to deal with invasive
In the case of introduced mammals, Baker (1990) states that many problem species of
naturalised mammals have been expensive to eradicate or control. Indeed, most attempts to
remove such introduced species have failed. For example the attempt to eradicate mink
Mustela vison was abandoned in 1970 after only 5 years (Thompson 1971). Those species
that have caused problems have generally been released, or have escaped, in large numbers;
the more individuals that escape at the same time, the greater are the chances of establishing
in the wild. Examples of such species are muskrats, coypu and mink, all of which escaped
from fur farms. Both muskrats and coypus are generalist herbivores, damaging a wide range
of native plants and crops. The longer term economic and environmental costs of damage
cannot always be assessed at the early stages of an introduction (Gosling and Baker, 1989),
and this affects whether or not prompt action is taken. It was known from Europe that
muskrats could cause serious damage, and staff at the British Museum were able to convince
Review of information, policy and legislation on species translocations
the Ministry of Agriculture and the Treasury that eradication should be funded (Sheail, 1988).
In the case of coypu, however, relatively little was known about their biology and potential
for damage and thus little was done to prevent their establishment in the 1930s.
There is a reluctance on the part of those who finance eradication operations to take action
unless: 1) they believe there is a problem, 2) they think it can be solved, and 3) they know
roughly what the costs will be. The muskrat campaign was undertaken because much of this
information was available from experiences abroad. However, chances of success were
greatly improved because prompt action was taken. If the population had spread more widely,
as in Europe, chances of success would have been remote. The coypu eradication campaign
was only undertaken following detailed assessments of the efforts and costs required and the
likely chances of success. Eradication of muskrat and coypu was possible because
populations were confined to reasonable small areas with no immigration (unlike mink).
Smal (1991), discussing mink control, feels that control only disturbs the equilibrium
achieved naturally by mink populations, and that removal of mink leads to replacement by
transient individuals from outlying areas.
A survey of 1952 estimated that rabbit damage cost the agricultural industry 40-50 million
pounds per annum (Sheail 1984). Research into control was carried out at the Bureau of
Animal Population at the University of Oxford, financed by the Agricultural Research
Council. However, no successful or acceptable means of wide-scale control were established.
When myxomatosis was first reported in Britain in 1953, it appeared that here was a method
for the complete extermination of the rabbit, but by 1958 it was apparent that some rabbits
could survive infection, and the population numbers have recovered subsequently.
In Britain, methods of control used against grey squirrels include drey-poking and shooting
between November and April, but this is time consuming and expensive. Another method,
trapping between April and July, is effective but again is expensive in labour. Kenward et al.
(1992) found that none of these methods controlled grey squirrels to any degree, but they did
find that warfarin was more effective. The MAFF Warfarin Order 1973 allowed the use of
warfarin for control of grey squirrels. However, this has not eradicated the problem. Kenward
& Holm (1989) and Usher et al. (1992) think that it is unlikely that the grey squirrel could be
eradicated, due to their fecundity and the wide availability of suitable habitat.
The introduced Canada Goose has few friends among conservationists and there seems little
objection to co-ordinated control. There are a number of ways of limiting numbers (Owen,
1990). Control of breeding by egg collection might be a feasible and perhaps cost-effective
way of limiting numbers. Chemical control using poisons would be a very effective method
of population control, but would be unlikely to be popular on large scale. Increased shooting
on wintering grounds could be used as well. This could be done by three methods, but none
of these are popular. 1. Increasing bag size. Most shooting of geese in Britain is carried out
by wildfowlers who are resistant to taking large bags, but one way of increasing the size of
the bag would be to permit the sale of dead wild geese (prohibited in 1967) and to allow
commercial exploitation. 2. Relaxation of refuge restrictions. Geese are concentrated on
protected areas and some of these could be opened for shooting. However, this would be
unacceptable to private owners, and increased shooting activity on SSSIs would go against
Review of information, policy and legislation on species translocations
the provisions of the Wildlife and Countryside Act 1981. 3. Extending the shooting season.
Spring shooting harms future productivity.
Fish and shellfish
Measures for the control of aquatic organisms, other than plants, are not widely reported. In
general, control programmes are not implemented even against species such as zander or
crayfish. However control measures against alien crayfish are recommended by JNCC
(Palmer 1994).
Control of Crassula helmsii has been attempted using three techniques (Dawson & Warman,
1987): 1. use of a dark geotextile material to shade out the plants; 2. use of grass carp to graze
the plants; and 3) physical removal and the inhibition of regrowth. Covering the plant was
successful in the short term, but the shading material was progressively colonised by the plant
after longer periods. Herbivorous fish will eat the plant, but the nutrient cycling probably aids
regrowth. Removal of the plant, either mechanically or by hand, proved unsuccessful.
However, a combination of removal, followed by the suppression of regrowth, seems
suitable. Approved herbicides may be required for effective control for commercial or private
water bodies, but are unsuitable for use in nature reserves.
Rhododendron ponticum control would require large-scale clearance, followed by spraying of
regrowth. An estimate of the cost of a rhododendron control programme in the Snowdonia
National Park by Gritten (1988) suggested that 30 million pounds might cover the cost of the
initial clearance. Follow-up spraying programmes would add to the cost.
Dispersal of Heracleum mantegazzianum is almost entirely by seed, and so a control
programme would need to prevent plants setting seed (Dodd et al. 1994). Due to extensive
seed banks and possible long-term viability of seeds, any control programme would need to
have follow-up monitoring for at least 7 years after the initial control measures. Clearance of
stands is usually by using herbicides. Selective herbicides which are safe for use along
watercourses are ineffective in controlling giant hogweed (Sampson 1994). Glyphosate may
be used as a spot treatment, but all plants in the area need to be treated to prevent reseeding.
Also, re-treatment may be necessary, but this can be impractical and prohibitively expensive.
Cutting the plant is ineffective, as it will regenerate from substantial root reserves. Cutting
during seed set aids spread of seeds. The plant could be eradicated if a committed and coordinated control programme, using appropriate techniques, was implemented (Dodd et al.
Control measures used against Fallopia japonica in Wales include the following (Gritten
1988). 1. Digging out plants and rhizomes. This is labour intensive and fails to control the
plant without other control methods. 2. Cutting is labour intensive, but is effective as a longterm measure. 3. Selective herbicides are successful where loss of broadleaved species is not
a problem, but to eradicate the plant respraying would be necessary. 4. Non-selective
herbicides have been successful, but are not suitable for use near watercourses or for use on
Review of information, policy and legislation on species translocations
vegetation of conservation value. There are very few examples of eradication of this plant,
and successful control requires perseverance.
3.14.5 Regulation and risk assessment of introductions
Risk assessment
Because of the existing risk of adverse ecological impacts of alien species, the potential for
increased risk due to future climate and other environmental change, and the difficulty in
naming the distinguishing characteristics of a potentially invasive species, two necessities are
clear in considering future introductions to the UK:
to assess every proposed introduction of alien species to the UK;
to use a case by case approach to risk assessment.
This approach is characterised in the phrase 'guilty until proved innocent', which was applied
by Ruesink et al. (1995) to alien species. While some organisations oppose introductions
outright (e.g. RSPB - concerned with birds only), others provide guidelines on good practice.
The conservation organisation guidelines covering the introduction of non-native species are
thorough, and suggested assessment procedures, such as those by NCC (1990), ICES (1995)
and IUCN (1987) could be followed closely to ensure that harmful organisms are screened
out before introduction. The DOE and MAFF information requirements for licences to
release, market or keep in uncontained conditions (i.e. with the risk of escape to the wild)
alien species under the Wildlife and Countryside Act are very similar to these, and the DOE
risk assessment procedure is similar to the detailed approach taken to GMO release (see
Chapter 4). It is important to note that the guidelines and legislation concern single releases,
each new release of an alien species in the UK must be separately licensed.
Information requirements to allow informed decisions to be made concerning the potential
impact of the introduction on native biota must be based on a thorough understanding of the
biology and ecology of each candidate for introduction, and knowledge of the native species
at the receptor site and in the UK as a whole, and should consider: the potential for escape of
the alien from the receptor site; the potential for the alien to establish and spread in the wild;
assessments of likelihood and consequences of hybridisation with native species; ecological
consequences of establishment and spread in the wild; and potential for control and risk
management. As an illustration of the caution necessary because of our poor ability to predict
the consequences of an introduction, except on a detailed case by case basis, DOE are not
contemplating any form of Fast Track licensing - as is being used for some GMO release
consents - for releases under the Wildlife and Countryside Act (W. Parish pers. comm.).
Omissions and proposed additions to the legislation concerning introductions
One criticism of the current procedures under the Wildlife and Countryside Act 1981 is that
the Act does not make provision for compulsory pre-introduction testing of potential
environmental impacts of introduced species. This type of screening would provide a great
deal of relevant information, and may be essential to assess properly the risks of some
Review of information, policy and legislation on species translocations
proposed introductions (IUCN 1987). However, DOE do incorporate their own risk
assessment into the licensing procedure, despite it not being required by the Act (W. Parish
pers. comm.), and the requirement for a further risk assessment may, in some circumstances,
be imposed.
The EC Birds and Habitats Directives, the Bern Convention and the Convention on
Biological Diversity all make statements concerning the introduction of alien species, with no
taxonomic restrictions. However, while the Wildlife and Countryside Act 1981 and the
Wildlife (Northern Ireland) Order 1985 prohibit the release of any animal not ordinarily
resident in the UK, they make no general provision against the release of alien plant species.
The only specific provisions in UK law against the introduction of alien plants relate to the
short lists of named plant species in the Schedules 9. There is no good conservation reason
for this omission; as we have shown, aliens plants pose no less a threat than animals
(although the omission probably results from a wish to not restrict introduction of plants for
farming and horticulture). There is also a massive influx of alien plant species to the UK.
About 40% of the species described in Stace's (1991) British flora were not native, and
Nelson (1994) reports that there are about 55,000 species or varieties of alien plants in British
gardens - over 15 times the total number of British native plant species. Nelson (1994) also
states that this number is increasing as more plants are imported. New cultivars and species of
crop and forestry plants are being introduced with no controls with respect to conservation;
whereas genetically modified crop plants are subject to strict controls by virtue of the
molecular techniques used, rather than any greater potential for harm from such plants (see
Chapter 4).
Problems with current procedures on introductions
There is great concern, especially with regard to alien birds (either non-resident or listed on
Schedule 9), that there are a large number of illegal releases or escapes from private
collections (e.g. RSPB - G. Williams pers. comm.). This is thought to be a fundamental
problem in that it maintains and establishes feral populations of birds in the UK, and this was
a major discussion point at the recent BOU/JNCC conference on Feral and Introduced Birds.
Thus, the problem is not in regulating imports or releases applied for under the Wildlife and
Countryside Act 1981, but enforcement of relevant legislation and action against releases
without a licence. Holmes & Simon (1995) report that they are aware of only one prosecution
under Section 14 of the Wildlife and Countryside Act. There seems to be a need for the
conservation organisations to consult with DOE regarding implementation of the legislation.
Schedule 9 of the Wildlife and Countryside Act 1981 does not constitute a complete list of all
problem alien species in Great Britain (certainly not of plant species, see above). Certain
non-native animal species have established breeding populations in the wild, but are not on
Schedule 9. Such species, e.g. the Muscovy duck Cairina moschata could be considered
'ordinarily resident' and therefore, a release into the wild would not be an offence under
Section 14 (see Holmes & Simon 1995). Holmes & Simon (1995) call for these to be added
to Schedule 9. G. Williams (pers. comm.) of RSPB calls for several other additions of birds
RSPB considers to be on the brink of, or have started, forming feral populations in the UK:
barnacle goose Branta leucopsis, greylag goose Anser anser, snow goose A. coerulescens,
pink-footed goose A. brachyrynchus and red-crested pochard Netta rufina.
Review of information, policy and legislation on species translocations
There is a feeling that the licensing procedure for Section 14 introductions is confused. The
licensing is split between MAFF Fisheries Division and DOE (until recently DOE European
Wildlife Division dealt with licensing, but the Toxic Substances Division has taken this over
(F. Grant pers. comm.)). Although DoE always consult MAFF over releases of non-native
species, MAFF do not consult DoE on releases of fish or shellfish (W. Parish, D. Linskey
pers. comm.). Both have comprehensive information requirements, but a more coordinated
approach would seem sensible. This may also aid the statutory conservation agencies, which
also require a more coordinated and structured approach to assessments of proposed
introductions of all types.
Another complaint is that ACRE consists of experts on GMOs who may not be the best
qualified to judge introductions of alien species. This is a fair criticism, and although one
world expert on plant introductions is on the committee (A. Gray of ITE), there are no experts
on insect biological control agents or vertebrate introductions.
3.14.6 Control of alien species - problems with legislation
Apart from releases to the wild, conservationists are concerned with the control of wild
populations of problem alien species. The EC Habitats Directive and the Convention on
Biological Diversity both include provision for control of alien populations. However, it is
here that problems arise with UK legislation. While it is an offence to release or allow to
escape the species described in section 14 of the Wildlife and Countryside Act, there is no
explicit provision for their control in the wild. In fact, once a bird is viewed as 'ordinarily
resident' in Great Britain it is protected under Section 1 of the Wildlife and Countryside Act.
Other animal species are only protected if they are named in Schedules, and so there is little
problem with the control of alien animals other than birds. The protection afforded to birds
involves a prohibition against killing, injuring or taking any wild bird, taking, damaging or
destroying the nest of a wild bird or taking or destroying the egg of any wild bird.
It is not certain when a feral bird species becomes 'ordinarily resident', but the general
interpretation is that this is when the species has bred in the wild (e.g. Holmes & Simon
1995). Control measures, which involve infringement of the Section 1 prohibitions, against
any resident bird species can only be carried out under a Section 16 licence. Such a licence
can be issued for a number of purposes, of which those relevant to conservation are the
protection of wild birds and the prevention of spread of disease. Therefore, licences could be
issued to allow control of resident aliens which are having such effects. However, there are
two problems with the current legislation.
1. The licence is issued to allow limited control to meet a specific objective. Section 16 is not
intended to allow for general management or culling of a species in Great Britain. This causes
problems in running a large-scale control programme against a problem species. However,
RSPB believe that licences issued allowing the killing of Canada geese are being used for a
large-scale culling programme (G. Williams, pers. comm.). Because the law does not
discriminate between feral birds and native species, RSPB are concerned that this sets a
precedent which could be applied to native birds, e.g. for economic reasons of control.
Review of information, policy and legislation on species translocations
2. A second problem arises from the limited list of purposes for Section 16 licence. The
reason for the wish to control ruddy duck is its effects on white-headed ducks in other
European countries. There is no perceived effect on the native bird populations in Britain and
thus, there can be no provision for its control under current law. A current experimental
control project is being carried out under a scientific licence.
Similar problems could apply to the control of resident alien plants, as there is a general
prohibition in the Wildlife and Countryside Act against uprooting any wild plant (although
one can do so with the landowner's permission for species not listed in the schedules).
The same problems apply under the Wildlife (Northern Ireland) Order in Northern Ireland.
Therefore, changes are needed in the legislative provision for control of problem feral bird
species (and, possibly, plants). The simplest change would involve an implementation of the
Habitats and Birds Directives and the Convention on Biological Diversity to allow general
protection of flora and fauna and their habitats, possibly by extending the provisions of
Section 16 licensing.
The Weeds Act 1959 forms a possible legislative basis for the control of named alien plant
species. It names five (native) plant species - Cirsium vulgare, C. arvense, Rumex crispus, R.
obtusifolius and Senecio jacobaea - as 'injurious weeds'. If any of these species are growing
on a piece of land the Minister of Agriculture Fisheries and Food (who may act jointly with
the Secretaries of State for Scotland or Wales; the Act does not apply in Northern Ireland)
may serve notice to the owner or occupier to take action to prevent the spread of the species.
It is an offence not to comply with the notice. The Minister can add other species to the list of
five. Therefore, it may be possible to add certain problem alien species to the Weeds Act
1959 and to provide a mechanism for their control. However, two problems may arise.
Firstly, there may be resistance to adding species which are not agricultural weeds, but cause
only conservation problems; i.e. such additions may be against the spirit of the Act. Secondly,
the onus for control is placed on the owner or occupier of the land and this may seem unfair if
the alien is an invasive species which is hard to control (e.g. Fallopia japonica)
3.14.7 Further research
The conclusion that it is hard to make general predictions concerning the characteristics of a
successful invader indicates that future research should concentrate on specific cases rather
than the investigation of general theories. This was, in fact, the conclusion of a reviewer of
the SCOPE synthesis (Gilpin 1990 reviewing Drake et al. 1989). However, general work,
such as that by Williamson & Fitter (1995a, b) should continue in order to pursue possibilities
for a predictive approach. The research on specific cases should be of two types.
Investigating the potential risk of proposed introductions, by answering the questions
asked in the risk assessments such as those given in the guidelines from NCC, IUCN
and ICES (see section 3.3), for the DOE or MAFF licensing procedures (see section
3.4.1), and for specific types of introduction such as biological control (see section
3.6.3) or fisheries (see section 3.5.3).
Review of information, policy and legislation on species translocations
This could use experimental introductions (as are used for GMOs - see Chapter 4),
assessment of comparable case studies, and modelling. Modelling of invasion and spread is
advanced (see Williamson 1989 for a thorough review) and of further use will be the current
developments in using GIS-based procedures to determine the distribution of appropriate
habitat for a species and thus model the anticipated rate and pattern of spread (e.g. Carey &
Brown 1994). Modelling of gene flow and thus possibilities of hybridisation is also advanced
(see Gliddon 1994).
Assessment of current changes in the status of alien species resident in the wild and
methods for control, if necessary.
The response to problem alien species in the UK is generally reactive - responding to a
problem that has arisen - rather than proactive - assessment of the potential for problems to
develop and taking action to contain the problem before it arises. Coblentz (1990) makes the
point that research into a problem with an alien species after it has arisen may sometimes
'provid[e] information only for the eulogy'. Models of invasions show that if action is taken
while populations are small, few and restricted in distribution, it is very much more effective
at restricting spread and persistence, and may mean the difference between success and
failure (Williamson 1989). The proactive approach restricts damage to biodiversity and is
probably cheaper - containment or control of the ruddy duck to prevent its release or
subsequent spread would have been easier and cheaper than the current programmes, and
there would have been no effects on the white-headed duck.
It may be a problem to obtain funding for such a approach. A paper by Feare at the at the
BOU/JNCC conference on Feral and Introduced Birds 1995 discussed the problems with
obtaining funding for research on the ring-necked parakeet because it is not yet perceived as a
problem. Major funding may only be forthcoming if the bird starts to cause major economic
or conservation problems.
One proactive approach would be to restrict or stop the further release of any alien species,
but as a general approach this may be against the interests of other members of society.
Another approach is to aim to control or eradicate all alien species established in the wild, but
this seems over-cautious and expensive and would probably meet with widespread
opposition. If one combines the precautionary principle with a policy of no action until a
problem is predicted or detected in it early stages, one can develop a sensible proactive
policy. However, this requires a great deal of work. One essential method would be the
monitoring of abundance and distribution of species. The BOU Records Committee wish to
develop the British and Irish bird list to take account of feral species (J. Marchant, BTO at the
BOU/JNCC conference on Feral and Introduced Birds 1995) and the BSBI is considering
development of an alien plant register (Ellis 1994). However, this may expose a problem only
rather late in the day. To allow early detection or, even prediction, of problems of spread
and/or negative effects on biodiversity, specific studies of the ecology of species and the
mechanisms of their spread into new areas are needed, along with assessments of their effects
on biodiversity. Population studies would also help the rapid development of control
measures which can be implemented immediately problems are perceived; thus avoiding the
loss of time in gathering ecological data after the species has become a problem.
Review of information, policy and legislation on species translocations
Such detailed studies would be impossible for all alien species in the UK. Therefore, a
detailed review of the ecology and status of all such species (as carried out by Eno (1995) on
UK marine organisms) and small-scale studies to gather necessary information, would allow
accurate targeting of potential problem species which would justify more detailed
investigation. The distribution records held by the Biological Records Centre could be used to
determine changes in the distributions of alien species (see Harding 1990).
3.15 Summary conclusions concerning introduction of species not native
to the UK
Alien species are, and have been, introduced to the wild in the UK through deliberate
releases or escapes from fish and shellfish stocking, biological control, wildfowl and
game stocking, amenity and ornamental planting or stocking, keeping of pets or
domestic animals, forestry and crop plantings and keeping of fur animals.
A further category is accidental introduction, comprising of species which have been
brought into the UK accidentally. This is usually through shipping and species are
transported in the ballast, in cargoes or as stowaways.
Introduced species include birds, mammals, herpetofauna, invertebrates, vascular and
lower plants, and micro-organisms. These inhabit the whole range of habitat types
found in marine, freshwater and terrestrial conditions.
Many species brought into the UK cannot establish populations in the wild, and many
of the species that do establish in the wild cause no clear adverse effects on
biodiversity. However a small proportion of introduced alien species cause
conservation problems, and some cause very great problems.
Adverse effects on population size and persistence of native flora and fauna can occur
through: direct effects through competition, herbivory, predation and the alien species
itself becoming a food item; indirect effects through habitat alteration; spread of novel
pathogens and parasites; and hybridisation with native species. Examples from other
countries show that adverse effects tend to be greater on small oceanic islands.
There are few legislative provisions in the UK for the control of alien species which
are established in the wild. The Wildlife and Countryside Act 1981 and the Wildlife
(Northern Ireland) Order 1985 cover some aspects, but there is a need for new
legislation, perhaps by implementing provisions in the Habitats and Birds Directives
and the Convention on Biological Diversity. Guidelines by conservation organisations
would be irrelevant without such changes.
It is unnecessary to control or eradicate all alien species in the UK. Only certain
species are causing conservation problems. However, other aliens resident in the UK
may cause problems in the future if they increase in numbers, perhaps due to
environmental and climate change. Therefore these changes must be monitored and
provision made in any legislation to include these aliens if necessary.
Review of information, policy and legislation on species translocations
Control or eradication of widespread and common problem aliens is extremely
expensive and difficult. Provision for prompt and early control of developing problem
species is necessary.
It is difficult to make generic predictions of which types of species may cause
conservation problems if newly introduced to the UK, or to new regions in the UK.
Each proposed new introduction to the UK or to parts of the UK must, therefore, be
assessed individually.
There are generally good legislative controls on the keeping and release of alien
species which are not already resident in the wild in the UK, but only for animals.
There should be similar regulation of alien plant introductions, perhaps through
implementation of provisions in the EC Birds and Habitats Directives, the Bern
Convention and the Convention on Biological Diversity.
The UK legislation also requires clearer procedures for assessing the risk from such
introductions, based on a standardised risk assessment approach. Such an approach is
being developed by DOE. A second improvement would be a requirement for preintroduction testing for potential impacts. Guidelines by the statutory conservation
agencies should follow similar criteria.
The controls on keeping and release of alien species which have established wild
populations (i.e. they are technically 'wild') are much weaker. Many such species
could be added to Schedule 9 of the Wildlife and Countryside Act 1981 (or Wildlife
(Northern Ireland) Order 1985). There is also a need for more rigorous enforcement of
the controls against releases of species already on Schedule 9.
Review of information, policy and legislation on species translocations
4.1.1. Background
Genetic modification through the use of recombinant DNA has developed gradually through
the latter part of this century, but the technology has undergone an explosion over the last 1520 years. New techniques, and refinements of existing techniques, of modification are
developing rapidly under the impetus of a massive investment into research by industry and
governments throughout the world. A DOE (1994a) review shows the number of publications
on the BIDS bibliographic database with titles containing the words transgenic or transgene
increased from 0 in 1981 to 1236 in 1993. Our own search found that this number increased
to 1465 in 1994 and was already 826 by the end of June 1995. The UK is one of the centres
of such developments.
These advances in 'biotechnology' have caused great concern among the public, government
and conservationists about the risks and impacts of the use of GMOs. The concerns are of
three types: ethical questions, risks to human health and impacts on the environment. Health
considerations and ethics are paramount to many people, but we shall consider only the
potential environmental hazards.
These concerns have led to a great variety of reviews (e.g. Hoffman 1988, CEP 1989,
Ellstrand & Hoffman 1990, Mooney & Bernardi 1990, Smit et al. 1992, FOEFL 1993,
Raybould & Gray 1993b, c, van der Meer 1993, Harding & Harris 1994, Rogers & Parkes
1995, and the whole issue of Molecular Ecology 1994 3:1) and a number of research
programmes (e.g. PROSAMO in the UK and BAP/BRIDGE in Europe) into possible
impacts. However, as we describe below, there are few data on the impacts of full releases (as
opposed to heavily controlled small-scale releases) of GMOs and therefore most discussion
concerns potential risks. The subject is also highly technical and a full assessment of possible
risks would require detailed analysis of all the current and developing procedures used in
producing GMOs. We will aim in this review to summarise the policies and legislative
controls pertaining to GMOs, to give an outline of the use of GMOs in the UK, and to
summarise the current discussion on the environmental risks. As a background to this
discussion, we will describe both the procedures used in developing GMOs and the types of
GMOs currently used.
Review of information, policy and legislation on species translocations
Genetic modification
The aim of genetic modification is to insert DNA into the nuclear genome of a eukaryote cell
(techniques are being developed for insertion into chloroplasts), or into a prokaryote cell, to
produce organisms which show stable inheritance and expression of the introduced
sequence(s). Techniques are used, and are continually being developed, to achieve great
precision and control of both the modification procedure and the final expression of the
inserted DNA. The methods involved are highly technical and for this review we will
summarise them very briefly (see also Old & Primrose 1985, Draper & Scott 1991, Raybould
& Gray 1994b). Genetic modification involves three steps: the isolation of appropriate DNA
sequences from the donor organism and the modification of the sequence(s) to ensure
appropriate expression; the introduction and integration of a DNA sequence(s) (the
'transgene') into the recipient organism ('recombination'); and production of a modified
Isolation of DNA sequences
The DNA of the donor organism is fragmented using restriction enzymes which cut the DNA
at specific sequences. Each fragment is inserted into a cloning vector, which is usually a
plasmid (bacterial DNA which can replicate independently of the chromosome) or a virus,
and then the recombinant vector undergoes amplification within a bacterial host. This
produces many copies, or clones, of the fragments which are screened to detect and isolate
the desired clone(s). The isolated clone(s) is sequenced to allow identification of its structure.
The isolated DNA sequence is often further manipulated to modify the expression of the
inserted gene, e.g. by fusion of control sequences from different genes.
Insertion of DNA sequences
Procedures may be used to integrate the DNA sequence into the recipient chromosome.
However, the DNA does not have to be integrated into the chromosome in prokaryotes, and
may be introduced into the bacterial cell within a vector such as a plasmid. This
'transformation vector' usually includes a marker gene to allow detection of transformed cells.
In eukaryotes the DNA must usually be integrated into the nuclear genome. As for bacteria,
the sequence is transferred within a transformation vector, usually with a marker gene. A
number of biological or physical vector systems can be used to integrate the new DNA.
Biological systems include transposition or the use of retroviruses, Agrobacterium or
bacterial prophages. These systems use the ability of the vector to transfer DNA into cells and
to incorporate it into the genome. Direct, 'physical' methods of transfer include: heat shock or
electric shock ('electroporation') to disrupt the cell membrane and, in plants, the cell wall to
allow physical introduction of the vector; chemical stimulation of uptake into cells (e.g. using
Review of information, policy and legislation on species translocations
polyethylene glycol); microinjection into the cell; and projectile bombardment through pores
in the cell membrane ('biolistic' methods). Most of these are used only on plant tissue where
the cell wall must be crossed. Microinjection, electroporation or viral infection are the
methods usually used to traverse animal cell membranes.
Production of a modified organism
Modification of a bacterial cell results in a modified organism. In animals the oocyte is
modified to produced a modified adult. Usually tissue culture is used on plants, although a
number of other methods are being investigated (see Raybould & Gray 1993b).
4.1.3 Types and uses of GMOs
A major consideration in the development of GMOs is that specific single traits can be
introduced into organisms without the need for back-crossing programmes to remove
unwanted genetic linkages. Most modifications have involved traits controlled by single
genes and have been developed for a variety of uses (some of the information below is
obtained from the DOE public register of consents to release). However only few of these
GMOs have been involved in releases of any form in the UK (see below).
Bacteria and viruses have been genetically modified for use in control of crop pests and
diseases and enhancement of plant growth. The ability to produce insecticidal proteins has
been introduced into Pseudomonas and Clavibacter bacteria from Bacillus thuringiensis, to
improve crop pest control. The virulence and host range of baculoviruses have been increased
by modification to improve their effectiveness as biopesticides. A gene causing production of
scorpion toxin was introduced into Autographa californica nuclear polyhedrosis virus to
improve its use as an insecticide. Pseudomonas syringae damages crop plants, and a control
agent produced by genetic modification was used to generate a non-damaging strain which
also outcompetes the wild-type. Future and developing uses of genetically modified
microorganisms may include treatment of effluents and pollution and mineral extraction (see
also Shorrocks & Coates 1993, van der Meer 1993). Viruses are being modified for use as
vaccines against a number of animal diseases (Fenner 1990)
Although many of the plants that have been modified are arable crop species, others include
plants used in forestry, horticulture, soft fruits and market gardening. The species commonly
used in the UK are potato Solanum tuberosum, sugar beet Beta vulgaris, oilseed rape
Brassica napus oleifera and maize Zea mays. Other species of which GMOs are being
developed are wheat Triticum aestivum, barley Hordeum vulgare, rye Secale cereale,
strawberry Fragaria × ananassa, blackcurrant Ribes nigrum, raspberries and blackberries
Rubus spp, carrot Daucus carota sativa, lettuce Lactuca sativa, tobacco Nicotiana tabacum,
tomato Lycopersicon esculentum, pepper Capsicum annuum, beans Phaseolus vulgaris and P.
coccineus, vegetable cucurbits Cucumis sativus, C. melo and Cucurbita pepo, sunflower
Review of information, policy and legislation on species translocations
Helianthus annuus ryegrass species Lolium spp, poplars Populus spp, elms Ulmus spp, and
various conifers (Raybould & Gray 1993b).
Modifications include: resistance to herbicides (to allow use of the herbicide in weed
control), insects, and viral, bacterial and fungal pathogens; changing the quality of seeds, fruit
other plant products; introduction of male sterility (to prevent gene escape through pollen);
changing the appearance and longevity of ornamental flowers, incorporating the production
of non-plant compounds; and increasing tolerance to various environmental stresses (see
Raybould & Gray 1993b, ACRE 1994a). Tomatoes have been modified to exhibit resistance
to the herbicide glyphosate and glufosinate-resistant oilseed rape has been produced. Insect
resistance has been established in poplars and potatoes using a gene from Bacillus
thuringiensis. Tomato has been transformed to express the coat protein of tomato mosaic
virus which renders it resistant to the virus. Potato has been made resistant to two viruses in
the same way. The type of oil produced from oilseed rape, soybean, safflower and other oil
seed plants has been modified using genes from other species or varieties. The softening of
tomatoes during ripening has been slowed by modifying the production of an enzyme in the
fruit ('Flavr Savr' tomatoes). Potato has been modified to divert carbohydrate metabolism
sugar into starch production. Tolerance of cadmium pollution has been incorporated into
oilseed rape.
The modifications involving resistance to herbicides, insects, and pathogens and changes to
characteristics of the crop product are the commonest and most advanced in the UK. This can
be seen in the description of releases approved by ACRE (see below).
Many fewer animal species have been modified than plants or microorganisms. Those that
have been modified include nematodes, sea urchins, insects, fish, amphibians and mammals.
These have been modified for economic use and experimental purposes. Fish which are
modified in the UK to increase their growth rate include: Carp Cyprinus carpio, salmon
Salmo salar, tilapia Oreochromis niloticus, and rainbow trout Oncorhynchus mykiss. Other
developing modifications of fish include tolerance of environmental factors such as
temperature, oxygen levels and salinity (see DOE 1994a, Berkowitz & Kryspin-Sorensen
1994). Insects such as Drosophila melanogaster have been modified for biochemical and
physiological research. Mice have been modified for laboratory studies. Rabbits, pigs, cows
and sheep have been modified for a variety of medical and experimental purposes. For
instance, sheep have been modified to produce a human protein for use in treating
haemophiliacs (Shorrocks & Coates 1993).
Summary of guidelines and policies
This section summarises the policies (statements of position) and guidelines (recommended
procedures) of conservation and other organisations. These are not implemented by law, but
represent, at most, agreed codes of conduct.
Review of information, policy and legislation on species translocations
In many cases the guidelines have been condensed although some sections are transcribed
verbatim. Some terms have been changed to conform with our definitions.
4.2.1 UK Guidelines
Nature Conservancy Council (1988a) Wildlife translocations in Great Britain
There are no comprehensive guidelines on GMO translocations by NCC or its successors.
The proposed guidelines for introductions in NCC (1990) were intended to cover the release
of GMOs, but made no separate provisions (see Chapter 3). However, the earlier draft policy
in NCC (1988a) made the extra provisions that for GMOs the following should be assessed:
the nature and stability of the genetic modification;
the ecology of the GMO;
the ability of the GMO to transfer genetic material to native species or races;
the host range or pathogenicity of the GMO.
International Guidelines
IUCN. (1987). The IUCN position statement on translocation of living organisms
In the section concerning introductions this document stated that where genetically altered
microorganisms are introduced into areas where they did not formerly exist, the same
procedures should be used as for other introductions (see Chapter 3).
Worldwide Fund for Nature (1995). Biosafety and use of genetically engineered organisms
This Position Statement was the product of discussions among parties to the Convention on
Biological Diversity (e.g. WWF 1994). WWF (1995) state concerns on the use of GMOs
(called GE(ngineered)Os by WWF) relating to:
ecological effects;
impacts on genetic diversity;
economic implications;
the appropriateness of GMO applications ('there is no evidence that the use of GMOs
is superior to non-GMO alternatives').
WWF stated that
'the component of high risk associated with biotechnology ... necessitates that the
precautionary approach should be adopted', and accordingly affirm the need for a
biosafety protocol for the Biodiversity Convention and to be adopted by all
Review of information, policy and legislation on species translocations
A biosafety protocol would be based on the following.
The principle of Advanced Informed Agreement (i.e. consent to GMO translocations
only with sufficient information on risks).
Designation of national authorities responsible for biosafety.
Environmental impact assessment, including an assessment of the need for GMOs,
and of non-GMO alternatives.
Risk assessment and management based on a case-by-case and step-by-step basis.
Reporting, inspection, monitoring, transparency and information exchange, including
labelling of GMOs and their products.
Provision of information to the public, and public participation in decisions.
A regime for assignment of liability and provision of compensation.
Support and assistance for national capacity building in the field of biosafety.
An international clearing-house to facilitate coordination between governments and
national authorities.
International Council for the Exploration of the Sea (1995). ICES code of practice on the
introductions and transfers of marine organisms 1994
ICES (see Chapter 3 for a background) included separate guidelines for the translocation of
genetically modified fish, molluscs, crustaceans and plants for marine aquaculture. They
included cell fusion into their definition of genetic modification. ICES state that there is little
information on the genetic, ecological or other effects of GMO release. Therefore:
member countries should legally regulate GMO translocations, and include mandatory
licensing for the utilisation of GMOs;
the ICES should be notified, and given an assessment of the environmental
consequences, of any GMO releases;
where feasible, initial releases of GMOs should use reproductively sterile organisms,
to minimise impacts on the genetic structure of other populations;
there should be research into the ecological effects of GMO release.
North-east Atlantic Commission (1995). Introductions and transfers including the
amendments proposed by the European Union
These guidelines for translocation of salmon Salmo salar for restocking include the following
Transgenic Atlantic salmon, and other salmonids containing genetic material from
Atlantic salmon, should not be permitted in the Commission area except in secure
self-contained facilities.
However, they state that there is no unanimous position at this stage within NEAC and the
Review of information, policy and legislation on species translocations
Summary of legislation
This section summarises UK and EC legislation and international conventions which are
relevant to controlling releases of GMOs.
International Conventions and European Legislation
The Convention on Biological Diversity calls on each party to regulate, manage and control
the risks associated with the use and release of 'living modified organisms resulting from
biotechnology'. These risks involve threats to biological diversity. There is an accepted need
for a Protocol to the Convention to deal with these issues (see de Klemm 1995) but no
agreement has been reached.
The Organisation for Economic Cooperation and Development (OECD) set out a basis for
risk assessment of GMO use in the 1986 document on Good Development Principles
The 1982 United Nations Convention on the Law of the Sea, enforced in 1994, requires that
the member states take all measures necessary to prevent and control the intentional or
accidental introduction of 'new species' (i.e. GMOs) which could cause harm to the marine
EC Directives 90/219, on contained use, and 90/220, on deliberate release, laid out a series of
regulations which were implemented in the UK by the Environmental Protection Act 1990
and its Regulations.
Recommendation R(92)9 of the Council of Europe on the Potential Ecological Impact of the
Contained Use and Deliberate Release of Genetically Modified Organisms built upon the EC
Directives 90/219 and 90/220, recommending that long-term ecological effects of GMOs be
researched and be taken account of in risk assessments (see also Tachmintzis 1994).
The Council of Europe 'Lugano' Convention on Civil Liability for Damage resulting from
Activities Dangerous to the Environment, signed in 1993, includes GMOs in its definition of
dangerous substances. This establishes a 'Polluter Pays' principle in relation to environmental
and other damage resulting from pollutants and establishes a series of procedures following
this principle.
UK Legislation - Environmental Protection Act 1990
Part IV of the Environmental Protection Act 1990, implementing EC Directives 90/219 and
90/220, has two main sets of regulations; the Genetically Modified Organisms (Contained
Review of information, policy and legislation on species translocations
Use) Regulations 1992 and the Genetically Modified Organisms (Deliberate Release)
Regulations 1992. The Genetically Modified Organisms (Deliberate Release) Regulations
1995 have recently been published and these implement EC Directive 94/15. Risk
assessments and notification to the Secretary of State are required by anyone wishing to
import, keep, release or market GMOs. The Secretary of State can prohibit dealings with the
GMO or require consent to be obtained before any of these actions are carried out. Any
deliberate release of GMOs requires the consent of the Secretary of State. It is the duty of the
applicant to notify the statutory conservation agencies separately. At the moment DOE
consults the statutory conservation agencies on all releases except fast track releases.
Release is defined as the situation where the organism(s) is no longer subject to physical,
chemical or biological barriers which prevent the organism(s) entering the environment, the
production of descendants which are not contained or which render the organism(s) or its
descendants harmless.
The precautionary principle is incorporated into the legislation. 'Harm' and 'damage to the
environment' can be given broad definitions to allow consideration of effects on the natural
The application for release must include information on the proposed release and its potential
effects, under almost 90 headings, and a risk assessment. Many applications are for one
release of a single type of GMO. However, under the 'simplified' procedures one application
can be made, and one consent given for releases of one or more types of GMO at one or more
sites within a specified time period (Genetically Modified Organisms (Deliberate Release)
Regulations 1992 and 1995). Consents given are always subject to conditions and are always
for a limited period of time. Public registers are held of notifications, applications, consents
and notices of termination of the release. Certain information on the release must also be
advertised After consent and release the Secretary of State must be kept informed of any risks
involved and which appear more serious than at the time of consent. The Secretary of State
can act immediately in the case of imminent danger to cause any GMO to be rendered
Each proposed application is assessed for risks to the environment and human health on a
case-by-case basis. The risk assessment aims to identify: if there are any hazards posed to the
environment and how they could be realised; the magnitude of harm if a hazard is realised;
and the likelihood of the hazard being realised. Consent applications are assessed by the
Advisory Committee on Releases to the Environment (ACRE) who advise the Secretary of
State on whether the release should be allowed and of any conditions that should be imposed.
ACRE consists of 12 members representing a variety of backgrounds, such as: university
academics, scientists from research institutes (e.g. Institute of Terrestrial Ecology, Scottish
Crop Research Institute, etc.), scientists from the private sector (e.g. Zeneca Seeds, the
Welcome Foundation) and the Director of the Green Alliance. The Secretariat of ACRE
consists of DOE staff.
Applications for consent to release GMOs fall into two categories: standard, which are
discussed in full committee; and fast track, dealt with by the Secretariat only.
Review of information, policy and legislation on species translocations
Fast track releases
Fast track procedures (ACRE 1994b) are the result of a decision by ACRE in 1993 that, in the
light of experience, they were able to provide generic advice for releases of certain types of
GMOs. These fall into three types.
Low hazard GMOs
These are all genetically modified plants which are considered to not pose a risk to the UK
environment and therefore special control measures are not needed. Seven named
modifications of maize Zea mays, tobacco Nicotiana tabacum, tomato Lycopersicon
esculentum, pepper Capsicum annuum, beans Phaseolus vulgaris and P. coccineus, vegetable
cucurbits Cucumis sativus, C. melo and Cucurbita pepo and sunflower Helianthus annuus are
allowed because these species are unable to form feral populations in the UK, there is no
potential for gene flow to wild relatives, and the modification does not affect these
characteristics or pose any other risk. Fewer modifications of potato Solanum tuberosum,
oilseed rape Brassica napus or sugar beet Beta vulgaris are allowed on the fast track
procedure because of the increased risk of gene flow and establishment of feral populations.
Low risk GMO releases
Certain types of GMO release are considered low risk if appropriately managed. If certain
criteria for the release are accomplished the following can be released: any plant species not
containing cloned viral nucleic acid, which cannot flower (due either to management or
environmental factors), and where the biological material is contained and disposed of at the
end of the release; or plants not containing cloned viral nucleic acid, and which have been
made male sterile and this cannot be reversed in the conditions of the release.
Repeat applications
A release qualifies if it is a repeat of one previously consented to (i.e. the description of the
GMO is the same and the same risk management procedures are in place). These are limited
in size so that releases over an area greater than 2ha do not qualify (W. Parish pers. comm.;
ACRE 1994b).
Applications to release GMOs
An application to release must contain extremely detailed information on the GMO, the
release and possible effects of the release or unexpected spread of the GMO. Guidelines on
releases and the information needed on applications are given in Guidance Notes on
microorganisms (ACRE 1994c), baculoviruses (ACRE 1995a) and plants (ACRE 1994a), and
one for fish is to be published shortly. Detailed information is required on each of the
following (see ACRE 1993, some of the headings are repetitive and we have deleted these):
Review of information, policy and legislation on species translocations
Information relating to the organisms
the applicant and qualifications and training of all individuals involved in the release;
characteristics of the 'donor organism' ('the organisms from which the genetic material
to be inserted is taken'), the 'parental organism' ('organisms used in cell fusion
experiments') - where relevant, the 'recipient organism' ('an organism which
undergoes genetic modification'), to include (where relevant) for each taxonomy,
phenotypic and genetic markers - e.g. genetic maps, genetic markers, etc.
identification and detection techniques and their sensitivity (particularly for
the geographic distribution, natural habitat, and biotic interactions,
potential for genetic exchange with other organisms,
verification of its genetic stability and factors affecting this,
pathological, ecological and physiological traits (hazards to human or other
health, generation time and lifecycle, survivability, detailed pathogenicity,
antibiotic resistance),
history of genetic modification if any organisms are GMOs;
the degree of relatedness between the donor and recipient or between the parental
characteristics of the vector (where genetic transformation is vector-mediated), to
include its nature and source,
a genetic or restriction map of the vector, transposons and other sequences
used to construct the GMO,
the genetic transfer capabilities and estimated frequency of mobilisation
with methods for determining the presence of the vector in the host,
the amount of DNA on the vector which is excess to that needed for the
intended function;
characteristics of the modification, including modification methods,
derivation of the insert and methods of introduction into the recipient,
details of sequences on the insert additional to those carrying out the
required function,
a genetic or restriction map of the insert;
characteristics of the GMO, including new traits or lost traits as a result of the modification,
the vector or donor sequences remaining in the GMO,
the stability of genetic traits,
the rate and degree of expression of the new genetic material,
the activity of the gene product,
techniques for identifying the GMO and their sensitivity,
history of previous releases of the GMO,
hazards to human health of the GMO or its products;
Review of information, policy and legislation on species translocations
Conditions of release
description and purpose of the proposed release dates,
site preparation,
site size,
release methods,
quantity of organisms to be released,
foreseeable disturbances to the site,
protection for workers,
post-release site treatment,
techniques for elimination or inactivation of the GMOs at the end of the
release period;
the environment of the release site location of the site,
proximity and other significant biota (including target wild species),
proximity to protected areas,
human factors (proximity to humans, size of local population, local
economic activities, distance to drinking water supplies),
regional climate,
local geography, geology, pedology,
local flora and fauna, domesticated and wild,
description of ecosystems of the release site and adjacent sites,
comparison of the release site with the natural habitat of the recipient
The organisms and the environment
features of the GMO affecting ability to survive, multiply and disperse and effects of
environmental change on these characteristics;
the ecology of the GMO, especially in relation to the unmodified donor;
the capability of post-release transfer of genetic material to or from GMOs;
the likelihood of natural selection leading to the expression of unexpected traits;
measures to monitor genetic stability;
potential routes of dispersal;
description of ecosystems to which the GMO could be dispersed;
potential environmental impact potential for excessive population growth,
competitive advantage in relation to unmodified recipient or parental
potential effects on other organisms,
likelihood of post-release shifts in biological interactions,
involvement in biogeochemical processes;
Review of information, policy and legislation on species translocations
Monitoring, etc.
methods (and their specificity, sensitivity and reliability) for tracing GMOs and
monitoring their effects;
methods for detecting transfer of genetic material;
monitoring duration and frequency;
methods to avoid GMO spread;
security measures against entry to the site by humans or other organisms;
type of waste from the release, and risks and planned treatments of the waste;
emergency response plans for controlling spread, eradicating and disposing of
organisms, isolating the area affected by the spread, etc.
This must be followed by a risk assessment of each potential hazard associated with the
release, which is based on the information given according to the above criteria. Examples of
risk assessments are given in ACRE (1994a, c).
Marketing GMOs
The above procedure governs spatially and temporally small-scale, usually experimental,
releases of GMOs. Widespread release of a GMO through sale to users (e.g. to farmers to
plant economic crops) requires a consent for marketing. Application must be made to a
European Community Member State's Competent Authority, which, in the UK, is DoE.
Applicants themselves choose the appropriate Competent Authority, depending on the GMO
and where it will be marketed; for example France is usually chosen for maize and the UK
for oilseed rape. In the UK ACRE consider the application first, and, if they are positive, the
Competent Authorities in other EC countries and Norway consider the application and
ACRE's advice. Marketing anywhere in the EC or in Norway must go through the same
process, starting with an application to the Competent Authority in one country. If a State
objects to the proposal the Member States may vote on it. As yet, very few consents to
market have been given, and none which involve selling GMOs in the UK. However a
number of such applications are awaiting votes in Brussels (W. Parish pers. comm.).
Environmental impacts of GMO introduction
Types of release and precautions
Scale of current releases
As we described above, the use of GMOs is heavily regulated. All releases outside contained
use conditions in the UK have so far been very small-scale and experimental and with strict
protocols to prevent escape. These releases have been purely experimental so far; either the
performance of the GMO is tested or potential environmental hazards are assessed. One could
therefore argue that there have been no real releases yet. The following representative
selection of releases on the public register of consents indicates the size of releases over the
last two years and the precautions adopted against any form of escape. It should also be noted
Review of information, policy and legislation on species translocations
that the procedures followed in all these releases were based upon previous releases of the
same or similar GMOs.
A glyphosate-tolerant sugar beet was planted in four plots of about 1400m2 each,
within a buffer zone such that the total area of the site was about 6ha. Escape was
controlled by harvesting before any flowering occurred, herbicide application after
harvesting and subsequent inspections for volunteers (i.e. plants arising after
Glufosinate-tolerant sugar beet was planted over an area of 360m2 and spread was
controlled by prevention of flowering, ploughing in after harvest and checking for
A potato with modified carbohydrate metabolism was planted over an area of 2000m2
within a buffer planting of non-modified potatoes. After harvest the shoots were
chemically killed and subsequent monitoring for volunteers was carried out. Gene
flow to other potato crops was minimised as the nearest crop was over 1km distant.
Virus resistant potatoes were planted over an area of 96m2 with guard rows of nonmodified potatoes. Above-ground shoots were chemically destroyed and post-harvest
cultivation was used to destroy any remaining tubers. Monitoring for volunteers was
carried out. Again it was ensured that the site was distant from other potato crops.
Oilseed rape with fungal resistance was planted over 400m2 in the middle of a large
field. After seed harvest the plants were killed with herbicides. Volunteers were
checked for over the whole site.
A rhizobium with a marker gene only was applied to seeds planted over 9m2 in the
middle of a large field. The aim was to assess the spread of the modified bacterium,
which was not considered to be potentially harmful.
The famous release of an insecticidal baculovirus (Autographa californica NPV) modified to produce a scorpion toxin - occurred within 32 enclosures each of 1m2,
situated over a 400m2 area. The virus could only escape by the spread of the host
insect and the enclosures were secured against escape of the infected insects, or
contact of the insects with those outside the enclosures, using fine netting (previously
perspex had been used). To minimise use of the site by other insects, and thus as an
extra safeguard against contact, vegetation was removed from the site and ditches
were dug to prevent ingress of ground-dwelling insects. The virus was applied by
spraying into the enclosures and heavy polythene sheeting was used at the time of
spraying as a safeguard against drift. After the experiment the site was disinfected
with formalin and persistence of the virus was determined with bioassays of field
Types of GMO currently released
It is also important to note that only very few types of GMO have passed through ACRE for
even these small-scale releases. ACRE issued 16 consents to release GMOs between March
1993 and January 1994, 14 of which were plants. These constituted:
baculovirus (one release)- as a biopesticide;
bacteria (two releases) - both with marker genes only;
Review of information, policy and legislation on species translocations
potato (seven releases) - one with viral resistance, two with modification of
carbohydrate metabolism, one with insect resistance, and three with both herbicide
tolerance and modified carbohydrate metabolism;
oilseed rape (three releases) - two with herbicide tolerance, one with fungal
sugar beet (one release) - herbicide tolerance;
wheat (one release) - marker gene only;
eucalyptus (one release) - marker gene only.
More consents (27) were given between February 1994 and February 1995. These were:
baculovirus (one release) - biopesticide;
bacteriophage (one release) - marker gene only;
bacteria (two releases) - both with marker genes only;
potato (three releases) - one with modification of carbohydrate metabolism, one with
insect resistance, and one with both virus and blight resistance;
chicory (one release) - herbicide tolerance and male sterility;
oilseed rape (ten releases) - one with modification of fatty acid composition, two with
herbicide tolerance, one with both herbicide tolerance and male sterility, and six with
modifications of male sterility;
tobacco (two releases) - one with pigment alteration, and one with male sterility;
sugar beet (three releases) - one with herbicide tolerance, and two with herbicide
tolerance and virus resistance;
maize (one release) - with herbicide tolerance and pollen control;
wheat (three releases) - one with a marker only, one with herbicide resistance, and one
with herbicide resistance and male sterility.
This reflects the situation in the rest of the world. Ahl Goy & Duesing (1995) found that four
crops - potato, oilseed rape, maize and tomato - constituted over 60% of the plant trials
carried out in North America and the European Union between 1986 and 1993. Of the other
popular species, sugar beet and tomato were common in the EU and tomato, soybean and
cotton were common in North America. Apart from marker genes only, the modifications of
the top four species have been for herbicide tolerance, product quality, and resistance to
insects, viruses, fungi or bacteria. The profile of USA releases of modified crop plants
between 1987 and 1994 looks much the same as in the UK: 30% with herbicide tolerance,
24% with product quality changes, 21% with insect resistance, 14% with viral resistance, 3%
with fungal resistance and 8% of other forms of modification (Stone 1994).
4.4.2 Future changes in the use and release of GMOs
Because of this background, many of the suggested risks are still hypothetical, although some
of the more tractable consequences have been tested in a number of experiments. However,
further assessment of environmental risks must consider the increase in the number of smallscale experimental releases of GMOs, which rose from 16 in 1993/94 to 27 in 1994/95
(ACRE 1994d, 1995b), and the fact that marketing, and thus widespread, less-controlled
Review of information, policy and legislation on species translocations
releases, of certain GMOs is likely to happen soon in the UK. In late 1994 the DOE received
the first UK application for a consent to market a GMO, a glufosinate herbicide-tolerant
oilseed rape. Another application followed shortly afterwards, for import for consumption
(i.e. not to grow) of soybean tolerant to glyphosate. ACRE gave favourable opinions on the
both applications which are now in the hands of the EC. The basis of the favourable opinion
on oilseed rape was that ACRE considered any spread of the modified plant, or of the tolerant
transgene to other oilseed rape, to not be harmful - these were likely to happen but were not
likely to result in harm to humans or the environment. Spread of the transgene to related
species was also considered to be low risk - with little harm expected from its spread (A.
Gray pers. comm.). Soybean cannot establish in the wild in the UK.
This increase in releases of GMOs is worldwide. In the USA five field tests of transgenic
crops in 1987 increased to 486 in 1994 (Stone 1994). Over 32 countries, covering the
Americas, the Pacific Rim, the Middle East, Africa and the European Union, the number of
field trials of transgenic plants increased from five in 1986 to 332 in 1993 (Ahl Goy &
Duesing 1995). Analysis of these changes by Ahl Goy & Duesing (1995) have also indicated
that the species used are changing. Although potato and oilseed rape have been the most
popular plants for modification, monocotyledonous species, such as maize and cereal crops
are being used increasingly. The authors also show that the number of different trait × crop
species combinations tested over the 32 countries increased linearly over the five years, from
under 10 in 1987 to about 200 in 1993. If this trend continues, they predict the number of
combinations could reach 400 by 2000.
4.4.3 Potential impacts of GMO release
Concerns over small- or large-scale releases of GMOs fall into two categories: effects of the
release of non-native species, and particular impacts of organisms which have been
genetically modified. This division disregards the taxonomically untenable argument that a
GMO is no longer the same species as the recipient organism.
Much discussion of the impacts of GMOs concerns the fact that alien species may be used
(Mooney & Drake 1990, van der Meer 1993). Therefore, comparison is made with impacts of
exotic species, which can be great (Chapter 3). However, such comparisons are of limited
value. Many of the GMOs used in the UK are of non-native species, but few (and no plants)
are of species never before released in this country in their unmodified form. As discussed
above, most GMOs released in the UK are of traditional crop plants. If DOE were to receive
an application for the release of a GMO of a species never before released in the UK they
would regard it as a potentially high risk release (W. Parish pers. comm.). This would be
primarily because of its unknown potential impact as a non-native species rather than as a
GMO (such impacts are discussed in Chapter 3). Indeed, this is the main criticism of the
release of the modified Autographa californica NPV, which was a baculovirus from the USA
(see Godfray 1995). However, it should be said that the experimental release of this GMO
was in heavily controlled conditions (see section 4.4.1). At the moment, other countries are
more likely to face this type of risk (van der Meer 1993), but the UK may have to face it in
the future.
Review of information, policy and legislation on species translocations
Aside from this question of alien species which are novel to the UK, GMOs may have two
main impacts on the environment: the GMO may itself establish or spread and form new
populations, and the transgene may spread into conspecifics or other species. Impacts of establishment and spread of the GMO
Establishment of the GMO in the wild cannot be considered a negative effect in itself. There
has been a great deal of discussion about the potential of genetically modified plants to
become weeds, i.e. to invade agricultural or semi-natural communities (Crawley 1990,
Mooney & Drake 1990, Harding & Harris 1994, Williamson 1994). There seems to be
confusion between the potential weediness of the plant species itself and the possible
increased weediness of the species when genetically modified. For plants or other taxa the
invasiveness of the species itself does not concern us in this review of GMOs (unless the
species has never before been released in the UK - see above). The question to consider here
is whether the genetic modification has changed the GMO such that it has a different impact
from unmodified individuals of the same species. For example, although introduction of
bacterial GMOs has been shown to cause changes in the bacterial and fungal populations in
the soil (Smit et al. 1992) and on plants (de Leij et al. 1994), a study by Thompson et al.
1995 showed that the impact of a modified bacterium on a microbial community was less
than that of an untransformed type.
Impacts must be considered in terms of the response of other species in the communities
colonised by the GMO in much the same way as we considered the effects of non-native
species. Thus, plants and animals could invade communities and affect other species. In the
same way, genetically modified bacteria could affect microbial communities (see Dwyer &
Timmis 1990). The GMO could also cause hazards within the release site itself, if negative
effects on the co-existing community of wild species are considered important. One of the
few obvious examples of this is the possibility of toxicity to bee pollinators of insect resistant
plants. This has attracted some discussion, and one solution has been to modify the GMO
such that the insect toxin is not expressed in the pollen or nectar (see Rogers & Parkes 1995).
Therefore, the possible impacts involve: competitive effects leading to declines and losses of
other species (especially of non-transgenic conspecifics); predation and infection of other
species; and changes in the community processes. Genetic effects are discussed below.
Expression of transgenic traits
To begin with, one must consider how the genetic modification affects the phenotype of the
recipient species. Phenotypic changes due to genetic change should be easier to detect in
GMOs than in varieties and strains developed by conventional breeding programmes. The
molecular techniques allow single genes for specific traits to be introduced to organisms.
Conventional techniques introduce a variety of linked genes and traits along with the desired
one, and therefore, despite removal of most linked genes by backcrossing, there will be more,
and less easily detectable, novel traits in the conventionally bred organism.
This argument may be weakened by the development of more complex modifications which
involve multigenic traits and/or multiple traits such that GMOs become very different to the
Review of information, policy and legislation on species translocations
non-transgenic types. This is more of an hypothetical problem at the moment, but may
become more common in the future (Dwyer & Timmis 1990).
A further complication is that the modification itself may not be stable. There is a tendency
for modified or hybrid plasmids not to be inherited stably in freely dividing bacterial
populations (i.e. in the wild). However, this seems to be caused usually by loss or inactivation
of the transgene, or deleterious effects on the survival of the bacterium itself, rather than
causing unpredictable genotypic and phenotypic changes (Dwyer & Timmis 1990). Increased
stability may be obtained by insertion of the transgene into the bacterial chromosome rather
than as separate plasmids (Dwyer & Timmis 1990).
Apart from these bacterial plasmids, the inserted sequences are as stable as the rest of the
genome in prokaryotes and eukaryotes. However, this does not mean transgenic traits are
stable. Most genetic traits, including transgenic traits, often show some intrapopulational
variation in the level of expression of a gene or its inheritance either between individuals or
in different environmental conditions (Rogers & Parkes 1995).
The actual form of integration may cause variation in expression (see Rogers & Parkes 1995).
Most sequences integrate more or less at random into the genome, although some techniques
allow replacement of homologous genes ('homologous recombination'). Transposition is the
most problematical technique in this regard. It involves the use of transposable elements as
vectors for the DNA sequences. Transposable elements are DNA sequences which can insert
into the recipient genome and the point of insertion is determined by the presence of specific
sequences in the genome which can be recognised by a particular transposase enzyme of the
vector. If there is more than one possible point of insertion, the transposable element can
move within the genome. Such movement is usually at very low frequency, but can lead to
variation in expression due to position effects (i.e. the position of the inserted sequence
relative to the rest of the genome can affect its expression).
Questions also arise concerning our ability to describe fully the phenotypic changes caused
by the modification. The position of a transgene in the recipient genome may affect its
expression (Bernardi et al. 1990). The transgene may interfere with the expression of other
genes in the recipient organism (epistasis) or it may affect more than one phenotypic trait
(pleiotropy) (Campbell 1990). Thus, small changes in the genomes may cause major
phenotypic changes.
It should be pointed out that questions as to stability of the modification and phenotypic
changes in response to modification are all considered in the risk assessment for any release
application in the UK (see above). It seems fair to say as well that these complications are not
likely to occur to a great degree in GMOs for which applications for release are made at the
moment. Such effects would be detected in the testing of a GMO, and the development of
GMOs with unexpected and unwanted phenotypic changes or great instability of expression
would be discontinued for economic reasons at the very least.
However, as GMOs are marketed and large-scale releases take place, the rare events
described above will occur frequently because of the large number of GMOs that will be
released. It is also possible that intraspecific gene flow between GMO populations with
Review of information, policy and legislation on species translocations
different modifications may occur. This may cause variations in stability and expression
which are hard to predict.
Changes in competitive ability against conspecifics
Crawley et al. (1993) compared a herbicide tolerant oilseed rape with a non-transgenic type
in the ability to establish populations in a variety of vegetation types and under a range of
climatic conditions. In none of the sites did the transgenic type show better establishment or
population growth, and in a number the non-transgenic type performed better. This
corresponds to the hypothesis that the introduction of a novel trait to an organism will,
through physiological trade offs with expression of other traits, decrease its fitness compared
to the organism in its unmodified state, unless there is a selective benefit of the novel trait.
However, this hypothesis of increased cost should not be interpreted too generally (Tiedje et
al. 1989, Crawley 1990, Regal 1994). Raybould & Gray (1993b, c) reviewed field trials of
transgenic crop species and found few studies showing slightly poorer performance in the
modified plants, but most showed few differences from unmodified plants in performance
measures such as phenology, growth rate and seed production. A similar lack of general
effects of modification on performance has been seen in bacteria (Smit et al. 1992).
However, if the modification confers such attributes as disease or pest resistance or tolerance
of herbicide, the GMO will probably perform much better than the non-transgenic types in
the presence of pathogens, pests or herbicide.
The ability of a GMO to outcompete a non-transgenic conspecific is only of environmental
concern if the conspecific is a wild species. The replacement of, e.g. a feral population of
unmodified oilseed rape by the transgenic type should be of little concern to conservationists
(ignoring for the moment effects on the rest of the community), but the spread of, e.g. a
modified cabbage into a wild cabbage population would be a problem.
Changes in competitive ability against other species
The more general concern is whether the competitive ability against other species is changed,
giving the GMO a greater invasiveness. Various lists have been produced of characters
which, if genetically modified or changed by epistatic or pleiotropic effects of modification,
might increase the ability of species to establish and spread in certain habitats (e.g. Crawley
1990, Mooney & Drake 1990, Harding & Harris 1994, Williamson 1994, Purrington &
Bergelson 1995). Those named for plants include: seed dispersal, seed size and number, seed
viability, seed dormancy controls, seedling survival and growth rate, days to maturity and
flowering, winter survival, susceptibility to pests and diseases, longevity and persistence
(including that of plant parts and perennating organs), and tolerance of certain environmental
factors (e.g. salinity, drought, frost, etc.).
Changes in certain characters could increase or decrease the invasiveness of the species. Of
the genetic modifications currently applied to plants (described above) pest and disease
resistance seem most likely to affect some of these characters. However, as discussed in
Chapter 3, no generalisations can be made as to what changes in which characters might
increase the invasiveness of a species into particular community types. Despite this widely
Review of information, policy and legislation on species translocations
accepted conclusion, Purrington & Bergelson (1995) criticise risk assessment procedures in
the USA for not making explicit the character changes that may increase weediness of crop
species. It is true that this inability to generalise may affect our ability to predict the
consequences of releases of certain GMOs, although, more realistically, it points to the need
for detailed analysis of risks on a case by case basis (see Mooney & Drake 1990).
The ruderal nature of arable crop species means that transgenic types are likely to invade only
disturbed areas, unless modification has caused radical changes to their ecology. Therefore
arable areas and other areas susceptible to disturbance, e.g. through human activity, may be
those more likely to be colonised by genetically modified crops with increased invasive
abilities. For example, conventional oilseed rape forms feral populations on roadside verges.
Crawley (1987) could find no natural or semi-natural community in which any crop plant is
seriously invasive, and a review by the Ecological Society of America (Tiedje et al. 1989)
states that it would be difficult to convert most crop plants into organisms that can survive
and reproduce without human support. By definition, these extremely disturbed areas are
often the least important for nature conservation, although agriculturists are concerned by the
possibility of transgenic crops becoming arable weeds (Harding & Harris 1994).
However, other types of GMOs, e.g. trees, pasture grasses, microorganisms or animals, may
invade other community types. Such possibilities have not been investigated in detail, and the
concentration of research on arable plants should not be taken to warrant a conclusion that
GMOs are generally unlikely to invade semi-natural and natural habitats. Raybould & Gray
(1993a) list ten UK agricultural plants which are known to form feral populations. Three of
these are not confined to disturbed areas: ryegrass Lolium spp, clover Trifolium spp and
cabbage Brassica oleracea.
Changes in predation, herbivory, parasitism, mutualism and pathogenicity
Modified bacteria and viruses could have effects on non-target species and communities.
Much of the controversy surrounding the development and release of Autographa californica
NPV as an insecticide concerned possible effects on a wide variety of insect species (e.g.
'Will the scorpion gene run wild?' New Scientist 25 June 1994). While this particular release
was considered to avoid these problems (see above), the more general question of the effects
of releasing microorganisms with introduced or increased pathogenicity or toxicity is clearly
worrying. It seems unlikely that large-scale releases (i.e. for marketing) of such organisms
will be approved by ACRE unless there are certain controls (A. Gray pers. comm.). Controls
to contain the spread of bacteria could include (Dwyer & Timmis 1990): use of unstable
inheritance to ensure the transgene cannot be passed onto daughter cells; modifications to
reduce the survival capacity of the GMO such that it has a limited survival time outside the
laboratory; addition of 'autodestruct' or 'kill' genes to bring about rapid death of the GMO
once it has fulfilled its purpose.
If the GMO invades new communities or establishes larger populations compared to the nontransgenic type it will cause a general perturbation to the community, as would changes in
abundance of any species. However, where it just replaces the non-transgenic type
community changes will occur if the GMO shows greater resistance to herbivores or toxicity
to pollinators or decomposers (plants), resistance to predators (animals), or pathogenicity to
Review of information, policy and legislation on species translocations
the host (microorganisms) (see Hoffman 1988, Angle 1995). Other more subtle - and, at the
moment, unpredictable - effects could arise through changes in other traits such as product
quality or herbicide tolerance in plants.
Effects on ecosystem processes
Microorganisms with altered biochemical pathways could alter processes such as
decomposition (e.g. if expression of a ligninase is altered) or nitrogen cycling (e.g. if
denitrification pathways are changed) (see also Tiedje et al. 1989, Smit et al. 1992). Toxicity
or other effects of modified plants on decomposers, mycorrhizae and other soil fauna and
flora are possible causes of changes in ecosystem processes (see Miller 1993, Angle 1994,
Morra 1994, Trevors et al. 1994), but all discussion of this subject is highly speculative. Spread of the transgene
The transgene may spread from released or feral populations of the GMO. The spread of the
transgene into other organisms, especially wild species, may be regarded as an impact in
itself, but there may also be effects on the ecology and genetics of these species. This seems
to be the most commonly perceived risk of GMO release at the moment (Abbot 1994).
Potential for gene transfer
Virally or bacterially mediated transfer of genes between unrelated eukaryotes is thought to
be very rare outside the laboratory (Caplan & van Montagu 1990) and therefore the mostly
likely route for gene transfer is by hybridisation. For hybridisation to allow spread of the
transgene, the hybrid must be fertile.
Raybould & Gray (1993b) provide a comprehensive summary of the possibilities for
hybridisation of genetically modified crop plants (including arable, horticultural, market
garden and tree crops) with wild species in the UK. Such hybridisation is controlled by a
series of factors: production of pollen containing the transgene; pollen spread and dispersal
from crop or feral populations; the presence of species with which the GMO can hybridise;
and the compatibility of the GMO with the wild species. It is possible to prevent
hybridisation completely by modifying the GMO so that it does not produce pollen (male
sterility) or so that the pollen does not contain the transgene.
Experiments on pollen dispersal from a number of crop species have been used to
recommend 'isolation' distances from other populations to minimise gene flow (e.g. Rogers &
Parkes 1995), but Raybould & Gray (1993b, c) warn against applying these too rigidly. These
can be used in containing risk of escape in some of the current small-scale releases (Ellstrand
& Hoffman 1990), but will largely be irrelevant for marketed GMOs.
Gene flow to non-transgenic types of the same species should usually only be restricted by
the ability of the transgene to come into contact with these organisms. However, the degree
of compatibility forms an extra constraint to gene flow into other species. Studies have been
carried out to predict rates and patterns of gene flow between modified crops and wild
Review of information, policy and legislation on species translocations
relatives (de Vries 1993, Jacot 1993, Gliddon 1994, Jorgensen & Andersen 1994, Paul et al.
1995). Raybould & Gray (1993b, c) distinguished three categories of genetically modified
crop plant, based on the probability of gene flow into wild relatives in the UK.
Group 1 plants show 'minimal' probability of gene flow because there are no wild
relatives with which to hybridise, or because there are complete breeding barriers.
These include potato, tomato, cucumber, maize, wheat, rye, sunflowers and
Group 2 species show 'low' probabilities of gene flow. These have no wild
conspecifics, but there are close wild relatives (usually congenerics) with which there
is limited compatibility. These include: oilseed rape, flax Linum usitatissimum,
raspberry, blackcurrant, lettuce and barley. For instance, oilseed rape can cross with a
number of Brassica species and some other crucifers, and flax has low compatibility
with some Linum species. Raybould & Gray (1993b, c) point out that hybridisation in
this group is difficult to assess because the nature and strength of the breeding barriers
vary greatly among species and, for a particular species over time and space and
among genotypes.
Group 3 represents a 'high' probability of gene flow and includes sugar beet, carrot,
cabbages Brassica oleracea, poplars and ryegrass, all of which have wild conspecifics
and/or high compatibility with wild relatives. Sugar beet Beta vulgaris vulgaris is
highly compatible with sea beet Beta vulgaris maritima, and cultivated cabbages are
the same species as wild cabbage as well as having low compatibility with wild
congenerics and other crucifers.
Ahl Goy (1993) applied this classification to the crop species used in field trials in Europe
between 1986 and 1992. 52% of the trials involved group 1 species, 26% used group 2
species and 18% had group 3 species.
While many fish species undergoing genetic modification are not native to the UK, salmon, if
released, might be expected to breed readily with wild conspecifics and may hybridise with
brown trout Salmo trutta (DOE 1994a). There is little work on the hybridisation of other
genetically modified animals.
Much less is known about gene transfer in microorganisms than in plants or animals.
Horizontal bacterial gene transfer (i.e. between species) falls into three types. Conjugation
necessitates contact between bacteria, but remote transfer can occur via bacteriophages
(transduction) or the taking up of naked DNA from the environment (transformation).
Transfer is not limited to closely related bacteria, and transfer of transgenic marker genes by
all three mechanisms has been demonstrated between a variety of bacteria (Fry & Day 1988,
Palacios et al. 1990, Shorrocks & Coates 1993). However, little is known about the frequency
and direction (i.e. between which species) of such transfers (Dwyer & Timmis 1990). The
main concerns are that transgenes could be passed to pathogenic, or potentially pathogenic,
Review of information, policy and legislation on species translocations
bacteria, and, in addition, that the expression of the gene may change or pleiotropic or
epistatic effects could arise (Palacios et al. 1990).
Dwyer & Timmis (1990) assessed methods of decreasing the possibility of transfer of
transgenes in bacteria, and considered that the linkage of additional genetic modifications to
inhibit transfer or to inactivate a bacterium ('kill' genes) were potential methods.
Effects of transgene transfer
Some GMOs show a high probability for gene transfer and this may occur even from smallscale releases. However, if a GMO is used widely and there are no imposed constraints to
gene flow the number of individuals of the GMO will ensure that gene transfer to
conspecifics or wild relatives is inevitable, even if the probability of successful transfer is
very low. Therefore, if particular GMOs are marketed, the only risk it is useful to consider is
the consequence of gene transfer rather than its possibility.
Much of the current concern with transfer of transgenes does not directly involve
conservation issues. Stone (1994) and many others discuss transfer of transgenes from crops
to wild relatives purely in terms of the possibility of the transgene increasing the potential for
the wild relatives to become agricultural weeds and this is one of the major risks considered
by ACRE (A. Gray pers. comm.). Transfer to non-transgenic crops or domesticated animals
is only of concern for conservationists if it acts as a 'stepping-stone' which facilitates
subsequent spread to wild relatives.
Potential conservation problems fall into two categories, reflecting two steps in the process of
transgene spread: changes in the genetic constitution of wild relatives, and ecological
consequences of these changes.
Changes in the genetic constitution of wild species
Ahl Goy (1993) considered the consequences of transgene transfer from European crops,
based on the 'potential capacity to confer selective advantage' to wild relatives.
Class I traits ('minimal advantage') are expected to confer little or no selective
advantage or even a selective disadvantage. These include marker genes and product
quality modifications.
Class II traits ('low advantage') may confer an advantage under the relevant selection
pressure. Herbicide or 'stress' tolerance and insect or disease resistance fall into this
Class III traits ('high advantage') are those which may confer advantage in most
conditions. Transgenes conferring enhanced growth and survival are in this class.
Genetically modified crops released in Europe between 1986 and 1992 were partitioned: 25%
in class I, 72% in class II and 3% in class III. Combining the consequence of transfer with the
possibility of transfer (see above), Ahl Goy (1993) arrived at three levels of 'potential
environmental risk' and classified the European releases accordingly: minimal risk (91% of
releases), low risk (9%) and high risk (0%).
Review of information, policy and legislation on species translocations
This classification is too simplified. If the single transgene is transferred then selective
advantage may accrue as stated and the hybrid will take over from the wild type. However,
there are two complications. One is that hybridisation may involve mingling of other parts of
the crop plant genome, and it seems that the general traits of the crop plant will often confer
selective disadvantage compared to the locally adapted wild type. A second possibility is that
there may be complications with the expression of the transgene due to instability and
pleiotropic or epistatic effects. With our current lack of knowledge or understanding of these
possibilities, for microorganisms and animals as well as plants, it is difficult to suggest the
consequences of such effects.
A further consequence may result from the development of a range of modifications within
individual species. For example, different crops of oilseed rape may be planted in a region,
one with glufosinate tolerance, another with a modified carbohydrate metabolism and another
with insect resistance, etc. If wild cabbage populations receive transgenes from these
different sources, unpredictable genetic instability and pleiotropic or epistatic effects may
However, genetic change is not governed solely by selection. If there is a large amount of
gene flow from GMO populations to wild relative 'swamping' may occur such that the high
number of copies of the transgene coming into a wild population cause the transgene to
become fixed even if it imposes a selective disadvantage (Gliddon 1994). This is more likely
if large numbers of the GMO are released, the wild population is small and/or there is a high
compatibility between the GMO and wild relative.
Ecological consequences of changes in the genetic constitution of wild species
The range of these potential consequences mirrors those suggested for the spread of the
GMOs themselves into natural and semi-natural communities, although, again, these have not
been tested. Competitive and invasive abilities may be changed and plant-herbivore,
predator-prey, host-pathogen and other interactions may be modified; all having
consequences for species survival and community structure (see Dale 1994). Ecosystem
processes may also be changed. These risks might be greater than those from GMO spread
because the GMO must invade into natural and semi-natural communities whereas the
establishment of the transgene into wild species requires only its genetic infiltration into an
established population of the wild species.
Viral coat protein transgenic plants
An extra impact on the environment may arise from the modification of plants using viral
coat protein genes to confer viral resistance. Small-scale releases of such plants have
occurred in the UK. It has been suggested that recombination between the viral transgene and
viruses that infect the plant could lead to the emergence of novel viruses or that the protein
expressed in the plant could form the coat protein of infecting viruses (heteroencapsidation or
transencapsidation). There is debate over the likelihood or consequences of such
recombination (see Tepfer 1993, ACRE 1993, Stone 1994, Rogers & Parkes 1995), but it is
too detailed and technical to rehearse here. In summary, analysis of our knowledge of viral
genetics and epidemiology and recent experiments (see Stone 1994) has given rise to a
Review of information, policy and legislation on species translocations
number of concerns. The concerns are particularly in terms of economic effects of new crop
diseases, but disease impacts on wild species are also discussed. In response to these concerns
it has been pointed out that such recombination and transencapsidation probably occurs
within non transgenic systems as well. The DOE exercises caution in the release of viral coat
protein transgenic plants in not allowing any such GMOs to enter the fast track procedure
(see section 4.3.2).
Does genetic modification pose different risks to conventional breeding?
Crawley (1990) reports a view that the ecology of GMOs and the risks of their release are
different to those of non-transgenic types. He then states that this view is mistaken. Many
authors consider that it is a fundamental misconception to think that the main environmental
concern about the development and release of GMOs is the new technology involved. Rather,
it is the phenotype of the GMO that should be considered (Tiedje et al. 1989, Levin 1990,
Huttner et al. 1992). Increased potential to invade and disrupt natural and semi-natural
communities and to transfer novel genes to wild relatives are found in conventionally bred
varieties and strains (see Chapter 3). Instability, epistasis and pleiotropy as a result of the
genetic modifications are likewise not restricted to GMOs and hybrids with GMOs. There is
also little evidence that the impacts of GMOs and GMO hybrids would be greater than those
involving conventional strains and cultivars.
This is probably all true, but it does not mean that GMO releases should be considered to be
risk-free - it actually makes the opposite point, that there is a case for assessment of the
dangers from conventionally bred varieties and strains. Despite the enormous research effort
and debate on GMOs, there is little information on many of the suggested risks. There have
been studies on the performance of GMOs and their ability to form feral populations
compared to non-transgenic types, and on the possibility of transfer of transgenes to wild
relatives. However, these are not impacts in themselves, unless in the latter case, one
considers an effect on the genetic structure of wild populations to constitute a loss of
biodiversity. The major consequences - changes in communities and wild species as a result
of invasion of the GMO or spread of the transgene - have not been investigated. Many
discussions in the absence of data try to use invasions of exotic species as an analogy for
effects of GMOs or try to predict which trait modifications might influence the effects of
GMO release. However, as we discussed above, these are too general to form a real
alternative to research on specific problems.
Certain of these suggested impacts seem to be particularly hazardous and deserving of
special attention.
Insecticidal viruses or bacteria resulting from genetic modification or horizontal
transfer of transgenes may have dramatic effects on insect communities, especially
since some, such as the Autographa californica NPV (see DOE notification of
release), are not very specific.
Review of information, policy and legislation on species translocations
There is a poor understanding of the extent and consequences of gene flow in
Containment of microorganisms in the field may be difficult - active transport of
bacteria in soil is limited but passive transport in wind, animals or percolating
water may form a greater hazard.
The likelihood and consequences of viral recombination or transencapsidation in
viral coat protein transgenic plants are poorly understood.
Direct changes of plant traits which may affect competitive ability or biotic
interactions - such as tolerance to environmental factors, resistance to herbivores
or pathogens - may disrupt communities.
Toxicity of plants to insects may have negative effects on certain important insect
Unpredictability of gene expression and phenotype caused by instability,
pleiotropy or epistasis may be a particular problem where: different transgene lines
are mixed due to gene flow between different GMOs or between GMOs and wild
relatives; modifications using multigenic traits or multiple traits are carried out; or
microorganisms are used.
The release of a GMO of a species never before released in the UK may have
unpredictable and potentially dramatic consequences.
These only seem to represent greater hazards; no definitive statement can be made without
further research. However, a more extensive and technical review than we have been able
to accomplish may throw further light on these questions.
We should also point out, as a footnote to this review of hazards, that we have not
considered some other suggested problems because they are either indirect - e.g. the
suggestion that the development of crop weeds with herbicide resistance may lead to the
use of different, and more dangerous chemicals (Hoffman 1988) - or because they are
vague - e.g. that such modifications interfere with, or usurp, the natural evolutionary
process (Hoffman 1988, Caplan & Montagu 1990, Regal 1994).
4.5.2 Risk assessment for GMO release
Models for risk assessment of GMOs are often discussed (Skalka 1990, Kingsbury 1990,
Smit et al. 1992, Gustafsson & Jansson 1993, Tzotzos 1995) and additional or new
protocols often suggested, based on detailed analysis of particular potential hazards
(Kareiva et al. 1994, Linder & Schmitt 1994). Van der Meer (1993) reviewed GMO risk
assessment protocols from a number of countries and organisations. He noted a number of
general principles:
GMOs should be assessed in terms of the impact of the specific modification, not
from a presumption that GMOs are inherently dangerous;
environmental risk is of two types the hazard posed by the GMO or its descendants, and
the possibility of transfer of transgenes to other organisms and the
hazard resulting from this;
Review of information, policy and legislation on species translocations
a step by step principle to safety should be applied, such that the risks of the
modification are tested and any problems are acted upon at each stage of the
development of the GMO from laboratory to marketing of the GMO and its
risk assessment should be allied to risk management, and in proportion to the level
of the risk.
The DOE/ACRE requirements for the application to release GMOs and the allied risk
assessment cover all these points in great deal. They require all the hazards mentioned in
this review to be assessed on the basis of both the possibility and the consequences of a
hazard. The use of generalisations that we have criticised is avoided and each release is
considered on a case by case basis. It is therefore difficult to suggest improvements in the
current assessment procedures.
As it is used at the moment, the DOE/ACRE risk assessment considers small-scale, often
tightly controlled, releases. The future changes in the scale of releases may require
detailed consideration of two problems: how this may affect the possibilities and
consequences of hazards; and how regulatory changes may lead to insufficient
management of risk.
Problems of risk assessment are illustrated by the difficulties in using small-scale studies
to predict risks of release at the large-scale. There is a widespread feeling that such
extrapolations will be insufficient and that large-scale field trials are needed in tests
covering many hectares, particularly to investigate transfer of transgenes to other species
(Stone 1994). However, this leads to the problem that large-scale trials may cause the
problems they are designed to investigate and that problems such as control of spread of
GMOs or transgenes may require totally different procedures from those used at the
moment. Such tests are already taking place in China and marketing of a number of
GMOs is already happening in the USA (the Flavr Savr tomatoes are already being sold to
the general public). Monitoring and review of the result of these large-scale releases in
other countries will help the assessment of hazards and thus allow decisions on new
assessment protocols.
Purrington & Bergelson (1995) report a trend in the USA regulatory body APHIS towards
deregulation of the marketing consents, which, they state, involves: consideration en
masse of multiple transgenic lines, cursory examination of additional lines similar to those
previously deregulated, and insufficient assessment of the invasive potential of GMOs into
novel habitats. While we are not in a position to assess the accuracy of this criticism,
deregulation does represent a move away from the case by case approach intrinsic to many
risk assessment procedures, including that of DOE, and seems to be rather premature
given the lack of knowledge about a number of risks.
4.5.3 Involvement of nature conservation bodies in GMO assessment
Many authors point out the potential for GMOs to be used to tackle a variety of
environmental problems such as: replacing the use of unrenewable resources, provision of
food, development of specific biopesticides, the potential to use low hazard herbicides,
Review of information, policy and legislation on species translocations
bioremediation and replacing the need for nitrogen fertilisation (e.g. Gustafsson & Jansson
1993, Rogers & Parkes 1995). Therefore, there are benefits to conservation as well as
industry in the development of GMOs. The result of this is that there is much to be gained
by the involvement of conservationists, especially the statutory conservation agencies, in
the development of GMOs and of appropriate assessment of environmental risks.
The statutory conservation agencies are notified of applications to release GMOs in the
UK and used to have an observer's role on ACRE. However, it is fair to say that they have
not played a strong role in either the control and policy of GMO releases or in highlighting
research areas. The statutory conservation agencies (A. Burn, M. Palmer and J. Tait pers.
comm.) are working towards a policy to improve the input to the decision making process.
The probable future changes in GMO releases and the potential for wider environmental
problems both indicate a real need for the involvement of the statutory conservation
Summary conclusions concerning introduction of genetically
modified organisms
Genetic modifications have been made to a range of plants, micro-organisms and
some invertebrate and vertebrate animals. Most GMO releases in the UK are of
crop plants, a few are of viruses and bacteria, and no animals have been released.
Current releases in the UK are very small-scale and experimental, with strict
protocols to prevent escape.
In the future a wider range of GMOs may be released, and certain GMOs will be
marketed, i.e. they will be used widely in the countryside with few direct controls
against escape.
Because no GMOs have escaped into the wild in this or other countries there are
no examples of the impacts of such escape, all assessments are based upon smallscale experiments or speculation.
Impacts of GMOs following release into the wider countryside must be
differentiated into effects of non-native species and effects of the modification. If
the species used has never before been released in the UK then it must be
considered as a novel alien species; if the species is one already released or used in
the UK (e.g. many crop plants) the consequences of the modification for impacts
on UK biota must be considered.
A GMO could have a different ecology to the unmodified type, resulting in
changed interactions with other species through alterations in e.g. competitive
ability, toxicity, or pathogenicity. Changes in ecosystem processes could also
occur e.g. in decomposition rates or nitrogen cycling.
Review of information, policy and legislation on species translocations
Certain genetic modifications may increase the likelihood of such effects, e.g.
changed growth rate, pest resistance or disease resistance.
Unexpected or unstable effects of the modification on the phenotype may lead to
unpredictable changes in the ecology of the species.
Gene flow may lead to transfer of the transgene to conspecifics or other
compatible species. Hybridisation with wild relatives is very likely for some
species. Such gene transfer may change the ecology of the wild relatives, with a
similar range of consequences to changes in the ecology of the GMO. However,
the precise effects of such gene transfer are hard to predict.
At the moment there is little evidence that these suggested impacts of GMOs and
GMO hybrids would be greater than those of strains and cultivars developed by
conventional breeding methods. However, the risks of such impacts should still be
assessed thoroughly. It may also be that the risks from future GMOs will be
The Environmental Protection Act 1990 strictly regulates all use of GMOs in the
UK and provides a comprehensive procedure for risk assessment. It applies a step
by step principle to safety, and the risks of the modification are tested and any
problems are acted upon at each stage of the development of the GMO from
laboratory to marketing of the GMO and its products.
An independent committee of experts, ACRE, assesses each proposed release with
reference to a lengthy risk assessment. A few, well researched, GMOs are subject
to a less rigorous risk assessment under the fast track procedure.
It would be difficult to improve on these procedures given the current small-scale
controlled releases in the UK. However, future changes in the scale of releases
may require detailed consideration of two problems: how this may affect the
possibilities and consequences of hazards; and how regulatory changes may lead to
insufficient management of risk.
The UK statutory conservation agencies have played little role in the development
of risk assessment procedures for GMO release, or in the regulatory process, and
they have no policy statement or guidelines. The possibility of future changes in
GMO releases and the consequent potential for environmental problems both
indicate a real need for the future involvement of the statutory conservation
This may require a more detailed review than we have been able to achieve of the
current status and the prospects for future changes in use of GMOs in the UK and
Review of information, policy and legislation on species translocations
The translocation of species assemblages is used to rescue communities which would
otherwise be destroyed by a change in land use at the donor site. The causes have been
civil engineering and excavation projects (road building, pipeline laying, construction of
buildings, quarrying, peat extraction, etc.) although they could be used to mitigate
ploughing up for agriculture or forestry. This technique is usually called 'habitat
translocation', but because this is an inaccurate use of 'habitat', we prefer the term
'community translocation'.
Usually in a translocation the vegetation is lifted as turves and re-laid at a prepared
receptor site. Occasionally, soil transfer is described as a method of translocating and reestablishing a community (e.g. Worthington & Helliwell 1987). Other projects - which we
will not consider to be forms of community translocation - have involved translocation of
pieces of turf either into existing grasslands in order to establish new species (e.g. Rawes
& Welch 1972) or to act as inocula in restoration projects (e.g. Wathern & Gilbert 1978)
(see Chapter 2).
The types of vegetation translocated have in most, if not all, cases been of conservation
value, and many have been SSSIs. The 'Habitat Transplant Site Register' was created by
the England Field Unit (EFU) at NCC in 1988 (although it has not been updated) in order
to create a record of all community translocation sites in England. Of the 76 sites on this
register, 47 were notified SSSIs or sites for which the SSSI notification was under way.
Translocations have been usually of herbaceous or dwarf shrub vegetation. The EFU
Habitat Transplant Site Register listed 48 translocations of grassland and marsh, 12 of
heathland, seven of woodland, seven of coastal vegetation, six of bog, mire, flush and
spring vegetation, six of marginal aquatic vegetation, four of fen, swamp and inundation
vegetation, and two of open water vegetation (Byrne 1990) (this sums to more than 76
because some sites contained more than one community type). The results or even the
techniques used in many of these translocations are not available. The only reported
method of translocating woodlands and their constituent trees was purely a planning
exercise which was not implemented (Down & Morton 1989).
Because community translocation is used for only one reason, this chapter has a different
structure to those on translocation of native and non-native species. After summaries of
relevant policy and legislation, we give the background to the major case studies and then
we discuss the problems illustrated by these studies.
Review of information, policy and legislation on species translocations
Summary of guidelines and policies
This section summarises the policies (statements of position) and guidelines
(recommended procedures) of UK conservation and other organisations. These are not
implemented by law, but represent, at most, agreed codes of conduct. There are no
international guidelines on translocations of species assemblages.
In many cases the guidelines have been condensed although some sections are transcribed
verbatim. Some terms have been changed to conform with our definitions.
Nature Conservancy Council (1988). Habitat translocation and the safeguard of seminatural habitats
This draft guidance note is the only full statement of a policy towards translocation of
species assemblages by NCC or its successor organisations. The document stated the need
to assess the advantages and disadvantages of a particular proposal as well as the
particular risks of failure in the translocation. The following principles were developed for
NCC responses to proposed translocations.
This form of translocation can be seen on a priori grounds as having unacceptable
risks of failure - therefore, it should not be seen as an alternative to in situ
A critical appraisal of a proposal should be based on evaluation of the scientific
interest of a site and account taken of potential causes of failure, including:
poor or no proper planning;
lack of sufficient finance and expert personnel;
insufficient analysis of donor site - e.g. soil chemistry and structure, hydrology,
microclimate, etc.;
poor or no evaluation and trial of techniques;
inability to identify an appropriate recipient site;
insufficient evaluation of recipient site;
poor site preparation;
presence of features which cannot be translocated;
presence of species incapable of withstanding translocation;
disruption of communities and community processes;
ineffective aftercare and management;
too short a time scale for the project;
other, uncontrollable risks.
(these risks are also discussed, in more detail in the NCC report, Byrne 1990)
NCC should not be initiating or implementing translocation schemes - other areas
of work will have greater conservation benefits.
There is a need for education and explanation to the public of an NCC position on
such translocations.
Review of information, policy and legislation on species translocations
Nature Conservancy Council (1990). Review of NCC policy on Species Translocations in
Great Britain
Translocations of species assemblages were treated briefly in this document. NCC stated
that such translocations must be seen as secondary to the protection of the assemblage in
situ. Procedures were as for single-species relocations (see Chapter 2).
English Nature (1993). Roads and nature conservation
This report provides guidance to developers on mitigation of the impacts of road building.
It states that translocation does not provide compensation for loss or damage to high
value, non-replaceable sites.
Department of Transport, Scottish Office Industry Department, The Welsh Office,
Department of the Environment Northern Ireland (1993). The wildflower handbook
This advice by Government departments for those involved in road building includes short
guidelines on translocations of species assemblages ('habitat transfer'). Recommendations
are given on:
suitable vegetation types for turfing;
appropriate methods of turf lifting, storage and re-laying.
Summary of legislation
This section summarises UK and EC legislation and international conventions which are
relevant: in halting developments when community translocation is not considered a
viable alternative to in situ conservation; in enforcing the use of translocation to mitigate
for environmental damage; or in imposing conditions and precautions upon the
translocation procedures.
Town and country planning
The relocation of species assemblages will usually be carried out in response to a change
in land use. Suggestions for relocations, objections to relocations, or proposed changes to
the relocation procedure may be considered while planning permission for a development
is being decided. The Town and Country Planning (Scotland) Act 1972 and the Town and
Country Planning Act 1990 set out the procedures by which the planning authority
(usually the District Council, with the involvement of the Secretary of State) can grant or
refuse planning permission or can grant permission subject to certain conditions. In this
process the planning authorities have a duty to have regard for conservation of the natural
heritage (Scotland) or of the natural beauty and amenity of the countryside (England and
Wales). The EC Directives on Habitats (Council of Europe 1992) and on Environmental
Assessment (85/337) both require account to be taken, during the planning process, of
Review of information, policy and legislation on species translocations
effects of developments on the natural environment and wildlife, and may require
assessments of environmental impact.
Protected areas
Local Nature Reserves, Areas of Outstanding Natural Beauty (both Natural Parks and
Access to the Countryside Act 1949), National Nature Reserves, Areas of Special
Protection (both Wildlife and Countryside Act 1981), Sites of Special Scientific Interest,
National Parks (both Natural Parks and Access to the Countryside Act, amended by
Wildlife and Countryside Act) and Natural Heritage Areas (Natural Heritage (Scotland)
Act 1991) all, to a varying degree, receive extra protection against development that may
damage the environment. Except for Local Nature Reserves (assessed by Local
Authorities), the statutory conservation agencies or the Countryside Commission (when,
in England, its role in the planning process replaces that of EN) must be consulted
concerning any planning applications for these protected areas. However, unless the site is
owned by the statutory conservation agency or is acquired by compulsory purchase, they
cannot ultimately stop any development or relocation of a species assemblage.
Sites protected under European legislation and international agreements should also
receive some security against development and the translocation of species assemblages.
These sites in the UK include the following. Special Protection Areas are designated under
the EC Directive on the Conservation of Wild Birds (the 'Birds Directive', 92/43) for the
conservation of rare bird species listed in its Annex I. Under the EC Habitats Directive
Special Areas of Conservation (SACs) will be created. Cases in the European court (see
Reid 1994, p192) have lead to an interpretation of the Directive that any action reducing
the size and quality of a SAC can be permitted only rarely. Ramsar sites, designated under
the Convention on Wetlands on International Importance (the 'Ramsar Convention') are
also protected against activities resulting in reducing their size unless there is an urgent
national interest.
Case studies
Because of the propensity for translocations to be of grassland or heathland, these are the
vegetation types covered by our case studies. However, other studies are summarised in
the discussion.
Dongas, Hampshire
This project, carried out by the ITE, involved the translocation of some 3000m2 of chalk
grassland turf from a site known as the Dongas (part of the St Catherine's Hill SSSI),
which lay on the route of a new section of the M3 Motorway between Bar End and
Compton near Winchester, Hampshire. The receptor site, the Arethusa Clump, was
Review of information, policy and legislation on species translocations
calcareous ex-arable land 1.6km from the Dongas. This project arose from a condition
imposed by the Inspector as a result of the Public Inquiry, although this was merely for the
restoration of an area of chalk downland in 'compensation' for the losses due to the
construction. There were no criteria as to the form or quality of the restoration. The
translocation, and a larger programme including seeding and planting of other areas on
Arethusa, were developed by ITE and accepted by the Department of Transport as a high
quality restoration programme (R. Snazell pers. comm.). Thus, the project was not strictly
one of community translocation, it was rather a restoration. However, because of the
methods used, it provides a useful case study of community translocation.
The project involved extensive planning (Thomas et al. 1992, Morris et al. 1994)
following ecological surveys (Snazell et al. 1991) and a review of possible techniques
(Exton et al. 1991).
The Arethusa Clump was prepared, following tests for soil nutrients, by removal of the
nutrient-rich topsoil to expose a thin nutrient-poor soil more suitable for chalk grassland
vegetation. Turf translocations were accomplished using 'Macroturfing' equipment (see
Pywell 1990). This uses tractor-mounted equipment which cuts large and deep turves
(0.25-0.3m depth) for the purposes of: avoiding freezing or desiccation of the turves in
transport, minimum damage to plant roots or soil disturbance, and the translocation of
burrowing invertebrates such as yellow meadow ants Lasius flavus. Translocation was
carried out in December, a season of dormancy or low activity for most plants and
animals. To reduce desiccation and freezing, the turves were out of the ground for less
than one hour. The turves were not placed in the receptor site in the same positions
relative to each other as they had the donor site, although this would have been preferred
(R. Snazell pers. comm.). The turves were placed with as narrow an interstitial gap as
possible and were rolled and tamped down after placing.
The receptor site was mown when necessary for the first two years, to keep the sward
short. In the winter of the second year cattle were introduced to the site, and these have
now been replaced with sheep. Sheep grazing was the standard management at the donor
The area of the Dongas to be turfed was monitored the year before the translocation to
provide baseline botanical and entomological data. Botanical and invertebrate monitoring
was planned for each year for the first four years, and then in years 6, 8 and 10. The data
are being organised and analysed using computer-based databases (Ward & Stevenson
1994a, b, Snazell et al. 1995a, b).
Review of information, policy and legislation on species translocations
Hand translocation
A trial translocation of a small area of turf, involving the same donor and receptor sites as
above, was carried out using a different technique, 'hand translocation'. Spades were used
to lift smaller and shallower turves and to re-lay them. Site preparation, management and
monitoring followed the same procedures as above.
Hockley Junction, Hampshire
This was another ITE project carried out due to the construction of the M3 Motorway
between Bar End and Compton near Winchester, Hampshire. 5000m2 of a species-rich
flood meadow in the Itchen Valley SSSI was translocated to an ESA site 25km distant.
The project planning and preparation followed the same procedures as for the Dongas
(Ward & Stevenson 1995). However, there was no consideration of mitigation for
Hockley Junction in the public inquiry. This programme was instigated subsequently by
EN after discussions with ITE, DTp and MAFF (R. Snazell pers. comm.).
Pre-translocation investigations showed that the receptor site was generally drier than the
donor site. In an attempt to remedy this, the receptor site was excavated to lower the soil
surface relative to the water table. The turves were translocated in October (to reduce
desiccation) using similar procedures as for the Dongas, except that the turves spent a
longer time, 2 hours, out of the ground.
Seasonal sheep and cattle grazing was introduced at the receptor site in the late summer of
the first year following translocation. Control of invasive species was also carried out. The
donor site had been sheep grazed.
Botanical monitoring is following the same protocol as for the Dongas, but no invertebrate
monitoring has taken place.
Hand translocation
As for the Dongas, a small area of turf was translocated using spades.
Review of information, policy and legislation on species translocations
Middlebere Heath, Dorset
This translocation of turves of humid/dry heath was carried out by Pywell (1993) as part
of a larger investigation into heathland restoration (see also Pywell et al. 1995). However,
it provides the most comprehensive study of heathland translocation in the UK to date.
1500m2 of turves were translocated from a site to be developed for mineral workings, to a
site 11km away. The receptor site was an acid grassland which had been heath until
agricultural development about 30 years earlier. Therefore, the soil fertility was low, but
addition of agricultural chemicals meant that the pH and certain nutrient concentrations
were higher than in a local heathland. The objective was to establish a heathland plant
community on the receptor site, although any changes in the plant species composition
were recorded.
A depth of soil was stripped from the receptor site to allow bedding down of the turves to
the same height as the remaining vegetation. As for the Dongas, Macroturfing was used to
supply large turves which extended to the depth of the eluvial mineral horizon (about
0.15m). The turves were out of the ground for 2-3 hours and were re-laid close together to
reduce gaps. Any gaps were filled with sand.
Translocation was carried out in July, and was followed by a severe drought. This
necessitated irrigation of the re-laid turves over the remaining summer months. Otherwise,
no management has been carried out over the next six years.
Detailed botanical monitoring was carried out for four years following translocation, and
further monitoring is planned in the sixth and subsequent years (R. Pywell pers. comm.).
Selar Farm, West Glamorgan
ITE carried out this trial of turf translocation techniques for British Coal who were
investigating mitigation methods in the development of coalfields. Selar Farm was an area
of traditionally managed species-rich pasture which contained several rare plant species.
138m2 of turf was translocated to a site 1km away which was chosen for its proximity, its
low levels of fertiliser inputs and its similar environment to the donor site, particularly the
similar aspect, altitude and soil profile and type and range of moisture levels which
resembled that at the donor site (Good et al. 1994). However, there were differences in
Review of information, policy and legislation on species translocations
slope and in soil pH; the receptor site was steeper and more acidic. The project was
carried out in consultation with NCC and British Coal and was planned after research into
translocation techniques, translocation trials and extremely detailed studies of the donor
and receptor sites (Good et al. 1992).
The receptor site was sprayed with herbicide to kill the vegetation, and then rotovated.
Most turf in the donor site was stripped to a depth of 0.08-0.1m using a machine bucket.
Turves from drier or wetter areas were laid into respectively drier or wetter areas of the
receptor site. The turves were pushed together to reduce gaps. The translocation took
place in late September.
The donor site was traditionally lightly grazed by cattle, but the translocated turves were
managed only by occasional strimming over the subsequent four years.
The donor site was surveyed and the translocated turves were monitored botanically for
four years.
A second method of turf translocation was used. Half of the 138m2 of turves were laid at
the receptor site and then rotovated over twice their original area to increase coverage.
Otherwise the techniques used were exactly those described above.
Brocks Farm, Devon
This project and the five following were initially monitored by the England Field Unit
(EFU) of NCC and then by EN as part of an intensive investigation into the outcome of
community translocations in different parts of England.* None of these translocations was
carried out by NCC themselves. Byrne (1990) gives some detail on the early results of
some of these projects.
0.4ha of a species-rich mesotrophic grassland were translocated to avoid its destruction by
the tipping of waste from a ball clay works and with the stated objective 'to safeguard the
botanical composition and ecological characteristics of the grassland' (Leach et al. 1995).
English China Clays Ltd translocated the turf after representations from NCC and the
We have not included another major site in the EFU project, Brampton Meadow, Cambridgeshire, because although reports
exist on the pre-transplant surveys by NCC/EN (Buckingham 1987, Winder & Robertson 1993), the three years of posttransplant monitoring were carried out by environmental consultants and the only information made available to us (with
thanks to H. Robertson) are the unanalysed data of the first year's survey (Anon 1992).
Review of information, policy and legislation on species translocations
local Wildlife Trust. The receptor site was 1km distant and had a slightly sloping
topography, similar to the donor site.
The topsoil and subsoil were stripped from the receptor site to provide a suitable bed for
the turves. Turves 2.4 × 1.2m and 0.16-0.2m deep were lifted in September/early October
using a machine bucket. Gaps between the re-laid turves were filled with soil and litter
from the donor site. Turves from a wetter part of the donor site were laid on a wet part of
the receptor site
The turves were rolled and watered after re-laying. The donor site had been traditionally
cut in the summer, but this had lapsed in recent years. A cut was made in the two summers
prior to translocation and this was resumed on the translocated turves in the first summer
after translocation. In recent years the turves have been lightly grazed by sheep in early
Monitoring of plant species was carried out using random 10cm square quadrats by
NCC/EN in the year prior to translocation, and then for another seven years at the receptor
site (Leach et al. 1995). The monitoring is continuing (S. Leach pers. comm.). An adjacent
undisturbed grassland SSSI with the same management as the donor site (pretransplantation) and receptor site (post-transplantation) is also monitored as a 'control'.
1.2ha of the donor site was transferred by 'blading'. The vegetation was close cut and then
the soil was rotovated to a depth of 0.5m. This material was collected and spread over an
area of the receptor site which had been prepared in the same way as for the turves.
Subsequent management was the same for the whole receptor site, except that cutting was
not resumed until the second summer after translocation.
Thrislington Plantation, Durham
This extremely important magnesian limestone grassland was translocated to make way
for an extension of a limestone quarry. This translocation is one of the most heavily
studied and it has resulted in a number of publications (Park 1988, 1989, Sheppard 1990,
Cullen & Wheater 1993). 8.5ha were translocated to two arable fields 0.8km away. The
quarrying company carried out the relocation with advice from NCC.
Review of information, policy and legislation on species translocations
The receptor site was stripped of soil to expose the limestone overburden. Translocation
took place over the winter months because the high soil moisture content in this season
allowed the turves to maintain their integrity (Park 1989). Trees and shrubs were cut at the
donor site and turf to a depth of 0.5m was lifted using a machine bucket. The turves were
re-laid within a few minutes. The gaps around each turf of about 0.25m were infilled with
soil. Separate plots of the 8.5ha of grassland were translocated over a period of seven
The donor site was probably unmanaged, but at the receptor site the turves were managed
for the first two years by controlling weeds and shrub regrowth and subsequently by
annual cutting.
Plant and invertebrate monitoring took place in each of the donor plots before
translocation. The invertebrates were re-surveyed three years after the translocation, and
plant monitoring has continued up to the present day (S. Hedley pers. comm.). Most
monitoring has been carried out by NCC/EN.
Ashington, Northumberland
1ha of a species-rich neutral grassland was translocated by the National Coal Board after
negotiations with NCC. The donor site was to be landscaped as part of a project to
develop a colliery. The receptor site was an arable field only a few hundred metres from
the donor site.
The receptor site was stripped of topsoil to decrease fertility. The turves were lifted using
a machine bucket and were of variable thickness. The donor site had a ridge and furrow
topography and, in an attempt to replicate this, an undulating surface was created when
laying the turves at the receptor site (Piekarczyk 1991).
The management of neither the donor or the receptor sites are described by Piekarczyk
Review of information, policy and legislation on species translocations
No results are available from monitoring of the vegetation, but, unusually, the beetle and
spider fauna was monitored intensively for 18 months after the translocation. Pretranslocation monitoring was insufficient, so the fauna of the receptor site was compared
to an undisturbed control over the same period (Piekarczyk 1991).
Newhall Reservoir, Nottinghamshire
An SSSI calcareous grassland (about 0.7ha) growing on a reservoir roof was removed by
the local Water Authority while repairs were carried out to the roof. The turves were not
transported to a new site, but were replaced on the roof after the repairs were completed
(Cox et al. 1992).
The turves were stripped using a commercial grass turf stripper to a depth of 0.07-0.1m.
These were stored nearby on polythene sheeting for 'several weeks'. The repaired roof was
prepared with a layer of gravel and soil and the turves were bedded onto this.
Management of the grassland before the repairs consisted of occasional cutting. This
continued uninterrupted after the repairs.
Botanical monitoring was carried out by NCC, using random 10cm square quadrats before
and for three years after the turfing (Cox et al. 1992).
Potatopot, Cumbria
8000m2 of species-rich acidic grassland were translocated by British Coal from an area to
be used for open-cast coalmining to a receptor site 2km away. The project planning was
rather confused (Byrne 1990), and earlier small-scale translocations had been carried out
by a variety of organisations. The eventual large-scale translocation was planned after
surveys and consultation between British Coal, NCC and environmental consultants, with
the objective of retaining the botanical interest of the vegetation. The receptor site was
considered to have similar soil, topographical and hydrological characteristics to the donor
Review of information, policy and legislation on species translocations
Deep turves were cut and lifted using a machine bucket. The receptor site was prepared by
levelling the soil to facilitate the laying down of turves. To maintain the original
vegetation distribution, the turves were re-laid at the receptor site in the same pattern as at
the donor site, although an inability to push the turves together meant that the large gaps
between some turves had to be filled with soil. The translocation was carried out in
August-September and the donor site was watered prior to turfing to improve soil
The donor site was previously grazed by sheep, but it was not grazed for four years prior
to the translocation, and the turves were not sheep grazed until two years after
translocation. Grazing has continued subsequently (Jerram 1993).
Botanical monitoring using random quadrats was carried out prior to translocation and
continued in the same quadrats for three years after transplantation. Monitoring was
carried out by NCC/EN and environmental consultants.
A second technique was used in an attempt to translocate some of the Potatopot plant
communities. Turf and soil was excavated by bulldozer to a maximum depth of 0.3m,
heaped into a truck, and spread out at the receptor site. The receptor site had been
prepared in the same way as for the turves. The material was driven over by a tracked
vehicle to produce a level surface. The management was the same as for the turves and
botanical monitoring followed the same procedure (Jerram 1993).
Ashcott Heath, Somerset
This project was different from the others described here. It used only the 'blading'
technique that was used partly at the Potatopot and Brocks Farm sites - rather than
translocate turves, the excavated soil and vegetation of the donor site was used to establish
vegetation on the receptor site. This was carried out because 'trials...indicated that it would
be difficult to pick up and transplant turves' (Cox et al. 1991). Although NCC advised of
the area to be translocated, they had no involvement with the actual translocation (Cox et
al. 1991).
The donor site was to be developed for peat extraction and material from 1200m2 of wet
acid grassland was moved to a receptor site 8km away on a nature reserve. This was a
worked-out peat extraction site which had undergone some revegetation, but consisted of
a thin peat layer over clay.
Review of information, policy and legislation on species translocations
The receptor site was lightly rotovated, but there was some regeneration of the vegetation
by the time of the translocation. The vegetation and soil at the donor site was excavated to
a depth of about 0.2m. It was piled for a short time at the receptor site before being spread
out and rolled flat over the existing vegetation.
The donor site was traditionally grazed, but the only management of the receptor site was
a single cut (without removal of the mowings) two years after the translocation.
Botanical monitoring was carried out by NCC on the donor site prior to the translocation
and on the receptor site until three years after the translocation.
Consequences of community translocation
The potential adverse effects on biodiversity fall into two categories: those involved in
any species translocation, and those particular to translocations of species assemblages.
Effects of species translocation
The first category is therefore a subset of the problems described in the chapters on native
and non-native species. Because the community at the receptor site is usually totally
destroyed it is irrelevant to consider effects of the translocated organisms on those resident
at the receptor site. However, spread of translocated species outside the receptor site,
especially of invasive species, may lead to: colonisation of species into regions where they
are not native; colonisation of locally native species into communities where they were
not previously resident; or supplementation of populations of resident species, possibly by
a quite different race. None of these possibilities have been studied in community
translocation projects, but changes in biodiversity at the colonised sites may occur
through: declines or losses of resident species by competition, predation, parasitism,
pathogenicity, or other interactions with the colonising species; changes in community and
ecosystem processes; changes in genetic structure or diversity of resident populations; or
hybridisation between colonising and resident species.
Although there are no studies of these possibilities, it is fair to say that none of the genetic
problems are likely while translocations remain at the local scale. For the same reason, it
is impossible to introduce new species to a region in this way, although the spread of nonnative species may be facilitated. Spread of locally native species into communities
previously unoccupied by that species may occur, but this is likely to be of minor
importance compared with the other consequences described below.
Review of information, policy and legislation on species translocations
Effects particular to translocations of species assemblages
The precise aim of community translocations is often poorly stated. By implication there
are two subtly different reasons.
1. To translocate, unchanged, all populations of every taxon (animals, plants, fungi and all
micro-organisms) of a community from within a prescribed area. 'Unchanged' entails the
long-term maintenance of population sizes and of ecological processes such as
interactions among species and ecosystem functions. This may be termed the
'conservation aim' and represents the ideal (to the conservationist) outcome of a
2. To translocate a plant community and, it is to be hoped, some of the other taxa
(especially invertebrates), such that the translocated community resembles the pretranslocated state. This may be termed the 'amenity aim'.
Much of the controversy surrounding community translocations concerns these
differences in the objectives. For instance Byrne (1990), in a review for the NCC, defines
'habitat transplantation' as ' the [translocation] of a complete assemblage of plants and
animals, with the aim of maintaining the habitat unaltered in its new location'. This is very
similar to our 'conservation aim'. Our assessment of consequences for biodiversity
necessarily reflects the conservation aim, but we shall return to this discussion later.
Two types of problem are well illustrated by the case studies: losses and changes in
abundance of species originating from the donor site, and the invasion of new species. We
discuss other potential problems, but there are no data available with which to assess their
impact or relevance. Poor translocation of particular species or changes in translocated populations
Certain species may not be translocated by the lifting and transferring of turves and will
therefore be lost from the community. If the proportion of the original number of
individuals in the turfed area that is translocated differs among species, this will change
the community structure. If only a small proportion of the population of a species is
translocated, the resulting small population size may lead to poor long-term population
survival at the receptor site. Poor transferral of certain species may occur if a majority of
the population is killed by the translocation or many individuals are missed by the
translocation technique (e.g. more mobile animal species escape, or deep-burrowing
animal species are not collected). There may also be problems of small population size if
the turfing does not encompass the whole site and the individuals in the turfed area
represent only a small proportion of a population which originally extended over the
larger area of the whole site.
After translocation, species may change in abundance or even be lost, resulting in effects
on community structure. These changes may be due to: differences between the donor and
Review of information, policy and legislation on species translocations
receptor sites in abiotic environmental conditions (soil type, topography, altitude,
hydrology, etc.); changes in ecosystem function such as, nutrient cycling, productivity,
decomposition rates, etc.; changes in biotic conditions as a consequence of a number of
factors such as, different management, alterations of interactions at all trophic levels,
especially through losses or changes in abundance of original species and the invasion of
novel species; or by rearrangement of spatial patterns, affecting population processes and
species interactions.
The case studies provide ample evidence of changes in plant communities after turfing or
soil transfer, and also give some information on the responses of invertebrate communities
to turfing.
Plant communities after turf translocation
For two years after translocation, the Dongas hand translocated turves showed a few
losses of typical chalk grassland plant species, including some deep-rooted orchid species
(Ward & Stevenson 1994a, b, L. Ward pers. comm.).
The Dongas machine translocated turves showed no losses of chalk grassland plant
species over the first two years, apart from the loss of two species which were extremely
rare in the first place, and which Ward & Stevenson (1994b) predict are probably showing
normal population fluctuations. However, this may illustrate a problem of translocating
small numbers of individuals which represent only a part of the population at the original
site. Otherwise, there were actually slight increases in numbers of chalk grassland species
(probably due to establishment from the seed bank following disturbance) over the first
three years. Because the turves were fairly deep, all the deep-rooted orchid species
survived (L. Ward pers. comm.). The lack of grazing management in the second year lead,
in some cases, to decreases in cover of some chalk grassland species (Ward & Stevenson
1994a, b).
The Hockley hand translocated turves showed continuing declines in the number of
wetland plant species over the first three years, with species of wetter habitats suffering
most. Other wetland species showed decreases in cover. These changes were thought to be
due to reduced grazing and the receptor site being drier than the donor site (Ward &
Stevenson 1995).
The machine translocated turves at Hockley showed better survival of wetland plant
species than the hand translocated turves, although there were slight declines in the
species numbers over three years. There were indications that certain coarser species were
increasing in cover, perhaps due to changes in water level and reductions in grazing.
For both types of translocation, it is thought that a more intensive grazing regime will
remedy some of these negative responses. However, there are clear and continuing
Review of information, policy and legislation on species translocations
changes in the topography of the receptor site caused by the translocation. Some areas are
becoming drier and others are sinking and becoming wetter as the disturbed soil and peat
are shifted by water movements and the hydrology remains unstable (L. Ward pers.
comm.). It is predicted that these changes will lead to changes in species abundances and
distributions in the receptor meadow (Ward & Stevenson 1994, 1995).
Some areas of the transferred heathland turf at Middlebere died during the summer
drought, although this did not result in the loss of any species from the whole site.
However, the disturbance caused by the translocation and the more freely draining soil at
the receptor site compared to the donor site resulted in gradual changes in the vegetation
in response to the drier conditions. In particular, cross-leaved heather Erica tetralix, which
was initially very common and is an indicator of humid or wet heathland, declined to a
frequency of near zero after four years.
Selar Farm
The turves at Selar Farm gradually changed over four years of monitoring away from the
original donor plant community to become transitional between the donor community and
the previous community at the receptor site (Good et al. 1994). There were also declines
in plant species richness. Between 30-50% of the original plant species were lost in the
transferred turves. These were not replaced by many invading species (see below). The
plant species at the donor site were categorised into various types, and over four years the
calcicolous species, mesotrophic forbs and those characteristic of wet and dry heath
declined, while mesotrophic grasses increased. The drier areas showed greater
deterioration than the wetter areas. These changes were probably caused by the different
soil (pH and fertility) and topographical conditions at the receptor site and the lack of
grazing. Good et al. (1994) suggest that either a better receptor site should have been
chosen, or techniques should have been developed to modify the receptor site. Grazing
should also have been introduced at the receptor site.
Brocks Farm
The transplanted turves at Brocks Farm showed some botanical changes compared with
the pre-transplantation state. Most of these changes were probably due to the
reinstatement of cutting before the translocation was carried out; both the control and
transplanted vegetation were showing some similar changes (Leach et al. 1995). For our
purposes this unfortunately masks any effects of the translocation. It seems that generally,
the changes due to translocation were relatively few, probably because of careful turf
transfer and prompt reinstatement of the management. However, Leach et al. (1995)
report that although there has been a slow increase in long-lived, slow-growing species,
such as the orchids Dactylorhiza praetermissa and Orchis morio, in the tranplanted turves,
this increase has been less than at the control site. On the transplant site they have
probably been less able to benefit from the reinstatement of cutting and grazing. These
effects are thought to be both short term (disturbance) and long term (environmental
changes at the receptor site).
Review of information, policy and legislation on species translocations
Thrislington Plantation
Although turves were translocated over several years at Thrislington, all seem to have
shown similar responses, with few changes in the plant species compositions compared to
undisturbed grassland. However, three calcareous grassland species were lost, including
Linum perenne anglicum, and there were some significant variations in species'
abundances. The major change (except for invasions in some areas, see below) was that a
few resident species invaded the interstitial gaps in the turves and that many gaps
remained for some years after the translocation (Byrne et al. 1991). A survey by Cullen &
Wheater (1993) on a number of plots translocated at different times showed similar
results, with little difference in species composition or diversity index among the different
plots (their claim for a complex temporal change in these values is not supported by their
data), although the bare ground in the more recently turfed plots lead to lower cover
values for a number of species.
More recent statistical analysis of changes in individual plant species up to 1994 (unpub.
report, EN Northumbria team, S. Hedley pers. comm.) has indicated a more negative
response to transplantation. Sesleria albicans (a key component of he sward) decreased in
two transplanted plots, compared to stability in undisturbed grassland, while the ruderal
Sonchus arvensis was increasing. S. Hedley (pers. comm.) states that S. albicans is
declining with a parallel increase in Bromus erectus, possibly because the grassland is
becoming rank. Grazing has been introduced to part of the receptor site to try to overcome
this. S. Hedley (pers. comm.) and Byrne et al. (1991) also report that the structure of the
grassland community has been disrupted by translocation. Vegetation patches have been
broken up and redistributed, leading to 'unnatural patterning' of the relocated grassland.
The Newhall 'translocation' differed from the others in that the turves were replaced onto
the original site. However, there were a number of environmental changes to the site after
the turfing operation. The soil profile was changed dramatically; a pre-turfing depth of
about 0.23m was changed to 0.1m of turf overlaying 0.2m of subsoil and 0.1m of rubble.
The site was flat before turfing, but had a slight south -west slope after the turfing. The
turves were not replaced in the same pattern as before lifting and there were also gaps
between the turves after re-laying. These lead to changes in species composition. A few
species were lost from the vegetation, particularly some deep-rooting orchids which may
have been damaged by the shallowness of the turves. Other species declined in the long
term, including some with high moisture demands; indicating a freer drainage after
turfing. A few species increased in frequency, some were ruderal and fast-growing
(indicating disturbance and nutrient release), and some were drought-tolerant (again
indicating freer drainage) (Cox et al. 1992).
At Potatopot many plant species were lost over the three years after translocation and
other characteristic plant species of the donor site showed declines in frequency (Jerram
1993). Some species increased in frequency, and these were typically tall species which
are competitively superior in lightly-grazed grassland. This was probably mostly due to
Review of information, policy and legislation on species translocations
the absence of grazing, an hypothesis which is supported by the fact that similar changes
were occurring at the ungrazed donor site before translocation (Leach et al. 1991). Jerram
(1993) perceives a need to increase grazing intensity, perhaps using cattle instead of
sheep. However, cattle would be threatened by injury from the occasional large gaps
between turves. Thus, the changes were probably not due to the transplantation itself, but
rather to the inappropriate management. However, Jerram (1993) also suggests that some
of these botanical changes could have been due to differences in the hydrology of receptor
and donor sites (perhaps explaining the increase in some species which prefer damper
conditions), and nutrient release caused by soil disturbance in the translocation.
Plant communities after soil transfer
Selar Farm
Plant community changes were greater in the 'spreading' treatment at Selar Farm than in
the turfing trial. The large area of bare soil allowed a great expansion of certain grass
species at the expense of other plant species. About 40-50% of the original species were
lost in this treatment.
Brocks Farm
The transfer of excavated soil and plant material at Brocks Farm ('blading') led to much
greater differences from the control plot than shown by the turf transplants. There were
initial losses or very low frequencies of many species, although a proportion were
recovering six years later (Leach et al. 1995). At this time a number of other species were
much more abundant than in the turf or control areas. However, S. Leach (pers. comm.)
reports seven years later that parts of the bladed area are developing into a species-rich
sward, with some areas resembling the control.
Blading at Potatopot lead to losses of many plant species compared to the donor site and
large fluctuations in species composition, particularly through an initial dominance of
ruderal species which established on the bare soil. Obviously the vegetation structure was
totally disrupted by the technique and bare areas still remained in the last survey, three
years after translocation. As in the turves, the vegetation was becoming dominated by tall
species, due to the poor grazing management (Jerram 1993).
The blading at Ashcott lead to the loss of many species seen previously at the donor site,
particularly slow-growing species (due to the massive disturbance and, possibly, the lack
of management) and species preferring wetter conditions (possibly due to the receptor site
being drier than the donor site). Many other characteristic species declined. A number of
ruderal species increased in the first year, due to the disturbance, but subsequently
declined as the vegetation closed up (Cox et al. 1991). Since these measurements the
receptor site has been flooded by the creation of a wetland on the nature reserve, and this
has almost certainly destroyed the developing community (S. Leach pers. comm.).
Review of information, policy and legislation on species translocations
Invertebrate communities after turf translocation
The fauna is rarely monitored in community translocations, but three studies indicate
losses of invertebrate species.
The invertebrate populations on the receptor area after translocation were compared with
those on the whole of the Dongas before translocation (of which the turfed area was only a
part). Therefore, it is not known if any species were lost in the translocation. This is
because, if species were found on the Dongas but not at the receptor site this may not
indicate losses of species, but rather the fact that they were not in the turves even before
translocation. However, certain invertebrate species found in the Dongas were also present
in the translocated turves after three years, indicating their successful translocation. These
included the chalkhill blue Lysandra corridon (translocated as caterpillars), yellow
meadow ants Lasius flavus and the 'Notable a' crab spider Xysticus acerbus. Many other
species colonised subsequently. In the first two years these were mostly early successional
and primary colonising species, but after three years other species were colonising.
Although it is difficult to say if species were lost in the translocation, comparison after one
year of the butterflies at the receptor site with those in the undisturbed vegetation adjacent
to the donor site showed nearly half of the species found in the latter were not present in
the former (possibly lost in the translocation), and most of the remaining species had
lower densities (Snazell et al. 1995a). One year later the translocated vegetation had
gained many of these lost species and most of the butterfly populations were increasing in
size (Snazell et al. 1995b).
Thrislington Plantation
Although there were few effects of translocation on the plant communities of the
Thrislington turves, the invertebrate fauna responded more negatively. Sheppard (1990)
found that only about half of the invertebrate (Aranae, Opiliones, Heteroptera, Homoptera,
Coleoptera) species found on a pre-transplantation plot remained three years after turfing
(although there were many new species, see below). Sheppard (1990) concluded that
translocated grassland was poorer in calcareous grassland species especially those
characteristic of wetter grassland. This indicated drying out of the turves.
Cullen & Wheater (1993) compared the numbers of individuals of 17 invertebrate orders
(covering arachnids, crustaceans, myriapods, molluscs and insects) on turves translocated
for different time periods at Thrislington. There were few differences in the relative
proportions of each order, although there seemed to be more ground-dwelling predators on
the younger plots. However, there were more individuals in both the youngest and oldest
turves. Cullen & Wheater (1993) suggested that this indicates disruption of the
invertebrate community by translocation (perhaps with early successional species in the
younger turves) which causes a decline in numbers, but this is followed by stabilisation of
Review of information, policy and legislation on species translocations
the community. Cullen & Wheater (1993) could not address the question of changes in the
species community before and after translocation.
Although there were few changes in invertebrate species numbers, the Ashington
translocation seemed to have resulted in a major change in the species compositions of the
beetle and spider communities (Piekarczyk 1991). Almost a quarter of species present on
the control site were missing from the translocated turves (many invasions occurred as
well, see below). The lost species tended to be either moisture-loving or intolerant of
disturbance, indicating increased drainage and disturbance of the translocated turves. Invasion of species not present at the donor site
The original communities may be changed by invasion of new species. New plant species
may establish from the donor or receptor seedbank or may invade from surrounding
communities. Animals may be present at the receptor site or also colonise from the new
surroundings of the translocated community. Invasions may be facilitated both by
differences in abiotic and biotic environmental conditions compared to the donor site, and
by the proximity of new sources of invaders. This may lead to negative effects on original
species through competition, predation, parasitism, pathogenicity, etc. and a change in the
community characteristics as the invading species establish.
In two of the case studies, Middlebere and Newhall, there was little or no invasion of plant
species, but the other studies all showed some invasion.
Plant communities after turf translocation
A number of undesirable 'weed' plant species, not seen in the donor site, established in the
Dongas hand translocated turves in the year following translocation. The number of such
species did not diminish over the first three years and certain species were increasing in
cover or had to be controlled (Ward & Stevenson 1994a, b).
Weedy plant species also established in the machine translocated turves on the Dongas,
although not as many as in the hand translocated turves. This was due to the revegetating
of areas left bare by the translocation. R. Pywell & M. Stevenson (pers. comm.)
categorised the vegetation of the donor site and of the three year old translocated turves
into chalk grassland species, generalist species and 'weedy' (ruderal) species, using NVC
lists. The chalk species number was fairly stable, but the number of generalist and weedy
species doubled after translocation. However, Ward & Stevenson (1994a, b) found that the
cover of these species declined greatly over the first three years.
Review of information, policy and legislation on species translocations
Although some ruderal plant species were found in the Hockley translocation, the lack of
many bare areas in the turves meant that these were not a major problem and were easily
controlled (Ward & Stevenson 1995).
Selar Farm
At Selar Farm the whole turves gained between three and four new species (Good et al.
1994). Some of these were fast-growing, invasive grass or herb species.
Brocks Farm
Monitoring at Brocks Farm (Leach et al. 1995) indicates a few new species invading the
turves, including low numbers of some ruderals. Some species, such as Carex hirta,
originally invaded from around the edge of the transplant and are still spreading.
Thrislington Plantation
Most of the Thrislington turves suffered extensive invasions from ruderals, especially
those which seemed to have been more 'roughly handled' (Byrne et al. 1991). It is not
clear whether these have decreased in cover subsequently.
Despite the large area of gaps between turves at Potatopot there was no colonisation of
any new invasive species. A few new species were found, but all were characteristic of the
vegetation type (Jerram 1993).
Plant communities after soil transfer
Selar Farm
The 'spread' turves at Selar Farm gained five new plant species (Good et al. 1994).
Brocks Farm
A number of new species, including some ruderals, are seen in the species lists at Brocks
Farm (Leach et al. 1995). Short-term invasions included Juncus bufonius and longer-term
ones Ulex europaeus, which was controlled by cutting and grazing.
Many new species established in the bladed areas at Potatopot. Some were ruderals which
subsequently declined, but others have become well established and indicate a significant
change from the donor site.
Review of information, policy and legislation on species translocations
Many ruderal species colonised the receptor site after the blading' at Ashcott, due to the
disturbance. These declined subsequently. A few other species which are characteristic in
conditions of high nutrients and light grazing invaded the sward and were continuing to
increase at the final census. However, some or all of these may have been present in the
original receptor site vegetation, which was not removed (Cox et al. 1991).
Faunal communities after turf translocation
At the Dongas site a number of early-successional invertebrates colonised the translocated
turves initially, but many were lost as the community developed, and many of the species
seen at the donor site recolonised the turves after three years (Snazell et al. 1995a, b).
Over a third of the beetle and spider species found on the Ashington translocated turves
after 18 months were not seen in the undisturbed grassland. These were species
characteristic of either lower moisture levels or greater disturbance (Piekarczyk 1991).
Sheppard (1990) found that about two fifths (51 of 128) of the invertebrate species
(arachnids, bugs and beetles) found on the turves three years after translocation were
apparently new compared to a survey before the turfing, and many of these 'colonising'
species (38) were apparently new to the Thrislington site. The new species were
characteristic of either tall, dry grassland or dry, open areas. This suggests responses to
both the unvegetated gaps between the turves and rank growth of the turves, and to a
possible drying out of the turves. Genetic effects
A reduction in effective population size after translocation due to poor translocation or
subsequent declines in the receptor site (see above) may lead to adverse genetic
consequences of small population size - inbreeding depression, genetic bottlenecks etc. which will lead to changes in genetic structure and may cause population extinction.
Environmental changes may lead to changes in selection pressures which will cause
changes in genetic structure.
These possibilities and those discussed for metapopulations and ecological landscape are
hypothetical and have never been studied in the context of community translocation.
Review of information, policy and legislation on species translocations Changes in metapopulation dynamics and dispersal among sites
If a species population at the donor site existed as part of a metapopulation - especially a
Levins-style or source-sink metapopulation - the translocation of the population may
cause disruption of the population itself and of other linked populations in the donor
landscape. Two worst-case scenarios could occur for source-sink metapopulations. If the
population at the donor site was a sink, its translocation may take it too far from a source
population for the level of immigration to allow persistence. Alternatively, the population
could be a source itself and its removal from the donor landscape may cause population
extinctions at other sites. This second case may be more likely. The communities that are
translocated are often of high conservation value because they hold reasonably large
populations of rare species, and therefore they are more likely to have source populations
for these species. A possible positive effect may transpire for the receptor landscape if the
new population either acts as a source for new populations or benefits the local
metapopulation dynamics.
Changes in metapopulation dynamics may also affect gene flow and patterns of genetic
variation among the metapopulations, with unknown consequences.
Some animals, especially larger vertebrates, will not use the donor habitat exclusively, but
will use it as a habitat patch within their range. The removal of such habitat patches may
result in population decline, or at least a disruption of the spatial dynamics of populations.
This will be especially true for species restricted to that community type.
Where such disruption occurs it may be a useful generalisation that the consequences will
be less severe where the receptor site is closer to the donor site. The metapopulation
dynamics or patchy habitat use may be little affected if the population is shifted only a
short distance. The case studies showed a range of translocation distances between a few
100m up to 25km. However, physical barriers to dispersal (e.g. roads) may negate the
positive effects of short translocation distances.
A different possibility arises if a species is lost from the translocated community or has a
small population size after translocation. Dispersal from other sites may allow
recolonisation of the species or may build up a small population. The Dongas study
showed exactly this happening. Many of the butterfly species lost or diminished in the
translocation process had built up populations two years later (Snazell et al. 1995b). It is
important to note that the receptor site was not far from the donor site and was close to
many other chalk grassland communities (J. Thomas pers. comm.). It is impossible to
extract this sort of information from the plant data of any of the case studies. Landscape effects
Ecological landscapes
The movement of communities inevitably disrupts the landscape pattern. This may have
ecological consequences. The continuity among landscape elements may be decreased
Review of information, policy and legislation on species translocations
(i.e. fragmentation is increased), especially if the site is part of a larger complex of patches
of a biotope. Fragmentation on a smaller scale may result if only a part of a site is
translocated, causing disintegration of a single site into two or more smaller and separated
parts. This happens in most, if not all, translocations. The NCC (1988a) guidelines showed
that in six translocations between 4-63% (average = 33%) of the original site was
translocated. Effects on metapopulation dynamics and population viability are discussed
above, but there are also possible negative consequences of edge effects. The smaller
biotopes, with a greater ratio of edge to area, may be more susceptible to invasion from
adjacent biotopes.
Aesthetic landscapes
This disruption of the historical pattern of biotopes in the landscape also has aesthetic and
cultural consequences. Every biotope patch, particularly those of high conservation status
that are the main target of translocation projects, has a value as part of the wider
countryside. The present and past systems of land use and management of the site and the
surrounding biotope patches have interacted with local environmental factors to produce a
unique and irreplaceable landscape pattern. Other than species and habitat (biotope)
conservation this aesthetic landscape perspective is another reason for the objections of
the conservation organisations (NCC 1988, Hopkins 1989, C. Pulteney pers. comm.) to
certain developments and the consequent need for mitigation by community translocation.
It is also arguable that such landscape considerations are more important to many
members of the public than conservation of particular species or habitats. Effects on the remainder of the original community
As shown above, most community translocations do not involve the relocation of the
whole of a site (i.e. a biotope patch). Sometimes only the area directly affected by the
development is moved, leaving some of the site intact. Often the amount moved is limited
by cost and the remainder is entirely lost to the development. Thus, only a portion of the
community is translocated. Some of the threats to the translocated portion may also apply
to the remaining portion. There will be fewer changes in the abiotic environment, although
disruption of certain processes such as hydrology and nutrient cycles may occur through
the removal of turves. However, disturbance of the untranslocated area during the
translocation and effects of decreasing both population size and habitat area of the resident
species may lead to changes in the community that remains. There are no studies that we
know of on the possibilities and consequences of these effects.
The arguments against community translocation can be divided into two categories:
arguments based on the risk of failure of the translocation; and 'ecological landscape' and
'aesthetic landscape' arguments against the translocation of any community, even if
completely successful. Both are rehearsed by the NCC (1988b) guidance document. The
aesthetic landscape argument is one of aesthetics rather than evidence. There are no
specific data to allow us to assess the risks to the translocated communities and the wider
landscape due to spread of translocated species outside the receptor site, genetic problems,
Review of information, policy and legislation on species translocations
disruption of metapopulation dynamics and dispersal among landscape patches, increased
habitat fragmentation and edge effects, and effects on the remainder of the community at
the donor site. However, there is information on some of the influences affecting the
successful translocation of species populations from the donor site to the receptor site.
5.6.1 Influences on the outcome of community translocation
Use of soil transfer and other inappropriate techniques
It is inaccurate to describe the excavation and transfer of soil and parts of vegetation
('blading' - see above) as community or habitat translocation. This method will kill many
plants and invertebrates, totally disrupt the soil structure and chemistry, the plant and
animal community structure and the ecosystem processes, and create an early successional
community which in some circumstances may develop into a very different community to
the donor one. EN guidance to road builders on environmental mitigation (EN 1993),
criticises the use of soil transfer as a method of community translocation for similar
reasons. Byrne (1990) found 14 projects on the Habitat Transplant Site Register which
used blading and showed that where blading and turfing had been used in the same
project, the blading always gave poorer results in terms of successful transfer of all plant
Worthington & Helliwell (1987) describe such a project and it is clear that this technique
is useful as a method of introducing soil and propagules of a good selection of plant
species to a bare landfill site, rather than as an attempt to conserve the donor community.
Another project carried out by Helliwell (1990) used similar techniques to mitigate for a
large construction project. A full assessment of either study is difficult because changes in
species composition were not analysed (although Hodgson 1990 points out that in the first
project the rarest plant species, Oenanthe fistulosa, was lost). Helliwell (1990) describes
both projects as successes, although he talks purely in terms of establishing a grassland
roughly resembling the original one.
The Ashcott Heath project, although perceived by the developers as a method of
community translocation, clearly falls into the same category. In addition, the failure to
remove the community at the receptor site would also have strongly influenced
community development. As described above, this method lead to the development of a
community very different to the original pre-translocation sward. The Ashcott example,
with the use of 'blading', the failure to immediately spread out the translocated material on
the receptor
site, the lack of management and the subsequent flooding of the site, could be seen as an
example of the worst way to carry out a community translocation.
The same problems apply, to a lesser degree, to the 'blading' treatments at Brocks Farm
and Potatopot. In both cases there were large changes in plant species composition and
relative abundances, including the invasion of new species. S. Leach (pers. comm.) points
out that the 'bladed' area at Brocks Farm has produced a high quality, plant species-rich
grassland which is improving yearly. However, this was only a success in introducing
Review of information, policy and legislation on species translocations
propagules and live plants to the receptor site and managing this correctly to produce a
new grassland of botanical interest. It is still very different from the original donor site or
the translocated turves (Leach et al. 1995), and the invertebrate fauna has not been
measured. The 'spreading' technique used at Selar Farm was also much less successful
than the turfing. Other authors describe such techniques in terms of restoration of plant
communities on damaged sites (see Chapter 2), rather than as methods of community
translocation (EAU 1988, Putwain & Gillham 1988, Pywell et al. 1995).
The woodland translocation scheme developed by Down & Morton (1989) involved: the
movement of individuals trees by excavating the rootball, cutting the trees back (the larger
ones to coppice stools), and replanting at the receptor site; planting of new trees; and the
collection of individual specimens of the ground flora, propagating them in a nursery, and
planting them out at the receptor site. Even more so than the soil transfer technique, it is
unlikely that such a method could allow the establishment of a community very similar to
that at the receptor site.
Turf translocation
Plant communities
All of our case studies involving turf translocation showed changes in the plant
communities which were attributable to the translocation projects. However, the Dongas,
Thrislington and, possibly, Brocks Farm showed fewer and more minor changes than the
others. Changes were due to loss of species, invasion of new species or changes in
abundance of surviving species, and were ascribed to the following factors.
The lifting, transport and re-laying of turves killed or damaged some
plants and often left gaps between the turves. This resulted in the loss or
decline of some species and allowed others to invade or spread into the
open areas (especially ruderals) or to take advantage of the nutrient
pulse caused by the disturbance (fast-growing species). This cause was
cited for the Dongas, Hockley, Selar Farm, Brocks Farm, Thrislington,
Potatopot and Newhall.
Turf depth
A form of disturbance, the severing of roots through too shallow a depth
of turf, was cited as a cause of the loss of orchid species from the hand
dug turves at the Dongas, and from the turves at Newhall.
The dry weather that followed turf translocation at Middlebere was
thought to have caused death of some turves.
Differences in the abiotic environment of the receptor and donor sites
were commonly thought to have lead to losses and gains of certain
species and to have changed relative abundances of species. In most
cases, a range of variables were probably influencing change, but,
because these variables were rarely measured, definitive statements on
the causes of change could not be made. A changed level of the water
table was a cited factor at Hockley, Middlebere and Potatopot. At Selar
Review of information, policy and legislation on species translocations
Farm differences in topography, soil type and soil pH were thought to be
important, while topography and soil profile were altered at Newhall.
In many cases the management of the translocated turves was different
to that of the donor site. Sometimes, the management of grazing or
cutting was relaxed initially to avoid damage or disruption of the turves,
but was then reinstated. In other cases, there were no plans to reinstate
the management. Such changes were thought to be causes of variations
in species composition and/or abundance at the Dongas, Hockley and
Potatopot. Brocks Farm forms an interesting counterpoint to these
results; the reinstatement of traditional management on the translocated
turves after the donor site had been neglected for some time lead to
changes in the plant community away from that recently found at the
donor site, but towards the type that was presumably traditionally seen
at the donor site.
Our case studies are not unusual. Other studies of a range of community types have all
shown floristic changes which can be attributed to the translocation. These changes
occurred despite the claims of 'success' by some authors. Lifting by machine bucket and
translocation of deep turves of a species-rich grassland in southern England resulted in
survival of 80% of the plant species and limited invasion by new species (17% of the total
species number) after three years (Anderson 1989). Humphries et al. (1995) report the
results of another two EFU translocation studies, Westhay Heath, Somerset and
Monkspath Meadow, West Midlands (see also Byrne 1990). The species-rich hay meadow
at Westhay Heath underwent dramatic changes after translocation, due to changes in the
water table. Monkspath Meadow was also a hay meadow with many rare plant species,
and the turves showed some changes in species abundances after translocation caused by
reductions in cutting and grazing.
EAU (1988, p69) report the translocation of heathland in Dorset which, over five years,
showed changes in species composition caused by the greater drainage at the receptor site.
Machine-lifting and translocation of turves in a sand prairie in Wisconsin, USA resulted in
the loss of 13% of forb species. Initially, resident weedy exotic species increased in cover
and a number of exotics also invaded the vegetation (Kearns 1986). Bragg (1986) found
that all plant species survived the translocation by hand of an area of tallgrass prairie in
Nebraska, USA, although ruderal species increased in cover and others, including invasive
exotics, colonised the sward. Stiegman & Ovenden (1986) translocated tallgrass prairie in
Texas, USA using a commercial sodcutter and found survival of 78% of plant species over
two years. Although there was little change in the dominant species, the relative
abundances of a number of translocated species changed in comparison to undisturbed
prairie, and a number of ruderal species invaded to become minor components of the
It must also be noted that monitoring of the plant community in most projects has
continued for a very short time. The case studies presented here have been monitored for
3-4 years on average (although monitoring has not ended in many of these studies). It is
very unlikely that the plant communities will have stabilised over this time period. There
is an inherent time lag in the response of plant communities to environmental change, as
Review of information, policy and legislation on species translocations
individual plants must complete their life cycles and new individuals must be recruited
and establish into the vegetation. Invasion and population increase of new species may
also take many years to stabilise. It is only once these population processes have settled
down to a new dynamic equilibrium that we can finally say what the consequence of the
translocation are. It is therefore important that the case histories with the longest period of
monitoring, Brocks Farm and Thrislington, are reported to be showing continuing trends
of changes in species abundances after six and eight years respectively. The community
changes reported here may not represent a complete portrait of the effects of translocation.
Animal communities
Invertebrate communities are rarely monitored in community translocation projects. Of
those projects that have included consideration of invertebrates, it is interesting that in
two, Thrislington and the Dongas, which showed fewer changes in their plant
communities than most projects, the invertebrate communities seemed to show large
changes. The apparent losses of species at Thrislington, Ashington and the Dongas
immediately followed the translocation. This indicates poor initial translocation success,
either due to mortality or escape (especially of winged insects or mobile ground-dwellers,
such as large carabid beetles) during translocation or the disruption of habitat immediately
following translocation. However, the development of rank vegetation at Thrislington,
caused by altered management, lead to changes in the invertebrate community.
These conclusions should be treated cautiously, because the efficiency of sampling of
different invertebrate taxa can be affected by the vegetation structure (Snazell et al.
1995a). Therefore, if the translocation affected this structure (e.g. bare areas or a different
vegetation height) the detected changes in the fauna may not reflect real changes.
Another study indicates the complex consequences of disruption of a food web by
translocation. After five years translocated salt marshes in Texas, USA had changed
animal communities compared with undisturbed marshes. There were lower amounts of
detritus, leading to lower numbers of polychaetes and amphipods which in turn caused
depressed densities of the predatory decapod crustaceans, although fish densities were
unaffected (Minello & Zimmerman 1992).
These results indicate that translocation can severely damage the invertebrate
communities, even if the plant communities are only slightly changed. The qualification
that the full extent of community changes are not known because of the short period of
post-translocation monitoring applies to invertebrates as well as plants.
5.6.2 Policy on the use of community translocation
The aim of community translocation
It is clear that community translocations can never or rarely fulfil the 'conservation aim'
that we described above: 'to translocate, unchanged, all populations of every taxon of a
community from within a prescribed area'. This is not an excessive requirement. As we
Review of information, policy and legislation on species translocations
showed, the NCC review of translocations (Byrne 1990), stated similar requirements: ' the
[translocation] of a complete assemblage of plants and animals, with the aim of
maintaining the habitat unaltered in its new location'. Such an aim must take account of
the fact that the composition and relative abundances of species in a community will
always fluctuate due to temporal variation in the environment (e.g. climate) and
demographic processes. Thus, the loss of two species from the Dongas translocation (see
section was credited to demographic fluctuations rather than an effect of the
translocation. Evidence of adverse effects of the translocation can only be from consistent
trends of community change which can be attributed to the translocation. These consistent
trends would include the loss of many species, invasion by many novel species, or large
and persistent changes in the relative abundances of species. Such changes were seen in
many of the case studies that we have assessed.
If we add to these established risks of poor translocation success the possible (and more or
less probable) risks of long term ecological and genetic changes in the translocated
community and the donor and receptor landscapes suggested above, the conservation
argument against community translocation as a viable alternative to in situ conservation is
convincing. The 'aesthetic landscape' argument also leads to the same conclusion, but this
could be dismissed on the grounds that it generally constrains development of certain
areas (however valid this argument may be), rather than being a criticism of community
translocation itself.
These arguments are less valid if one creates different aims and criteria for success for
community translocation. For example, EAU (1988) report the success of five turfing
projects on heathland in Dorset. However, this 'success' was stated only in terms of turf
survival (in one other project many of the turves died); there were no measures of
vegetation change or stability. While not being so general, the 'amenity aim' described
above ('to translocate a plant community and, it is to be hoped, some of the other taxa,
such that the translocated community resembles the pre-translocated state) sets a less
stringent criterion for success than the 'conservation aim'. All our case studies, apart from
Ashcott, unequivocally fulfilled this aim.
The 'amenity aim' is also similar to the usual aim of habitat/community restoration or
creation projects: 'to establish semi-natural vegetation communities which in some way
resembles the semi-natural original, although not necessarily to re-create their full
diversity' (Buckley 1989). The volume edited by Buckley (1989) includes community
translocation as one of the techniques of community restoration (called 'habitat
reconstruction'). While translocation (including soil transfer) could be seen as a very
effective method of creating a new plant community on a damaged site (e.g. Pywell et al.
1995) or compensation area, it is much less effective as a method of preserving, with few
changes, the full community of the donor site.
Translocation as rescue or preservation
An article by an environmental consultant (Tyldesley 1993) states that 'translocation is not
a substitute for in situ conservation. It cannot avoid demonstrable harm, it cannot
compensate for the loss to nature conservation and it cannot remove the proposal's conflict
with national, strategic and local planning policies'. An EN report providing guidance to
Review of information, policy and legislation on species translocations
developers on mitigation of the impacts of road building (EN 1993) states that
translocation 'does not provide compensation for loss or damage to high value, nonreplaceable sites'. The NCC guidelines (1988) described above, are based on this
argument. The Canadian Botanical Association have issued a policy statement (see Fahselt
1988) in which it states a strong opposition to the idea that community translocation can
be used as a reliable method of conserving rare species. This is based on concerns about
the risks involved, especially those due to changes in environmental conditions. The
Independent on Sunday (18 June 1995) summarises this view by describing one such
project as 'a consolation prize'.
Therefore, community translocation should be used only to mitigate the total loss of the
plant and animals of a community (a 'rescue') and not as viable alternative to in situ
conservation (i.e. 'preservation' of an intact community). Indeed, the Dongas translocation
was seen by the ITE primarily as a good technique for re-creation of chalk downland in
the compensation area rather than as a way of preserving the Dongas community
(although this was an incentive, R. Snazell pers. comm.). This distinction may be great
and seem clear; however, it is not made by some developers and others. For example, the
only drawback to community translocation given in the guidance notes by the Department
of Transport, Scottish Office Industry Department, The Welsh Office and Department of
the Environment Northern Ireland (1993) is its high cost. A recent newspaper article on
plans to carry out another translocation at Brocks Farm ('Plan to move field turf by turf
resisted by conservationists' Guardian July 31 1995) contains a quote from a spokesman
for English China Clay: 'We have moved similar herb-rich grassland before, and we know
it works'.
To be fair to developers some, such as the Highways Agency (R. Kent pers. comm.),
realise the difficulties involved in translocating turves so that species loss and changes in
composition are minimised. However, because of this confusion the NCC (1988b)
guidelines express a concern that the involvement of staff in planning translocation
projects for rescue motives could be seen as endorsing the view of its suitability as use for
The role of community translocation in the planning process
The confusion between the use of translocation for 'rescue' as opposed to its use for
'preservation' has important implications for the planning process. 'Preservation' implies
that development of a site of conservation value is of little consequence because the
community can be moved out of harm's way, whereas 'rescue' gives the different message
of a damage limitation exercise to save what one can from a community that is to be
otherwise destroyed. Obviously the second interpretation reflects the reality from a
conservation perspective.
This difference can be made clear in the planning process. Tyldesley (1993) transcribes
the Secretary of State's refusal in 1992 of planning permission for a housing development
on an SSSI. This refusal takes the view of translocation as an imperfect technique. 'In my
view, the objective of any translocation scheme ... must be to replicate as closely as
possible the nature conservation interest and value of [the site]'. He interprets this as
including the successful establishment of colonies of a particular rare butterfly as well as
Review of information, policy and legislation on species translocations
survival of the 'same wide range of plants and species-rich vegetation'. However, 'given
the acknowledged difficulties of translocation of the key species ... I consider the prospect
of successfully achieving this wider objective to be poor'. After statements sympathising
with the view that translocation is not a substitute for in situ conservation and that even a
highly successful translocation may create new habitat but will not re-create the SSSI he
concludes: 'As a means of allowing development I see it at best as a technique which may
be applicable where the chances of a successful translocation are sufficient to tip a finely
balanced case in favour of allowing development to go ahead. To my mind to regard it as
any more than this would seriously undermine the intent of national and local policy to
protect the nature conservation value of SSSIs'. Despite this statement, development on
SSSIs continues.
To avoid the promotion of translocation as a substitute for in situ conservation Tyldesley
(1993) suggests that the possibility of community translocation should not be considered
until the planning decision has been made. By this he means that the development has to
be considered to be of such importance as to outweigh the importance of a site for nature
conservation. One criticism of this is that the lack of consideration of translocation in the
planning permission process may lead to no or insufficient provisions for translocation if
permission is granted. This would apply to the planning process behind the M3 Motorway
between Bar End and Compton (the Dongas and Hockley projects described above).
When the construction project was approved, the requirement for compensation made only
vague provision for the restoration of an area of chalk downland and gave no conditions or
objectives for this. The highly technical and expensive compensation work, which
included translocation, was only carried out after subsequent consultation between DTp,
ITE and the Hampshire Wildlife Trust (with some input from EN).
It seems more appropriate to consider translocation and other mitigation procedures while
deciding on the granting of planning permission, but, in the light of the above
reservations, to consider translocation only as a method to mitigate in some way for the
unavoidable loss of the donor community. The granting of planning permission must
therefore include the concession that unalterable damage will be done to the donor
community. This process can allow the addition of conditions to the permission to compel
both translocation and the
use of techniques to enhance the value of translocation as a rescue. Expert judgement,
such as the statutory conservation agencies could provide, would be essential at this stage.
Increasing the value of a translocation
If translocation is to be carried out as a form of rescue, its value will be increased by
limiting the effects on the translocated community and the surrounding landscape to a
minimum. This can be achieved by considering the following points (see also the
guidelines described above and Byrne 1990).
Review of information, policy and legislation on species translocations
Type of community
Certain vegetation types will transfer with fewer changes than others. Dry grasslands or
dry heaths seem to transfer most successfully, and the shallow rendzina soils of chalk
grasslands facilitate the lifting of the whole soil profile. Wetter communities are more
difficult, and there are great risks associated with the disruption of hydrological patterns,
water tables and soil structures. Tall vegetation, such as woodlands or tall shrub
communities, with their complex structure and deep rooting, will also be difficult to
translocate without great damage.
Translocation techniques
As discussed above, turfing will almost invariably result in fewer community changes
than methods such as soil transfer. Deeper turves will allow better preservation of the soil
structure and decrease damage to roots and soil invertebrates. At the Dongas and Hockley
the deeper turves gave better translocation results. Bragg (1986) found that translocation
of deeper turves of tallgrass prairie increased the survival of a number of species. Larger
turves and smaller gaps between re-laid turves both seem to reduce disturbance and both
are achieved by the use of machinery for lifting as well as re-laying (e.g. the Dongas and
Hockley). Kearns (1986) tried a number of methods for translocating prairie vegetation,
and found that machine lifting of turves gave the best result for survival of plants,
maintenance of species and avoidance of invasion. Macroturfing (see Pywell 1991) has
been developed to allow the use of large and deep turves to reduce disturbance of the
turves and to enable operators to place the turves close together, and it has been used in a
number of projects on heathland, meadows and chalk grassland (Pywell 1993). Its use on
the Dongas allowed the successful translocation of buried nests of several ant species (J.
Thomas pers. comm.)
Choosing the appropriate time of year (usually autumn/winter) may also be important.
Choice and preparation of the receptor site
Surveys of the donor and receptor site can indicate environmental differences and allow
decision to be made as to whether a proposed receptor site is appropriate (soil type,
fertility, topography, drainage etc.) and/or the appropriate preparation of the receptor site
(e.g. soil removal, changing drainage or water table, manipulating pH). There seems to be
little consideration of alternative sites in translocation projects. Good et al. (1994)
suggested that the translocation at Selar Farm would have been more successful if they
had been able to asses and choose between alternative sites.
A second, much less discussed, consideration is the position of the receptor site. The
potential problems due to disruption of metapopulation dynamics and inter-patch dispersal
and disturbance of landscape pattern will be reduced if the receptor site is close to the
donor site. For instance, the Dongas turves were moved to another, nearby part of the
same downland area and this probably allowed rapid colonisation of a number of
Review of information, policy and legislation on species translocations
invertebrate species (R. Snazell, J. Thomas pers. comm.). If only part of the donor
community is moved and the receptor site is part of, or is joined to the donor site the
negative consequences of dividing the populations may be avoided or reduced.
Subsequent management of translocated community
We showed above that many community changes were blamed on an alteration in
management following translocation, rather than on direct effects of the translocation
itself. A review of community translocation, summarised by Humphries et al. (1995),
suggested that major factors which increased change in grassland plant communities after
translocation were removal or variation of grazing and/or cutting and inappropriate
environmental conditions at the receptor site.
Surveying and monitoring
To assess fully the consequences of translocations and to refine methods to improve the
preservation of communities, full surveys of the environment, management and plant and
animal communities of the donor and receptor sites are needed. As we state above, the
changes in plant and animal communities may take many years to stabilise following
translocation. Therefore, monitoring of the translocated communities should continue to
assess these changes. Many translocations are inadequately monitored and where
monitoring occurs this may only be for a short time. The intensive botanical and
invertebrate monitoring by ITE on the Dongas, which will continue for ten years, and the
continued monitoring of Brocks Farm and Thrislington Plantation by EN are examples of
the effort needed to assess the outcome of translocations. The ITE project provides a
model approach, although this covers a relatively short period. Monitoring of the
remaining community at the donor site or a similar undisturbed area, (as at Brocks Farm
and Thrislington) may allow an assessment of whether community changes can be linked
to the translocation, or are part of larger-scale trends (e.g. due to climatic fluctuation).
The monitoring project set up by the EFU has been maintained fitfully by EN. Monitoring
of plant communities has continued in only some sites and in some cases this is not
detailed and/or has not been fully analysed. This has lead to problems with inconsistency
in species identification and quality control. The Thrislington project also suffers from a
lack of funding and personnel to carry out analysis of data (S. Hedley pers. comm.).
While many community translocation projects involve monitoring of the plant
community, very few have included consideration of the animal communities. While
vertebrates will rarely, if ever, be translocated it is worthwhile to monitor the use of the
translocated community by birds, mammals, reptiles, amphibia or fish. The poor
monitoring of
invertebrate communities reflects the difficulties of monitoring rather than a lack of
appreciation of the importance of invertebrates. Twenty six of the 77 sites on the EFU
Habitat Transplant Site Register were also on the Invertebrate Site Register (see English
Nature 1994c), being of interest for their invertebrate communities (Byrne 1990). This
major part of the conservation interest of a community should not be ignored.
Review of information, policy and legislation on species translocations
Monitoring programmes at a larger spatial scale will allow assessment of the potential
problems discussed above. Effects on metapopulation dynamics, inter-patch dispersal,
landscape fragmentation, population genetic structure, the remainder of the donor
community and the other processes described above can only be measured by intensive
empirical studies on a landscape scale.
Summary conclusions concerning translocation of species
Community translocation is widely used in attempts to rescue communities which
are to be destroyed by a change in land use, or to carry out a restoration in
mitigation for such a destruction.
Turf-lifting is the most common technique used, although soil transfer is an
occasional method.
Many translocated communities are of high conservation value, e.g. from SSSIs,
and are usually of herbaceous or dwarf shrub vegetation.
A review of case studies shows that changes in the communities following
translocation are almost ubiquitous. Losses of species and invasion of novel
species occurred in most translocations for plants, and also for invertebrate species
where these were monitored. These changes were greater and more persistent than
would be caused by the species fluctuations seen in most (untranslocated)
ecological communities.
These changes were probably related to the disturbance involved in transferring
the community and to environmental and management differences at the receptor
site. Improved translocation (minimising of community changes) will result from:
minimising disturbance during transfer (e.g. turf-lifting is superior to soil transfer);
appropriate choice and preparation of the donor site (e.g. appropriate topography,
or soil removal to decrease fertility); and appropriate management of the
translocated community.
There are other possible factors which may affect the success of the translocation,
but which have not been investigated. These are: genetic effects of small
population size of translocated species (e.g. genetic bottlenecks); effects of
fragmentation on the translocated community; and disruption of metapopulation
dynamics (e.g. causing dislocation from a source population and subsequent
extinction of a species in the community).
There may be effects on the wider landscape which, again, have not been
investigated in community translocations. Disruption of the environment and of
populations may lead to changes in the remainder of the original community (i.e.
the part of the donor site which was not translocated). Metapopulations and large236
Review of information, policy and legislation on species translocations
scale dynamics of species in the landscape of the donor site may be disrupted by
the removal of the translocated community (especially as an SSSI may be an
important site for certain species). Community translocation also involves the
disruption of the historical pattern of biotopes in the landscape and thus has
aesthetic and cultural consequences.
Research should be carried out to look at these wider problems and to refine
techniques to achieve greater success with community translocations. This should
include long-term monitoring in the translocated community of invertebrates and,
possibly, vertebrates, as well as the plant species. Monitoring of consequences for
the wider landscape and the remaining community at the donor site would be
Because of these effects, it is clear that community translocation will rarely
achieve the translocation of a complete community such that it remains
fundamentally unchanged from its pre-translocation state. With care however, one
should be able to use this technique to create a community which resembles the
pre-translocated state.
It should be a general policy to emphasise that community translocation can never
be a substitute for in situ conservation, and that to develop a site is to destroy the
community. A translocation will create a new community rather than preserve the
original community.
Such a policy may inform the early part of the planning process, when
development of a site of conservation value is being considered. However, if
development is to go ahead, community translocation may be the best option for
mitigation and the statutory conservation agencies should be able to advise on the
best methods to increase the value of a translocation.
Detailed guidelines on best practice to increase the value of a translocation should
be drawn up. This can use the results of this review and the ongoing studies of the
EFU, and can be updated if the suggested research is carried out.
It should be a policy to educate the public and developers: 1. about this best
practice; and 2. that translocation cannot preserve a community.
Review of information, policy and legislation on species translocations
Each chapter includes a discussion and conclusions. We will not repeat these here, but will
address some of the wider issues made apparent by this review.
Translocations and their effects on biodiversity
The forms of translocation covered in this review fall into two types; and type of
translocation influences the concerns expressed as to its effects on biodiversity.
Conservation translocations
Conservation translocations are carried out specifically to maintain or increase
biodiversity. The first half of Chapter 2 concerned conservation translocations of UK
native species and covered both the maintenance of biodiversity - e.g. by supplementing
declining populations or relocating populations (which would otherwise be lost) - and the
enhancement of biodiversity, usually by attempting to rectify past losses - e.g. by restoring
communities or reintroducing species. Community translocations, covered in Chapter 5,
are also conservation translocations because they are attempts to maintain biodiversity by
salvaging some elements of communities that would otherwise be lost.
The two concerns with conservation translocations effects on biodiversity are therefore:
what are the best methods for maintaining or enhancing biodiversity?; and does the
translocation have an associated risk of damage to biodiversity?
Non-conservation translocations
Translocations carried out for reasons other than conservation are covered in the
remainder of the review. The translocations covered in Chapters 3 and 4 - introductions of
species alien to the UK and of GMOs - do not have primary conservation motives, but are
carried out for a variety of reasons, such as crop development, pest control, stocking for
country sports or ornament. However, some may have some conservation motives as part
of their ultimate aim, such as reduction in the use of pesticides, more efficient and less
damaging agricultural systems, or natural history education. These indirect conservation
motives may need to be taken into account in the final assessment of the advisability of
such translocations.
Many translocations of UK natives, covered in the second half of Chapter 2, are also
carried out for non-conservation reasons and for similar primary motives as those of nonnative species. However, additional motives involve aesthetics, which includes misguided
releases with the intention of enhancing biodiversity.
Therefore, the single concern for conservationists with non-conservation translocations is
the risk of damage to biodiversity.
Review of information, policy and legislation on species translocations
Effects of translocations
We have reviewed the effects on biodiversity of these various types of translocation using
the comprehensive list of types of biodiversity developed in section 1.3. These effects are
summarised in the conclusions for each of the chapters, but here we will give an overview
of the primary concerns of conservationists and those aspects of biodiversity which are
poorly covered by the existing data.
Success of conservation translocations
The aim is to establish, enhance, or maintain one or more populations of one or more
targeted species. This involves objectives of increasing or conserving the population size
and/or the number of populations and/or the geographical range of certain species.
Techniques for ensuring that populations establish and persist at the receptor site are well
researched in general (although certain projects would benefit from greater consideration
of these techniques), and cover the majority of the information available on translocation
success. In summary, they include consideration of a receptor site with appropriate
habitat, good habitat management, use of sufficient individuals and an appropriate
population structure to begin a population, and ensuring no pathogens or parasites are
infecting the stock. However, other aspects of successful translocation are more poorly
Genetic considerations for a successful translocation are widely discussed, but poorly
researched. The Plymouth Pear Pyrus cordata example in section is one of the few
projects involving genetic studies in the UK. Genetic concerns include: use of a group of
individuals which will not suffer inbreeding depression or, in the case of re-enforcements,
outbreeding depression; and the use of individuals adapted to the receptor site
environment. The general idea that use of a 'large' number of individuals will reduce the
possibility of inbreeding depression is vague. Inbreeding may not be a problem for some
species, either because the species has existed naturally in the wild in populations with
low genetic diversity, or because inbreeding will not necessarily threaten population
persistence (see Section 1.6, Thornhill 1993, Gray 1995). Even if species are susceptible
to inbreeding depression, more definite techniques are needed to avoid these problems.
The appropriate method would be to study the genetic structure of wild populations and to
use the data to inform the choice of individuals, appropriate programmes of artificial
propagation, and sufficient individuals to achieve the level of variation shown by wild
populations in the translocated populations. Care must be taken however, to ensure that
the wild populations are not themselves declining because of low genetic diversity.
Further work could be used to understand whether or not it is important to maintain
certain levels of genetic variation to ensure long-term population persistence. Problems
associated with outbreeding depression can also be assessed by genetic studies.
Ensuring genetic adaptation to local conditions may be facilitated by the use of local
populations, but, as discussed in Chapter 2, genetic studies are needed to ensure that this is
important and that the following factors are not operating to negate the relevance of local
genetic adaptation. 1. Although there is fine-tuned local adaptation, phenotypic plasticity,
and/or the ability of the translocated population to genetically adapt to the local conditions
Review of information, policy and legislation on species translocations
through selection, will allow the population to persist. 2. There is little local adaptation,
perhaps through widespread gene flow or because the population has a wide ecological
amplitude. The Natterjack toad Bufo calamita example in section illustrates just
such a lack of local adaptation.
An additional aim, which could enhance the success of a conservation translocation by
facilitating population establishment and maintaining or increasing geographical ranges,
involves consideration of regional spatial dynamics. Some species show metapopulation
dynamics and other species may need to use a number of sites in a region in order to
persist (e.g. large vertebrate predators or insects which use different resources during the
life cycle). In Chapter 5 (Community translocations) it was pointed out that the dislocation
of a community from its context in the landscape could disrupt ecological and genetic
metapopulation processes or exploitation of nearby sites and could, possibly, lead to
eventual failure of some populations. The same applies to the relocation of individual
populations of single species which was discussed in section 2.5.6. The creation of new
populations or communities of UK native species, discussed in the remainder of Section
2.5, could meet similar problems if species require a metapopulation or particular
landscape structure (i.e. juxtaposition of different habitats) in order to persist at the
receptor site.
Conversely, study of metapopulation dynamics and the movement of individuals between
sites could allow one to choose receptor sites (i.e. close to other populations or
communities and/or within existing metapopulations) which allow movement to continue
and thus increase the chances of success of the translocation. This could be taken further
to actually enhance the biodiversity of the landscape around the receptor site if
metapopulation dynamics are enhanced by the addition of a new population, especially if
the new population can act as a source and provide colonisers for new populations in the
landscape. Also, a number of populations or foci could be created in a single programme
to create a metapopulation within a region. These considerations are apparent in some of
the EN Species Recovery Programme projects (R. Mitchell pers. comm.).
Impacts of conservation translocations
Possible negative or positive effects of conservation translocations on the population size
or persistence of other species at the receptor site of in the surrounding area or on the
donor site (effects on species biodiversity and genetic biodiversity) are widely discussed.
However, there is little empirical work on this. This is probably because the aims of
conservation translocation are focused on the translocated species rather than its effects on
the environment. The beaver Castor fiber is one species for which a more broadly ranging
assessment is being carried out (section
The question of genetic effects on conspecific populations is discussed in detail in Chapter
2, but the idea that use of non-local sources of individuals will cause outbreeding
depression or more vague impacts such as 'genetic contamination' provides a prime
example of pronouncements and decisions being made with little or no data to support
them. It also brings up the broader question of whether maintenance of the genetic
structure of existing populations or the pattern and amount of genetic differences among
Review of information, policy and legislation on species translocations
conspecific populations can remain a relevant concern when there are so few data or
studies to assess the importance of these biodiversity measures. While it may be a good
precautionary principle to use local, or nationally native, populations where possible, it is
necessary to realise that this may sometimes be the wrong approach from a genetic
biodiversity point of view (e.g. the case of the chequered skipper Carterocephalus
palaemon - section and, more generally, it may force compromises that decrease
the chances of success of the translocation (e.g. use of a inbred or poorly adapted donor
population, or one from which only few individuals can be taken) or increase damage to
other aspects of biodiversity (e.g. damage to a sensitive donor population). It may even
mean that an otherwise beneficial translocation will not be carried out. Without more
studies on the genetics of reintroductions it is unwise to make general statements about
which aspects of genetic biodiversity are important.
Impacts of non-conservation translocations
Impacts of such translocations of UK native, UK non-natives and GMOs are much better
researched than impacts of conservation translocations. To this must be added the caveat
that certain impacts of GMO introductions are well researched, whereas others are only
speculated upon. Some ecological impacts of non-conservation translocations of UK
natives and non-natives are fairly well researched, and the relevant chapters should be
consulted for an assessment of areas which would benefit from further research. However,
in general, genetic impacts are, again, less well understood. It seems justified to view
hybridisation (interspecific gene transfer) between non-native and native species of the
UK as a negative effect on biodiversity in its own right, because it changes the character
of the native species fundamentally, and amounts - if hybrids are viewed as no longer the
same as the native species - to a decline in the abundance of a species. If new varieties or
strains are introduced of a non-native species which is already present in the UK, and they
do not show an increased propensity to hybridise with native species, one cannot say that
this is a greater threat to biodiversity in itself. This argument is particularly important with
respect to GMOs (see Section 4.5.1).
Intraspecific gene flow between wild and translocated populations has been studied, to a
limited extent, in non-conservation translocations of nationally native species (see section
2.6.1). These have found some negative effects on biodiversity of the type suggested for
conservation translocations and indicate problems with using non-native races which have
also undergone artificial selection. Because of the greatly modified nature of such
organisms, these studies are not good models of possible genetic problems with
conservation translocations.
Costs of translocations
One subject not covered in this review is the financial cost of translocations. There are a
few attempts to calculate the costs of a translocation; e.g. Stevenson et al. (1995)
estimated the costs of different seeding rates in a restoration; and Pywell (1993) costed a
number of heathland community translocation projects. However, the reason for our
omission is that there is a general lack of information on the costs of specific projects. For
a conservation translocation, such costs may include the equipment, facilities and people
Review of information, policy and legislation on species translocations
employed: to assess and plan the project (e.g. reviewing the status of the species, genetic
and ecological analysis of extant populations, assessment of potential receptor sites, etc.);
to purchase or rent a suitable receptor site; to carry out the translocation (e.g.
capture/collection of individuals, captive breeding, turf-lifting and transport, etc.); to carry
out habitat management at the receptor site; to carry out post-translocation monitoring and
analysis of trends (perhaps including genetic analysis) of the species at the receptor and
donor sites; to act on problems (e.g. subsequent supplementation of the translocated
populations if ecological or genetic problems arise, or control of a population that is
expanding too rapidly); and to publicise the translocation and its results. Costs of nonconservation translocations will rarely be an issue for conservationists, except in the
planning and assessment stages, but costings for control of problem species and of
rectification of adverse effects on biodiversity may be necessary.
To assess the benefits and impacts of translocations and to choose between different
courses of action, will require reliable estimates of costs. An extensive review of cost of
past and current translocations would be valuable. This would allow some idea of the
figures involved. It is likely that these figures differ greatly between projects, and for this
reason it would be misleading and dangerous to report only a few examples. However,
standard accounting procedures could be developed to ensure that all components are
A second benefit of a review of costs would be to allow an assessment of the truth of
suggestions concerning the relative costs of translocations and other courses of action.
Two common statements are that: 1. conservation translocations of single species or for
community restoration are very costly compared to in situ conservation of species and
their habitats; and 2. rescue translocations of communities or single species in mitigation
for land development may, if designed to minimise the ecological damage, be so costly
that it would be cheaper for the developer to relocate the development. These suggestions
seem, at the moment, to be based on 'gut feelings' rather than on any real financial
Monitoring and databases
Information on translocations is accruing continually and databases of information and/or
bibliographies will help to coordinate the available information. This is particularly
important for gathering information on the areas described above for which knowledge is
lacking. The bibliography we have supplied partially fulfils this requirement, but it would
require more work to produce a complete and fully functioning database.
In carrying out this review we have noticed that much information is unpublished or in
reports with restricted access. It is difficult to say exactly what translocations are
occurring at this moment in the UK without talking to a wide range of people. The insect
establishment recording scheme set up by JCCBI still exists but is largely unused now (M.
Warren pers. comm.), although it provides an example of the necessary approach to
coordinating information on translocations. T. Gent (pers. comm.) of EN has stated that
there is a need for a database of great-crested newt Triturus cristatus translocations in
Review of information, policy and legislation on species translocations
A database of translocation projects in the UK would also provide information on
distribution of projects (to allow assessment of spreads and outbreaks, or, simply,
revisions of UK distributions of species), as well as up-to-date information on the status of
such projects. The Biological Records Centre of NERC is a possible coordinating body for
such a database, but a feasibility study would be required in order to decide on the
attributes and structure of such a database.
The Reintroduction Specialist Group of the IUCN (see below) have also set up a
bibliographic database on reintroductions and wish to start a database on reintroduction
projects. There is a possibility that a UK database could be incorporated into that of the
RSG (M. Maunder pers. comm.). H. Frost (pers. comm.) of the Oxford Forestry Institute
has also investigated the feasibility of a world weeds database (the report is on the World
Wide Web
Assessments of risks and of benefits - formats for guidelines
Although the discussion of the limits of current knowledge indicates essential areas for
future research, it also makes clear that assessment of a proposed translocation should be
based on a clear exposition of possible risks and benefits, and how general theory and
specific study of the species involved provide evidence for or against these possibilities.
The risks and benefits should also be based on actual consequences for biodiversity,
possibly using the set of definitions in section 1.3. Thus the vague perception that certain
events - such as translocation of a non-local population, establishment of non-native races
or species in the UK countryside, or replacement of feral populations of a crop with its
genetically modified derivative - are in some way harmful will be avoided.
Risk assessment seems to be the best method for scrutinising all the potential hazards to
biodiversity posed by a translocation project. With its clear layout and detailed structure,
risk assessment is becoming a popular form of appraisal for conservation projects. We
have given some detail on this approach in the review, but DOE (1995) have produced a
thorough guide (see also Burgman et al. 1993). A form of risk assessment is in place in
the DOE procedures for appraising applications for GMO release (see Chapter 4), and a
similar approach is being developed for the appraisal of licences for introductions of nonnative species (see Chapter 3).
One difficulty with a risk assessment approach is that they may be slow to accomplish. If
a response is needed rapidly (e.g. to a licence application), a risk assessment by the
assessors (e.g. the statutory conservation agencies) may be impossible. However, if a risk
assessment by the applicant is required as part of the planning procedure (e.g. as part of
the licence application), as is carried out for GMO release applications (see section 4.4),
the lengthy work will be put into the hands of the applicant.
The detailed information requirements suggested in guidelines on conservation
translocations of nationally native species (Chapter 2) and community translocations
(Chapter 5) provide the basis for risk assessment procedures, but would need reformatting to provide the specific questions required for a risk assessment. This may, in
fact, answer some of the complaints we have reported concerning the generality and
Review of information, policy and legislation on species translocations
difficulty in application of guidelines such as IUCN (1995) and NCC (1990). While
generality is necessary, intentions will be clarified if the guidelines are structured as a risk
assessment. This would also show where differences in approach may arise due to
taxonomic, ecological, or other possible differences between projects.
While classical risk assessment would be sufficient for non-conservation translocations,
the two concerns with conservation translocations would merit a modified approach.
While risks of negative effects on biodiversity can be assessed as for non-conservation
translocations, there should be additional assessment of the potential for success of the
project. This 'benefits assessment' would involve a similar procedure to the risk
assessment, but questions would be asked concerning the procedures used to increase the
chances of success, whether further measures could be included and whether alternative
procedures would increase the benefit of the translocation. These questions could be based
on the existing guidelines and the output would give one confidence that the best methods
are being used, or whether resources would be better employed elsewhere to maximise
biodiversity conservation.
In both the risk and benefits assessments an important output would be a decision as to
whether any of the gaps in knowledge that have been exposed should be filled before the
translocation goes ahead, or whether the extra work would either impose an unnecessary
delay which is damaging to the success of the project or have little effect on the chances
of success or risk of damage.
Co-ordination among organisations
The UK
There is a plethora of guidelines and policy statements for some types of translocation.
This is particularly true of conservation translocations of native species. Our review of
guidelines uncovered sixteen on conservation translocations of native species by UK
organisations, with seven more concerning seed mixes and plantings for restoration
projects. Introductions of UK aliens had fewer sets of guidelines (four), but the Wildfowl
and Wetlands Trust and RSPB have had a recent meeting to decide a policy on feral birds
in the UK (B. Hughes pers. comm.). Conversely, policy on community translocations is
poorly developed and there are no policies by UK organisations specifically for GMOs.
The areas lacking policies should be addressed, but it is questionable whether a simple
revision of existing policy on native and non-native translocations would be necessary or
It would seem to be more useful to encourage coordination among the different
organisations involved in translocations to produce a common policy statement and set of
guidelines (possibly using the risk assessment approach) for each type of translocation.
Conservation organisations to be involved could include JNCC, DOE(NI), SNH, CCW,
EN, the Countryside Commission, the British Butterfly Conservation Society, the Royal
Society for the Protection of Birds, the Wildfowl and Wetlands Trust, Plantlife, the
Botanical Society of the British Isles, the Joint Committee for the Conservation of British
Insects, the British Herpetological Society, the National Trust and the Wildlife Trusts.
This list is not intended to be comprehensive, and the success of the venture would be
Review of information, policy and legislation on species translocations
dependant on the involvement of all interested parties. Some of these organisations have
responded favourably to such an idea, e.g. BBCS (M. Warren pers. comm.) and BSBI (D.
Pearman pers. comm.). The process would be more relevant if there was consultation
with: government departments administrating the legislation (MAFF, DAFS, SOAFD,
WOAD, DOE, etc.); other important organisations, e.g. the NRA (fish and shellfish
translocations), the Highways Agency (restorations and community translocations), seed
merchants (restorations), and environmental consultants (usually relocations and
restorations); and academics and research institutes.
A coordinated approach would also have the benefit that assessments of proposed
translocations could be carried out by panels of representatives from these different
organisations (either conservation organisations only, or including other experts) using
agreed procedures and criteria - a 'Joint Panel on Translocations'. The proposed
translocations would come either via government (for translocations regulated by law) or
voluntarily from organisations or individuals. It might be thought that such a panel may
hamper the rapid responses needed to provide routine advice to government. However,
this problem could be circumvented by two approaches: 1. delegation of certain functions
to certain organisations, which would report to the panel; and/or 2. the use of modern
Information Technology to allow rapid exchange of information among panel members.
An international perspective
Many international or European organisations have also produced guidelines on certain
aspects of translocations. Those covered in this review are the International Commission
for Exploration of the Sea, the International Maritime Organisation, the North East
Atlantic Commission, the Worldwide Fund for Nature, European Inland Fisheries
Advisory Committee, and the IUCN. As well as the Reintroduction Specialist Group of
the IUCN, which produced the recent IUCN (1995) guidelines, the Captive Breeding
Specialist Group and the Invasive Species Specialist Group have relevant interests within
the IUCN.
Coordination with these and any other relevant organisations would seem to be important.
This would be primarily because some projects are international and require international
controls, e.g. use of non-native races in reintroductions (e.g. chequered skipper), or
problems with international translocation or spread of non-native species (e.g. ruddy
duck). A secondary benefit would result if information exchange on projects and
techniques is facilitated.
The Reintroduction Specialist Group (RSG) of the IUCN is keen on such an approach (M.
Maunder pers. comm.). As well as production of the IUCN (1995) guidelines RSG
provides a forum for worldwide exchange of information on reintroductions. It has formal
links with some national conservation agencies in other countries, but not in the UK (the
only, informal, links are through individual membership of staff in EN, SNH, etc.). More
formal links would be welcomed (M. Maunder pers. comm.). There is a perceived need
for a European network - given that the taxa used in projects in different countries are
often the same or closely related, donor populations may occur in other countries,
reintroductions may have cross-border effects, and more legislation is becoming Europewide.
Review of information, policy and legislation on species translocations
The Captive Breeding Specialist Group of the IUCN coordinates animal, and now plant,
captive breeding. It is a networking organisation for the IUCN which advises and initiates
ex situ conservation project. It works mostly in developing countries, but, again, it would
like to strengthen its links with the UK (O. Byers pers. comm.). The Invasive Species
Specialist Group has been set up recently to coordinate information exchange on invasive
species and has expressed interest in creating links with the UK conservation
organisations (J. Cooper pers. comm.).
A European framework of legislation
Linkages between international and UK organisations on the question of translocation
could come to reflect a commonalty of legislation. The Council of Europe perceives a
need to harmonise the national legislation of member states on translocations (i.e.
introductions, reintroductions, supplementations and release of GMOs) in view of the
'obvious transfrontier nature of the problem' (Council of Europe 1993a); e.g. spread over
borders of translocated species or races, international problems of controlling alien species
and the need for international cooperation and coordination in certain reintroduction
programmes. Recommendation R(84)14 asks states to carry out studies of the
consequences of planned introductions of non-native species and to submit them for an
opinion to the European Committee for the Conservation of Nature (CDSN). Under
Recommendation R(85)15 the CDSN wishes also to be informed of any reintroduction
projects. Both Recommendations also suggest that, where necessary, governments of
neighbouring countries be informed of reintroductions and intentional and accidental
introductions. These Recommendations are not binding, although they are expected to be
taken in good faith.
The legal controls on translocations of member states differ widely in range and in
strictness (de Klemm 1995). A Group of Experts on the legal aspects of such
translocations have been meeting to discuss this issue, and the possible role of the Bern
Convention on Conservation of European Wildlife and Natural Habitats in facilitating
such harmonisation. The meetings have resulted in proposals for the setting up of a system
of reciprocal consultation involving the contracting parties to the Bern Convention and to
study procedures for harmonising national legislation (Council of Europe 1993a), but at
this moment in time little has been implemented. De Klemm (1995) carried out a study of
introductions of non-native species and GMOs for this Group of Experts and proposed the
setting up of a European 'group of experts on introductions' which could take over some of
the tasks entrusted to the CDSN by recommendations R(84)14 and R(85)15. This could
keep a European register of introductions, organise exchange of information, frame a
general policy and guidelines concerning introductions, and identify problem species and
sensitive areas (e.g. islands) requiring priority attention.
However, European legislation can also act to weaken controls on translocations,
particularly by the impositions of Directives facilitating free trade. For instance, the EC
Fish Health Directive (91/67) relaxes quarantine rules for European trade in fish and
shellfish (Eno 1993, Kerry 1993).
Review of information, policy and legislation on species translocations
Education and legislation
Although changes in legislation and guidelines are necessary, it is clear that neither will be
fully effective without a comprehensive policy of education. Many of the problems with
translocations described in this review are due to a lack of knowledge among members of
the public or certain organisations of the consequences of their actions. These actions
deliberate releases of native butterflies for aesthetic or pseudo-conservation
reasons by breeders,
deliberate releases of rehabilitated barn owls Tyto alba and other species by
members of the public,
accidental and deliberate releases of native and non-native wildfowl by breeders or
birds of prey by falconers,
planting of interesting or aesthetically pleasing plant species by botanists,
the use by developers of community translocation as an alternative to in situ
the sale and use of non-native seed sources or agricultural varieties by certain seed
merchants and certain users,
unnecessary restocking of salmon Salmo salar in aquaculture,
the release of unwanted pets,
escape and release of plants and animals from collections,
the accidental import of a wide variety of species by all sections of society.
It is unlikely that many people carry out these releases or allow escapes with a full
knowledge of the harm they could be causing, although the example of the butterfly
breeder who delights in fooling 'experts' (section 2.6.6) indicates some malicious elements
in society. A comprehensive and targeted policy of education and dissemination of
information to interested and involved parties, perhaps with the added incentive of
improved legislation and improved enforcement of legislation, could drastically reduce the
problems caused by such activities.
Education would also facilitate the success of a Joint Panel on Translocations, as it could
encourage individuals and organisations to have their proposals officially vetted.
Summary general conclusions of the review
Two particular areas of research whose development could aid the assessment and
design of translocations in general are: 1. ecological genetics - the extent and
importance of local adaptation, the consequences of changes in genetic diversity,
and the rates and consequences of gene flow; and 2. the metapopulation and largescale dynamics of species in the wider landscape.
There is a general need for the creation of national databases for the different
forms of translocation; to hold information on both the details of the programmes
Review of information, policy and legislation on species translocations
and on the results. In addition, there is a need for a greater effort to carry out
detailed and long-term monitoring of the outcome of translocations.
The financial costs of translocations are not generally assessed, either for the
whole of conservation translocation projects, or for planning, assessment or
amelioration of adverse effects of non-conservation translocation. A review of
costs would be valuable.
For each planned translocation it should be a priority to carry out a full assessment
of its consequences to biodiversity, in the light of the best possible information.
The risk assessment approach forms the best method for determining all the
potential adverse effects of a translocation on biodiversity, and the likelihood of
their occurrence.
Conservation translocations should have beneficial effects on biodiversity, and a
'benefits assessment' of such a project could be used to ensure that the most
appropriate techniques will be used.
Guidelines and policy are at different stages of development for the different types
of translocation. However, any new guidelines would benefit greatly and have
much greater applicability and impact if they were produced in consultation and
coordination with a range of UK organisations. These would be the parties that
would be carrying out, advising on, or licensing translocations.
These parties could form a Joint Panel on Translocations which would assess
proposed translocations of all types, either together or, more profitably, in
specialist panels for particular types of translocation.
Formal links with international organisations concerned with translocations,
especially the IUCN, would allow exchange of information, development of
coordinated policy, and joint action in international programmes.
International links would also allow development of responses to common
legislation in Europe and to international conventions which are concerned with
There is a clear need for a policy of education, targeted at the public and particular
organisations, in order to publicise: the benefits of certain types of translocation to
conservation; and, more importantly, the damage caused by poorly planned and
unregulated translocations of any type.
Review of information, policy and legislation on species translocations
Many experts from a range of organisations provided us with information for this review.
This included: providing us with policy statements and unpublished material; informing us
of published material; and giving opinions on certain areas covered in this review. Their
help was indispensable.
John Akeroyd
Plantlife/Plant Talk
Colin Bannister
Ministry of Agriculture
Fisheries and Food
Scottish Natural Heritage
Phil Boon
Paul Bright
Alistair Burn
Joanna Bury
Onnie Byers
Chris Cheesman
Peter Clement
Royal Holloway and Bedford
New College
English Nature
Humphries Kirk Solicitors
IUCN Captive Breeding
Specialist Group
Central Science Laboratory
English Nature
IUCN Invasive Species
Specialist Group
National Rivers Authority
(Southern Region)
Clive Cummins
Roger Daniels
Willie Duncan
Institute of Terrestrial Ecology
Institute of Terrestrial Ecology
Environment Conservation and
Management Division, Council
of Europe
Scottish Natural Heritage
Clare Eno
Joint Nature Conservation
Joint Nature Conservation
Scottish Natural Heritage
Oxford Forestry Institute
English Nature
English Nature
National Rivers Authority
Department of the Environment
European Wildlife Division
Institute of Terrestrial Ecology
John Cooper
Ian Evans
Vin Fleming
Hugh Frost
Tony Gent
Mary Gibson
Stephen Gledhill
Felicity Grant
Alan Gray
Plant translocations and use of
wildflower seed
Crustacea releases
SNH policy on freshwater
Mammal translocations
EN policy on GMOs
English legal system
IUCN policy and research on captive
Badger translocation
EN licensing issues under the
Wildlife and Countryside Act
IUCN policy and research on
Licensing under Section 30 of the
Salmon and Freshwater Fisheries Act
Herpetofauna translocations
Plant translocations
Council of Europe policy and
legislation on translocations
SNH policy on freshwater
Marine introductions
Bird reintroductions
SNH Species Action Programme
World Weeds Database
Herpetofauna translocations
Fisheries translocations
Fisheries translocations
Section 14 administration
Research on GMOs
Review of information, policy and legislation on species translocations
David Hallam
Graham Harris
Stuart Headley
Keith Hiscock
John Holmes
Ian Holt
Phillip Horton
Liz Howe
Baz Hughes
Neil Humphries
Peter Hutchinson
Anton Ibbotson
Andy Jackson
Maddy Jago
Richard Jefferson
Roger Kent
Robert Kenward
Sandy Kerr
Richard Keymer
Simon Leach
Diana Linskey
David MacDonald
Georgina Mace
Mike Maunder
Selwyn McGrorty
Donald McIntyre
Ian McLean
Nick Michael
Chris Mills
Roger Mitchell
Vicky Morgan
Ministry for Agriculture
Fodder Plant Seeds Regulations
Fisheries and Food Plant Variety
Rights Office and Seeds
National Rivers Authority
Monitoring of translocated fish
English Nature
EN monitoring of community
Joint Nature Conservation
Marine introductions
Joint Nature Conservation
Policy and research on vertebrate
Scottish Office Environment
Section 16 licensing
Humphries Rowell Associates
Community translocation
Countryside Council for Wales CCW translocations policy and
recovery programmes
Wildfowl and Wetlands Trust
WWT policy on bird translocations
Humphries Rowell Associates
Community translocation
North Atlantic Salmon Advisory Research and policy on salmon
Institute of Freshwater Ecology Freshwater fish translocations
Royal Botanic Gardens
Plymouth Pear reintroduction
Countryside Commission
Countryside Stewardship Scheme
English Nature
Wildflower seed mixes
Highways Agency
Highways Agency policy on
community translocation and habitat
Institute of Terrestrial Ecology Bird and mammal reintroductions
Scottish Natural Heritage
SNH policy and Species Action
English Nature
EN monitoring of community
English Nature
EN monitoring of community
Ministry for Agriculture
Section 16 licences of fish
Fisheries and Food
University of Oxford
Mammal reintroductions
Institute of Zoology
IUCN Red List categories
IUCN Re-Introduction Specialist IUCN policy and implementation of
Institute of Terrestrial Ecology Shellfish ecology and translocations
Emorsgate Seeds
Wildflower seed mixes
English Nature
EN translocations policy
English Nature
Habitat restoration
National Rivers Authority
Monitoring of translocated fish
English Nature
EN Species Recovery Programme
Botanical Society of the British BSBI policy on plant translocations
Review of information, policy and legislation on species translocations
Pat Morris
Robert Moss
Greg Mudge
Manfred Nauke
Tony Owen
Royal Holloway and Bedford
New College
Institute of Terrestrial Ecology
Scottish Natural Heritage
International Maritime
National Rivers Authority
(Southern Region)
Margaret Palmer Joint Nature Conservation
Department of the Environment
Bill Parish
Toxic Substances Division
Botanical Society of the British
David Pearman
Ministry for Agriculture
Ted Potter
Fisheries and Food
Stewart Pritchard Scottish Natural Heritage
Charles Pulteney English Nature
Richard Pywell
Paul Raven
Alan Raybould
Tim Rich
Art Schwartz
Rowley Snazell
Pritpal Soorae
Brian Spencer
Mark Stanley
Sylvia Swabe
Joyce Tait
Richard Tapper
Fran Tattershall
Tom Tew
Jeremy Thomas
Lena Ward
Mammal translocations
Gamebird translocations
Bird reintroductions
Policy on introductions by ships
Licensing under Section 30 of the
Salmon and Freshwater Fisheries Act
JNCC translocations policy
Consents for GMO releases and
Section 16 licences
BSBI policy on plant translocations
Licensing, policy & research on
salmonid translocations
Translocations in Scotland
EN monitoring of community
Institute of Terrestrial Ecology Community translocations and
habitat restoration
National Rivers Authority
NRA research on fisheries
Institute of Terrestrial Ecology Research on GMOs
Charles Blandford Associates
Community translocations and plant
English Nature
EN monitoring of community
Canadian Circumpolar Institute, Community translocation
Institute of Terrestrial Ecology Community translocations and
habitat restoration
IUCN Re-Introduction Specialist IUCN policy and implementation of
Ministry for Agriculture
Shellfish translocations
Fisheries and Food
IUCN Re-Introduction Specialist IUCN policy and implementation of
Plymouth Marine Laboratory
Fish translocations
Scottish Natural Heritage
SNH policy on GMOs
Worldwide Fund for Nature
WWF policy on GMOs
University of Oxford
Beaver reintroduction
Joint Nature Conservation
Policy and research on vertebrate
Institute of Terrestrial Ecology Butterfly reintroductions
Institute of Terrestrial Ecology Community translocations and
habitat restoration
Review of information, policy and legislation on species translocations
Martin Warren
Tony Watts
British Butterfly Conservation
Ministry for Agriculture
Fisheries and Food Plant Variety
Rights Office and Seeds
Institute of Terrestrial Ecology
Policy and implementation of
butterfly translocations
Fodder Plant Seeds Regulations
Community translocations and
habitat restoration
Institute of Terrestrial Ecology Wildflower seed mixes
Terry Wells
Department of the Environment Translocations policy and legislation
Richard Weyl
(Northern Ireland) Environment in Northern Ireland
Ministry for Agriculture
Lobster translocations
John Wickens
Fisheries and Food
Royal Society for the Protection RSPB policy on bird translocations
Gwyn Williams
of Birds
Introductions and GMO releases
Mark Williamson York University
Nigel Webb
Review of information, policy and legislation on species translocations
This bibliography contains the references used in the review and additional publications
concerning translocations in the UK and abroad. These were obtained by literature searches,
personal recommendation, and searches on BIDS and CD-ROM Silver Platter. We thank
Janet Dobson, Liz Guerin and Becky Edwards for their help in compiling this list.
Abbott, R.J. 1994. Ecological risks of transgenic crops. Trends in Ecology & Evolution, 9: 280-282.
ACRE. 1993. The regulation and control of the deliberate release of genetically modified organisms. London,
Department of the Environment.
ACRE. 1994a. Guidance for experimental releases of genetically modified plants. London, Department of the
ACRE. 1994b. Fast track procedures for certain GMO releases. London, Department of the Environment.
ACRE. 1994c. Guidance for experimental releases of genetically modified microorganisms (excluding viruses
and similar agents). London, Department of the Environment.
ACRE. 1994d. Annual report. 1993/94. London, Department of the Environment.
ACRE. 1995a. Guidance for experimental releases of genetically modified baculoviruses. London, Department
of the Environment.
ACRE. 1995b. Annual Report. 1994/5. London, Department of the Environment.
Adams, C.E. 1994. The fish community of Loch Lomond, Scotland: its history and rapidly changing status.
Hydrobiologia, 290: 91-102.
Adams, C.E. & Maitland, P.S. 1991. Evidence of further invasions of Loch Lomond by nonnative fish species
with the discovery of a roach x bream, Rutilus rutilus (L) x Abramis brama (L), hybrid. Journal of Fish
Biology 38: 961-963.
Agassiz, D.J.L., Nash, D.R., Godfray, H.C.J. & Lawton, J.H. 1994. Estimating the spread of small invading
organisms using information from the public. Global Ecology and Biogeography Letters, 4: 1-8.
Agounke, D., Agricola, U. & Bokonon-Ganta, H.A. 1988. Rastrococcus invadens Williams (Hemiptera,
Pseudococcidae), a serious exotic pest of fruit trees and other plants in West Africa. Bulletin of
Entomological Research, 78: 695-702.
Ahl Goy, P. 1993. Gene transfer from genetically modified plants to wild relatives: worldwide field trial activity
and CIBAs own experience. In: Gene transfer: are wild species in danger?, edited by J. Youlande, K.
Ammann & F. Pythoud, 24-27. Bern: Federal Office of Environment, Forests and Landscape.
Ahl Goy, P. & Duesing, J.H. 1995. From pots to plots: genetically modified plants on trial. Bio Technology, 13:
Ahmed, N. & Roy, D.C. 1991. Lantana camara poisoning. Livestock Advisory, 16: 21-23.
Aizen, M.A. & Feinsinger, P. 1994. Habitat fragmentation, native insect pollinators, and feral honey bees in
Argentine "Chaco Serrano". Ecological Applications, 4: 378-392.
Akeroyd, J. (1994): Seeds of destruction? Non-native wildflower seed and British floral diversity. Plantlife.
Akeroyd, J. & Wyse Jackson, P. 1995. A Handbook for Botanic Gardens on the Reintroduction of Plants to the
wild. Surrey: Botanic Gardens Conservation International.
Alberch, P. 1993. Museums, collections and biodiversity inventories. Trends in Ecology & Evolution, 8: 372375.
Alexander, D.J. 1988. Newcastle disease: methods of spread. In: Newcastle disease, ed. by D.J. Alexander, 256272.
Allaby, M. 1994. The concise Oxford dictionary of ecology. Oxford: Oxford University Press.
Allan, J.R., Kirby, J.S. & Feare, C.J. 1996. The biology of Canada Geese Branta canadensis in relation to the
management of feral populations. In press.
Allen, W.H. 1994. Reintroduction of endangered plants. Bioscience, 44: 65-66.
Altieri, M.A. 1991. Classical biological control and social equity [guest editorial]. Bulletin of Entomological
Research, 81: 365-369.
Altukhov, Y.P. & Salmenkova, E.A. 1990. Introductions of distinct stocks of chum salmon, Oncorhynchus keta
(walbaum),into natural populations of the species. Journal of Fish Biology, 37: 25-33.
Alutkov, Y.P. 1981. The stock concept from the viewpoint of population genetics. Canadian Journal of
Fisheries and Aquatic Science, 38: 1523-1538.
Review of information, policy and legislation on species translocations
Amarasekare, P. 1994. Ecology of introduced small mammals on Western Mauna-Kea, Hawaii. Journal of
Mammology, 75: 24-38.
Amori, G. & Zima, J. 1994. Threatened rodents in Europe: species status and some suggestions for conservation
strategies. Folia Zoologica, 43: 1-9.
Anderson, B. 1993. The Philippine snail disaster. Ecologist, 23: 70-72.
Anderson, P. 1989. Modelling and shaping new habitats in landscaping works. In: Biological habitat
reconstruction, ed. by G.P. Buckley, 1st ed., 235-250. London, Belhaven Press.
Andow, D.A. 1994. Community response to transgenic plant release - using mathematical theory to predict
effects of transgenic plants. Molecular Ecology, 3: 65-70.
Andow, D.A. & Imura, O. 1994. Specialization of phytophagous arthropod communities on introduced plants.
Ecology, 75: 296-300.
Andrews, J.D. 1980. A review of introductions of exotic oysters and biological planning for new importations.
Marine Fisheries Review, 42: 1-11.
Angermeier, P.L. 1994. Does biodiversity include artificial diversity? Conservation Biology, 8: 600-602.
Angermeier, P.L. 1995. Ecological attributes of extinction-prone species: loss of freshwater fishes of Virginia.
Conservation Biology, 9: 143-158.
Angle, J.S. 1994. Release of transgenic plants: Biodiversity and population-level considerations. Molecular
Ecology, 3: 45-50.
Anon. 1988a. Heathland restoration : A handbook of techniques. Southampton, The Camelot Press plc.
Anon. 1988b. Bird reintroduction schemes in Britain. Ornithology Note, 14: 1-5.
Anon. 1992. Brampton Meadow SSSI, phase III:site quality monitoring. Chis Blandford Associates.
Anon. 1993. International Oxyjura jamiacensis workshop. Summary and recommendations. IWWRB, DOE,
JNCC. 1-2 March 1993, Arundel, United Kingdom.
Anon. 1994a. Convention on biological diversity. Nairobi, ACTS Press.
Anon. 1994b. Biodiversity: the UK Action Plan. 1994. London, HMSO.
Anon. 1994c. Biodiversity Challenge. Sandy, RSBP.
Anon. 1994d. Wild flower plants and seeds. Guidance leaflet produced by BSBI, JNCC, EN, Plantlife, RSNC
and WWF.
Anon. 1994e. International technical meeting on Oxyjura leucocephala and Oxjura jamaicensis in the
Palaearctic region. Conclusions and recommendations. Cordoba (Andalusia, Spain) 29th & 30th
September 1994.
Anon. 1995. Genetically-modified animals. Atla-Alternatives to Laboratory Animals, 23: 264.
Antonovics, J. 1984. Genetic variation within small populations. In: Perspectives on plant population ecology,
ed by R. Dirzo & J. Sarukhan, 229-241. Sunderland, Mass: Sinauer.
Aplet, G.H., Anderson, S.J. & Stone, C.P. 1991. Association between feral pig disturbance and the composition
of some alien plant assemblages in Hawaii Volcanoes National Park. Vegetatio, 95: 55-62.
Arano, B., Llorente, G., Garciaparis, M. & Herrero, P. 1995. Species translocation menaces Iberian waterfrogs.
Conservation Biology, 9: 196-198.
Armour, C.J. & Thompson, H.V. 1955. Spread of myxomatosis in the first outbreak in Great Britain. Annals of
applied Biology, 43: 511-518.
Armstrong, D.P. 1995. Effects of familiarity on the outcome of translocations. 2. A test using New Zealand
robins. Biological Conservation, 71: 281-288.
Armstrong, D. & Perrott, J. 1995. Testing for food limitation following translocation, New Zealand. ReIntroduction News, 10: 9.
Arnold, H.R. 1989. Atlas of Mammals in Britain. London, HMSO.
Arnould, J., Gouyon, P.H., Lavigne, C. & Reboud, X. 1993. GMO - a theory to evaluate the risks. Biofutur, 4550.
Arthington, A.H. 1991. Ecological and genetic impacts of introduced and translocated fresh-water fishes in
Australia. Canadian Journal of Fisheries and Aquatic Science, 48: 33-43.
Arthington, A.H. & Mitchell, D.S. 1986. Aquatic invading species. In: Ecology of biological invasions, ed by
R.H. Groves & J.J. Burdon, 34-53. Cambridge: Cambridge University Press.
Ash, H.J., Bennett, R. & Scott, R. 1992. Flowers in the grass. Peterborough: English Nature.
Ashton, P.J., Appleton, C.C. & Jackson, P.B.N. 1986. Ecological impacts and economic consequences of alien
invasive organisms in southern African aquatic ecosystems. In: The ecology and management of
biological invasions in Southern Africa, ed by I.A.W. Macdonald, F.J. Kruger & A.A. Ferrar, 247-257.
Capetown, Oxford University Press.
Atkinson, I.A.E. & Cameron, E.K. 1993. Human influence on the terrestrial biota and biotic communities of
New Zealand. Trends in Ecology & Evolution, 8: 447-451.
Review of information, policy and legislation on species translocations
Autrique, A., Stary, P. & Ntahimpera, L. 1989. Biological control of pest aphids by hymenopterous parasitoids
in Burundi. FAO Plant Protection Bulletin, 37: 71-76.
Avery, M., Gibbons, D.W., Porter, R., Tew, T., Tucker, G. & Williams, G. 1994. Revising the British Red Data
list for birds: the biological basis of UK conservation. Ibis, 137, S232-S239.
Bailey, E.P. 1992. Red foxes, Vulpes vulpes, as biological control agents for introduced Arctic foxes, Alopex
lagopus, on Alaskan islands. Canadian Field Naturalist, 106: 200-205.
Bain, M.B. 1993. Assessing impacts of introduced aquatic species - grass carp in large systems. Environmental
Management, 17: 211-224.
Baker, H.G. 1965. Characteristics and modes of origin of weeds. In: The genetics of colonising species, ed by
H.G. Baker & G.L. Stebbins, 147-172. New York, Academic Press.
Baker, H.G. 1986. Patterns of plant invasion in North America. In: Ecology of biological invasions of North
America and Hawaii, ed by H.A. Mooney & J.A. Drake, 44-57. Berlin, Spinger-Verlag.
Baker, R.J. 1994. Some thoughts on conservation, biodiversity, museums, molecular characters, systematics,
and basic research. Journal of Mammology, 75: 277-287.
Baker, S.J. 1990. Escaped exotic mammals in Britain. Mammal Review, 20: 75-96.
Balharry, E., Staines, B.W., Marquiss, M. & Kruuk, H. 1994. Hybridisation of British mammals. JNCC Report,
No. 154.
Ballou, J.D. 1993. Assessing the risks of infectious diseases in captive breeding and reintroduction programs.
Journal of Zoo and Wildlife Medicine, 24: 327-335.
Balodis, M. 1995. Reintroduction of beavers in Latvia: Dynamics and results. Paper presented at the 2nd
European congress of mammalogy. 27th March-1st April 1995, Southampton University, England.
Balon, E.K. 1993. Dynamics of biodiversity and mechanisms of change: a plea for balanced attention to form
creation and extinction. Biological Conservation, 66: 5-16.
Balon, E.K. & Bruton, M.N. 1986. Introduction of alien species or why scientific advice is not heeded.
Environmental Biology of Fishes, 16: 225-230.
Baltz, D.M. 1991. Introduced fishes in marine systems and inland seas. Biological Conservation, 56: 151-177.
Barbault, R. & Hochberg, M.E. 1992. Population and community level approaches to studying biodiversity in
international research programs. Acta Oceologica, 13: 137-146.
Barlow, N.D. & Goldson, S.L. 1993. A modeling analysis of the sucessful biological control of Sitona
discoideus (Coleoptera, Curculionidae) by Microctonus aethiopoides (Hymenoptera, Braconidae) in
New Zealand. Journal of Applied Ecology, 30: 165-178.
Barnes, A.M. 1993. A review of plague and its relevance to prairie dog populations and the black-footed ferret.
In: Proceedings of the symposium on the management of prairie dog complexes for the reintroduction
of the black footed ferret, ed by J.L. Oldemeyer, D.E. Biggins, B.J. Miller & R. Crete, 28-37.
Barr. 1993. Countryside survey 1990. Main report. Merlewood, Institute of Terrestrial Ecology.
Barratt, E.M., Deavill, R.C. & Gurnell, J. 1995. Conservation of the red squirrel (Sciurus vulgaris) in Britain- a
genetic approach. Paper presented at the 2nd European congress of mammalogy. 27 March- 1st April
1995, Southampton University, England.
Barrett, N.E. & Niering, W.A. 1993. Tidal marsh restoration: Trends in vegetation change using a geographical
information system (GIS). Restoration. Ecology, 1: 18-28.
Barrett, S.C.H. 1989. Waterweed invasions. Scientific American, 261: 66-73.
Barrett, S. 1994. The biodiversity supergame. Environmental Resource Economics, 4: 111-122.
Barrett, S.C.H. & Kohn, J.R. 1991. Genetic and evolutionary consequences of small population size in plants:
implications for conservation. In: Genetics and conservation of rare plants, ed by D.A. Falk & K.E.
Holsinger, 23-42. Oxford, University Press.
Barrett, S.C.H. & Richardson, B.J. 1986. Genetic attributes of invading species. In: Ecology Of biological
invasions, ed by R.H. Groves & J.J. Burdon, 21-33. Cambridge, Cambridge University Press.
Bartha, S., Csapody, I., Danos, B., Fekete, G., Galle, L., Holly, L., Horvath, F., Jenser, G., Kereszty, Z., Kovacs,
G., Nemeth, F., Papp, L., Simon, T., Suranyi, D., Szabo, A.T., Szocs, Z. & Varga, Z. 1994.
Foundations for developing a national strategy of biodiversity conservation, by the Committee on
Ecology, Biology Section, Hungarian Academy of Sciences. Acta Zoologica Hungarica, 40: 289-327.
Beck, B., Cooper, M. & Griffith, B. 1993. Infectious-disease considerations in reintroduction programs for
captive wildlife. Journal of Zoo and Wildlife Medicine, 24: 394-397.
Beerling, D.J., Bailey, J.P. & Conolly, A.P. 1994. Fallopia japonica Houtt.; Polygonum cuspidatum. Biological
flora of the British Isles. Journal of Ecology, 82: 959-979.
Beerling, D.J. & Perrins, J.M. 1993. Impatiens glandulifera Royle (Impatiens roylei Walp.). Biological flora of
the British Isles. Journal of Ecology, 81: 367-382.
Beerling, D.J. & Woodward, F.I. 1994. Climate change and the British scene. Journal of Ecology, 82: 391-397.
Review of information, policy and legislation on species translocations
Beggs, J.R. & Wilson, P.R. 1991. The kaka Nestor meridionalis, a New Zealand parrot endangered by
introduced wasps and mammals. Biological Conservation, 56: 23-38.
Bell, B.D. 1994. A review of the status of New Zealand Leiopelma species (Anura: Leiopelmatidae), including a
summary of demographic studies in Coromandel and on Maud Island. New Zealand Journal of
Zoology, 21: 341-349.
Belliars, S.M. & Bell, D.T. 1993. Seed stores for the restoration of species-rich shrubland vegetation following
mining in Mid-western Australia. Restoration Ecology, 1: 231-240.
Bendel, P.R. & Therres, G.D. 1994. Movements, site fidelity and survival of Delmarva Fox- Squirrels following
translocation. American Midland Naturalist, 132: 227-233.
Benford, G. 1992. Saving the "library of life". Proceedings of the National. Academy of Science USA, 89: 10981101.
Bennett, F.D. 1993. Do introduced parasitoids displace native ones. Florida Entomologist, 76: 54-63.
Berger, J.J., ed. 1990. Environmental restoration. Washington, Island Press.
Berger, J.J. 1993. Ecological restoration and nonindigenous plant species: A review. Restoration Ecology, 1: 7482.
Berkowitz, D.B. & Kryspin-Sorensen, I. 1994. Trangenic fish: safe to eat? Bio Technology, 12: 247-252.
Berman, J. & Carlton, J.T. 1991. Marine invasion processes - interactions between native and introduced marsh
snails. Journal of Experimental Marine Biology and Ecology, 150: 267-281.
Berrebi, P. 1995. Speciation of the genus Barbus in the north Mediterranean basin: Recent advances from
biochemical genetics. Biological Conservation, 72: 237-249.
Berry, R.J. 1978. Genetic variation in wild house mice: where natural selection and history meet. American
Scientist, 66: 52-60.
Berry, R.J. 1981. Town mouse, country mouse: adaptation and adaptability in Mus domesticus (M. musculus
domesticus). Mammal Review, 11: 92-131.
Berry, R.J. 1985. Evolutionary and ecological genetics of the bank vole and wood mouse. Symposium of the
Zoological Society of London, 55: 1-31.
Berry, R.J., Triggs, G.S., King, P., Nash, H.R. & Noble, L.R. 1991. Hybridization and gene flow in house mice
introduced into an existing population on an island. Journal of Zoology, 225: 615-632.
Bertram, B.C.R. & Moltu, D.P. 1986. Reintroducing red squirrels into Regent's Park. Mammal Review, 16: 8188.
Bevanger, K. & Aelbu, O. 1986. Decrease in a Norwegian feral mink Mustela vison population -- a response to
acid precipitation? Biological Conservation, 38: 75-78.
Beveridge, M.C.M., Ross, L.G. & Kelly, L.A. 1994. Aquaculture and Biodiversity. Ambio, 23: 497-502.
Bianco, P.G. 1995. Mediterranean endemic freshwater fishes of Italy. Biological Conservation, 72: 159-170.
Biberklemm, S. 1995. Legal aspects of the conservation of endemic freshwater fish in the northern
Mediterranean region. Biological Conservation, 72: 321-334.
Bilio, M., Rosenthal, H. & Sindermann, C.J. 1981. Realism in aquaculture: achievements, constraints,
perspectives. In: World Conference on Aquaculture, 375-394.
Billington, H.L. & Mortimer, A.M. 1988. Divergence and genetic stucture in adjacent grass populations.1.
Quantitative genetics. Evolution, 42: 1267-1277.
Birkinshaw, C.R. 1991. Guidance notes for translocating plants as part of recovery plans. Nature Conservancy
Council, CSD Report.
Bishop, H.C. 1993. Economic efficiency, sustainability, and biodiversity. Ambio, 22: 69-73.
Black, J.M. 1991. Reintroduction and restocking: guidelines for bird recovery programmes. Bird Conservation
International, 1: 329-334.
Black, J.M. 1995. The Nene Branta sandivicensis recovery initiative: research against extinction. Ibis, 137,
Blackshaw, R.P. 1991. Mortality of the earthworm Eisenia fetida (Savigny) presented to the terrestrial planarian
Artioposthia triangulata (Dendy) (Tricladia:Terricola). Annals of applied Biology, 118: 689-694.
Bloomer, J.P. & Bester, M.N. 1992. Control of feral cats on sub-Antarctic Marion Island, Indian Ocean.
Biological Conservation, 60: 211-219.
Blossey, B. 1993. Herbivory below ground and biological weed control life history of a root boring weevil on
purple loosestrife. Oecologia, 94: 380-387.
Blossey, B. & Schroeder, D. 1995. Host specificity of 3 potential biological weed control agents attacking
flowers and seeds of Lythrum salicaria (purple loosestrife). Biological Control, 5: 47-53.
Blossey, B., Schroeder, D., Hight, S.D. & Malecki, R.A. 1994. Host specificity and environmental impact of
two leaf beetles (Galerucella calmariensis and G. pusilla) for biological control of purple loosestrife
(Lythrum salicaria). Weed Science, 42: 134-140.
Review of information, policy and legislation on species translocations
Boag, B., Palmer, L.F., Neilson, R. & Chambers, S.J. 1994. Distribution and prevalence of the predatory
planarian Artioposthia triangulata (Dendy) (Tricladida:Terricola) in Scotland. Annals of applied
Biology, 124: 165-171.
Boavida, C., Neuenschwander, P. & Herren, H.R. 1995. Experimental assessment of the impact of the
introduced parasitoid Gyranusoidea tebygi noyes on the mango mealybug Rasrtococcus invadens
williams by physical exclusion. Biological Control, 5: 99-103.
Bock, C.E., Bock, J.H., Jepson, K.L. & Ortega, J.C. 1986. Ecological effects of planting African love-grasses in
Arizona. National Geographical Research, 2: 456-463.
Bock, J.H. & Bock, C.E. 1992. Vegetation responses to wildfire in native verses exotic Arizona grassland.
Journal of Vegetation Science, 3: 439-446.
Boersma, P.D. 1991. Status of wild and captive penguin populations. Trends in Ecology & Evolution, 6: 381382.
Boitani, L. 1976. Reintroduction a controversial issue in ecosystem management. In: Reintroductions:
techniques and ethics. ed by L. Boitani, 35-42. Rome, World Wildlife Fund.
Boitani, L. 1976. Reintroductions: techniques and ethics. Rome, World Wildlife Fund.
Boitani, L. 1992. Wolf research and conservation in Italy. Biological Conservation, 61: 125-132.
Booth, T.H. 1990. A climatic analysis method for expert systems assisting tree species introductions.
Agroforestry Systems, 10: 33-45.
Booth, T.H., Nix, H.A., Hutchinson, M.F. & Jovanovic, T. 1988. Niche analysis and tree species introduction.
Forest Ecology and Management, 23: 47-59.
Bossard, C.C. 1991. The role of habitat disturbance, seed predation and ant dispersal on establishment of the
exotic shrub Cytisus scoparius in California. American Midland Naturalist, 126: 1-13.
Bossard, C.C. & Rejmanek, M. 1994. Herbivory, growth, seed production, and resprouting of an exotic invasive
shrub Cyisus scoparius. Biological Conservation, 67: 193-200.
Botanical Society of the British Isles. 1991. Guidelines for the transfer of rare vascular plants. Conservation
Committee of the Botanical Society of the British Isles.
Bouman, I., Bouman, J. & Boyd, L. 1994. Reintroduction. Przwalski's horse. Suny series in Endangered
Species. 255-263, New York, Suny.
Bowen, B.W., Conant, T.A. & Hopkins-Murphy, S.R. 1994. Where are they now? The Kemp's ridley headstart
project. Conservation Biology, 8: 853-856.
Bowles, M., Flakne, R., McEachern, K. & Pavlovic, N. 1993. Recovery planning and reintroduction of the
federally threatened Pitcher's thistle Cirsium pitcheri in Illinois. Natural Areas Journal, 13: 164-176.
Bowles, M.L. & Whelan, C.J. 1994. Restoration of endangered species. Cambridge, University Press.
Boyd, J.M. 1992. Sycamore - a review of its status in conservation in Great Britain. Biologist, 39: 29-31.
Boyer, D.C. & Boyer, H.J. 1989. The status of alien invasive plants in the major rivers of the Namib Naukluft
Park. Madogna, 16: 51-58.
Bradford, D.F., Graber, D.M. & Tabatabai, F. 1994. Population declines of the native frog, Rana muscosa, in
Sequoia and Kings-Canyon National Parks, California. Southwestern Naturalist, 39: 323-327.
Bradshaw, A.D. 1983. The reconstruction of ecosystems. Journal of Applied Ecology, 20: 1-17.
Bradshaw, A.D. 1993. Restoration ecology as a science. Restoration Ecology, 1: 71-73.
Bradshaw, A.D. 1994. Expanding the scope of restoration ecology. Restoration Ecology, 2: 137-146.
Bradshaw, A.D. & Chadwick, M.J. 1980. The restoration of land: The ecology and reclamation of derelict land
and degraded land. Oxford, Blackwell Scientific Publications.
Bragg, T.B. 1986. Prairie transplants: preserving ecological diversity. Proceedings of the tenth North American
Prairie Conference.
Bramwell, D. 1988. Are botanic gardens doing the right conservation research. Species, 11: 30-31.
Bramwell, D. 1991. Botanic gardens in conservation: reintroduction into the wild. In: Tropical botanic gardens:
Their role in conservation and development, ed by V.H. Heywood & P.S. Wyse-Jackson, 209-216,
London, Academic Press.
Brandt, C.A. & Rickard, W.H. 1994. Alien taxa in the north American shrub-steppe four decades after cessation
of livestock grazing and cultivation agriculture. Biological Conservation, 68: 95-105.
Brattstrom, B.H. 1990. Biogeography of the Islas Revillagigedo, Mexico. Journal of Biogeography, 17: 177183.
Breitenmoser, U. 1983. Reintroduction and distribution of the lynx in Switzerland. Schweizerische Zeitschrift
fur Forstwesen, 134: 207-222.
Bridgewater, P.B. 1988. Synthetic plant communities: problems in definition and management. Flora German
Democratic Republic, 180: 139-144.
Review of information, policy and legislation on species translocations
Bright, P.W. & Harris, S. 1993. Reintroduction of the Pine Marten: feasibility Study. Peterborough, English
Nature. (Research Report No. 84.)
Bright, P.W., Mitchell, P. & Morris, P.A. 1994. Dormouse distribution: Survey techniques, insular ecology and
selection of sites for conservation. Journal of Applied Ecology, 31: 329-339.
Bright, P.W. & Morris, P.A. 1994. Animal translocation for conservation: performance of dormice in relation to
release methods, origin and season. Journal of Applied Ecology, 31: 699-708.
Brookes, B.S. 1981. The discovery, extermination, translocation and eventual survival of Schoenus ferrugineus
in Britain. In: The Biological aspects of rare plant conservation, ed by H. Sygne, 421-428. London,
Wiley & sons Ltd.
Brothers, T.S. & Spingarn, A. 1992. Forest fragmentation and alien plant invasion of central Indiana old-growth
forests. Conservation Biology, 6: 91-100.
Brown, J.A. & Cheeseman, C.L. 1995. A study of the effects of translocating a social group of badgers (Meles
meles), with particular reference to foraging and ranging behaviour. Animal Welfare, In press.
Brown, J.H. 1989. Patterns, modes and extents of invasions by vertebrates. In: Biological Invasions, ed by J.A.
Drake, H.A. Mooney, F. di Castri, R.H. Groves, F.J. Kruger, M. Rejmánek & M. Williamson, 85-109.
Chichester, John Wiley & Sons.
Brown, K.C. 1986. Animals, plants and micro-organisms introduced to the British Isles. London, Department of
the Environment.
Brown, K.C. & Williamson, M.H. 1986. Risks of introducing novel species into the UK - Phase II. London,
Department of the Environment.
Brown, V.K., Hendrix, S.D. & Dingle, H. 1987. Plants and insects in early old-field succession: Comparison of
an English site and an American site. Biological Journal of the Linnean Society, 31: 59-74.
Bruckart, W.L. & Dowler, W.M. 1901. Evalution of exotic rust fungi in the United States for classical biological
control of weeds. Microbiological Control of Weeds, 34: 11-14.
Bruggeman, E.P. 1993. Environmental safety issues for genetically modified animals. Journal of animal
science, 71: 47-50.
Brussard, P.F. 1985. Minimum viable populations: How many are too few. Restoration & Management Notes,
3: 21-25.
Bruton, M.N. & As, J. 1986. Faunal Invasions in southern Africa, with suggestions for their management. In:
The ecology and managment of biological invasions in Southern Africa, ed by I.A.W. Macdonald, F.J.
Kruger & A.A. Ferrar, 47-61. Cape Town, Oxford University Press.
Buchting, A.J. 1995. Experiences from release experiments with rhizomania resistant sugar beet.
Zuckerindustrie, 120: 138-142.
Buckingham, H.J. 1987. Monitoring grassland transplants. Brampton Meadow Base-line Phase III Survey.
Peterborough, Nature Conservancy Council.
Bullock, D.J. 1986. The ecology and conservation of reptiles on Round Island and Gunner's Quoin, Mauritius.
Biological Conservation, 37: 135-156.
Bullock, J.A. 1991. The distribution of taxon is that of its students and the diversity of a site is a matter of
serendipity? Antenna, 15: 6-7.
Bullock, J.M., Clear Hill, B. & Dale, M.P. 1994. An experimental study of the effects of sheep grazing on
vegetation change in a species-poor grassland and the role of seed recruitment into gaps. Journal of
Applied Ecology, 31: 493-507.
Bunin, J.S. & Jamieson, I.G. 1995. New approaches toward a better understanding of the decline of takahe
(Pophyrio mantelli) in New Zealand. Conservation Biology, 9: 100-106.
Burgman, M.A., Ferson, S. & Akçakaya, H.R. 1993. Risk assessment in conservation biology. London,
Chapman and Hall.
Burkhead, N.M. & Williams, J.D. 1991. An intergeneric hybrid of a native minnow, the golden shiner,and an
exotic minnow, the rudd. Transactions of the American Fisheries Society, 120: 781-795.
Burns, C. & Sauer, J. 1992. Resistance by natural vegetation in the San Gabriel mountains of California to
invasion by introduced conifers. Global Ecology and Biogeography Letters, 2: 46-51.
Butman, C.A., Carlton, J.T. & Palumbi, S.R. 1995. Whaling effects on deep-sea biodiversity. Conservation
Biology, 9: 462-464.
Byrne, S. 1990. Habitat transplantation in England. A review of the extent and nature of practice and the
techniques employed. Peterborough, Nature Conservancy Council. (England Field Unit.)
Byrne, S.A., Buckingham, H.G., Leach, S.J. & Keymer, R.J. 1991. Progress reports on monitoring of grassland
transplant sites. Thrislington plantation, Durham. Peterborough, English Nature. (EFU Project No
Review of information, policy and legislation on species translocations
Cade, T.J. 1983. Hybridisation and gene exchange among birds in relation to conservation. In: Genetics and
conservation, ed by C.M. Schonewald-Cox, S.M. Shambers, B. MacBryde & W.L. Thomas, 288-309.
London, Benjamin Cummings.
Cade, T.J. 1984. Reintroduction as a method of conservation. In: Raptor Conservation in the Next 50 Years.
Proceedings of a Conference held at Hawk Mountain Sanctuary, Kempton, Pennsylvania, USA. 14
October 1984, ed by S.E. Senner, C.M. White & J.R. Parrish, 72-84. Raptor Research Report No. 5.
Cade, T.J. 1986. Using science and technology to reestablish species lost in nature. In: Biodiversity, ed by E.O.
Wilson, 279-288. Washington, National Acadamy Press.
Cade, T.J. & Jones, C.G. 1993. Progress in restoration of the Mauritius kestrel. Conservation Biology, 7: 169175.
Cade, T.J. & Temple, S.A. 1995. Management of threatened bird species: evaluation of the hands-on approach.
Ibis, 137, S161-S172.
Cairns, J., Jr. 1988. Can the global loss of species be stopped? Speculations in Science and Technology, 11: 189196.
Caldwell, L.K. 1984. International envionment policy: emergence and dimensions. Durham, Duke University
Caldwell, P.M. & Kluge, R.L. 1993. Failure of the introduction of Actinote anteas (Lep, Acraeidae) from Costa
Rica as a biological control candidate for Chromolaena odorata (Asteraceae) in South Africa.
Entomophaga., 38: 475-478.
Calley, M., Braithwaite, R.W. & Ladd, P.G. 1993. Reproductive biology of Ravenala madagascariensis Gmel.
as an alien species. Biotropica, 25: 61-72.
Caltagirone, L.E. & Huffaker, C.B. 1980. Benefits and risks of using predators and parasites for controlling
pests. In: Environmental protection and biological forms of control of pest organisms, ed by B.
Lundholm & M. Stackerud, 103-109. Stockholm, Ecological Bulletins.
Campbell, A. 1988. Epistatic and pleiotrophic effects on genetic manipulation. In: Introduction of genetically
modified organisims into the environment, ed by H.A. Mooney & G. Bernardi, 27-31. Chichester, John
Wiley & Sons.
Campbell, E.W. 1991. The effect of introduced roof rats on bird diversity of Antillean Cays. Journal of Field
Ornithology, 62: 343-348.
Caplan, A. & Van Montagu, M. 1988. Evolutionary consequences of modifying cultivated plants. In:
Introduction of genetically modified organisims into the environment, ed by H.A. Mooney & G.
Bernardi, 57-68. Chichester, John Wiley & Sons.
Carey, P.D. & Brown, N.J. 1994. The use of GIS to idintify sites that will become suitable for a rare orchid,
Himantoglossum hircinium L., in a future changed climate. Biodiversity Letters, 2: 117-123.
Carl, K.P. 1985. Biological control successes in the tropics. Giessener Beitrage zur Entwicklungsforschung, I
Symposien, 12: 27-35.
Carlton, J.T. 1989. Man's role in changing the face of the ocean: biological invasions and implications for
conservation of near-shore environments. Conservation Biology, 3: 265-272.
Carlton, J.T. & Geller, J.B. 1993. Ecological roulette: the global transport of nonindigenous marine organisms.
Science, 261: 78-82.
Carpenter, J.W., Gabel, R.R. & Goodwin, J.G., Jr. 1991. Captive breeding and reintroduction of the endangered
masked bobwhite. Zoo Biology, 10: 439-449.
Carr, G.W., Robin, J.M. & Robinson, R.W. 1986. Environmental weed invasion of natural ecosystems:
Australia's greatest conservation problem. In: Ecology Of Biological Invasions, ed by R.H. Groves &
J.J. Burdon, 150-172. Cambridge, Cambridge University Press.
Carruthers, R.I. & Onsager, J.A. 1993. Perpsective on the use of exotic natural enemies for biological control of
pest grasshoppers (Orthoptera, Acrididae). Environmental Entomology, 22: 885-903.
Carss, D.N. 1993. Shags Phalacrocorax aristotelis at cage fish farms in Argyll, Western Scotland. Bird Study,
40: 203-211.
Carter, C.I. 1983. Some new aphid arrivals to Britain's forests. Proceedings and Transactions of the British
Entomological and Natural History Society, 16: 81-87.
Case, T.J. & Bolger, D.T. 1991. The role of introduced species in shaping the distribution and abundance of
island reptiles. Evolution, 5: 272-290.
Cayford, J. & Percival, S.M. 1992. Born captive, die free. New Scientist, 133: 29-33.
Center, T.D., Frank, J.H. & Dray, F.A. 1995. Biological invasions: Stemming the tide in Florida. Florida
Entomologist, 78: 45-55.
CEP 1989. Royal commission on environmental pollution: The release of genetically modified organisms into
the enviroment. London, HMSO.
Review of information, policy and legislation on species translocations
Cezilly, F. & Johnson, A.R. 1992. Exotic flamingoes in the Western Mediterranean region - a case for concern.
Colonial Waterbirds., 15: 261-263.
Chambers, P.A., Barko, J.W. & Smith, C.S. 1993. Evaluation of invasions and declines of submersed aquatic
macrophytes. Journal of Aquatic Plant Management, 31: 218-220.
Chapman, J.W. & Carlton, J.T. 1991. A test of criteria for introduced species - the global invasion by the isopod
Synidotea laevidorsalis (Miers, 1991). Journal of Crustacean Biology, 11: 386-400.
Chapman, N., Harris, S. & Stanford, A. 1994. Reeve's Muntjac Muntiacus reevsii in Britain: their history,
spread, habitat selection and the role of human intervention in accelerating their dispersal. Mammal
Review, 24: 113-160.
Chapman, S.B. & Rose, R.J. 1994. Changes in the distribution of Erica ciliaris L. and E. watsonii Benth. in
Dorset, 1963-1987. Watsonia, 20: 89-95.
Chapuis, J.L., Bousses, P. & Barnaud, G. 1994. Alien mammals, impact and management in the French
subantarctic islands. Biological Conservation, 67: 97-104.
Chen, C., Bauske, E.M., Musson, G., Rodriguezkabana, R. & Kloepper, J.W. 1995. Biological control of
fusarium wilt on cotton by use of endophytic bacteria. Biological Control, 5: 83-91.
Choquenot, D., Kay, B. & Lukins, B. 1990. An evaluation of warfarin for the control of feral pigs. Journal of
Wildlife Management, 54: 353-359.
Christensen, O.M. & Mather, J.G. 1995. Colonisation by the land planarian Artioposthia triangulata and impact
on lumbricid earthworms at a horticultural site. Pedobiologia, 39: 144-154.
Clark, T.W. & Westrum, R. 1989. High-performance teams in Wildlife Conservation: a species reintroduction
and recovery example. Environmental Management, 13: 663-670.
Clarke, J.A. & Johnson, R.E. 1990. Biogeography of white-tailed ptarmigan (Lagopus leuurus) - implications
from an introduced population in the Sierra Nevada. Journal of Biogeography, 17: 649-656.
Clout, M.N. 1984. Improving exotic forests for native birds. New Zealand Journal of Forestry, 29: 193-200.
Clutton-Brock, J. 1989. Five thousand years of livestock in Britain. Biological Journal of the Linnean Society,
38: 31-37.
Clutton-Brock, J. 1991. Canadian beaver, Castor canadensis. In: The handbook of British mammals, ed by G.B.
Corbet & S. Harris, 572. Oxford, Blackwell Scientific Publications.
Coblentz, B.E. 1990. Exotic organisms: A dilemma for conservation biology. Conservation Biology, 4: 261-265.
Cohen, A.N., Carlton, J.T. & Fountain, M.C. 1995. Introduction, dispersal and potential impacts of the green
crab Carcinus maenas in San Francisco Bay, California. Marine Biology, 122: 225-237.
Cole, F.R., Loope, L.L., Medeiros, A.C., Raikes, J.A. & Wood, C.S. 1995. Conservation implications of
introduced game birds in high- elevation Hawaiian shrubland. Conservation Biology, 9: 306-313.
Cole, F.R., Medeiros, A.C., Loope, L.L. & Zuehlke, W.W. 1992. Effects of the Argentine ant on arthropod
fauna of Hawaiian high-elevation shrubland. Ecology., 73: 1313-1322.
Collier, K. 1993. Review of the status, distribution, and conservation of fresh-water invertebrates in New
Zealand. New Zealand Journal of Marine and Freshwater Research, 27: 339-356.
Colwell, R.R. 1994. Biodiversity and release of genetically engineered organisms: A partnership of value. Curr.
Opin. Biotechnol., 5: 244-246.
Conant, S. 1988. Saving endangered species by translocation: are we tinkering with evolution? Bioscience, 38:
Connell, J.H. & Orias, E. 1964. The ecological regulation of species diversity. American Naturalist, 98: 399414.
Connor, H.E., Edgar, E. & Bourdot, G.W. 1993. Ecology and distribution of naturalized species of stipa in New
Zealand. New Zealand Journal of Agricultural Research, 36: 301-307.
Conservation Committee of the British Herpetological Society. 1983. Herpetofauna translocations in Britain - a
policy. British Journal of Herpetology, 6: 314-316.
Cooke, A.S. & Coldham, R.S. 1995. Establishment of populations of the common frog, Rana temporia and
common toad,Bufo bufo, in e newly created reserve following translocation. Herpetological Journal, 5:
Cooper, J. & Brooke, R.K. 1986. Alien plants and animals on South Africa continental and oceanic islands:
Species richness, ecological impacts and management. In: The ecology and management of biological
invasions in Southern Africa, ed by I.A.W. MacDonald, F.J. Kruger & A.A. Ferrar, 133-142. Cape
Town, Oxford University Press.
Cooper, M.E. 1993. Legal Implications for the management of infectious disease in captive breeding and
reintroduction programs. Journal of Zoo and Wildlife Medicine, 24: 296-303.
Copp, G.H., Vaughan, C. & Wheeler, A. 1993. 1st occurance of the North-American white sucker Catostomus
commersoni in Great Britain. Journal of Fish Biology, 42: 615-617.
Review of information, policy and legislation on species translocations
Corbett, K.F. 1994. Pilot Study for Sand Lizard UK Recovery Programme. English Nature Research Reports No
102. Peterborough: Englich Nature.
Cordo, H.A., Deloach, C.J. & Ferrer, R. 1995. Host-range of the Argentine root borer Carmenta haematica
(Ureta) (Lepidoptera, Sesiidae), a potential biocontrol agent for snakeweeds (Gutierrezia spp in the
United States. Biological Control, 5: 1-10.
Corlett, R.T. 1988. The naturalized flora of Singapore. Journal of Biogeography, 15: 657-663.
Corlett, R.T. 1992. The ecological transformation of Singapore, 1819-1990. Journal of Biogeography, 19: 411420.
Corlett, R.T. 1992. The naturalized flora of Hong Kong: A comparison with Singapore. Journal of
Biogeography, 19: 421-430.
Costapierce, B.A. 1992. Review of the spawning requirements and feeding ecology of silver
carp(Hypophthalmichthys molitrix) and the reevaluation of its use in fisheries and aquaculture. Reviews
in Aquatic Sciences, 6: 257-273.
Costello, M.J. 1993. Biogeography of alien amphipods occurring in Ireland, and interactions with native species.
Crustaceana, 65: 287-299.
Coulson, J.R. 1992. Documentation of classical biological-control introductions. Crop Protection, 11: 195-205.
Council of Europe. 1979a. Convention on the conservation of European wildlife and natural habitats. (Bern
Convention). Strasbourg, Council of Europe. (European treaty series 104.)
Council of Europe. 1979b. Council directives of 2 April 1979 on Conservation of wild birds 79/409/EEC.
Official Journal of the European Communities, L103: 1-18.
Council of Europe. 1988. Council directive amending directive 77/93/EEC on protective measures against
introduction into the member states of organisms harmful to plants or plant products. Strasbourg,
Council of Europe.
Council of Europe. 1992. Council Directive 92/43/EEC. Official Journal of the European Communities, L206/7.
Council of Europe. 1993a. Secretariat memorandum. Group of experts on legal aspects of introduction and
reintroduction of wildlife species. Strasbourg, Council of Europe.
Council of Europe. 1993b. Convention on civil liability for damage resulting from activities dangerous to the
environment. Strasbourg, Council of Europe. (European treaty series 150. Lugano, July 1993.)
Council of Europe. 1994. Model act on the protection of the environment. Strasbourg.
Council of Europe. 1995. Replies to the questionnaire in national amd community legislation and practice
applicable to the introduction and reintroduction of wild species. Strasbourg, Council of Europe.
Countryside Commission. 1995. Countryside stewardship: handbook and application form. London, HMSO.
(CCP 453.)
Courant, A.V., Holbrook, A.E., Van-der-Reijden, E.D. & Chew, F.S. 1994. Native pierine butterfly (Pieridae)
adapting to naturalized crucifer? Journal of the Lepidoptera Society, 48: 168-170.
Cowie, I.D. & Werner, P.A. 1993. Alien plant species invasive in Kakadu National Park, tropical northern
Australia. Biological Conservation, 63: 127-135.
Cox, J.H.S., Leach, S.J., Byrne, S.A., Blake, C.P. & Buckingham, H.G. 1991. Progress reports on monitoring of
grassland transplant sites. Ashcott heath, Somerset 1987-1990. Peterborough, English Nature. (EFU
Project No 105/01.)
Cox, J.H.S., Leach, S.J., Byrne, S.A., Blake, C.P. & Buckingham, H.G. 1992. Progress reports on monitoring of
grassland transplant sites. Newhall reservoir, Nottinghamshire 1987-1990. Peterborough, English
Nature. (EFU Project NO 105/05.)
Cranston, D.M. & Valentine, D.H. 1983. Transplant experiments on rare plant species from Upper Teesdale.
Biological Conservation, 26: 175-191.
Crawley, M.J. 1986. The population biology of invaders. In: Quantitative aspects of the ecology of biological
invasions, ed by H. Kornberg & M.H. Williamson, 209-228, London, Royal Society.
Crawley, M.J. 1987. What makes a community invasible? In: Colonization, succession and stability, ed by A.J.
Gray, M.J. Crawley & P.J. Edwards, 429-453. Oxford, Blackwell Scientific Publications.
Crawley, M.J. 1988. The ecology of genetically engineered organisms: assessing the environmental risks. In:
Introduction of genetically modified organisims into the environment, ed by H.A. Mooney & G.
Bernardi, 133-150. Chichester, John Wiley & Sons.
Crawley, M.J. 1989. The successes and failures of weed biocontrol using insects. Biocontrol News and
Information, 10: 213-223.
Crawley, M.J., Hails, R.S., Rees, M., Kohn, D. & Buxton, J. 1993. Ecology of transgenic oilseed rape in natural
habitats. Nature, 363: 620-623.
Cresswell, J.E. 1994. A method for quantifying the gene flow that results from a single bumblebee visit using
transgenic oilseed rape, Brassica napus L CV westar. Transgenic Research, 3: 134-137.
Review of information, policy and legislation on species translocations
Critchley, A.T., Nienhuis, P.H. & Verschuure, K. 1987. Presence and development of populations of the
introduced brown alga Sargassum muticum in the southwest Netherlands. In: Twelfth International
Seaweed Symposium, ed by M.A. Ragan & C.J. Bird, 245-255.
Crivelli, A.J. 1995. Are fish introductions a threat to endemic freshwater fishes in the northern Mediterranean
region? Biological Conservation, 72: 311-319.
Cross, J.R. 1975. Rhododendron ponticum L. (Biological flora of the British Isles, no.137). Journal of Ecology,
63, 345-364.
Crossman, E.J. 1991. Introduced fresh-water fishes - a review of the North-American perspective with emphasis
on Canada. Canadian Journal of Fisheries and Aquatic Science, 48: 46-57.
Crouch, J.H., Lewis, B.G., Lydiate, D.J. & Mithen, R. 1995. Genetic diversity of wild, weedy and cultivated
forms of Brassica rapa. Heredity, 74: 491-496.
Crozier, R.H. 1992. Genetic diversity and the agony of choice. Biological Conservation, 61: 11-15.
Crozier, W.W. 1993. Evidence of genetic interaction between escaped farmed salmon and wild Atlantic salmon
Salmo salar L. in a Northern Irish river. Aquaculture, 113: 19-29.
Cruz, F., Cruz, J. & Lawesson, J.E. 1986. Lantana camara a threat to native plants and animals. Noticias de
Galapagos, 42: 10-11.
Cruz, J.B. & Cruz, F. 1987. Conservation of the Dark-rumped Petrel Pterodroma phaeopygia in the Galapagos
Islands, Ecuador. Biological Conservation, 42: 303-311.
Cruz, J.B.C., F. 1987. Conservation of the Dark-rumped Petrel Pterodroma phaeopygia in the Galapagos
Islands, Ecuador. Biological Conservation, 42: 303-311.
Csermely, D. & Corona, C.V. 1994. Behavior and Activity of Rehabilitated Buzzards (Buteo buteo) Released in
Northern Italy. Journal of Raptor Research, 28: 100-107.
Cullen, W.R. & Wheater, C.P. 1993. The flora and invertebrate fauna of a relocated grassland at Thrislington
Plantation, County Durham, England. Restoration Ecology, 1: 130-137.
Cully, J.F., Jr. 1993. Plague, prairie dogs, and black-footed ferrets. In: Proceedings of the symposium on the
management of prairie dog complexes for the reintroduction of the black footed ferret, ed by J.L.
Oldemeyer, D.E. Biggins, B.J. Miller & R. Crete, 38-49.
Cummings, al. 1995. Evaluation of rejex-it AG-36 as a canada goose grazing repellent. Journal of Wildlife
Management, 59: 47-50.
Curry, G.N. 1991. The influence of proximity to plantation edge on diversity and abundance of exotic bird
species in north-eastern New South Wales. Wildlife Research, 18: 299-314.
Dale, P.J. 1994. The impact of hybrids between genetically-modified crop plants and their related species general-considerations. Molecular Ecology, 3, 31-36.
Dalrymple, N.K., Dalrymple, G.H. & Fanning, K.A. 1993. Vegetation of restored rock-plowed wetlands of the
East Everglades. Restoration Ecology, 1: 220-225.
Daly, K. 1989. Eradication of feral goats from small islands. Oryx, 23: 71-75.
Damascos, M.A. & Gallopin, G.G. 1992. Ecology of an introduced shrub (Rosa rubiginosa-L = Rosa eglanteria
L) - invasion risks and effects of the Andean-Patagonic region of Argentina. Revista Chilena de
Historia Natural, 65: 395-407.
Daniels, T.J. & Bekoff, M. 1989. Feralization: The making of wild domestic animals. Behavioural Processes,
19: 79-94.
D'Antonio, C.M. 1993. Mechanisms controlling invasion of coastal plant communities by the alien succulent
Carpobrotus edulis. Ecology, 74: 83-95.
D'Antonio, C.M., Odion, D.C. & Tyler, C.M. 1993. Invasion of maritime chaparral by the introduced succulent
Carpobrotus edulis. The roles of fire and herbivory. Oecologia, 95: 14-21.
D'Antonio, C.M. & Vitousek, P.M. 1992. Biological invasions by exotic grasses, the grass fire cycle, and global
change. Annual Review of Ecology and Systematics, 23: 63-87.
Darmency, H. 1994. The impact of hybrids between genetically-modified crop plants and their related species introgression and weediness. Molecular Ecology, 3, 37-40.
Darrigran, G. & Pastorino, G. 1995. The Recent Introduction of a Freshwater Asiatic Bivalve, Limnoperna
Fortunei (Mytilidae) into South America. Veliger, 38: 171-175.
Dasilva, J.M.C.O., DC & Oren, D.C. 1990. Introduced and invading birds in Belem, Brazil. Wilson Bulletin,
102: 309-313.
Davies, J.C. 1991. Global support and coordination: Conserving germplasm of world crop species and their
relatives. Biological Journal of the Linnean Society, 43: 61-71.
Davis, B.N.K. 1989. Habitat creation for butterflies on a landfill site. Entomologist, 108: 109-121.
Davis, M.H. 1983. Post-release movements of introduced marten. Journal of Wildlife Management, 47: 59-66.
Review of information, policy and legislation on species translocations
Dawson, F.H. & Warman, EA. 1987. Crassula helmsii (T.Kirk) Cockayne: Is it an aggressive alien aquatic plant
in Britain? Biological Conservation, 42: 247-272.
Day, W.H., Prokrym, D.R., Ellis, D.R. & Chianese, R.J. 1994. The known distribution of the predator Propylea
quattuordecimpunctata (Coleoptera, Coccinellidae) in the United States, and thoughts on the origin of
this species and 5 other exotic lady beetles in eastern North America. Entomological News, 105: 244256.
Greatorex-Davies, J.N. 1991. The introduction of the silver-washed fritillary (Argynnis paphia) into Waresley
Wood, Cambridgeshire, in 1989, and its subsequent monitoring. Report of the Huntingdonshire Moth
and Butterfly Group, 1: 12-20.
Deacon, J. 1986. Human settlement in South Africa and archaeological evidence for alien plants and animals.
In: The Ecology And Management Of Biological Invasions In Southern Africa, edited by I.A.W.
Macdonald, F.J. Kruger & A.A. Ferrar, 3-19. Cape Town: Oxford University Press.
Dean, S.J., Holmes, P.M. & Weiss, P.W. 1986. Seed biology of invasive alien plants in South Africa and South
West Africa/Namibia. In: The ecology and management of biological invasions in Southern Africa, ed
by I.A.W. MacDonald, F.J. Kruger & A.A. Ferrar, 157-170. Cape Town, Oxford University Press.
Dean, W.R.J., Milton, S.J., Ryan, P.G. & Moloney, C.L. 1994. The role of disturbance in the establishment of
indigenous and alien plants at Inaccessible and Nightingale Islands in the South Atlantic Ocean.
Vegetatio, 113: 13-23.
De Goffau, L.J.W. 1989. Problems caused by indigenous and exotic Scarabaeidae (Coleoptera) in the
Netherlands. Proceedings of the Section of Experimental and Applied Entomology, Netherlands
Entomological Society, 1: 152-157.
Degroot, S.J. 1985. Introductions of non-indigenous fish species for release and culture in the Netherlands.
Aquaculture, 46: 237-257.
Dejeant-Pons, M. 1993. European biodiversity: the Bern convention of 19 September 1979 on the conservation
of European wildlife and natural habitats. Strasbourg, Council of Europe.
De Jongh, A. 1995. Otter Seminar. Naturopa, 77: 9.
De Klemm, C.1995. Introductions of non-native organisms into the environment. Council of Europe Group of
experts on legal aspects of introduction and reintroduction of wildlife species.Strasbourg, Council of
Delany, S. 1993. Introduced and escaped geese in Britain in summer 1991. British Birds, 86: 591-599.
De Leij, F.A.A.M., Sutton, E.J., Whipps, J.M. & Lynch, J.M. 1994. Effect of a genetically-modified
pseudomonas-aureofaciens on indigenous microbial-populations of wheat. Fems Microbiology
Ecology, 13: 249-257.
De Marais, B. & Minckley, W.L. 1993. Genetics and morphology of Yaqui chub Gila purpurea, an endangered
cyprinid fish subject to recovery efforts. Biological Conservation, 66: 195-206.
Demoor, F.C. 1992. Factors influencing the establishment of aquatic insect invaders. Transactions of the Royal
Society of South Africa, 48: 141-158.
Dempster, J.P. 1989. Insect introductions: natural and population persistence in insects. Entomologist., 108: 513.
Dennill, G.B. & Donnelly, D. 1991. Biological control of Acacia longifolia and related weed species (Fabaceae)
in South Africa. Agriculture, Ecosystems and Environment, 37: 115-135.
Dennill, G.B., Donnelly, D. & Chown, S.L. 1993. Expansion of host-plant range of a biocontrol agent
Trrichilogaster acaciaelongifoliae (Pteromalidae) released against the weed Acacia longifolia in South
Africa. Agriculture, Ecosystems and Environment, 43: 1-10.
Dennis, R.L.H. & Williams, W.R. 1986. Butterfly diversity - regressing and a little latitude. Bulletin of the
Royal Entomological Society London, 10: 108-112.
Denton, J.S., Hitchings, S. & Beebee, T.J.C. (1995): Natterjack Toad recovery programme 1992-1995. Final
report. School of Biological Sciences, University of Sussex.
Department of the Environment. 1981. Wildlife and Countryside Act. London, HMSO.
Department of the Environment. 1985. Wildlife and Countryside Act (Amendment). London, HMSO.
Department of the Environment. 1991. A Guide to the Environmental Protection Act 1990. London, HMSO.
Department of the Environment. 1992. Genetically Modified Organisms. Revised proposals for new regulations.
Health and Saftey Commission. Scottish Office, Welsh Office. London, HMSO.
Department of the Environment. 1994a. Genetic modification of fish - A UK perspective. London, HMSO.
Department of the Environment. 1995. A guide to risk assessment and risk management for enviromental
protection. London, HMSO.
Department of Transport. 1993. The wildflower handbook. London, HMSO.
Review of information, policy and legislation on species translocations
De Pietri, D.E. 1992. Alien shrubs in a national park: Can they help in the recovery of natural degraded forest?
Biological Conservation, 61: 127-130.
De Waal, L.C., Wade, P.M. & Brock, J.H., eds. 1994. Ecology and management of invasive riverside plants.
Chichester, John Wiley & Sons.
Dewald, L. & Wilzbach, M.A. 1992. Interactions between native brook trout and hatchery brown trout - effects
on habitat use, feeding and growth. Transactions of the American Fisheries Society, 121: 287-296.
Diamond, J. 1987. Reflections on goals and on the relationship between theory and practice. Restoration
Ecology, 329-336.
Diaz, G. & Honrubia, M. 1995. Effect of native and introduced arbuscular mycorrhizal fungi on growth and
nutrient uptake of Lygeum spartum and Anthyllis cytisoides. Biologia Plantarum, 37: 121-129.
Diefenbach, D.R., Baker, L.A., James, W.E., Warren, R.J. & Conroy, M.J. 1993. Reintroducing bobcats to
Cumberland Island, Georgia. Restoration Ecology, 1: 241-247.
Dobson, A.P. & Lyles, A.M. 1989. The population dynamics and conservation of primate populations.
Conservation Biology, 3, 362-380.
Dodd, F.S., de Waal, L.C., Wade, P.M. & Tiley, G.E.D. 1994. Control and management of Heracleum
mantegazzainum (giant hogweed). In: Ecology and management of invasive riverside plants, ed by
L.C. de Waal, L.C. Child, P.M. Wade & J.H. Brock, 101-126. Chichester, John Wiley & Sons.
Donath, H. 1989. Considerable losses in bumble bees and other flower-visiting insects due to exotic species of
lime (Tilia tomentosa Moench, Tilia euchlora C. Koch). Archiv fur Naturschutz und
Landschaftsforschung, 29: 117-120.
Dowell, R.V. & Gill, R. 1989. Exotic invertebrates and their effects on California. Pan Pacific Entomologist,
65: 132-145.
Down, G.S. & Morton, A.J. 1989. A case study of whole woodland transplanting. In: Biological habitat
reconstruction, edited by G.P. Buckley, 251-257. London, Belhaven.
Drake, J.A. & Mooney, H.A. 1989. Biological invasions: a global perspective. Chichester, John Wiley & Sons.
Drake, J.A., Mooney, H.A., Di Castri, F., Groves, R.H., Kruger, F.J., Rejmanek, M. & Williamson, M., eds.
1988. Biological invasions: a global perspective. Chichester, John Wiley & Sons.
Drake, M. 1994. Effects of wild flower seed harvesting on invertebrates. In: Species conservation handbook, 12. Peterborough, English Nature.
Driestadt, S.H. & Hagen, K.S. 1994. Classical biological-control of the acacia psyllid, Acizzia uncatoides
(Homoptere, Psyllidae) and predator-prey-plant interactions in the San Francisco Bay area. Biological
Control, 4: 319-327.
Dunwell, J.M. & Paul, E.M. 1990. Impact of genetically modified crops in agriculture. Outlook On Agriculture,
19: 103-109.
Dutton, J. 1994. Introduced mammals in Sao-Tome-and-Principe - possible threats to biodiversity. Biodiversity
and Conservation, 3: 927-938.
Dyer, A.F. 1994. Natural soil spore banks - Can they be used to retrieve lost ferns? Biodiversity and
Conservation, 3: 160-175.
Dyer, C. & Richardson, D.M. 1992. Population genetics of the invasive Australian shrub Hakea sericea
(Proteaceae) in South Africa. South Africa Journal of Botany, 58: 117-124.
Ebenhard, T. 1988. Introduced birds and mammals and their ecological effects. Swedish Wildlife Research, 13:
Edwards, C., ed. 1993. Monitoring genetically manipulated micro-organisms in the environment. Chichester:
Edwards, P.J., May, R.M. & Webb, N.R., eds. 1993. Large-scale ecology and conservation biology. London,
Blackwell Scientific Publications.
Ehler, L.E. 1992. Guild analysis in biological control. Environmental Entomology, 21: 26-40.
Ehler, L.E. & Hall, R.W. 1982. Evidence for competitive-exclusion of introduced natural enemies in biological
control. Environmental Entomology, 11: 1-4.
Ehler, L.E. & Hall, R.W. 1984. Evidence for competitive-exclusion of introduced natural enemies in biologicalcontrol - an addedum. Environmental Entomology, 13, R5-R7.
Ehrlich, P.R. 1986. Which animal will invade? In: Ecology of biological invasions of North America and
Hawaii. Ecological studies 58, ed by H.A. Mooney & J.A. Drake, 79-95. New York, Springer-Verlag.
Elliot, J.M. & Mills, C.J. 1989. Release of captive bred specise: freshwater fish. In: Release of captive bred
species: genetic interactions with wild relatives, ed by A.J. Gray, 21-36, Unpublished ITE Report to
the Department of the Environment.
Elliott, G., Dennis, R., Love, J., Pienkowski, M. & Broad, R. 1991. A future for the White-tailed Eagle in
Britain. RSPB Conservation. Review, 5: 41-46.
Review of information, policy and legislation on species translocations
Ellis, R.G. 1987. A cautionary tale. Botanical Society of the British Isles News, 46: 13-14.
Ellstrand, N.C. & Hoffman, C.A. 1990. Hybridization as an avenue of escape for engineered genes. Bioscience,
40: 438-442.
Elmes, G.W. 1989. Contamination of wild honey-bee stocks by genetic exchange with imported domesticated
races. In: Release of captive bred species: genetic interactions with wild relatives, ed by A.J. Gray, 4251, Unpublished ITE Report to the Department of the Environment.
Elton, C.S. 1958. The ecology of invasions by animals and plants. London, Chapman and Hall.
Elvira, B. 1995. Native and exotic freshwater fishes in Spanish river basins. Freshwater Biology, 33: 103-108.
Emmet, A.M. & Heath, J. 1989. The moths and butterflies of Great Britain and Ireland. Vol. 7. Colchester,
Harley Books.
England. 1974. A further review of the problem of "escapes". British Birds, 67: 177-196.
English Nature. 1993. Roads and nature conservation. Peterborough, English Nature.
English Nature. 1994a. Translocations and Establishments. In: Species Conservation Handbook. Peterborough,
English Nature.
English Nature. 1994b. Species Recovery Programme Newsletter. Issue 6. Peterborough, English Nature.
English Nature. 1994c. Species Conservation Handbook. Peterborough, English Nature.
English Nature. 1994d. Translocations: rationale and objectives. Discussion document for species recovery
programme for the sand lizard. Peterborough, English Nature.
Eno, N.C. 1995. Non-native marine species in British waters. Peterborough, Joint Nature Conservation
Erwin, T.L. 1991. An evolutionary basis for conservation strategies. Science, 253: 750-752.
Esler, A.E. 1987. The naturalization of plants in urban Auckland, New Zealand. 1. The introduction and spread
of alien plants. New Zealand Journal of Botany, 25: 511-522.
Esler, A.E. 1988a. The naturalisation of plants in urban Auckland, New Zealand. 4. The nature of the naturalised
species. N.Z. J. Bot., 26: 345-385.
Esler, A.E. 1988b. The naturalisation of plants in urban Auckland, New Zealand. 5. Success of the alien species.
N.Z. J. Bot., 26: 565-584.
Esler, A.E. 1988c. The naturalisation of plants in urban Auckland, New Zealand. 6. Alien plants as weeds. N.Z.
J. Bot., 26: 585-618.
Esler, A.E. & Astridge, S.J. 1987. The naturalisation of plants in urban Auckland, New Zealand. 2. Records of
introduction and naturalisation. N.Z. J. Bot., 25: 523-537.
European Inland Fisheries Advisory Committee. 1983. Report of the EIFAC working party on stock
enhancement. FAO. (EIFAC technical paper 44.)
Evans, H.F., Stoakley, J.T., Leather, S.R. & Watt, A.D. 1991. Development of an intergrated approach to
control of pine beauty moth in Scotland. Forest Ecology and Management, 39: 19-28.
Evans, I. 1996. Reintroduction as a conservation tool. In: Feral and introduced birds. A British Ornithologists'
Union & Joint Nature Conservation Committee Conference, 1995. Peterborough, Joint Nature
Conservation Committee.
Evans, I.M., Love, J.A., Galbraith, C.A. & Pienkowski, M.W. 1994. Population and Range Restoration of
Threatened Raptors in the United Kingdom. Raptor Conservation Today, 447-457.
Evans, I.M. & Pienkowski, M.W. 1991. A background to the experimental reintroduction to England and
Scotland. British Birds, 84: 171-187.
Evans, I.M. & Pienkowski, M.W. 1991. World status of the red kite. A background to the experimental
reintroduction to England and Scotland. British Birds, 84: 171-187.
Everett, M.J. 1978. Birds of Prey on Royal society for the Protection of Birds Reserves. In: Bird of Prey
Management Techniques, ed by T.A. Geer, 16-22. Oxford, The British Falconers' Club.
Eversham, B.C. & Arnold, H.R. 1992. Introductions and their place in British wildlife. In: Biological recording
of changes in British wildlife, ed by P.T. Harding, . (ITE Symposium No. 26). London, HMSO.
Ewel, J.J. 1986. Invasibility: Lessons from South Florida. In: Ecology of biological invasions of North America
and Hawaii, ed by H.A. Mooney & J.A. Drake, 214-230. New York, Springer-Verlag.
Exton, D., Morris, M.G., Thomas, J.A. & Webb, N.R. 1991. M3 Bar End to Compton habitat reconsruction: a
review. Wareham, Institute of Terrestrial Ecology.
Fa, J.E. 1994. Herbivore intake/habitat productivity correlations can help ascertain re-introduction potential for
the Barbary macaque. Biodiversity and Conservation, 3, 309-317.
Fahselt, D. 1988. The dangers of transplantation as a conservation technique. Natural Areas Journal, 8: 238244.
Fairey, J.S. & Smal, C.M. 1987. Feral house mice in Ireland. Irish Naturalists Journal, 22: 284-290.
Faith, D.P. 1992. Conservation evaluation and phylogenetic diversity. Biological Conservation, 61: 1-10.
Review of information, policy and legislation on species translocations
Falk, D.A. 1990. Restoration of endangered species: A strategy for conservation. In: Environmental restoration
science and strategies for restoring the earth, ed by J.J. Berger, 328-334. Washington, Island Press.
Falk, D.A. & Holsinger, K.E., eds. 1991. Genetics and conservation of rare plants. Oxford, Oxford University
Falk, D.A. & Olwell, p. 1992. Scientific policy considerationss in restoration and reintroduction of endangered
species. Rhodora., 94: 287-315.
FAO. 1984. Report of the EIFAC working party on stock enhancement. FAO. (EIFAC technical paper.)
Farrell, L. & Fitzgerald, R. 1989. The Nature Conservancy and orchid conservation. In: Modern methods in
orchid conservation: the role of physiology, ecology and management, ed by H.W. Pritchard, 147-152.
Cambridge, Cambridge University Press.
Faust, R., ed. 1994. Bearded Vulture reintroduction into the Alps (Annual report 1994). Wien, Austria,
Foundation for the Conservation of the Bearded Vulture.
Feare, C.J.M., Y. 1990. The Status and Management of the House Crow Corvus splendens (Vieillot) in
Mauritius. Biological Conservation, 51: 63-70.
Fenner, F. 1988. Patterns of establishment and the spread of animal viruses. In: Introduction of genetically
modified organisims into the environment, ed by H.A. Mooney & G. Bernardi, 99-116. Chichester,
John Wiley & Sons.
Fensham, R.J., Fairfax, R.J. & Cannell, R.J. 1994. The invasion of Lantana camara L in Forty Mile Scrub
National Park, north Queensland. Australian Journal of Ecology, 19: 297-305.
Ferguson, M.M. 1990. The genetic impact of introduced fishes on native species. Canadian Journal of Zoology Journal Canadien de Zoologie, 68: 1053-1057.
Ferreira, M.C., Guimaraes, J.M., Corinta-Ferreira, M. & Monteiro-Guimaraes, J. 1984. Exotic agricultural pests
that threaten the Portuguese continent and adjacent islands. Boletim da Sociedade Portuguesa de
Entomologia, II-16: 169-183.
Ferry, B. (1994): Species Recovery Programme - Stinking Hawksbeard Crepis foetida Final report - November
1994.. Peterborough, English Nature. (Project No F72-21-23.)
Fielder, P.L. & Jain, S.K. 1992. Conservation Biology- The theory and practice of nature conservation
preservation and management. London, Chapman and Hall.
Fisher, S.W. & Briggs, J.D. 1988. Environmental and ecological problems in the introduction of alien
microorganisms in the soil. Agricuture, Ecosystems, Environment, 24: 325-335.
Fitzgerald, B.M. 1990. Is cat control needed to protect urban wildlife? Environmental Conservation, 17: 168169.
Fitzgerald, J.P. 1993. The ecology of plague in Gunnison's prairie dogs and suggestions for the recovery of
black-footed ferrets. In: Proceedings of the symposium on the management of prairie dog complexes
for the reintroduction of the black footed ferret, ed by J.L. Oldemeyer, D.E. Biggins, B.J. Miller & R.
Crete, 50-59.
Flannery, T.F. & White, J.P. 1991. Animal translocation. Research & Exploration, 7: 96-113.
Flecker, A.S. & Townsend, C.R. 1994. Community-wide consequences of trout introduction in New Zealand
streams. Ecological Applications, 4: 798-807.
Fleming, I.A., Jonsson, B., Gross, M.R. & Lamberg, A. in press. An experimental study of the reproductive
behaviour and success of farmed and wild Atlantic salmon Salmo salar. Journal of Applied Ecology, .
Fontaine, T.D. & Stewart, D.J. 1992. Exploring the effects of multiple management-objectives and exotic
species on Great-Lakes food webs and contaminant dynamics. Environmental Management, 16: 225229.
Forcella, F. & Harvey, S.J. 1988. Patterns of weed migration in northwestern U.S.A. Weed Science, 36: 194201.
Forcella, F. & Wood, J.T. 1984. Colonization potentials of alien weeds are related to their "native" distributions:
Implications for plant quarantine. Jounal of the Australian Institute of Agricultural Science, 50: 35-40.
Forey, P.L., Humphries, C.J. & Vane-Wright, R.I., eds. 1994. Systematics and conservation evaluation. Oxford,
Clarendon Press.
Fowler, H.G., Schlindwein, M.N. & de Medeiros, M.A. 1994. Exotic ants and community simplification in
Brazil: a review of the impact of exotic ants on native ant assemblages. In: Exotic ants: biology,impact
and control of introduced species, ed by D.F. Williams, 151-162. California, Westview Press.
Fox, M.D. 1988. The ecological status of alien plant species. In: Weeds on public land - an action plan for
today. Proceedings of a Symposium presented by the Weed Science Society of Victoria Inc. and the
School of Environmental Science, Monash University, ed by R.G. Richardson, 42-45. Victoria,
Australia, Weed Science Society of Victoria Inc.
Review of information, policy and legislation on species translocations
Frank, J.H. & McCoy, E.D. 1993. The behavioral ecology of introduction - the introduction of insects into
Florida. Florida Entomologist, 76: 1-53.
Frank, J.H. & McCoy, E.D. 1995. Introduction to insect behavioral ecology: The good, the bad, and the
beautiful: Non-indigenous species in Florida. Florida Entomologist, 78: 1-15.
Frankham, R. & Loebel, D.A. 1992. Modeling problems in conservation genetics using captive Drosophila
populations: Rapid genetic adaptation to captivity. Zoo Biology, 11: 333-342.
Franklin, J.F. 1993. Preserving biodiversity: Species, ecosystems, or landscapes? Ecological Applications, 3:
Franklin, J. & Steadman, D.W. 1991. The potential for conservation of Polynesian birds through habitat
mapping and species translocation. Conservation Biology, 5: 506-521.
Franzreb, K.E. 1990. An analysis of options for reintroducing a migratory, native passerine, the endangered least
Bell's vireo Vireo bellii pusillus in the Central Valley, California. Biological Conservation, 53: 105123.
Fraser, M.W. & Crowe, T.M. 1990. Effects of alien woody plant invasion on the birds of Mountain Fynbos in
the Cape of Good Hope Nature Reserve. South Africa Journal of Zoology, 25: 97-108.
Frazer, J.F.D. 1964. Introduced species of amphibians and reptiles in mainland Britain. British. Journal. of.
Herpetology., 3: 145-150.
Freeland, W.J. 1990. Large herbivorous mammals: Exotic species in northern Australia. Journal of
Biogeography, 17: 445-449.
Frey, H. & Lexmond, M.B.van. 1994. The Reintroduction of the Bearded Vulture, Gypaetus barbatus aureus
into the Alps. Raptor Conservation Today.
Frissell, C.A. 1993. Topology of extinction and endangerment of native fishes in the Pacific Northwest and
California (U.S.A.). Conservation Biology, 7: 342-354.
Fritema de Vries, F.T. 1993. A study of the real changes for the gene flow from cultivated plants to the wild
flora of the Netherlands. In: Gene transfer: are wild species in danger?, ed by J. Yolande, K. Ammann
& F. Pythoud, 11-13. Bern, Federal Office of Environment, Forests and Landscape.
Fritts, S.H., Bangs, E.E. & Gore, J.F. 1994. The relationship of wolf recovery to habitat conservation and
biodiversity in the northwestern United States. Landscape Urban Planning, 28: 23-32.
Fritts, S.H., Wiliam, J.P. & Mech, L.D. 1984. Movements of translocated wolves in Minnesota. Journal of
Wildlife Management, 48: 709-721.
Fromanowicz, D.R. & Brodie, E.D. 1985. Unpalatability and toxicity of an introduced species (cinnabar moth
larvae) to native predators. American Midland Naturalist, 113: 401-403.
Fry, J.C. & Day, M.J. 1992. Release of genetically engineered and other microorganisms. Cambridge,
Cambridge University Press.
Fuester, R.W., Taylor, P.B., Day, W.H., Hendrickson, R.M., Jr. & Blumenthal, E.M. 1984. Introduction of
exotic parasites for biological control of the birch leafminer (Hymenoptera: Tenthredinidae) in the
Middle Atlantic states. Journal of Economic Entomology, 77: 1565-1570.
Fyfe, R.W. 1978. Reintroducing endangered birds to the wild. In: Endangered birds: Management techniques
for preserving threatened species, ed by S.A. Temple, 323-329. London: Croom Helm.
Galasun, P.T., Andryushchenko, A.I. & Grusevich, V.V. 1984. Biological principles of introducing new species
for aquaculture(Ictalurus punctatus and Ictalurus cyprinellus) into Ukranian waters. Aquaculture, 42:
Galbraith, C.A., Grice, P.V., Mudge, G.P., Parr, S. & Pienkowski, M.W. 1994. The role of the statutory bodies
in ornithological conservation within the U.K. Ibis, 137, S224-S231.
Game Conservancy Advisory Service. 1992. Wild Partridge Management. Fordingbridge, The Game
Conservancy Trust.
Garrott, R.A. & Siniff, D.B. 1992. Limitations of male-oriented contraception for controlling feral horse
populations. Journal of Wildlife Management, 56: 456-464.
Garson, P.J., Young, L. & Kaul, R. 1992. Ecology and conservation of the cheer pheasant Catreus wallichii :
Studies in the wild and the progress of a reintroduction project. Biological Conservation, 59: 25-35.
Gaston, A.J. 1994. Status of the ancient murrelet, Synthiboramphus antiquus, in Canada and the effects of
introduced predators. Canadian Field Naturalist, 108: 211-222.
Geddes, M.C., Musgrove, R.J. & Campbell, N.J.H. 1993. The feasibility of re-establishing the River Murray
crayfish, Euastacus armatus, in the lower River Murray. In: Freshwater Crayfish IX, ed by D.M.
Holdich & G.F. Warner, 368-379. Lafayette, USA, University of Southwestern Louisiana.
Gerlach, J. 1993. Invasive melastomataceae in Seychelles. Oryx, 27: 22-26.
Gerrish, G. & Mueller-Dombois, D. 1980. Behavior of native and non-native plants in two tropical rain forests
on Oahu, Hawaiian Islands. Phytocoenologia, 8: 237-295.
Review of information, policy and legislation on species translocations
Gibson, C.W.D., Watt, T.A. & Brown, V.K. 1987. The use of sheep grazing to recreate species-rich grassland
from abandoned arable land. Biological Conservation, 42: 1-19.
Gibson, M. 1995. Introduction of exotic fish. Species conservation handbook. Peterborough, English Nature.
Gilbert, G.S., Parke, J.L., Clayton, M.K. & Handelsman, J. 1993. Effects of an introduced bacterium on
bacterial communities in roots. Ecology, 74: 840-854.
Gilbert, O. 1994. Breckland Lichen Transplant Project. Third Resurvey 5/7/1994. Peterborough, English Nature.
(Species Recovery Programme Report.)
Gilfedder, L. & Kirkpatrick, J.B. 1993. Germinable soil seed and competitive relationships between a rare native
species and exotics in a seminatural pasture in the midlands, Tasmania. Biological Conservation, 64:
Gillham, D.A. 1977. Restoring moorland disturbed by pipeline installation. Landscape. Design., 119: 34-36.
Gilpin, M. 1990. Biological invasions - a global perspective - Drake, JA. Science, 248: 88-89.
Gilpin, m & Hanski, I., eds. 1991. Metapopulation dynamics: empirical and theoretical investigations. London,
Academic Press.
Ginn, H. 1983. The ecology and conservation of amphibian and reptile species endangered in Britain.
Peterborough, Nature Conservancy Council.
Ginsberg, J. 1995. Disposal of confiscated animals: developing guidelines for the placement of confiscated
animals. Re-Introduction News, 10: 4-5.
Gliddon, C. 1994. The impact of hybrids between genetically-modified crop plants and their related species biological models and theoretical perspectives. Molecular Ecology, 3: 41-44.
Glova, G.J. & Sagar, P.M. 1991. Dietry and spatial overlap between stream populations of a native and 2
introduced fish species in New Zealand. Australian Journal of Marine and Freshwater Research, 42:
Godfray, H.C.J. 1995. Field experiments with genetically manipulated insect viruses: ecological issues. Trends
in Ecololgy and Evolution, 10: 465-469.
Gogan, P.J.P. 1990. Considerations in the reintroduction of native mammalian species to restore natural
ecosystems. Natural Areas Journal , 10: 210-217.
Goldsmith, F.B., ed. 1991. Monitoring for conservation and ecology. London, Chapman and Hall.
Good, J.E.G., Stevens, P.A., Davis, B.N.K., Frost, A.J. & Plant, R.A. 1992. Selar farm grassland reestablishment project. Bangor, Institute of Terrestrial Ecology.
Good, J.E.G., Stevens, P.A., Wallace, H.L., Radford, G.L., Davis, B.N.K. & Frost, A.J. 1994. Selar farm
grassland re-establishment project. Final Report, December 1994. Bangor, Institute of Terrestrial
Goodman, D. 1993. The demography of chance extinction. In: Viable populations for conservation, ed by M.
Soule, 4th ed., 11-34. Cambridge, Cambridge University Press.
Goodyear, N.C. 1992. Spatial overlap and dietry selection of native rice rats and exotic black rats. Journal of
Mammology, 73: 186-200.
Goodyear, N.C. & Lazell, J. 1994. Status of a relocated population of endangered Iguana pinguis on Guana
Island, British Virgin Islands. Restoration Ecology, 2: 43-50.
Gordon, A.J. & Kluge, R.L. 1991. Biological control of St-Johns wort, Hypericum perforatum (Clusiaceae), in
South Africa. Agriculture, Ecosystems and Environment, 37: 77-90.
Gordon, D.R. 1994. Translocation of species into conservation areas. Natural Areas Journal, 14: 31-37.
Gosliner, T.M. 1995. Introduction and spread of Philine auriformis (Gastropoda: Opisthobranchia) from New
Zealand to San Francisco Bay and Bodega Harbor. Marine Biology, 122: 249-255.
Gosling, L.M. & Baker, S.J. 1989. The eradication of muskrats and coypus from Britain. Biological Journal of
the Linnean Society, 38: 39-51.
Gosling, L.M., Watt, A.D. & Baker, S.J. 1981. Continuous retrospective census of the East Anglian coypu
population between 1970 and 1979. Journal of Animal Ecology, 50: 885-901.
Gowdy, J.M. 1993. Economic and biological aspects of genetic diversity. Society and Natural Resources, 6: 116.
Graczyk, R. 1979. Reintroduction of the beaver (Castor fiber) in the Notec-Warta forests. Sylwan, 123: 53-64.
Granda, M.M. & Fuentes, V.R. 1986. Introduction and culture of exotic medicinal plants in Cuba. Newsletter,
Medicinal and Aromatic Plants, 2: 36-41.
Grant, P.R. & Grant, B.R. 1994. Phenotypic and genetic-effects of hybridization in Darwins finches. Evolution,
48: 297-316.
Gray, A.J. 1989. Release of captive bred species: non-technical review. In: Release of captive bred species:
genetic interactions with wild relatives, ed by A.J. Gray, 2-13, Unpublished ITE Report to the
Department of the Environment.
Review of information, policy and legislation on species translocations
Gray, A.J. 1989. Release of captive bred species: some general genetic and theoretical constraints. In: Release of
captive bred species: genetic interactions with wild relatives, ed by A.J. Gray, 14-32, Unpublished ITE
Report to the Department of the Environment.
Gray, A. 1995. The genetic basis of conservation biology. In: Conservation Biology, ed by I.F. Spellerberg, 107121. London, Longman.
Green, B.H. 1981. A policy on introductions to Britain. In: The biological aspects of rare plant conservation, ed
by H. Sygne, 403-412. London, Wiley & Sons.
Greig, J.C. 1979. Principals of genetic conservation in relation to wildlife management in Southern Africa.
Southern African Journal of Wildlife Research, 9: 57-78.
Grice, P. 1994. England's freshwater fish in context. In: Species conservation handbook.: Peterboriough, English
Griffith, B., Scott, J.M., Carpenter, J.W. & Reed, C. 1989. Translocation as a species conservation tool: status
and strategy. Science, 245: 477-480.
Griffith, B., Scott, J.M., Carpenter, J.W. & Reed, C. 1993. Animal translocations and potential disease
transmission. Journal of Zoo and Wildlife Medicine, 24: 231-236.
Griffiths, O., Cook, A. & Wells, S.M. 1993. The diet of the introduced carniverous Snaileuglandina rosea in
Mauritius and its implications for threatened island gastropod faunas. Journal of Zoology, 229: 79-89.
Grisham, J. 1994. Attack of the fire ant. Bioscience, 44: 587-591.
Gross, P. 1991. Influence of target pest feeding niche on success rates in classical biological control.
Environmental Entomology, 20: 1217-1227.
Groves, R.H. 1986. Plant invasions of Australia: An overview. In: Ecology of biological invasions, ed by R.H.
Groves & J.J. Burdon, 137-149. Cambridge, Cambridge University Press.
Groves, R.H. & Burdon, J.J. 1986. Ecology of biological invasions. Cambridge, Cambridge University Press.
Grumbine, E. 1990. Protecting biological diversity through the greater ecosystem concept. Natural Areas
Journal, 10: 114-120.
Gueirosfilho, F.J. & Beverley, S.M. 1994. On the introduction of genetically-modified leishmania outside the
laboratory. Experimental Parasitology, 78: 425-428.
Guerrant, E.O., Jr. 1992. Genetic and demographic considerations in the sampling and reintroduction of rare
plants. In: Conservation Biology The Theory And Practice Of Nature Conservation, Preservation, And
Management, ed by P.L. Fiedler & S.K. Jain, 321-344. New York, Chapman and Hall.
Guinon, M. & Allen, D. 1990. Restoration of dune habitat at Spanish Bay. In: Environmental restoration.
Science and strategies for restoring the earth, ed by J.J. Berger, 70-80. Washington, Island Press.
Gustafsson, K. & Jansson, J.K. 1993. Ecological risk assessment of the deliberate release of geneticallymodified microorganisms. Ambio, 22: 236-242.
Hadfield, M.G., Miller, S.E. & Carwile, A.H. 1993. The decimation of endemic Hawaiian tree snails by alien
predators. American Zoologist, 33: 610-622.
Haig, S.M., Ballou, J.D. & Derrickson, S.R. 1990. Management options for preserving genetic diversity:
Reintroduction of Guam rails to the wild. Conservation Biology, 4: 290-300.
Haigh, N. 1991. Manual of environmental policy the EC and Britain. London, Longman.
Hails, R.S. & Crawley, M.J. 1991. The population dynamics of an alien insect - Andicus quercuscalicis
(Hymenoptera, Cynipidae). Journal of animal Ecology, 60: 545-562.
Halley, D. 1995. The proposed reintroduction of the beaver to Britain-in reply. Re-Introduction News, 10: 17-18.
Halliday, T. 1985. Immigrant killers - introduced predators and the conservation of birds in New Zealand.
Nature, 317: 119.
Hall-Martin, A.J. & Penzhorn, B.L. 1977. Behaviour and recruitment of translocated black rhinoceros Diceros
bicornos. Koedoe., 20: 147-162.
Hamann, O. 1993. On vegetation recovery, goats and giant tortoises on Pinta Island, Galapagos, Ecuador.
Biodiversity and Conservation, 2: 138-151.
Hambler, C. 1994. Giant Tortoise Geochelone gigantia translocation to Curieuse Island (Seychelles): Success or
failure? Biological Conservation, 69: 293-299.
Hammar, J., Dempson, J.B. & Verspoor, E. 1991. Natural hybridization between Arctic char (Salvelinus
alpinus) and brook trout(S fontinalis) - evidence from northen Labrador. Canadian Journal of
Fisheries and Aquatic Science, 48: 1437-1445.
Handel, S.N., Robinson, G.R. & Beattie, A.J. 1994. Biodiversity resources for restoration ecology. Restoration
Ecology, 2: 230-241.
Hanna, L. 1992. The possible impacts of releasing captive bred barn owls in Britain. JNCC Report, No. 124.
Harding, K. & Harris, P.S., eds. 1994. Risk assessment of the release of genetically modified plants: a review.
London, Ministry of Agriculture, Fisheries and Food.
Review of information, policy and legislation on species translocations
Harding, P.T. 1990. Biological recording of changes in British wildlife. London, HMSO.
Harper, D.M., Mavuti, K.M. & Muchiri, S.M. 1990. Ecology and management of Lake Naivasha, Kenya< in
relation to climate change, alien species introductions, and agriculture-development. Environmental
Conservation, 17: 328-336.
Harris, G.S. 1978. Salmon Propagation in England and Wales. London, National Water Council.
Harris, R.J., Thomas, C.D. & Moller, H. 1991. The influence of habitat use and foraging on the replacement of
one introduced wasp species by another in New Zealand. Ecological Entomology, 16: 441-448.
Hart, R.C. 1983. Some considerations on the introduction of exotic species. Journal of the Limnological Society
of Southern Africa, 9: 49-51.
Hartman, G. 1994. Long-term population development of a reintroduced beaver (Castor fiber) population in
Sweden. Conservation Biology, 8: 713-717.
Hasselrot, B., Andersson, I. & Hultberg, H. 1984. Ecosystem shifts and reintroduction of Arctic char (Salvelinus
salvelinus (L.)) after liming of a strongly acidified lake in southwestern Sweden. Report of the Institute
of Freshwater Resources, Drottningholm, 78-92.
Heatwole, H. & Walker, T.A. 1989. Dispersal of alien plants to coral cays. Ecology., 70: 787-790.
Hedrick, P.W., Brussard, P.F., Allendorf, F.W., Beardmore, J.A. & Orzack, S. 1986. Protein variation, fitness,
and captive propagation. In: Proceedings of the workshop on genetic management of captive
populations, ed by K. Ralls & J.D. Ballou, 91-99.
Heidemann, G. 1976. Problems in reintroduction of the otter Lutra lutra. In: Reintroductions: techniques and
ethics, ed by L. Boitani, 67-69. Rome, World Wildlife Fund.
Helliwell, D.R. 1989. Soil transfer as a method of moving grassland and marshland vegetation. In: Biological
habitat reconstruction, ed by G.P. Buckley, 258-263. London, Belhaven.
Helliwell, R. 1990. Moving semi-natural vegetation. Landscape Design, 186: 36-38.
Henderson, L. 1989. Invasive alien woody plants of Natal and the north-eastern Orange Free State. Bothalia, 19:
Henderson, L. & Wells, M.J. 1986. Alien plant invasions in the grassland and savanna biomes. Cape Town,
Oxford University Press.
Hengeveld, R. 1994. Biodiversity - the diversification of life in a non-equilibrium world. Biodiversity Letters, 2:
Herbold, B. & Moyle, P.B. 1986. Introduced species and vacant niches. American Naturalist, 128: 751-760.
Herpetofauna International Ltd. 1991. Proposed guidelines for the translocation of crested newts Triturus
cristatus at "wild" sites. Herpetofauna News, 2: 5-6.
Hert, E. 1990. Factors in habitat partitioning in Pseudotropheus aurora (Pisces, Cichlidae), an introduced
species to a species-rich community of Lake Malawi. Journal of Fish Biology, 36: 853-865.
Hewitt, C.L. 1991. Native community susceptibility to invasion by introduced species. American Zoologist, 31,
Heywood, V.H. & Wyse Jackson, P.S. 1987. Botanic gardens in conservation, reintroduction into the wild.
Tropical Botanic Gardens, 209-216.
Hill, M.O., Wright, S.M., Dring, J.C., Firbank, L.G., Manchester, S.J. & Croft, J.M. 1994. The potential for
spread of alien species in England following climatic change. Final report. Peterborough, English
Nature. (English Nature Research Report no.90.)
Hindar, K., Ryman, N. & Utter, F. 1991. Genetic effects of cultured fish on natural fish populations. Canadian
Journal of Fisheries and Aquatic Science, 48: 945-957.
Hirsch, P.R. & Spokes, J.D. 1994. Survival and dispersion of genetically-modified rhizobia in the field and
genetic interactions with native strains. Fems Microbiology Ecology, 15: 147-159.
Hobbs, I.I.I., HH, Jass, J.P. & Huner, J.V. 1989. A review of global crayfish introductions with particular
emphasis on two North American species (Decapoda, Cambaridae). Crustaceana, 56: 299-316.
Hodgson, J.G. 1989. Selecting and managing plant materials used in habitat constuction. In: Biological habitat
reconstruction, ed by G.P. Buckley, 45-67. London, Belhaven Press.
Hoffman, C. 1990. Ecological risks of genetic engineering of crop plants. Bioscience, 40: 434-437.
Hoffman, J.H. 1991. Special issue - biological-control of weeds in South Africa - Introduction. Agriculture,
Ecosystems and Environment, 37: 1-3.
Hoffmann, J.H. & Moran, V.C. 1988. The invasive weed Sesbania punicea in South Africa and prospects for its
biological control. South African Journal of Science, 84: 740-742.
Hoffman, J.H. & Moran, V.C. 1995. Localized failure of a weed biological control agent attributed to insecticide
drift. Agriculture, Ecosystems and Environment, 52: 197-203.
Holcik, J. 1991. Fish introductions in Europe with particular reference to its central and eastern part. Canadian
Journal of Fisheries and Aquatic Science, 48: 13-23.
Review of information, policy and legislation on species translocations
Holditch, D.M. & Reeve, I.D. 1991. Distribution of freshwater crayfish in the British Isles, with particular
reference to crayfish plague, alien introductions and water quality. Aquatic conservation - marine and
freshwater ecosystems, 1: 139-158.
Holmes, J.S. & Simons, J.R. 1996. The Wildlife and Countryside Act and bird introductions. In: Feral and
introduced birds. A British Ornithologists' Union & Joint Nature Conservation Committee Conference,
1995. Peterborough, Joint Nature Conservation Committee.
Holmes, P.M. 1988. Implications of alien Acacia seed bank viability and germination for clearing. South African
Journal of Botany, 54: 281-284.
Holmes, P.M., Dennill, G.B. & Moll, E.J. 1987. Effects of feeding by native alydid insects on the seed viability
of an alien invasive weed, Acacia cyclops. South African Journal of Science, 83: 580-581.
Hone, J. 1992. Modelling of poisoning for vertebrate pest control, with emphasis on poisoning feral pigs.
Ecological Modelling, 62: 311-327.
Hone, J. 1995. Spatial and temporal aspects of vertebrate pest damage with emphasis on feral pigs. Journal of
Applied Ecology, 32: 311-319.
Honig, M.A., Cowling, R.M. & Richardson, D.M. 1992. The invasive potential of Australian banksias in South
African fynbos: A comparison of the reproductive potential of Banksia ericifolia and Leucadendron
laureolum. Australian Journal of Ecology, 17: 305-314.
Hoover, J.P., Root, C.R. & Zimmer, M.A. 1984. Clinical evaluation of American river otters in a reintroduction
study. Journal of the American Veterinary Medicine Association, 185: 1321-1326.
Hope-Simpson, J.F. & Pring, M.E. 1954. Experimental sowing and transplants of rare species near Bristol.
Proceedings of the Botanical Society of the British Isles, 1: 562-563.
Hopkins, J.J. 1989. Prospects for habitat creation. Landscape Design, 189: 19-23.
Hopper, K.R., Roush, R.T. & Powell, W. 1993. Management of genetics of biological-control introductions.
Annual Review of Entomology, 38: 27-51.
Horton, P.J. & Branscombe, J. 1994. Case study: Great Crested Newt project. In: Conservation and
management of great crested newts, ed by T. Gent & R. Bray, 104-110. Peterborough: English Nature.
Houston, D.B. & Schreiner, E.G. 1995. Alien species in national parks: Drawing lines in space and time.
Conservation Biology, 9: 204-209.
Howells, R.G. & Garrett, G.P. 1992. Status of some exotic sport fishes in texas waters. Texas Journal of
Science, 44: 317-324.
Hudson, P.J. & Rands, M.R.W. 1986. Ecology and management of gamebirds. Edinburgh, BSP Professional
Huenneke, L.F. & Thomson, J.K. 1995. Potential interference between a threatened endemic thistle and an
invasive nonnative plant. Conservation Biology, 9: 416-425.
Huenneke, L.F. & Vitousek, P.M. 1990. Seedling and clonal recruitment of the invasive tree Psidium
cattleianum : Implications for management of native Hawaiian forests. Biological Conservation, 53:
Huffaker, C.B. 1986. Introduction to symposium on biological control. Agriculture, Ecosystems and
Environment, 15: 85-93.
Hughes, R.D., Hughes, M.A., Aeschlimann, J.P., Woolcock, L.T. & Carver, M. 1994. An attempt to anticipate
biological-control ofDiuraphis noxia (Hom, Aphididae). Entomophaga., 39: 211-223.
Humphries, C., Vane-Wright, D. & Williams, P. 1991. Biodiversity reserves: Setting new priorities for the
conservation of wildlife. Parks, 2: 34-37.
Humphries, R.N., Horton, P.J. & Benyon, P.R. 1995. Transplantation of grasslands: I The importance onf
traditional management practices. In: Decades later: a time for reassessment, Proceedings of the 12th
Annual National Meeting of the American Society for Surface Mining and Reclamation, ed by G.E.
Schuman & G.F. Vance, 186-193. Princeton, American Society for Surface Mining and Reclamation.
Hunter, G.G. & Douglas, M.H. 1984. Spread of exotic conifers on South Island rangelands. New Zealand
Journal of Forestry, 29: 78-96.
Hutchings, J.A. 1991. The threat of extinction to native populations experiencing spawning intrusions by
cultured Atlantic salmon. Aquaculture, 98: 119-132.
Hutchings, P. 1992. Ballast water introductions of exotic marine organisms into Australia - current status and
management options. Marine Pollution Bulletin, 25: 196-199.
Hutchison, M.J. 1989. The invasion of south-western Australia by the orange palmdart (Cephrenes augiades
sperthias (Felder), Lepidoptera, Hesperiidae) and its positive effect on species richness. Journal of
Biogeography, 16: 131-139.
Hutchison, M.J. & Armstrong, P.H. 1993. The invasion of a south-western Australian river system by Perca
fluviatilis - history and probable causes. Global Ecology and Biogeography Letters, 3: 77-89.
Review of information, policy and legislation on species translocations
Huttner, S.L., Arntzen, C., Beachy, R., Breuning, G., Nestor, E., Qualset, C. & Vidaver, A. 1992. Revising
oversight of genetically modified plants. Bio Technology, 10: 967-971.
International Council for the Exploration of the Sea. 1988. Codes of practice and manual of procedures for
consideration of introductions and transfers of marine and freshwater organisms.
International Maritime Organisation. 1993. Guidelines for preventing the introduction of unwanted aquatic
organisms and pathogens from ships' ballast water and sediment discharges. Resolution A.774(18).
International Maritime Organisation. 1994. Unwanted organisms in ballast water. Report of the ballast water
working group MEPC 36.
IUCN. 1987. The IUCN position statement on translocation of living organisms. Gland, Switzerland, IUCN.
IUCN 1994. IUCN red list categories. Gland, Switzerland, IUCN Species Survival Commission.
IUCN. 1995. Guidelines for reintroductions. Gland, Switzerland, IUCN/SSC Reintroduction specialist group.
Ivbijaro, M.F., Udensis, N., Ukwela, U.M. & Anno-Nyako, F.V. 1992. Geographical distribution and host range
in Nigeria of the mango mealy bug, Rastrococcus invadens Williams, a serious exotic pest of
horticulture and other crops. Insecta Scientia Applicata, 13: 411-416.
Jackson, A. 1994. Species recovery - combining political, management and practical skill to conserve rare
species. International Dendrological Society Conference on Conservation of Temperate Trees.
Jackson, A. 1995. The Plymouth pear - the recovery programme for Britain's rarest wild tree. British Wildlife, 6:
Jacobson, E.R. 1993. Implications of Infectious diseases for captive propagation and introduction programs of
threatened and endangered reptiles. Journal of Zoo and Wildlife Medicine, 24: 245-255.
Jacot, Y., Ammann, K. & Pythoud, F., eds. 1993. Gene transfer:are wild species in danger? Bern, Federal
Office of Environment, Forests and Landscape.
Jaksic, F.M. & Yanez, J.L. 1983. Rabbit and fox introductions in Tierra del Fuego - history and attempts at
biological-control of the rabbit infestation. Biological Conservation, 26: 367-374.
James, R.R., McEvoy, P.B. & Cox, C.S. 1992. Combining the cinnabar moth (Tyria jacobaeae) and the ragwort
flea beetle (Longitarsus jacbaeae) for control of ragwort (Senecio jacobaeae) - an experimental
analysis. Journal of Applied Ecology, 29: 589-596.
Jansson, R.K. 1993. Introduction of exotic entomopathogenic nematodes (Rhabditida: Heterorhabditidae and
Steinernematidae) for biological control of insects: Potential and problems. Florida Entomology, 76:
Jarvis, P.J. 1990. Urban cats as pests and pets. Environmental Conservation, 17: 169-171.
Jefferies, D.J. & Mitchell-Jones, A.J. 1993. Recovery plans for British mammals of conservation importance,
their design and value. Mammal Review, 23: 155-166.
Jeffries, D.J., Wayre, P., Jessop, R.M. & Mitchell-Jones, A.J. 1986. Reinforcing the native otter Lutra lutra
population in East Anglia: an analysis of the behaviour and range development of the first release
group. Mammal Review, 16: 65-79.
Jepson, P.C., Croft, B.C. & Pratt, G.E. 1994. Test systems to determine the ecological risks posed by toxin
release from Bacillus thuringiensis genes in crop plants. Molecular Ecology, 3: 81-89.
Jerram, R. 1992. Five years of monitoring grassland transplantation at Potatopot, West Cumbria. Peterborough,
English Nature.
Jessop, R. 1991. The status of Otters in East anglia with particular reference to re-establishment. In: The wildlife
trusts partnership, ed by Norfolk Naturalists Trust. Norwich: Norfolk Naturalists Trust.
Jhala, Y.V. & Giles, R.H., Jr. 1991. The status and conservation of the wolf in Gujarat and Rajasthan, India.
Conservation Biology, 5: 476-483.
Johns, D.M. 1993. Landscape-scale restoration: The Wildlands Project. Restoration Management Notes, 11: 1819.
Johnson, H.B. & Mayeux, H.S. 1992. Viewpoint - a view on species additions and deletions and the balance of
nature. Journal of Range Management, 45: 322-333.
Joint Committee for the Conservation of British Insects. 1986. Insect re-establishment-code of conservation
practice. Antenna, 10: 13-81.
Joint Nature Conservation Committee. 1994a. The experimental reintroduction of the Red Kite Milvus milvus to
Scotland and England. Vertebrate Note, 6.
Joint Nature Conservation Commitee. 1994b. Conservation of the Barn Owl Tyto alba in the United Kingdom.
Vertebrate Note, 7.
Joint Nature Conservation Commitee. 1994c. Reintroduction Programmes. Vertebrate Note 5.
Jones, C.G., Heck, W., Lewis, R.E., Mungroo, Y., Slade, G. & Cade, T. 1995. The restoration of the Mauritius
Kestrel Falco punctatus population. Ibis, 137, S173-S180.
Review of information, policy and legislation on species translocations
Jordan, W.R., III, Peters, R.L., II & Allen, E.B. 1988. Ecological restoration as a strategy for conserving
biological diversity. Environmental Management, 12: 55-72.
Jorden, W.R., Gilpin, M.E. & Aber, J.D., eds. 1992. Restoration Ecology. 2nd ed. Cambridge, Cambridge
University Press.
Jorgensen, R.B. & Andersen, B. 1994. Spontaneous hybridization between oilseed rape (Brassica napus) and
weedy brassica-campestris (Brassicaceae) - a rise of growing genetically-modified oilseed rape.
American Journal of Botany, 81: 1620-1626.
Jude, D.J., Reider, R.H. & Smith, G.R. 1992. Establishment of Gobiidae in the Great Lakes Basin. Canadian
Journal of Fisheries and Aquatic Science, 49: 416-421.
Julien, M.H. 1992. Biological control of weeds: A world catologue of agents and their target weeds. Australian
Centre for International Agricultural Research.
Julien, M.H., Kerr, J.D. & Chan, R.R. 1984. Biological control of weeds: an evaluation. Protection Ecology, 7:
Juniper, T. 1994. Conservation on the global stage - the Habitats Directive, the Biodiversity Convention and the
UK. British Wildlife, 6: 99-105.
Kapteijns, A.J.A.M. 1993. Risk assessment of genetically-modified crops - potential of 4 arable crops to
hybridize with the wild flora. Euphytica, 66: 145-149.
Kapuscinski, A.R.D. & Lannan, J.E. 1984. Application of a conceptual fitness model for managing Pacific
salmon fisheries. Aquaculture, 43: 135-146.
Kardell, L., Steen, E. & Fabiao, A. 1986. Eucalyptus in Portugal--A threat or a promise? Ambio., 15: 6-13.
Kareiva, P., Morris, W. & Jacobi, C.M. 1994. Studying and managing the risk of cross-fertilisation between
transgenic crops and wild relatives. Molecular Ecology, 3: 15-23.
Karesh, W. 1993. Cost-evaluation of infectious disease monitoring and screening programs for wildlife
translocation and reintroduction. Journal of Zoo and Wildlife Medicine, 24: 291-295.
Kay, Q. 1993. Genetic differences between populations of rare plants. BSBI News, 64: 54-56.
Keddy, P.A. 1983. Transplanting rare plants to protect them. Canadian Botanical Association Bulletin, 16: 1315.
Kearns, S. 1984. A comparison of transplanting times and methods for salvaging prairie forbs and grasses.
Proceedings of the ninth North American Prairie Conference, 197-200.
Kenward, R.E. 1974. Mortality and fate of trained birds of prey. Journal of Wildlife Management, 38: 751-756.
Kenward, R.E. 1979. Winter predation by goshawks in lowland Britain. Bird Study, 72: 64-73.
Kenward, R.E. 1981. What happens to goshawks trained for falconry. Journal of Wildlife Management, 45: 802806.
Kenward, R.E. 1983. Bark stripping by grey squirrels in Britain and North America: why does the damage
differ? In: Mammals as pests, ed by R.J. Putman, 145-154. London, Chapman and Hall.
Kenward, R.E. 1989. Release of captive bred species: raptors, rodents and lagomorphs. In: Release of captive
bred species: genetic interactions with wild relatives, ed by A.J. Gray, 71-78, Unpublished ITE Report
to the Department of the Environment..
Kenward, R.E. 1995. The use of radio-tagging to monitor released animals. In: Restoration of Houbara Bustard
populations in Saudi Arabia, ed by P. Gaucher and Y. van Heezik. Saudi Arabia, National
Commission for Wildlife Conservation and Development.
Kenward, R.E. & Holm, J.L. 1989. What future for British red squirrels? Biological Journal of the Linnean
Society, 38: 83-89.
Kenward, R.E. & Holm, J.L. 1993. On the replacement of the red squirrel in Britain - a phytotoxic explanation.
Proceedings of the Royal Society of London Series B-Biological Sciences, 251: 187-194.
Kenward, R.E., Parish, T. & Robertson, P.A. 1992. Are tree species mixtures too good for grey squirrels? In:
The ecology of mixed-species stands of trees, ed by M.G.R. Cannell, R. Malcolm & P.A. Robertson.
British Ecological Society.
Keward, R.E., Hodder, K.H. & Rose, R.J. 1995. Conservation of red squirrels. Final Report. Wareham, Institute
of Terrestrial Ecology.
Kile, G.A. 1989. Infection of exotic and Tasmanian native tree and shrub species by the vascular stain fungus
Chalara australis. European Journal of Forest Pathology, 19: 98-104.
Kimmerer, W.J., Gartside, E. & Orsi, J.J. 1994. Predation by an introduction clam as the likely cause of
substantial declines in zooplankton of San Francisco Bay. Marine Ecology-Progress Series, 113: 8193.
King, M.L. 1976. The reintroduction of the swallowtail butterfly to Wicken Fen. Journal of the Cambridge
Association for Environmental Education, 1-3.
Review of information, policy and legislation on species translocations
Kingsbury, D.T. 1988. Regulation of biotechnology: a perspective on the US coorinated framework. In:
Introduction of genetically modified organisims into the environment, ed by H.A. Mooney & G.
Bernardi, Vol. SCOPE 44: 1st ed., 162-189. Chichester, John Wiley & Sons.
Kleiman, D.G. 1989. Reintroduction of captive mammals for conservation. Bioscience, 39: 152-161.
Klemens, M.W. 1995. Repatriation of confiscated tortoises:conscience-clearing expediency or sound wildlife
management. Re-Introduction News, 10: 5-6.
Klingman, D.L. & Coulson, J.R. 1982. Guidelines for Introducing Foreign Organisms into the United States for
Biological Control of Weeds. Weed Science, 30: 661-667.
Kloot, P.M. 1987. The invasion of Kangaroo Island by alien plants. Australian Journal of Ecology, 12: 263-266.
Kluge, R.L., Zimmermann, H.G., Cilliers, C.J. & Harding, G.B. 1986. Integrated control for invasive alien
weeds. In: The ecology and management of biological invasions in Southern Africa, ed by I.A.W.
MacDonald, F.J. Kruger & A.A. Ferrar, 295-303. Cape Town, Oxford University Press.
Knight, R.S. 1986. A comparative analysis of fleshy fruit displays in alien and indigenous plants. In: The
ecology and management of biological invasions in Southern Africa, ed by I.A.W. MacDonald, F.J.
Kruger & A.A. Ferrar, 171-178. Cape Town, Oxford University Press.
Kohler, C.C. & Courtenay, W.R. 1986a. American Fisheries Society position on introductions of aquatic
species. Fisheries, 11: 39-42.
Kohler, C.C. & Courtenay, W.R., Jr. 1986b. Regulating introduced aquatic species: A review of past initiatives.
Fisheries, 11: 34-38.
Kooijman, A.M., Beltman, B. & Westhoff, V. 1994. Extinction and reintroduction of the bryophyte Scorpidium
scorpioides in a rich-fen spring site in the Netherlands. Biological Conservation, 69: 87-96.
Kornberg, H. & Wiliamson, M.H., eds. 1987. Quantitative Aspects of the Ecology of Biological Invasions,
Proceedings of a Royal Society Discussion Meeting. London, The Royal Society.
Kozhova, O.M. & Izhboldina, L.A. 1993. Spread of Elodea canadensis in Lake Baikal. Hydrobiologia, 259:
Krajewski, C. 1994. Phylogenetic measures of biodiversity: A comparison and critique. Biological
Conservation, 69: 33-39.
Kraus, M. 1995. EU directive on GMOs exonerated. Biofutur, 32-34.
Kruger, F.J., Richardson, D.M. & van-Wilgen, B.W. 1986. Processes of invasion by alien plants.In: The ecology
and management of biological invasions in Southern Africa, ed by I.A.W. MacDonald, F.J. Kruger &
A.A. Ferrar, 145-155. Cape Town, Oxford University Press.
Krushinska, N.L., Rychlik, L. & Pucek, Z. 1994. Agonitic interactions between resident and immigrant
sympatric water shrews - Neomys fodiens and N anomalus. Acta Theriologica, 39: 227-247.
Kruuk, H. 1989. Release of captive bred species: carnivores. In: Release of captive bred species: genetic
interactions with wild relatives, ed by A.J. Gray, 83-87, Unpublished ITE Report to the Department of
the Environment..
Kruuk, H. & Conroy, J.W.H. 1987. Surveying otter Lutra lutra populations: a discussion of problems with
spraints. Biological Conservation, 41: 179-183.
Kurir, A. 1981. The introduction of exotic forest and timber insects of economic importance and their
integration into the central European fauna during the last three decades. Acta Entomologica
Jugoslavica, 17: 51-54.
Kurzejeskl, E.W. & Root, B.G. 1988. Survival of reintroduced ruffed grouse in north Missouri. Journal of
Wildlife Management, 52: 248-252.
Lacy, R.C. 1993. Vortex: a computer simulation model for population viability analysis. Wildlife Research, 20:
Lai, P.Y. & Funasaki, G.Y. 1985. Introductions for biological control in Hawaii, 1981 and 1982. Proceedings of
the Hawaiian Entomological Society, 25: 83-86.
Laird, L.M. & Needham, E.A. 1985. Salmon farming and the future of the Atlantic salmon. In: The status of the
Atlantic salmon in Scotland. ITE Symposium no. 15, ed by D. Jenkins & W.M. Shearer, 66-72. London,
Lake, P.S.O. & Dowd, D.J. 1991. Red crabs in rain-forest, Christmas Island - biotic resistance to invasion by an
exotic snail. Oikos, 62: 25-29.
Lamarre, E. & Cochran, P.A. 1992. Lack of host species selection by the exotic parasitic crustacean, Argulus
japonicus. Journal of Freshwater Ecology, 7: 77-80.
Lamb, D. 1992. Restoration for nature conservation. IVth world congress on national parks and protected
areas. Gland, Switzerland, IUCN.
Landwer, A.J., Ferguson, G.W., Herber, R. & Brewer, M. 1995. Habitat use of introduced and native angles
(Iguanidae: Anolis) along the northern coast of Jamaica. Texas Journal of Science, 47: 45-52.
Review of information, policy and legislation on species translocations
Langslow, D.R. 1992. Managing changes in British wildlife: the effects of nature conservation. In: Biological
recording of changes in British wildlife, ed by P.T., 65-78. Harding. London, HMSO.
Larsen, J.B. 1995. Ecological stability of forests and sustainable silviculture. Forest Ecology and Management,
73: 85-96.
Latin, J.D., Christie, A. & Schwartz, M.D. 1994. The impact of nonindigenous crested wheatgrasses on native
black grass bugs in North America: A case for ecosystem management. Natural Areas Journal, 14:
Latin, J.D., Christie, A. & Schwartz, M.D. 1995. Native black grass bugs (Irbisia labops) on introduced
wheatgrasses: Commentary and annotated bibliography (Hemiptera: Heteroptera: Miridae).
Proceedings of the Entomological Society of Washington, 97: 90-111.
Lawton & Brown. 1986. The population and community ecology of invading insects. In: Quantative aspects of
the ecology of biological invasions, ed by H. Kornberg & M. Williamson, 105-113. London, The Royal
Leach, S.J., Cox, J.H.S., Blake, C.P., Byrne, S.A. & Porley, R.D. 1992. Progress reports on monitoring of
grassland transplant sites. Brocks Farm 1988-1992. Peterborough, English Nature.
Leach, S.J., Pulteney, C.M., Butcher, M. & McDouall, A. 1995. Progress report on botanical monitoring.
Brocks Farm SSSI, update 1994. Peterborough, English Nature.
Leader-Williams, N., Smith, R.I.L. & Rothery, P. 1987. Influence of introduced reindeer on the vegetation of
South Georgia: results from a long-term exclusion experiment. Journal of Applied Ecology, 24: 801822.
Leatherhead, J.F. 1993. Field observations on reproductive and developmental dysfunction in introduced and
native salmonids from the Great-Lakes. Journal of Great Lakes Research, 19: 737-751.
Leberg, P.L. & Vrijenhoek, R.C. 1994. Variation among Desert Topminnows in Their Susceptibility to Attack
by Exotic Parasites. Conservation Biology, 8: 419-424.
Leclant, F. & Renoust, M. 1986. The Lagerstroemia aphid, a new pest for our flora. Phytoma, 375: 49-50.
Ledieu, M.S. & Bartlett, P.W. 1983. Leaf miners of chrysanthemum. London, Ministry of Agriculture, Fisheries
and Food.
Ledig, F.T. 1992. Human impacts on genetic diversity in forest ecosystems. Oikos, 63: 87-108.
Lee, S.Y. 1993. The management of traditional tidal ponds for aquaculture and wildlife conservation in
southeast Asia - problems and prospects. Biological Conservation, 63: 113-118.
Lees, D. 1989. Practical considerations and techniques in the captive breeding of insects for conservation
purposes. Entomologist, 108: 77-96.
Leppakoski, E. 1984. Introduced species in the Baltic Sea and its coastal ecosystems. Ophelia, S3: 123-135.
Leppakoski, E.J. 1991. Introduced species - resource or threat in brackish water seas - examples from the Baltic
Sea and the Black Sea. Marine Pollution Bulletin, 23: 219-223.
Lesica, P., Ahlenslager, K. & Desanto, J. 1993. New vascular plant records and the increase of exotic plants in
Glacier National Park, Montana. Madrono, 40: 126-131.
Letcher, A.J., Purvis, A., Nee, S. & Harvey, P. 1994. Patterns of overlap in the geographic ranges of palearctic
and British mammals. Journal of animal Ecology, 63: 871-879.
Leveque, C. 1995. Role and consequences of fish diversity in the functioning of African freshwater ecosystems:
A review. Aquatic Living Resources, 8: 59-78.
Lever, C. 1977. The Naturalised Animals of the British Isles. London, Paladin.
Lever, C. 1980. No beavers for Britain. New Scientist, 80: 812-814.
Lever, C. 1984. Conservation success for two Bermudan bird species. Oryx, 8: 138-143.
Lever, C. 1994. The proposed reintroduction of the beaver to Britain. Re-Introduction News, 9: 14-15.
Levin, S. 1988. Ecological issues related to the release of genetically modified organisms into the environment.
In: Introduction of genetically modified organisims into the environment, ed by H.A. Mooney & G.
Bernardi, 151-159. Chichester, John Wiley & Sons.
Li, H.W. & Moyle, P.B. 1981. Ecological Analysis of Species Introductions into Aquatic Systems. Transactions
of the American Fisheries Society, 110: 772-782.
Lieberman, A., Rodriguez, J.V., Paez, J.M. & Wiley, J. 1993. The reintroduction of the Andean condor into
Colombia, South America: 1989-1991. Oryx, 27: 83-90.
Lindeque, M. & Scheepers, J.L. 1992. Use of Datura innoxia by ungulates in the Hoanib River, Namibia. South
African Journal of Wildlife Research, 22: 45-48.
Linder, C.R. & Schmitt, J. 1994. Assessing the risks of transgene escape through time and crop-wild hybrid
persistence. Molecular Ecology, 3: 23-30.
Livdahl, T.P. & Willey, M.S. 1991. Prospects for an invasion - competition between Aedes Albopictus and
native Aedes triseriatus. Science, 253: 189-191.
Review of information, policy and legislation on species translocations
Lloyd, B.D. & Powlesland, R.G. 1994. The decline of Kakapo Strigops habroptilus and attempts at conservation
by translocation. Biological Conservation, 69: 75-85.
Lockwood, J.A. 1993. Environmental issues involved in biological control of rangeland grasshoppers
(Orthoptera: Acrididae) with exotic agents. Environmental Entomology, 22: 503-518.
Lodge, D.M. 1993. Biological invasions: Lessons for ecology. Trends in Ecology and Evolution, 8: 133-137.
Lonsdale, W.M. 1994. Inviting trouble: Introduced pasture species in northern Australia. Australian Journal of
Ecology, 19: 345-354.
Lorence, D.H. & Sussman, R.W. 1986. Exotic species invasion into Mauritius wet forest remnants. Journal of
Tropical Ecology, 2: 147-162.
Lowe, R.A.W. 1994. Principles of collecting plants as a conservation exercise. In: The Common Ground of Wild
and Cultivated plants, ed by A.R. Perry & R. Gwynn Ellis, 89-93. Cardiff, National Museum of Wales.
Lowe, V.P.W. 1983. Is the British squirrel (Sciurrus vulgaris leucourus Kerr) British? Mammal Review, 13: 5767.
Lubbering, J.M., Stuth, J.W., Mungall, E.C. & Sheffield, W.J. 1991. An approach for stategic-planning of
stocking rates for exotic and native ungulates. Applied Animal Behaviour Science, 29: 483-488.
Lubchenco, al. 1991. The Sustainable Biosphere Initiative: an Ecological Research Agenda. Ecology, 72:
Lugo, A.E. 1992. More on exotic species. Conservation Biology, 6: 6.
Luken, J.O. & Goessling, N. 1995. Seedling distribution and potential persistence of the exotic shrub Lonicera
maakii in fragmented forests. American Midland Naturalist, 133: 124-130.
Luken, J.O. & Mattimiro, D.T. 1991. Habitat-specific resilience of the invasive shrub Amur honeysuckle
(Lonicera maackii) during repeated clipping. Ecological Applications, 1: 104-109.
Lunney, D. & Leary, T. 1988. The impact on native mammals of land-use changes and exotic species in the
Bega district, New South Wales, since settlement. Australian Journal of Ecology, 13: 67-92.
Lyles, A. & May, R.M. 1987. Problems in leaving the ark. Nature, 326: 245-246.
Lyster, S. 1985. International Wildlife Law: an analysis of treaties concerned with conservation of wildlife.
Cambridge, Grotius.
Macan, T.T. 1965. The effect of the introduction of Salmo trutta into a moorland fishpond. [Description of a
demonstration.]. Mitteilungen der Internationalen Vereinigung fur theoretische und angewandte
Limnologie, 13: 194-197.
Macdonald, D.W., Brown, E. & Tattersall, F.H. 1995. Re-introducing the beaver to Britain: nostalgic meddling
or restoring biodiversity? Mammal Review, 25, in press.
MacDonald, D.W. & Halliwell, E.C. 1994. The rapid spread of red foxes Vulpes vulpes, on the Isle of man.
Global Ecology and Biogeography Letters, 4: 9-16.
Macdonald, I.A.W. 1983. Alien trees, shrubs and creepers invading indigenous vegetation in the HluhluweUmfolozi Game Reserve Complex in Natal. Bothalia, 14: 949-959.
MacDonald, I.A.W. 1988. Wildlife conservation and the invasion of nature reserves by exotic species: a global
perspective. In: Biological invasions: a global perspective, ed by al. Drake. Chichester, John
Wiley & Sons.
Macdonald, I.A.W., Clark, D.L. & Taylor, H.C. 1989. The history and effects of alien plant control in the Cape
of Good Hope Nature Reserve 1941-1987. South African Journal of Botany, 55: 56-75.
MacDonald, I.A.W. & Frame, G.W. 1988. The invasion of introduced species into nature reserves in tropical
savannas and dry woodlands. Biological Conservation, 44: 67-93.
MacDonald, I.A.W. & Gertenbach, W.P.D. 1988. A list of alien plants in the Kruger National Park. Koedoe, 31:
MacDonald, I.A.W., Graber, D.M., Debenedetti, S., Groves, R.H. & Fuentes, E.R. 1988. Introduced species in
nature reserves in Mediterranean-type climatic regions of the world. Biological Conservation, 44: 3766.
MacDonald, I.A.W., Ortiz, L., Lawesson, J.E. & Nowak, J.B. 1988. The invasion of highlands in Galapagos by
the red quinine-tree Cinchona succirubra. Environmental Conservation, 15: 215-220.
MacDonald, I.A.W., Powrie, F.J. & Siegfried, W.R. 1986. The differential invasion of southern Africa's biomes
and ecosystems by alien plants and animals. In: The ecology and management of biological invasions
in Southern-Africa, ed by I.A.W. MacDonald, F.J. Kruger & A.A. Ferrar, 209-225. Cape Town, Oxford
University Press.
MacDonald, I.A.W., Thebaud, C., Strahm, W.A. & Strasberg, D. 1991. Effects of alien plant invasions on native
vegetation remnants on La Reunion (Mascarene Islands, Indian Ocean). Environmental Conservation,
18: 51-61.
Review of information, policy and legislation on species translocations
MacDonald, I.A.W. & Wissel, C. 1992. Determining optimal clearing treatments for the alien invasive shrub
Acacia saligna in the southwestern Cape, South Africa. Agriculture, Ecosystems, Environment, 39:
Mack, R.N. 1986. Alien plant invasion into the Intermountain West: A case history. In: Ecology of biological
invasions of North America and Hawaii, ed by H.A. Mooney & J.A. Drake, 191-213. New York,
Mack, R.N. 1991. The commercial seed trade - an early dispenser of weeds in the United States. Economic
Botany, 45: 257-273.
Mackie, G.L. 1991. Biology of the exotic zebra mussel, Dreissena-polymorpha, in relation to native bivalves
and its potential impact in Lake St-Clair. Hydrobiologia, 219: 251-268.
MacNab, J. 1983. Wildlife management as scientific experimentation. Wildlife Society Bulletin, 11: 397-401.
Madsen, J.D., Sutherland, J.W., Bloomfield, J.A., Eichler, L.W. & Boylen, C.W. 1991. The decline of native
vegetation under dense Eurasian watermilfoil canopies. Journal of Aquatic Plant Management, 29: 9499.
Magurran, A.E. 1988. Ecological diversity and its measurement. London, Croom Helm.
Maitland, P.S. 1985. The potential impact of fish culture on wild stocks of Atlantic salmon in Scotland. In: The
status of the Atlantic salmon in Scotland. ITE Symposium no. 1, ed by D. Jenkins & W.M. Shearer, 7378. London, HMSO.
Maitland, P.S. 1987. Fish Introductions and translocations-their impact in the British Isles. In: Angling and
Wildlife in Fresh Waters, ITE Symposium No.19, ed by P.S. Maitland & A.K. Turner, 57-65. London,
Maitland, P.S. 1989. The genetic impact of farmed Atlantic salmon on wild populations. Edinburgh, Nature
Conservancy Council.
Maitland, P.S. 1995. The conservation of freshwater fish: Past and present experience. Biological Conservation,
72: 259-270.
Maitland, P.S. & Turner, A.K. 1987. Angling and wildlife conservation-are they compatible? In: Angling and
Wildlife in Fresh Waters, ITE Symposium No.19, ed by P.S. Maitland & A.K. Turner, 76-81. London,
Male, B. 1995. Recovery action for threatened species - an Australian perspective. Ibis, 137, S204-S208.
Malecki, R.A., Blossey, B., Hight, S.D., Schroeder, D., Kok, L.T. & Coulson, J.R. 1993. Biological control of
purple loosestrife. Bioscience, 43: 680-686.
Mansard, P. 1995. European directive restricts movement of animals. Re-Introduction News, 10: 18.
Mardon, D.K. 1993. Reinstatement of rare plants at Ben Lawers NNR. BSBI News, 64: 48-51.
Marion, J.L., Cole, D.N. & Bratton, S.P. 1986. Exotic vegetation in wilderness areas. In: Proceedings National
wilderness research conference:-current research, ed by R.C. Lucas, 114-120.
Marquiss, M. 1981. The Goshawk in Britain-its provenance and current status. In: Understanding the Goshawk,
ed by R.E. Kenward & I.M. Lindsay, 43-57. Oxford, The International Association for Falconry and
Conservation of Birds of Prey.
Marshall, B.E. 1991. The impact of the introduced sardine Limnothrisssa miodon on the ecology of Lake
Kariba. Biological Conservation, 55: 151-165.
Martell, M.S., Tordoff, H.B. & Redig, P.T. 1994. The introduction of three native raptors into the Midwestern
United States. Raptor Conservation Today, 465.
Martin, C.M. 1995. Recovering endangered species and restoring ecosystems: conservation planning for the
twenty-first century in the United states. Ibis, 137, S198-S203.
Martin, M.O. 1995. Biological conservation strategies: Optimizing in situ and ex situ approaches. Trends in
Ecology & Evolution, 10: 227-228.
Mason, C.F. 1991. Decline of the otter Lutra lutra in East Anglia. In: The wildlife trusts partnership, ed by
Norfolk Naturalists Trust, 2-4. Norwich, Norfolk Naturalists Trust.
Maunder, M. 1992. Plant reintroduction: An overview. Biodiversity and Conservation, 1: 51-61.
Maunder, M. & Ramsey, M. 1994. The reintroduction of plants to the wild: an integrated approach to the
conservation of native plants. In: The Common Ground of Wild and Cultivated Plants, ed by A.R. Perry
& R. Gwynn Ellis, 81-88. Cardiff, National Museum of Wales.
May, D.E., Webber, P.J. & May, T.A. 1982. Success of transplanted alpine tundra plants on Niwot Ridge,
Colorado. Journal of Applied Ecology, 19: 965-976.
McClintock, D. 1977. J.E.Lousley and plants alien in the British Isles. Watsonia, 11: 287-290.
McClure, M.S. 1980. Competition between exotic species - scale insects on hemlock. Ecology, 61: 1391-1401.
McGrady, M.J., Orr-Ewing, D.C. & Stowe, T.J. 1994. The reintroduction of the Red Kite Milvus milvus into
Scotland. Raptor Conservation Today, 471-472.
Review of information, policy and legislation on species translocations
McGrorty, S., Goss-Custard, J.D. & Clarke, R.T. 1993. Mussel Mytilus edulis (Mytilacea) dynamics in relation
to environmental gradients and intraspecific interactions. Netherlands Journal of Aquatic Ecology, 27:
McIntyre, S., Ladiges, P.Y. & Adams, G. 1988. Plant species-richness and invasion by exotics in relation to
disturbance of wetland communities on the Riverine Plain, NSW. Australian Journal of Ecology, 13,
McIntyre, S. & Lavorel, S. 1994. Predicting richness of native, rare, and exotic plants in response to habitat and
disturbance variables across a variegated landscape. Conservation Biology, 8: 521-531.
McKillup, S.C., Allen, P.G. & Skewes, M.A. 1988. The natural decline of an introduced species following its
initial increase in abundance - an explanation for Ommatoiulus morelettii in Australia. Oecologia, 77:
McKinnon, K. 1978. Competition between red and grey squirrels. Mammal Review, 8: 185-190.
McKnight, B.N., ed. 1993. Biological pollution: The control and impact of invasive exotic species. Indianapolis,
Indiana Academy of Sciences.
McMahan, L. 1990. Propagation and reintroduction of imperiled plants, and the role of botanical gardens and
arboreta. Endangered Species Update, 8: 4-7.
McMurtry, J.A. 1992. Dynamics and potential impact of generalist phytoseiids in agroecosystems and
possibilities for establishment of exotic species. Experimental & applied acarology, 14: 371-382.
McNeely, J.A. 1992. Parks for life: report of the 4th World Congress on National Parks and protected areas.
Cambridge, IUCN.
McOrist, S. & Kitchener, A.C. 1994. Current threats to the European wildcat, Felis silvestris in Scotland.
Ambio., 23: 243-245.
M.E. Cooper, L.L.B. 1987. An Introduction to Animal Law. London, Academic Press.
Michalakis, Y., Derancourt, F., Noeel, V. & Espiau, C. 1993. A first estimation of the mating system of
Onopordum illyricum (L.) and its relevance to the invasion of Australia. Acta Oecologica, 14: 539-545.
Mikkola, K. & Lafontaine, J.D. 1994. Recent introductions of riparian noctuid moths from the Palaearctic
Region to North America, with the first report of Apamea unanimis (Huebner) (Noctuidae:
Amphipyrinae). J. Lepid. Soc., 48: 121-127.
Milberg, P. & Tyrberg, T. 1993. Naive birds and noble savages - a review of man-caused prehistoric extinctions
of island birds. Ecography, 16: 229-250.
Millar, C.I. & Libby, W.J. 1994. Disneyland or Native Ecosystem: Genetics and the Restorationist. Restoration
& Management Notes, 7: 18-24.
Miller, B.J., Anderson, S.H., DonCarlos, M.W. & Thorne, E.T. 1988. Biology of the endangered black-footed
ferret and the role of captive propagation in its conservation. Canadian Journal of Zoology, 66: 765773.
Miller, B. & Mullette, K.J. 1985. Rehabilitation of an endangered Australian bird: The Lord Howe Island
woodhen Tricholimnas sylvestris (Sclater). Biological Conservation, 34: 55-95.
Miller, B., Reading, R., Conway, C., Jackson, J., Hutchins, M., Snyder, N., Forrest, S. & Frazier, J. 1994. A
model for improving endangered species recovery programs. Environmental Management, 18: 637645.
Miller, H.I., Huttner, S.L. & Beachy, R. 1993. Risk assessment experiments for genetically-modified plants. Bio
Technology, 11: 1323-1324.
Miller, P.S. & Hedrick, P.W. 1993. Inbreeding and fitness in captive populations: Lessons from Drosophila. Zoo
Biology, 12: 333-351.
Miller, R.M. 1993. Nontarget and ecological effects of transgenically altered disease resistance in crops possible effects on the mycorrhizal symbiosis. Molecular Ecology, 2: 327-335.
Mills, C.E. & Sommer, F. 1995. Invertebrate introductions in marine habitats: Two species of hydromedusae
(Cnidaria) native to the Black Sea, Maeotias inexspectata and Blackfordia virginica, invade San
Francisco Bay. Marine Biology, 122: 279-288.
Mills, E.L., Leach, J.H., Carlton, J.T. & Secor, C.L. 1993. Exotic species in the Great Lakes - a history of biotic
crises and anthropogenic introductions. Journal of Great Lakes Research, 19: 1-54.
Mills, E.L., Leach, J.H., Carlton, J.T. & Secor, C.L. 1994. Exotic species and the integrity of the great-lakes lessons from the past. Bioscience, 44: 666-676.
Mills, L.S. & Smouse, P.E. 1994. Demographic consequences of inbreeding in remnant populations. American
Naturalist, 144: 412-431.
Milner, C. 1979. The feral goats of Snowdonia. (Wildlife in Snowdonia 2). Cambridge, Institute of Terrestrial
Review of information, policy and legislation on species translocations
Minckley, W.L. 1995. Translocation as a tool for conserving imperiled fishes: Experiences in western United
States. Biological Conservation, 72: 297-309.
Minello, T.J. & Zimmerman, R.J. 1992. Utilization of natural and transplanted Texas salt marshes by fish and
decapod crustaceans. Marine Ecology Progress Series, 90: 273-285.
Ministry of Agriculture Fisheries and Food. 1993. Application for a licence to introduce into the wild fish or
eggs of fish, shellfish including crustacea not normally resident in Great Britain or which are listed in
Schedule 9 to the Wildlife and Countryside Act 1981. WCAF 2. London, MAFF.
Ministry of Agriculture Fisheries and Food. 1994. A guide to shellfish health controls. London, MAFF.
Minns, C.K., Cairns, V.W., Randall, R.G. & Moore, J.E. 1994. An index of biotic integrity (ibi) for fish
assemblages in the littoral-zone of great-lakes areas of concern. Canadian Journal of Fisheries and
Aquatic Science, 51: 1804-1822.
Mistretta, O. 1994. Genetics of species re-introductions: Applications of genetic analysis. Biodiversity and
Conservation, 3: 184-190.
Mitchell, C.P. 1986. Effects of introduced grass carp on populations of 2 species of small native fishes in a small
lake. New Zealand Journal of Marine and Freshwater Research, 20: 219-230.
Molin, S. & Kjelleberg, S. 1993. Release of engineered microorganisms - biological containment and improved
predictability for risk assessment. Ambio, 22: 242-245.
Moll, E.J. & Trinder-Smith, T. 1992. Invasion and control of alien woody plants on the Cape Peninsula
Mountains, South Africa--30 years on. Biological Conservation, 60: 135-143.
Moller, H., Tilley, J.A.V., Thomas, B.W. & Gaze, P.D. 1991. Effect of introduced social wasps on the standing
crop of honeydew in New Zealand beech forests. New Zealand Journal of Zoology, 18: 171-179.
Mollison, D. 1986. Modelling biological invasions: chance, explantion, prediction. In: Quantative aspects of the
ecology of biological invasion, ed by H. Kornberg & M.H. Williams, 173-189. London, The Royal
Mooney, H.A. & Bernardi, G., eds. 1990. Introduction of genetically modified organisms into the environment.
Chichester, John Wiley & Sons.
Mooney, H.A. & Drake, J.A. 1988. The release of genetically designed organisms in the environment: lessons
from the study of the ecology of biological invasions. In: Introduction of genetically modified
organisims into the environment, ed by H.A. Mooney & G. Bernardi, 117-129. Chichester, John Wiley
& Sons.
Mooney, H.A., Hamburg, S.P. & Drake, J.A. 1986. The invasions of plants and animals into California. In:
Ecology of biological invasions of North-America and Hawaii, ed by H.A. Mooney & J.A. Drake,
250-272. Berlin, Springer-Vertig.
Mooney, H.A. & Risser, P.G. 1989. The release of genetically engineered organisms:A perspective from the
Ecological Society of America. Ecology, 70: 297-298.
Moors, P.J. 1983. Predation by mustelids and rodents on the eggs and chicks of native and introduced birds in
Kowhai Bush, New Zealand. Ibis, 125: 137-154.
Morra, M.J. 1994. Assessing the impact of transgenic plant-products on soil organisms. Molecular Ecology, 3:
Morris, M. 1980. Cotoneaster integerrimus - a conservation exercise. Nature in Wales, 17: 19-22.
Morris, M.G., Thomas, J.A., Ward, L.K., Snazell, R.G., Pywell, R.F., Stevenson, M.J. & Webb, N.R. 1994. Recreation of early-successional stages for threatened butterflies-an ecological engineering approach.
Journal of Environmental Management, 42: 119-135.
Morris, P. 1986. An introduction to introductions. Mammal Review, 16: 49-52.
Morris, P. 1993. A red data book for British mammals. London, Mammal Society.
Morris, P.A., Meakin, K. & Sharafi, S. 1993. The behaviour and survival of rehabilitated hedgehogs. Animal
Welfare, 2: 53-66.
Morris, P.A. & Warwick, H. 1994. A study of rehabilitated juvenile hedgehogs after release into the wild.
Animal Welfare, 3: 163-177.
Morse, M.P. & Thorne, B.L. 1994. Science as a way of knowing - Biodiversity. American Zoologist, 34: 3-4.
Morton, J.K. 1982. Preservation of endangered species by transplantation. Canadian Botanical Association
Bulletin, 15: 32.
Moss, R. 1989. Capercaillie in native woodlands and plantations. Newsletter: Native Woodlands Discussion
Grp. no.14.
Moss, R. & Picozzi, N. 1994. Management of Forests for Capercaillie in Scotland. Forestry Commission
Bulletin 113. London, HMSO.
Moyle, P.B. & Williams, J.E. 1990. Biodiversity Loss in the Temperate Zone: Decline of the Native Fish Fauna
of California. Conservation Biology, 4: 275-282.
Review of information, policy and legislation on species translocations
Munn, L.C. 1993. Effects of prairie dogs on physical and chemical properties of soils. In: Proceedings of the
symposium on the management of prairie dog complexes for the reintroduction of the black footed
ferret, ed by J.L. Oldemeyer, D.E. Biggins, B.J. Miller & R. Crete, 11-17.
Munro, A.L.S. 1986. Transfer and introductions: do the dangers justify greater public control? In: Realism in
aquaculture: achievements, constraints, perspectives, ed by M. Bilio, H. Rosenthal & C.J. Sindermann,
Munson, A. 1995. Should a biosafety protocol be negotiated as part of the biodiversity convention? Global
Environmental Change - Human and Policy Dimensions, 5: 7-26.
Munson, L. & Cook, R.A. 1993. Monioring, investigation, and surveillance of diseases in captive wildlife.
Journal of Zoo and Wildlife Medicine, 24: 281-290.
Munton, P. 1988. The role, impact and management of the translocation of living organisms in the context of
European Wildlife resources. Gland, Switzerland, Introduction Specialist Group of the Species
Survival Commission of IUCN.
Murray, J.B. 1970. Escaped American red-tailed hawk nesting with buzzard in Midlothian. British Birds, 6: 3437.
Murton, R.K., Thearle, R.J.P. & Coombs, C.F.B. 1974. Ecological studies of the feral pigeon (Columba livia
var.) III. Reproduction and plumage polymorphism. Journal of Applied Ecology, 11: 841-854.
Musil, C.F. 1993. Effect of invasive Australian acacias on the regeneration, growth and nutrient chemistry of
South African lowland fynbos. Journal of Applied Ecology, 30: 361-372.
Nadel, H., Frank, J.H. & Knight, R.J. 1992. Escapes and accomplices - the naturalization of exotic ficus and
their associated faunas in Florida. Florida Entomologist, 75: 29-38.
Nafus, D.M. 1993. Movement of introduced biological-control agents onto nontarget butterflies, Hypolimnas
spp (Lepidoptera, Nymphalidae). Environmental Entomology, 22: 265-272.
Nagy, S. & Sporka, F. 1990. Macrozoobenthos of Danube branch of plesiopotamal type and its changes under
the influence of artificial fish introduction. Biologia, 45: 781-790.
National Rivers Authority. 1993. NRA Fisheries Strategy. Bristol, National Rivers Authority.
Nature Conservancy Council. 1983. Consultation procedures on wildlife introductions in Great Britain.
Peterborough, Nature Conservancy Council. (Policy and procedure guidelines 1/83.)
Nature Conservancy Council. 1985. The Sea Eagle. Peterborough, Nature Conservancy Council..
Nature Conservancy Council. 1987. Consultation procedures on wildlife introductions in Great Britain.
Peterborough, Nature Conservancy Council. (Policy and Procedure Guidelines 1/83 [revised August
Nature Conservancy Council. 1988a. Wildlife translocations in Great Britain. Peterborough, Nature
Conservancy Council. (Advisory Committee on Science.)
Nature Conservancy Council. 1988b. Habitat translocation and the safeguard of semi-natural habitats.
Peterborough, Nature Conservancy Council. (NCC draft guidance note.)
Nature Conservancy Council. 1988c. Native trees and shrubs for wildlife in the United Kingdom - Farm
Woodland Scheme. Peterborough, Nature Conservancy Council.
Nature Conservancy Council. 1989. Bird Reintroduction Schemes in Britain. Ornithology Note 14.
Nature Conservancy Council. 1990. Review of NCC policy on Species Translocations in Great Britain.
Peterborough, Nature Conservancy Council. (NCC BD P90 21.)
Nelson, C.N. 1994. Ergasiophygophytes in the British Isles-plants that jumped the garden fence. In: The
common ground of wild and cultivated plants, ed by A.R. Perry & R.G. Ellis, 17-30. Cardiff, National
Museum of Wales.
Neser, S. & Kluge, R.L. 1986. The importance of seed-attacking agents in the biological control of invasive
alien plants. In: The ecology and management of biological invasions in Southern-Africa, ed by I.A.W.
MacDonald, F.J. Kruger & A.A. Ferrar, 285-293. Cape Town, Oxford University Press.
Neth, P. & Barnhart, G. 1983. The landlocked salmon. Conservationist, 38: 24-33.
New, T.R., Pyle, R.M., Thomas, J.A., Thomas, C.D. & Hammond, P.C. 1995. Butterfly conservation
management. Annual Review of Entomology, 40: 57-83.
Newlands, W.A. 1976. Reintroduction of capercaillie Tetrao uregallus. In: Reintroductions: techniques and
ethics, ed by L. Boitani, 197-199. Rome, World Wildlife Fund.
Newmark, P. 1986. Approval for first British virus release experiment. Nature, 320, .
Newsome, A.E. & Noble, I.R. 1986. Ecological and physiological characters of invading species. In: Ecology of
biological invasions, ed by R.H. Groves & J.J. Burdon, 1-20. Cambridge, Cambridge University Press.
Newton, I. 1988. Reintroduction and its relation to the management of raptor populations. In: Proceedings of the
International Symposium on Raptor Reintroduction, ed by D.K.Garcelon & G.W.Roemer, 1-15. Arcata,
California, Institute for Wildlife Studies.
Review of information, policy and legislation on species translocations
Newton, I., Davis, P.E. & Moss, D. 1994. Philopatry and population growth of red kites, Milvus milvus, in
Wales. Proceedings of the Royal Society London. Series B, 257: 317-323.
Niemela, J. & Spence, J.R. 1991. Distribution and abundance of an exotic ground beetle (Carabidae) - a test of
community impact. Oikos, 62: 351-359.
North Atlantic Salmon Conservation Organisation. 1995. Report of the north-east Atlantic commission working
group on introductions and transfers of salmonids. Edinburgh, North Atlantic Salmon Conservation
Novellie, P.A. & Knight, M. 1994. Repatriation and translocation of ungulates into South African national
parks: An assessment of past attempts. Koedoe, 37: 115-119.
Nunney, L. & Elam, D.R. 1994. Estimating the effective population size of conserved populations. Conservation
Biology, 8: 175-184.
Oates, M.R. & Warren, M.S. 1990. A review of butterfly introductions in Britain. Goldalming, World Wide
Fund for Nature.
O'Brien, P. 1989. Introduced animals and exotic disease - assessing potential risk and appropriate response.
Australian Veterinary Journal, 66: 382-385.
O'Connor, R.J. 1987. Biological characteristics of invaders among bird species in Great Britain. In: Quantative
aspects of the ecology of biological invasions, ed by H. Kornberg & M.H. Williamson, 583-598.
London, The Royal Society.
Office of Technological Assessment. 1993. Harmful non-indigenous species in the US. Washington, US Office
of Technology Assessment.
Ogutuohwayo, R. 1990. The decline of the native fishes of Lake Victoria and Kyoga (East Africa) and the
impact of inroduced species, especially the Nile perch, Lates niloticus, and the Nile tilapia,
Oreochromis niloticus. Environmental Biology of Fishes, 27: 81-96.
Ogutuohwayo, R. 1990. The reduction in fish species diversity in lakes Victoria and Kyoga (East Africa)
following human exploitation and introduction of nonnative fishes. Journal of Fish Biology, 37: 207208.
Olckers, T. & Hulley, P.E. 1991. Impoverished insect herbivore faunas on the exotic bugweed Solanum
mauritianum Scop. relative to indigenous Solanum species in Natal/KwaZulu and the Transkei.
Journal of the Entomological Society of South Africa, 54: 39-50.
Olckers, T. & Hulley, P.E. 1995. Importance of preintroduction surveys in the biological control of Solanum
weeds in South Africa. Agriculture, Ecosystems and Environment, 52: 179-185.
Old, R.W. & Primrose, S.B. 1985. Studies in microbiology. Principles of gene manipulation. Vol. 2. 3rd ed.
Oxford, Blackwell Scientific Publications.
Oldham, R.S., Musson, S. & Humphries, R.N. 1991. Translocation of crested newt populations in the UK.
Herpetofauna News, 2: 3-5.
Oliver, J.H., Hayes, M.P., Kierans, J.E. & Lavender, D.R. 1993. Establishment of the Foreign Parthogenetic
Tick Amblyomma rotundatum (Acari: Ixodidae) in Florida. Journal of Parasitology, 79: 786-790.
Olney, P.J.S., Mace, G.M. & Feistner, A.T.C. 1994. Creative Conservation - Interactive management of wild
and captive animals. London, Chapman & Hall.
Owen, D.F. & Smith, D.A.S. 1989. Utilization of alien Asclepiadaceae as larval food-plants by Danaus
plexippus (L.) (Lepidoptera: Danaidae) on the Atlantic islands. Entomologist., 108: 158-164.
Owen, H.J. & Norton, D.A. 1995. The diet of introduced bushtail possums Trichosurus vulpecula in a low
diversity New Zealand nothofagus forest and possible implications for conservation management.
Biological Conservation, 71: 339-345.
Owen, M. 1990. The damage-conservation interface illustrated by geese. Ibis, 132: 238-252.
Packard, S. 1994. Successional restoration: Thinking like a prairie. Restoration Management Notes, 12: 32-39.
Padgett, D.J. & Crow, G.E. 1993. Some unwelcome additions to the flora of New Hampshire. Rhodora, 95: 348351.
Page, C.N. & Gardner, M.F. 1994. Conservation of rarer temperate rainforest conifer tree species: a fastgrowing role for arboreta in Britain and Ireland. In: The Common ground for Wild and Cultivated
Plants, ed by A.R. Perry & R. Gwynn Ellis, 119-144. Cardiff, National Museum of Wales.
Paillet, F.L. & Rutter, P.A. 1989. Replacement of native oak and hickory tree species by the introduced
American chesnut (Castanea dentata) in southwestern Wisconsin. Canadian Journal of Botany Journal Canadien de Botanique, 67: 3457-3469.
Palacios, R., Martinez, E., Flores, M., Romero, D., Brom, S., Davila, G. & Pinero, D. 1988. Organisation and
dynamics of the Rhizobium genome. A basis for introducing novel arrangements of genetic information
into the environment. In: Introduction of genetically modified organisims into the environment, ed by
H.A. Mooney & G. Bernardi, 69-78. Chichester, John Wiley & Sons.
Review of information, policy and legislation on species translocations
Palmer, J. 1990. The biology and control of invasive plants.London, BES Industrial Ecology Group.
Palmer, J.P. 1994. Fallopia japonica (Japanese knotweed) in Wales. In: Ecology and management of invasive
riverside plants, ed by L.C. de Waal, L.E. Child, P.M. Wade & J.H. Brock, 32-42. Chichester, John
Wiley & Sons.
Palmer, M. 1994. Action plan for the conservation of the native freshwater crayfish Austropotamobius pallipes
in the United Kingdom. JNCC Report, No.193
Palmer, M. 1995. A UK plant conservation strategy. Peterborough, Joint Nature Conservation Committee.
Pandey, H.N. & Dubey, S.K. 1989. Growth and population dynamics of an exotic weed Parthenium
hysterophorus Linn. Proceedings of the Indian Academy of Science. Plant Science, 99: 51-58.
Papworth, D.S. 1980. Registration requirements in the UK for bacteria, fungi and viruses used as pesticides. In:
Environmental protection and biological forms of control of pest organisms, ed by B. Lundholm & M.
Stackerud, 135-143. Stockholm.
Park, D.G. 1989. Relocating magnesian limestone grassland. In: Biological habitat reconstruction, ed by G.P.
Buckley, 264-279. London, Belhaven.
Parker, D.M. 1995. Habitat creation - a critical guide. Peterborough, English Nature.
Parker, I.M., Mertens, S.K. & Schemske, D.W. 1993. Distribution of seven native and two exotic plants in a
tallgrass prairie in southeastern Wisconsin: The importance of human disturbance. American Midland
Naturalist, 130: 43-55.
Parkes, J.P. 1990. Feral goat control in New Zealand. Biological Conservation, 54: 335-348.
Parkes, J.P. 1993. Feral goats: Designing solutions for a designer pest. New Zealand Journal of Ecology , 17:
Parnell, J.A.N., Cronk, Q., Jackson, P.W. & Strahm, W. 1989. A study of the ecological history, vegetation and
conservation management of Ile aux Aigrettes, Mauritius. Journal of Tropical Ecolology, 5: 355-374.
Parsons, P.A. 1994. The energetic cost of stress. Can biodiversity be preserved? Biodiversity Letters, 2: 11-15.
Part, T. 1995. The importance of local familiarity and search costs for age- and sex-biased philopatry in the
collared flycatcher. Animal Behaviour, 49: 1029-1038.
Passera, L. 1994. Characteristics of tramp species. In: Exotic ants: biology, impact and control of introduced
species, ed by D.F. Williams, 23-43. California, Westview Press.
Paul, E.M., Thompson, C. & Dunwell, J.M. 1995. Gene dispersal from genetically-modified oil seed rape in the
field. Euphytica, 81: 283-289.
Paulay, G. 1994. Biodiversity on oceanic islands - its origin and extinction. American Zoologist, 34: 134-144.
Pavlik, B.M., Nickrent, D.L. & Howald, A.M. 1993. The recovery of an endangered plant. I. Creating a new
population of Amsinckia grandiflora. Conservation Biology, 7: 510-526.
Peakall, D.B. 1990. Prospects for the peregrine falcon, Falco peregrinus, in the nineties. Canadian Field
Naturalist, 104: 168-173.
Pennington, W. 1969. The history of British vegetation. London, English Universities Press.
Pepper, H. & Stocker, D. (1993): Grey squirrel control using modified hoppers. Surrey, Forestry Authority
Resarch Information Note.
Perring, F.H. & Farrell, L. 1983. British red data book: vascular plants. London, RSNC.
Perrins, J., Williamson, M. & Fitter, A. 1992. A survey of differing views of weed classification: implications
for regulation of introductions. Biological Conservation, 60: 47-56.
Perry, A.R. & Ellis, R.G., eds. 1994. The common ground of wild and cultivated plants. Cardiff, National
Museum of Wales.
Philippart, J.C. 1995. Is captive breeding an effective solution for the preservation of endemic species?
Biological Conservation, 72: 281-295.
Philippart, J.C., Baras, E. & Rimbaud, G. 1990. First biological data on juvenile Atlantic salmon (Salmo salar
L.) reintroduced in the River Ourthe (Belgium). Cahiers Ethologie Applique, 10: 421-440.
Picozzi, N. 1989. Release of captive bred species: gamebirds. In: Release of captive bred species: genetic
interactions with wild relatives, ed by A.J. Gray, 92-100, Unpublished ITE Report to the Department of
the Environment.
Piekarczyk, J. 1991. Changes in the beetle and spider faunas of unimproved neutral grassland (Ashington,
Northumberland) after transplantation. Peterborough, NCC.
Pienkowski, M.W. 1993. A contribution to the development of a system to assess nature conservation quality
and to set targets for the national action plan required by the Convention on Biological Diversity.
Peterborough, JNCC.
Pimental, D. 1991. Diversification of biological-control strategies in agriculture. Crop Protection, 10: 243-253.
Review of information, policy and legislation on species translocations
Pimentel, D. 1986. Biological invasions of plants and animals in agriculture and forestry. In: Ecology of
biological invasions of North America and Hawaii, ed by H.A. Mooney & J.A. Drake, 149-162. Berlin,
Pimm, S.L. 1987. Determining the effects of introduced species. Trends in Ecology & Evolution, 2: 106-108.
Pinder, N. 1981. Conservation and Introduced Species. 30th ed. London, British Association of Nature
Conservationists/UCL Ecology & Conservation Unit.
Pitt, W.C. & Jordan, P.A. 1994. A Survey of the nematode parasite Parelaphostrongylus tenuis in the WhiteTailed Deer, Odocoileus virginianus, in a region proposed for Caribou, Rangifer tarandus Caribou, reIntroduction in Minnesota. Canadian Field Naturalist, 108: 341-346.
Polhemus, D.A. 1993. Conservation of aquatic insects - worldwide crisis or localised threats. American
Zoologist, 33: 588-598.
Porter, A. 1994. Implications of introduced garlic mustard (Alliaria petiolata) in the habitat of Pieris
virginiensis (Pieridae). Journal of the Lepidoptera Society, 48: 171-172.
Porter, R., Wynne, G., Avery, M., Thomas, G. & Williams, G. 1994. 1. Into the future: The RSPB's conservation
priorities for the UK. Sandy, RSPB. (Conservation. Review 8.)
Posey, M.H. 1987. Changes in community composition resulting from the introduction of an exotic seagrass.
American Zoologist, 27, A159.
Potts, G.R. 1986. The Partridge. London, Collins.
Powell, J.A. 1992. Recent colonization of the San Francisco Bay area, California, by exotic moths (Lepidoptera,
Tineoidea, Gelechioidea, Tertricoidea, Pyraloidea. Pan-Pacific Entomologist, 68: 105-121.
Prellwitz, D.M., Erickson, K.M. & Osborne, L.M. 1995. Translocation of piping plover nests to prevent nest
flooding. Wildlife Society Bulletin, 23: 103-106.
Price, A.J. 1994. Aspects of rearing Red Grouse Lagopus lagopus scoticus (Lath.) in captivity and their release
onto moorland. PhD Thesis. University of Aberdeen.
Price, E.C. 1992. Adaptation of captive-bred cotton-top tamarins Saguinus oedipus to a natural environment.
Zoo Biology, 11: 107-120.
Probert, P.K. 1993. 1st record of the introduced fouling tubeworm Ficopotamus enigmaticus (Polychaeta,
Serpulidae) in Hawke Bay, New Zealand. New Zealand Journal of Zoology, 20: 35-36.
Pullin, A.S. & Woodell, S.R.J. 1987. Response of the fen violet, Viola persicifolia Schreber, to different
management regimes at Woodwalton Fen National Nature Reserve, Cambridgeshire, England.
Biological Conservation, 41: 203-217.
Puntieri, J.G. 1991. Vegetation response on a forest slope cleared for a ski-run with special reference to the herb
Alstroemeria aurea Graham (Alstroemeriaceae), Argentina. Biological Conservation, 56: 207-221.
Putwain, P.D. & Gillham, D.A. 1988. Restoration of heather moorland. Landscape Design, 172: 51-56.
Pysek, P. & Prach, K. 1993. Plant invasions and the role of riparian habitats: a comparison of four species alien
to central Europe. Journal of Biogeography, 20: 413-420.
Pywell, R.F. 1990. Heathland translocation and restoration. In: Proceedings of Heathlands Conference II, ed by
M.H.D. Auld, 1st ed. : RSPB.
Pywell, R.F. 1993. The restoration of heathland on farmland in southern Britain. PhD Thesis, University of
Pywell, R.F., Webb, N.R. & Putwain, P.D. 1995. A comparison of techniques for restoring heathland on
abandoned farmland. Journal of Applied Ecology, 32: 400-411.
Quinn, R.M., Lawton, J.H., Eversham, B.C. & Wood, S.N. 1994. The biogeography of scarce vascular plants in
Britain with respect to habitat preference, dispersal ability and reproductive biology. Biological
Conservation, 70: 149-157.
Rabb, G.B. 1994. The changing roles of zoological parks in conserving biological diversity. American
Zoologist, 34: 159-164.
Rahbek, C. 1993. Captive breeding - a useful tool in the preservation of biodiversity? Biodiversity and
Conservation, 2: 426-437.
Ranwell, D.S. 1981. Introduced coastal plants and rare species in Britain. In: The biological aspects of rare
plant conservation, ed by H. Sygne, 413-419. Chichester, John Wiley & Sons Ltd.
Rasmy, A.H. & Ellaithy, Y.M. 1988. Introduction of Phytoseiulus persimilis for twospotted spider mite control
in greenhouses in Egypt (Acari: Phytoseiidae, Tetranychidae). Entomophaga., 33: 435-438.
Ratner, M. 1990. Survey and opinions - barriers to field-testing genetically modified organisms. Bio
Technology, 8: 196-198.
Rawes, M. 1972. Trials to recreate floristically-rich vegetation by plant introduction in the Northen Pennines,
England. Biological Conservation, 4: 135-139.
Review of information, policy and legislation on species translocations
Raybould, A.F. & Gray, A.J. 1993a. The impact of genetically modified crops on wild species in the United
Kingdom. In: Gene transfer: are wild species in danger?, ed by J. Yolande, K. Ammann & F. Pythoud,
19-23. Bern, Federal Office of Environment, Forest and Landscape.
Raybould, A.F. & Gray, A.J. 1993b. Genetically modified crops and their wild relatives- A UK perspective.
London, HMSO.
Raybould, A.F. & Gray, A.J. 1993c. Genetically modified crops and hybridisation with wild relatives: a UK
perspective. Journal of Applied Ecology, 30: 199-219.
Raybould, A.F. & Gray, A.J. 1994. Will hybrids of genetically modified crops invade natural communities.
Trends in Ecology & Evolution, 9: 85-88.
Reader, R.J. & Bricker, B.D. 1994. Barriers to establishment of invading, non-forest plants in deciduous forest
nature reserves. Environmental Conservation, 21: 62-66.
Reading, R.P., Grensten, J.J., Beissinger, S.R. & Clark, T.W. 1993. Attributes of black-tailed prairie dog
colonies in north-central Montana, with management recommendations for the conservation of
biodiversity. In: Proceedings of the symposium on the management of prairie-dog complexes for the
reintroduction of the black footed ferret, ed by J.L. Oldemeyer, D.E. Biggins, B.J. Miller & R. Crete, 910.
Reading, R.P. & Kellert, S.R. 1993. Attitudes toward a proposed reintroduction of black-footed ferrets (Mustela
nigripes). Conservation Biology, 7: 569-580.
Regal, P.J. 1993. The true meaning of exotic species as a model for genetically engineered organisms.
Experientia, 49: 225-234.
Regal, P.J. 1994. Scientific principles for ecologically based risk assessment of transgenic organisms. Molecular
Ecology, 3: 5-13.
Reichholf, J. 1976. The reintroduction of the beaver Castor fiber to Bavaria: some preliminary results. In:
Reintroductions: techniques and ethics, ed by L. Boitani, 49-53. Rome, World Wildlife Fund.
Reid, C. 1994. Nature Conservation Law. Edinburgh, Sweet & Maxwell.
Reverbori, H. 1993. GMOs in the bioindustry. Biofutur, 19-21.
Reynolds, J.C. 1985. Details of the geographic replacement of the red squirrel Sciurus vulgaris by the grey
squirrel Sciurus carolinensisin Eastern England. Journal of animal Ecology, 54: 149-162.
Reynolds, S.C.P. 1992. Distribution of alien and adventive plants at ports and on roadsides in Ireland in 1989.
Ir. Nat. J., 24: 59-65.
Rhodes, O.E., Jr. & Chesser, R.K. 1994. Genetic concepts for habitat conservation: The transfer and
maintenance of genetic variation. Landscape Urban Planning, 28: 55-62.
Ricci, M. & Eaton, L. 1994. The rescue of Wahlenbergia larrainii in Robinson Crusoe Island, Chile. Biological
Conservation, 68: 89-93.
Richardson, D.M. & Bond, W.J. 1991. Determinants of plant distribution - evidence from pine invasions.
American Naturalist, 137: 639-668.
Richardson, D.M., Williams, P.A. & Hobbs, R.J. 1994. Pine invasions in the Southern Hemisphere:
determinants of spread and invadability. Journal of Biogeography, 21: 511-527.
Rieder, N. & Rohrer, P. 1982. The possibility of reintroducing the beaver in SW Germany. Carolinea, 40: 9198.
Robertson, D.J., Robertson, M.C. & Tague, T. 1994. Colonization dynamics of four exotic plants in a northern
Piedmont natural area. Bulletin of the Torrey Botanical Club, 121: 107-118.
Robinson, G.R., Yurlina, M.E. & Handel, S.N. 1994. A century of change in the Staten Island flora - ecological
correlates of special losses and invasions. Bulletin of the Torrey Botanical Club, 121: 119-129.
Rodriguez, A., Barrios, L. & Delibes, M. 1995. Experimental release of an Iberian Lynx (Lynx pardinus).
Biodiversity and Conservation, 4: 382-394.
Rogers, H.J. & Parkes, H.C. 1995. Transgenic plants and the environment. Journal of Experimental Botany, 46:
Roland, J. & Embree, D.G. 1995. Biological-control of the winter moth. Annual Review of Entomology, 40: 475492.
Roper, S.A. 1988. Aspects of the ecology of the European Badger Meles meles on farmland. MPhil Thesis,
University of Exeter.
Rose, C.I. 1979. Nature conservation and species introductions. London, University College London.
(Discussion Papers in Conservation. Ecology & Conservation Unit.)
Rosecchi, E., Crivelli, A.J. & Catsadorakis, G. 1993. The establishment and impact of Pseudorasbora parva, an
exotic fish species introduced into Lake Mikri Prespa (north-western Greece). Aquatic conservation marine and freshwater ecosystems, 3: 223-231.
Review of information, policy and legislation on species translocations
Ross, K.G. 1994. Exotic ants - biology, impact and control of introduced species - Williams,DF. Science, 265:
Ross, S.T. 1991. Mechanisms structuring stream fish assemblages - are there lessons from introduced species.
Environmental Biology of Fishes, 30: 359-368.
Rudolph, C.D., Conner, R.N., Carrie, D.K. & Schaefer, R.R. 1992. Experimental reintroduction of red-cockaded
woodpeckers. Auk, 109: 914-916.
Rueetschi, J. 1988. Reintroduction-experiment with Colias palaeno europome Esper. Nota Lepidopterologica,
11: 223-230.
Ruesink, J.L., Parker, I.M., Groom, M.J. & Karieva, P.M. 1995. Reducing the risks of nonindigenous species
introductions. Bioscience, 45: 465-477.
Rutherford, M.C., Pressinger, F.M. & Musil, C.F. 1986. Standing crops, growth rates and resource use
efficiency in alien plant invaded ecosystems. In: The ecology and management of biological invasions
in Southern Africa, ed by I.A.W. MacDonald, F.J. Kruger & A.A. Ferrar, 189-199. Cape Town, Oxford
University Press.
Ryan, P.G. 1991. The impact of commercial lobster fishery on seabirds at the Trestan da Cunha Islands, South
Atlantic Ocean. Biological Conservation, 57: 339-350.
Ryan, R.B. 1990. Evaluation of biological-control, introduced parasites of larch casebearer (Lepidoptera,
Coleophoridae) in Oregon. Environmental Entomology, 19: 1873-1881.
Ryves, T.B. 1988. Supplementary list of wool-alien grasses recorded from Blackmoor, North Hants., 19591976. Watsonia, 17: 73-79.
Said, K. & Brittondavidian, J. 1991. Genetic differentiation and habitat partition of Robertsonian house mice
populations of Tunisia. Journal of Evolutionary Biology, 4: 409-427.
Sale, J.B., Mahendi-Andau, P. & Hiong, L.K. 1995. The capture and translocation of orangutans in Sabah,
Malaysia. Re-Introduction News, 10: 12-14.
Sale, J.B. & Singh, S. 1987. Reintroduction of greater Indian rhinoceros into Dudhwa National Park. Oryx, 21:
Salmon Advisory Commitee. 1991. Assessment of stocking - The salmon management strategy. London,
Saltz, D. & Rubenstein, D.I. 1995. Population dynamics of a reintroduced Asiatic wild ass (Equus hemionus)
herd. Ecological Applications, 5: 327-335.
Sampson, C. 1994. Cost and impact of current control methods used against Haracleum mantegazzianum (giant
hogweed) and the case for instigating a biological control programme. In: Ecology and management of
invasive riverside plants, ed by L.C. de Waal, L.E. Child, P.M. Wade & J.H. Brock, 55-65. Chichester,
John Wiley & Sons.
Sands, D.C. & Miller, R.V. 1993. Evolving strategies for biological-control of weeds with plant pathogens.
Pesticide Science, 37: 399-403.
Savidge, J.A. 1987. Extinction of an island forest avifauna by an introduced snake. Ecology, 68: 660-668.
Savidge, J.A. 1991. Population characteristics of the introduced brown tree snake(Boiga irregularis) on Guam.
Biotropica, 23: 294-300.
Sawyers, C. 1989. Native plants under siege. Garden USA, 13: 12-3,15,32.
Scanlan, J.C. 1986. Woody weeds control in northern Australia. In: Proceedings of the third Australian
conference on tropical pastures. Brisbane Australia, Tropical Grassland Society of Australia.
Schierenbeck, K.A., Mack, R.N. & Sharitz, R.R. 1994. Effects of herbivory on growth and biomass allocation in
native and introduced species of Lonicera. Ecology, 75: 1661-1672.
Schiffman, P.M. 1994. Promotion of exotic weed establishment by endangered giant kangaroo rats (Dipodomys
ingens) in a California grassland. Biodiversity and Conservation, 3: 524-537.
Schifinowittmann, M.T., Lau, A.H. & Simioni, C. 1994. The genera Vicia and Lathyrus (Leguminosae) in Rio
Grande do Sul (southern Brazil) - cytogenetics of native,naturalized and exotic species. Revista
Brasileira de Genetica, 17: 313-319.
Schmidt, R.E. & McGurk, J. 1982. Biology of the European bitterling Rhodeus sericeus (Pisces, Cyprinidae) in
the Bronx River, New York, USA - an apparently benign exotic species. Biological Conservation, 24:
Schmitt, J. & Linder, C.R. 1994. Will escaped transgenes lead to ecological release. Molecular Ecology, 3: 7174.
Schofield, E.K. 1989. Effects of Introduced Plants and Animals on Island Vegetation: Examples from the
Galapagos Archipelago. Conservation Biology, 3: 227-238.
Review of information, policy and legislation on species translocations
Schonrogge, K., Stone, G.N. & Crawley, M.J. 1995. Spatial and temporal variation in guild structure:
Parasitoids and inquilines of Andricus quercuscalicis (Hymenoptera: Cynipidae) in its native and alien
ranges. Oikos, 72: 51-60.
Schreiner, I. 1989. Biological control introductions in the Caroline and Marshall islands. Proceedings Hawaii
Entomological Society, 29: 57-70.
Schreiner, I.H. & Nafus, D.M. 1992. Changes in a moth community mediated by biological-control of the
dominant species. Environmental Entomology, 21: 664-668.
Schuele, W. 1993. Mammals, vegetation and the initial human settlement of the Mediterranean islands: a
palaeoecological approach. Journal of Biogeography, 20: 399-412.
Schwalbe, C.P. & Mastro, V.C. 1988. Multispecific trapping techniques for exotic-pest detection. Agriculture,
Ecosystems and Environment, 21: 43-51.
Scott, J.K. & Panetta, F.D. 1993. Predicting the Australian weed status of southern African plants. Journal of
Biogeography, 20: 87-93.
Scottish Natural Heritage. 1994. SNH policy on translocation of species. Interim guidance note.
Seidler, R.J. & Levin, M. 1994. Potential ecological and nontarget effects of transgenic plant gene-products on
agriculture, silviculture, and natural ecosystems -general introduction. Molecular Ecology, 3: 1-3.
Senner, S.E., White, C.M. & Parish, J.R. 1984. Raptor conservation in the next 50 years. Raptor Research
Servheen, c, Kasworm, W.F. & Thier, T.J. 1995. Transplanting Grizzly Bears Ursus arctos horriblis as a
management tool-results from the cabinet mountains, Montana, USA. Biological Conservation, 71:
Shafland, P.L. 1986. A review of Florida's efforts to regulate, assess and manage exotic fishes. Fisheries, 11:
Sharp, R. 1995. Bird conservation and the UK Biodiversity Action Plan. Ibis, 137, S219-S223.
Sharypova, L.A. & Simarov, B.V. 1994. Prospects of construction and release of genetically-modified rhizobia
into the environment. Genetika, 30: 1445-1457.
Sheail, J. 1984. Exploited animals: The rabbit. Biologist, 31: 135-140.
Sheail, J. 1988. The extermination of the muskrat (Ondatra zibethicus) in inter-war Britain. Archives of Natural
History, 15: 155-170.
Sheail, J. 1991. The management of an animal population: changing attitudes towards the wild rabbit in Britain.
Journal of Environmental Management, 33: 189-203.
Sheil, D. 1994. Naturalized and invasive plant species in the evergreen forests of the East Usambara Mountains,
Tanzania. African Journal of Ecology, 32: 66-71.
Shepherd, R.C.H. 1989. Insect fauna of the Australian noxious weed Emex australis Steinheil in the Western
Cape, South Africa. Australia Entomologists Magazine, 16: 9-15.
Sheppard, D.A. 1990. Changes in the fauna magnesian limestone grassland after transplantation. In: Calcareous
grasslands - ecology and management, ed by S.H. Hillier, D.W.H. Walton & D.H. Wells, 169-170.
Huntingdon, Bluntisham Books.
Sheppard, D. 1995. Guidance notes for invertebrate translocations and introductions. In: Species conservation
handbook. Peterborough, English Nature.
Shine, R. 1991. Strangers in a strange land - ecology of the Australian colubrid snakes. Copeia., 1: 120-131.
Shorrocks, B. & Coates, D., eds. 1993. The release of genetically engineered organisms. British Ecological
Society Ecological Issues No.4. Shrewsbury, Field Studies Council.
Short, C. 1994. Implications of game management for woodland management, landscape conservation and
public recreation. Cirencester, Centre for Rural Studies.
Short, J., Bradshaw, S.D., Giles, J., Prince, R.I.T. & Wilson, G.R. 1992. Reintroduction of macropods
(Marsupialia: Macropodoidea) in Australia - A review. Biological Conservation, 62: 189-204.
Simberloff, D. 1989. Which insect introductions succeed and which fail. In: Biological invasions, ed by J.A.
Drake, H.A. Mooney, F. di Castri, R.H. Groves, F.J. Kruger, M. Rejmánek & M. Williamson, 61-75.
Chichester, John Wiley & Sons.
Simberloff, D. 1992. Extinction, survival and effects of birds introduced to the Mascarenes. Acta Oecologica International Journal of Ecology, 13: 663-678.
Simberloff, D. & Boecklen, W. 1991. Patterns of extinction in the introduced Hawaiian avifauna - a
reexamination of the role of competition. American Naturalist, 138: 300-327.
Simpson, D.A. 1984. A short history of the introduction and spread of Elodea Michx in the British Isles.
Watsonia, 15: 1-9.
Sinicrope, T.L., Hine, P.G., Warren, R.S. & Niering, W.A. 1990. Restoration of an impounded salt marsh in
New England. Estuaries., 13: 25-30.
Review of information, policy and legislation on species translocations
Sjoasen, T. 1995. Re-introduction of European Otters in Sweden. Re-Introduction News, 10: 10-11.
Skalka, A.M. 1990. Risk assessment for genetic experimentation and application. In: Introduction of genetically
modified organisims into the environment, ed by H.A. Mooney & G. Bernardi, 3-8. Chichester, John
Wiley & Sons.
Smal, C.M. 1987. The diet of the barn owl Tyto alba in southern Ireland, with reference to a recently introduced
prey species - the bank vole Clethrionomys glareolus. Bird Study, 34: 113-125.
Smal, C.M. 1991. Population studies on feral American mink Mustela vison in Ireland. Journal of Zoology,
224: 233-249.
Smallwood, K.S. 1994. Site invasibility by exotic birds and mammals. Biological Conservation, 69: 251-259.
Smallwood, K.S. & Salmon, T.P. 1992. A rating system for potential exotic bird and mammal pests. Biological
Conservation, 62: 149-159.
Smit, E., Vanelsas, J.D. & Vanveen, J.A. 1992. Risks associated with the application of genetically modified
microorganisms in terrestrial ecosystems. Fems Microbiology Reviews, 88: 263-278.
Smith, R. 1992. Repopulation projects and butterfly conservation. Butterfly Conservation News, 51: 53-59.
Smith, R.E.N., Webb, N.R. & Clarke, R.T. 1991. The establishment of heathland on old fields in Dorset,
England. Biological Conservation, 57: 221-234.
Smith, R.H. & Holloway, G.J. 1989. Population genetics and insect introductions. Entomologist, 108: 14-27.
Snazell, R.G., Rispin, W.E., Thomas, J.A. & Elmes, G.W. 1995a. Invertebrate monitoring: Arethusa Clump
chalk grassland restoration 1993. Unpublished ITE Report to Mott MacDonald Civil Ltd.
Snazell, R.G., Rispin, W.E. & Thomas, J.A. 1995b. M3 Bar End to Compton. Invertebrate monitoring: Arethusa
Clump chalk grassland restoration. Unpublished ITE Report to Mott MacDonald Civil Ltd.
Snazell, R.G., Ward, L.K., Rispin, W.E., Thomas, J.A., Morris, M.G. & Webb, N.R. 1991. M3 Bar End to
Compton ecological survey 1991. Unpublished ITE Report to Mott MacDonald Civil Ltd.
Snyder, N.F.R., Koenig, S.E., Koschman, J., Snyder, H.A. & Johnson, T.B. 1994. Thick-billed parrot releases in
Arizona. Condor, 96: 845-862.
Society for the Promotion of Nature Reserves. 1970: Policy on introductions to nature reserves. SPNR
Technical Publication, No 2.
Soderback, B. 1995. Replacement of the native crayfish Astacus astacus by the introduced species Pacifastacus
leniusculus in a Swedish lake: Possible causes and mechanisms. Freshwater Biology, 33: 291-304.
Soeyrinki, N. 1991. On the alien flora of the province of Buenos Aires, Argentina. Annales Botanici Fennici, 28:
Soulé, M.E. 1986. The Science of scarcity and diversity. Massachusetts, Sinauer Associates Inc.
Soulé, M.E., ed. 1987. Viable populations for conservation. Cambridge, Cambridge University Press.
Soutar, R.G. & Peterken, G.F. 1988. Regional lists of native trees and shrubs for use in afforestation schemes.
Arboriculture Journal.
Species Survival Commission & Reintroduction Specialist Group. 1992. Guidelines for Introductions. ReIntroduction News, 4: 2-3.
Spier, R.E. 1990. Genetically modified organisms in commercial use. Enzyme and Microbial Technology, 12:
Spinney, L. 1995. Return to the wild. New Scientist, Jan, 35-38.
Spittler, H. 1994. Reintroduction trails of capercaillie (Tetrao uragallus L) in Hochsauerland. Zeitschrift Fur
Jagdwissenschaft, 40: 185-199.
Stace, C. 1991. New flora of the British Isles. Cambridge, Cambridge University Press.
Staines, B. 1989. Release of captive bred species: deer. In: Release of captive bred species: genetic interactions
with wild relatives, ed by A.J. Gray, 121-125. Unpublished ITE Report to the Department of the
Stanley, J.G., Peoples, R.A. & McCann, J.A. 1991. United States federal policies, legislation and responsibilities
related to importation of exotic fishes and other aquatic organisms. Canadian Journal of Fisheries and
Aquatic Science, 48: 162-166.
Stanley Price, M.R. 1990. Animal reintroductions: The Arabian Oryx in Oman. Cambridge, Cambridge
University Press.
Stanley-Price, M.R. 1991. A review of Mammal re-introductions and the role of the Re-introduction Specialist
Group of IUCN/SCC. Symposium of the Zoological Society of London, 62: 9-25.
Stary, P. & Gonzalez, D. 1992. Field acceptance of exposed exotic aphids by indigenous natural enemies
(Homoptera: Aphidinea: Aphididae). Entomologica Generalis, 17: 121-129.
Stary, P., Michelina, J.M. & Melia, A. 1985. Lysiphlebus testaceipes (Cresson, 1880), an exotic parasite of
aphids and a biological control agent in Spain (Hym., Aphidiidae). Graellsia, 41: 131-135.
Review of information, policy and legislation on species translocations
Steele, D. 1993. The new UK rules on releasing genetically-modified organisms. Chemistry & Industry, 542544.
Steigman, K.L. & Ovenden, L. 1986. Transplanting tallgrass prairie with a sodcutter. Proceedings of the tenth
North American Prairie Conference, .
Sterne, J.K. 1995. Supplementation of wild salmon- stocks: A core for the hatchery problem or more problem
hatcheries? Coastal Management, 23: 123-152.
Stevenson, M.J., Bullock, J.M. & Ward, L.K. 1995. Re-creating semi-natural communities: The effect of sowing
rate on the establishment of calcareeous grassland vegetation. Restoration Ecology, In press.
Stevenson, M.J., Snazell, R.G. & Ward, L.K. 1994. M3 bar end to Compton. Unpublished ITE Report to Mott
MacDonald Civil Ltd.
Stewart, J.E. 1991. Introductions as factors in diseases of fish and aquatic invertebrates. Can. J. Fish. Aquat.
Sci., 48: 110-117.
Stewart, J. 1993. Propagation and translocation of orchids. BSBI News, 64: 51-53.
Stone, R. 1994. Large plots are next test of transgenic crop safety. Science, 266: 1472-1473.
Story, J.M., Boggs, K.W., Good, W.R., Harris, P. & Nowierski, R.M. 1991. Metzneria paucipunctella zeller
(Lepidoptera - Gelechiidae), a moth introduced against spotted knapweed - its feeding strategy and
impact on 2 introduced Urophora spp (Diptera, Tephritidae). Canadian Entomologist, 123: 1001-1007.
Stubbs, A. 1995. Quinquennial Review of Wildlife & Countryside Act, 1981. Control of release of butterflies
proposal by Butterfly Conservation. Invertebrate Conservation News, 16: 3-7.
Stubbs, D. 1988. Towards an introductions policy. London, Wildlife Link.
Stuewe, M. & Nievergelt, B. 1991. Recovery of alpine ibex from near extinction: the result of effective
protection, captive breeding, and reintroductions. Applied Animal Behaviour Science, 29: 279-387.
Suckling, C.W. 1991. Identifying hazards in the release of genetically modified organisms. Biotechnology
Education, 2: 51-55.
Sukopp, H. & Sukopp, U. 1993. Ecological long-term effects of cultigens becoming feral and of naturalization
of nonnative species. Experientia, 49: 210-218.
Sutherland, W.J. & Allport, G. 1991. The distribution and ecology of naturalized Egyptian geese Alopochen
aegyptiacus in Britain. Bird Study, 38: 128-134.
Sykes, J.M., Lowe, V.P.W. & Briggs, D.R. 1989. Some effects of afforestation on the flora and fauna of an
upland sheepwalk during 12 years after planting. Journal of Applied Ecology, 26: 299-320.
Tachmintzis, J. 1994. EEC regulatory provisions for the release of transgenic plants. Biocontrol Science and
Technology, 4: 591-595.
Tanner, C.C., Clayton, J.S. & Coffey, B.T. 1990. Submerged-vegetation changes in Lake Rotoroa (Hamilton,
New Zealand) related to herbicide treatment and invasion by Egeria densa. New Zealand Journal of
Marine and Freshwater Research, 24: 45-57.
Taylor, O.R., Jr. 1985. African bees: Potential impact in the United States. Bulletin of the Entomological Society
of America, 31: 15-24.
Taylor, R.J. 1991. The origin of Lamium hybridum - a case-study in the search for the parents of hybrid species.
Northwest Science, 65: 116-124.
Tear, T.H. & Forester, D. 1992. Role of social theory in reintroduction planning: A case study of the Arabian
oryx in Oman. Soc. Natural Resources, 5: 349-374.
Tegelstrom, H. & Sjoberg, G. 1995. Introduced Swedish Canada geese (Branta canadensis) have low levels of
genetic variation as revealed by DNA fingerprinting. Journal of Evolutionary Biology, 8: 195-207.
Templeton, A.R. 1986. Coadaptation and outbreeding depression. In: Conservation Biology, ed by M.E. Soulé,
105-116. Sunderland, Sinauer.
Tepfer, M. 1993. Viral genes and transgenic plants. Bio Technology, 11: 1125-1133.
Terrasse, M., Bagnolini, C., Bonnet, J., Pinna, J.-L. & Sarrazin, F. 1994. Reintroduction of the Griffon Vulture
Gyps fulvus in the Massif Central, France. Raptor Conservation Today, 479-491.
Tewes, E. 1994. The European Black Vulture Aegypius monachus Project in Mallorca. Raptor Conservation
Today, 493-498.
Thebaud, C. & Abbott, R.J. 1995. Characterization of invasive Conyza species (Asteraceae) in Europe:
Quantitative trait and isozyme analysis. American Journal of Botany, 82: 360-368.
Thill, D.C., Lish, J.M., Callihan, R.H. & Bechinski, E.J. 1991. Intergrated weed management - a component of
integrated pest management - a critical review. Weed Technology, 5: 648-656.
Thomas, J.A. 1984. The conservation of butterflies in temperate countries: past efforts and lessons for the future.
In: Biology of Butterflies. Symposium of the Royal Entomological Society of London, ed by R.I. VaneWright & P. Ackery, 333-335. London, Academic Press.
Review of information, policy and legislation on species translocations
Thomas, J.A. 1989a. Rare species conservation:case studies of European butterflies. In: The Scientific
management of Temperate Communities for Conservation. The 31st Symposium of the British
Ecological Society, Southampton 1989, ed by I.F. Spellerberg, F.B. Goldsmith & M.G. Morris, 149197. Oxford, Blackwell Scientific Publications.
Thomas, J.A. 1989b. Ecological lessons from the re-introduction of Lepidoptera. Entomologist, 108: 56-68.
Thomas, J.A. 1989c. Genetic effects of the release of captive bred lepidoptera into the wild. In: Release of
captive bred species: genetic interactions with wild relatives, ed by A.J. Gray, 134-141. Unpublished
ITE Report to the Department of the Environment.
Thomas, J.A. 1995a. The ecology and conservation of Maculinea arion and other European species of large blue
butterfly. In: Ecology and Conservation of Butterflies, ed by A.S. Pullin, 181-197. London, Chapman
and Hall.
Thomas, J.A. 1995b. Why small cold blooded insects pose different conservation problems to birds in modern
landscapes. Ibis, 137, S112-S119.
Thomas, J.A. & Morris, M.G. 1994. Patterns, mechanisms and rates of extinction among invertebrates in the
United Kingdom. Philosophical Transactions of the Royal Society of London Series B, 344: 47-54.
Thomas, J.A., Ward, L.K., Snazell, R.G. & Pywell, R. 1992. M3 suggestions for the restoration/re-vegetation of
the moterway banks and the compensation area at the Arethusa Clump. Unpublished ITE report to
Mott McDonald Civil Ltd.
Thomas, M.C. 1995. Invertebrate pests and the Florida Department of Agriculture and Consumer Services.
Florida Entomologist, 78: 39-44.
Thompson, I.P., Ellis, R.J. & Bailey, M.J. 1995. Autecology of a genetically-modified fluorescent pseudomonad
on sugar-beet. Fems Microbiology Ecology, 17: 1-13.
Thompson, J.D. 1990. Spartina anglica, characteristic feature or invasive weed of coastal salt marshes?
Biologist, 37: 9-12.
Thompson, J.D. 1991. The biology of an invasive plant. Bioscience, 41: 393-401.
Thompson, S. 1988. Range expansion by alien weeds in the coastal farmlands of Guyana. Journal of
Biogeography, 15: 109-118.
Thomson, A.D. 1981. New plant disease record in New Zealand: cucumber mosaic virus in Mysotidium
hortensia (Decne) Baill. New Zealand Journal of Agricultural Research, 24: 401-402.
Thornhill, N.W. 1993. The natural history of inbreeding and outbreeding - theoretical and empirical
perspectives. Chicago, University of Chicago Press.
Tiedje, J.M. et al. 1989. The planned introduction of genetically engineered organisms:ecological considerations
and recommendations. Ecology, 70: 298-316.
Tittensor, A.M. 1977. Red squirrel Sciurus vulgaris. In: The handbook of British mammals, ed by G.B. Corbet
& H.N. Southern, 153-164. Oxford, Blackwell Scientific Publications.
Tolin, S.A. & Vidaver, A.K. 1989. Guidelines and regulations for research with genetically modified organisms
- a view from academe. Annual Review of Phytopathology, 27: 551-581.
Tomlin, A.D. 1994. Transgenic plant release - comments on the comparative effects of agriculture and forestry
practices on soil fauna. Molecular Ecology, 3: 51-52.
Towns, D.R., Daugherty, C.H. & Atkinson, I.A.W. 1989. Ecological restoration of New Zealand islands.
Wellington, Department of Conservation.
Townsend, C.R. 1991. Exotic species management and the need for a theory of invasion ecology. New Zealand
Journal of Ecology , 15: 1-3.
Townsend, C.R. & Crowl, T.A. 1991. Fragmented population structure in a native New Zealand fish - an effect
of introduced brown trout. Oikos, 61: 347-354.
Townsend, C.R. & Winterbourn, M.J. 1992. Assessment of the environmental risk posed by an exotic fish -The
proposed introduction of channel catfish (Ictalurus-punctatus) to New-Zealand. Conservation Biology,
6: 273-282.
Trevors, J.T., Kuikman, P. & Watson, B. 1994. Transgenic plants and biogeochemical cycles. Molecular
Ecology, 3: 57-64.
Trinquelle, I. 1991. Legal aspects of the introduction and reinroduction of wildlife species in Europe. Council of
Europe, Strasbourg. (Council of Europe Paper T-PVS (92) 7)
Turner, G.E. 1988. Codes of practice and manual of procedures for consideration of introductions and
transfers. International Council for the Exploration of the Sea. (Cooperative research report)
Turner, M.G. 1987. Effects of grazing by feral horses, clipping, trampling, and burning on a Georgia salt marsh.
Estuaries, 10: 54-60.
Tyldesley, D. 1993. Translocation - or wildlife on the move. Nature's Place, September, 11-12.
Review of information, policy and legislation on species translocations
Tyndalebiscoe, M. & Vogt, W.G. 1991. Effects of adding exotic dung beetles to native fauna on bush fly
breeding in the field. Entomophaga, 36: 395-401.
Tyser, R.W. 1992. Vegetation associated with two alien plant species in a fescue grassland in Glacier National
Park, Montana. Great Basin Naturalist, 52: 189-193.
Tyser, R.W. & Worley, C.A. 1992. Alien Flora in Grasslands Adjacent to Road and Trail Corridors in Glacier
National Park, Montana (U.S.A.). Conservation Biology, 6: 253.
Tzotzos, G.T., ed. 1995. Genetically modified organisms. A guide to biosafety. Wallingford, CAB International.
UK Committee for International Nature Conservation. 1979. Wildlife introductions to Great Britain. London,
Nature Conservancy Council.
Uresk, D. 1993. Relation of black-tailed prairie dogs and control programs to vegetation, livestock, and wildlife.
In: Proceedings of the symposium on the management of prairie-dog complexes for the reintroduction
of the black-footed ferret, ed by J.L. Oldemeyer, D.E. Biggins, B.J. Miller & R. Crete, 8.
Usher, M.B. 1987. Scope-international synthesis symposium on the ecology of biological invasions, held at the
East-West Center, Honolulu, Hawaii, USA. Environmental Conservation, 14: 84-85.
Usher, M.B., Crawford, T.J. & Banwell, J.L. 1992. An American invasion of Great Britain: the case of the
native and alien squirrel (Sciurus) species. Conservation Biology, 6: 108-115.
Uygun, N., Ulusoy, M.R., Sekeroglu, E., Ohnesorge, B. & Gozel, U. 1994. Interactions between 2 introduced
species of whiteflies in the Mediterranean area of Turkey - Dialeurodes citri (ashmead) and
Parabemisia myricae (kuwana) (Hem, Aleyrodidae). Journal of Applied Entomology - Zeitschrift fur
Angewandte Entomologie, 118: 365-369.
Vanas, J.G. 1985. Parasitological consequence of introducing foreign fish species. South African Journal of
Science, 81: 44.
Van der Meer, P. 1993. Potential long-term ecological impact of genetically modified organisms. A survey of
literature, guidelines and legislation. Strasbourg, Council of Europe Press. (Nature and environment,
No.6. )
Van Driesche, R.G. 1994. Classical biological control of environmental pests. Florida Entomology, 77: 20-33.
Vane-Wright, R.I., Humphries, C.J. & Williams, P.H. 1991. What to protect?- Systematics and the agony of
choice. Biological Conservation, 55: 235-254.
Van Hulst, R. 1987. Invasion models of vegetation dynamics. Vegetatio, 69: 123-131.
Vansplunder, I., Coops, H. & Schoor, M.M. 1994. Tackling the bank erosion - problem - (re)introduction of
willows on riverbanks. Water Science and Technology, 29: 379-381.
Van Wilgen, B.W. & Richardson, D.M. 1985. The effects of alien shrub invasions on vegetation structure and
fire behaviour in South African fynbos shrublands: A simulation study. Journal of Applied Ecology,
22: 955-966.
Veblen, T.T., Mermoz, M., Martin, C. & Kitzberger, T. 1992. Ecological impacts of introduced animals in
Nahuel Huapi national park, Argentina. Conservation Biology, 6: 71-83.
Velkov, V.V. 1994. Introduction of genetically-modified microorganisms into the environment - prospects and
risk. Genetika, 30: 581-592.
Vickery, J. 1994. Genetic biodiversity: analysing the data. Trends in Ecology & Evolution, 9,
Vidano, C. 1983. Actual and potential insect pests of Robinia pseudoacacia. L'Apicoltore Moderno, 74: 181185.
Villwock, W. 1993. The lake Titicaca region in the Peruvian and Bolivian highlands and the consequences of
introducing fishes for endemic Ichthyfauna and their natural habitat. Naturwissenschaften, 80: 1-8.
Vinicombe, K., Marchant, J. & Knox, A. 1993. Review of status and categorisation of feral birds on the British
List. British Birds, 86: 605-614.
Virk, P.S., Fordlloyd, B.V., Jackson, M.T. & Newbury, H.J. 1995. Use of RAPD for the study of diversity
within plant germplasm collections. Heredity, 74, Part 2: 170-179.
Vitousek, P.M. 1986. Biological invasions and ecosystem properties: Can species make a difference? In:
Ecology of biological invasions of North America and Hawaii, ed by H.A. Mooney & J.A. Drake, Vol.
58: 163-176. Berlin, Spinger-Verlag.
Vitousek, P.M. 1990. Biological invasions and ecosystem processes: Towards an integration of population
biology and ecosystem studies. Oikos, 57: 7-13.
Vitousek, P.M., Loope, L.L. & Stone, C.P. 1987. Introduced species in Hawaii - biological effects and
opportunities for ecological research. Trends in Ecology & Evolution, 2: 224-227.
Vlok, J.H.J. 1988. Alpha diversity of lowland fynbos herbs at various levels of infestation by alien annuals.
South Africa Journal of Botany, 54: 623-627.
Review of information, policy and legislation on species translocations
Von Broembsen, S.L. 1989. Invasions of natural ecosystems by plant pathogens. In: Biological invasions, ed by
J.A. Drake, H.A. Mooney, F. di Castri, M. Rejmánek & M. Williamson, 77-83. Chichester, John Wiley
& Sons.
Wali, M.K., ed. 1992. Ecosystem rehabilitation. Vol. 2: Ecosystem analysis and synthesis. SPB Academic
Walker, P.R. & Vitousek, P.M. 1991. An invader alters germination and growth of a native dominant tree in
Hawaii. Ecology, 72: 1449-1455.
Wallace, M.P. 1994. Control of behavioral development in the context of reintroduction programs for birds. Zoo
Biology, 13: 491-499.
Walls, R.M. 1993. Plant translocations - for. BSBI News, 63: 19-20.
Walls, S.S. & Kenward, R.E. 1995. Experimental releases of Buzzards in 1994. Wareham, Institute of
Terrestrial Ecology.
Walters, M. 1993. Wild and garden plants. London, The New Naturalist, Harper Collins.
Wanink, J.H. & Goudswaard, K. 1994. Effects of Nile perch Lates niloticus into Lake Victoria, East Africa, on a
diet of pied kingfishers Ceryle rudis. Hydrobiologia, 280: 367-376.
Waples, R.S. & Do, C. 1994. Genetic risk associated with supplementation of Pacific salmonids: Captive
broodstock programs. Canadian Journal of Fisheries and Aquatic Sciences, 51: 310-329.
Ward, L.K. & Stevenson, M.J. 1994a. Botanical monitoring of translocated turf and seeded areas on the
Arethusa Compensation Area, Summer 1994. Unpublished ITE Report to Mott McDonald Civil Ltd.
Ward, L.K. & Stevenson, M.J. 1994b. Botanical monitoring: Flood meadow translocation and restoration
1993. Unpublished ITE Report to Mott McDonald Civil Ltd.
Ward, L.K. & Stevenson, M.J. 1995. Botanical monitoring: Flood meadow translocation and restoration 1994.
Unpublished ITE Report to Mott McDonald Civil Ltd.
Ward, M. 1993. Do UK regulations of GMOs hamper industry. Bio Technology, 11: 1213.
Ward, M. 1994. EU plans to streamline GMO regulations. Bio Technology, 12: 864.
Waring, G.H., Loope, L.L. & Medeiros, A.C. 1993. Study on use of alien versus native plants by nectarivorous
forest birds on Maui, Hawaii. Auk, 110: 917-920.
Warren, M.S. 1993. A review of butterfly conservation in central southern Britain: I. Protection, evaluation and
extinction on prime sites. Biological Conservation, 64: 25-35.
Warren, M.S. 1995. Re-establishment plan for bringing the chequered skipper back to England, 1995-8.
Wareham, Butterfly Conservation.
Wathern, P.G. 1978. Artificial diversification of grasslands with native herbs. Journal of Environmental
Management, 7: 29-42.
Watson, J., Warman, C., Todd, D. & Laboudallon, V. 1992. The Seychelles magpie robin Copsychus
sechellarum: Ecology and conservation of an endangered species. Biological Conservation, 61: 93-106.
Watt, A.D., Leather, S.R. & Evans, H.F. 1991. Outbreaks of the pine beauty moth on pine in Scotland influence on host plant-species and site factors. Forest Ecology and Management, 39: 211-221.
Webb, D.A. 1985. What are the criteria for presuming native status? Watsonia, 15: 231-236.
Webb, J.H., Hay, D.W., Cunningham, P.D. & Youngson, A.F. 1991. The spawning behaviour of escaped
farmed and wild Atlantic salmon (Salmo salar L.) in a northern Scottish river. Aquaculture, 98: 97-100.
Webb, N.R. 1989. Studies on the invertebrate fauna of fragmented heathland in Dorset, UK, and the
implications for conservation. Biological Conservation, 47: 153-165.
Webb, N.R. & Hopkins, P.J. 1984. Invertebrate density on fragmented heathlands. Journal of Applied Ecology,
21: 921-933.
Webb, N.R. & Vermaat, L. 1990. Changes in vegetational diversity on remnant heathland fragments. Biological
Conservation, 53: 253-264.
Webster, S. & Felton, M. 1993. Targeting for nature conservation in agricultural policy. Land Use Policy, 10:
Wein, R.W., Wein, G., Bekret, S. & Cody, W.J. 1992. Northward invading non-vascular plant species in and
adjacent to wood, Buffalo National Park, Canada. Canadian Field - Naturalist, 106: 216-224.
Weiss, P.W. 1986. Invasion of coastal plant communities by Chrysanthemoides monilifera. In: Ecology of
biological invasions, ed by R.H. Groves & J.J. Burdon, 162. Cambridge, Cambridge University Press.
Welch, R.C. 1981. Insects on exotic broadleaved trees of the Fagaceae, namely Quercus borealis and species of
Nothofagus. In: Forest and woodland ecology: an account of research being done in ITE, ed by
F.T.Last & A.S.Gardiner, 110-115. (ITE symposium no.8). Cambridge, Institute of Terrestrial Ecology.
Welch, R.C. 1994. Rhynchaenus quercus (L.) and R. fagi (L.) (Col., Curculionidae) mining introduced oaks in
Britain. Entomologists Monthly Magazine, 130: 49-54.
Review of information, policy and legislation on species translocations
Welcome, R.L. 1988. International introductions of inland aquatic species. Fisheries technical paper. Rome,
Welling, C.H. & Becker, R.L. 1990. Seed bank dynamics of Lythrum salicaria L.: Implications for control of
this species in North America. Aquatic Botany, 38: 303-309.
Wells, G.A.H., Keymer, I.F. & Barnett, K.C. 1989. Suspected Aleutian disease in a wild otter (Lutra lutra).
Veterinary Record, 125: 232-235.
Wells, J.D. & Greenberg, B. 1994. Resource by an introduced and native carrion flies. Oecologia, 99: 181-187.
Wells, M.J., Poynton, R.J., Balsinhas, A.A., Musil, K.J., Joffe, H., van-Hoepen, E. & Abbott, S.K. 1986. The
history of introduction of invasive alien plants to southern Africa. In: The ecology and management of
biological invasions in Southern Africa, ed by I.A.W. Macdonald, F.J. Kruger & A.A. Ferrar, 21-35.
Cape Town, Oxford University Press.
Wells, T.C.E. 1989. The re-creation of grassland habitats. Entomologist, 108: 97-108.
Wells, T.C.E. 1990. Establishing chalk grassland on previosly arable land using seed mixtures. In: Calcareous
grasslands - ecology and management, ed by S.H. Hillier, D.W.H. Walton & D.H. Wells, 169-179.
Huntingdon, Bluntisham Books.
Wells, T.C.E., Bell, S.A. & Frost, A. 1981. Creating attractive grasslands using native plant species.
Shrewsbury, Nature Conservancy Council.
Wells, T.C.E., Cox, R. & Frost, A. 1989. Diversifying grasslands by introducing seed and transplants into
existing vegetation. In: Biological habitat reconsruction, ed by Buckley, 283-298. London, Belhaven
Wells, T.C.E., Sheail, J., Ball, D.F. & Ward, L.K. 1976. Ecological studies on the Portland Ranges. Journal of
Ecology, 64: 568-626.
Westbrooks, R.G. 1991. Plant protection issues .1. A commentary on new weeds in the United States. Weed
Technology, 5: 232-237.
Westman, W.E. 1990. Park management of exotic plant species: Problems and issues. Conservation Biology, 4:
Whalley, P. 1989. Principles and outcome of introductions. Entomologist., 108: 69-76.
White, M. (1995): Response to Butterfly Conservations' proposal to add RDB and notable butterflys to WCA,
Schedule 9. Letter to Invertebrate Conservation News.
Whitehead, D.R. & Wheeler, A.G., Jr. 1990. What is an immigrant arthropod? Annals of the Entomological
Society of America, 83: 9-14.
Whittaker, R.H. 1972. Evolution and measurement of species diversity. Taxon, 21: 213-251.
Whitten, A.J. 1990. Recovery: a proposed programme for Britain's protected species. Nature Conservancy
Council, CSD report.
Wilcove, D.S. 1987. Recall to the wild: Wolf reintroduction in Europe and North America. Trends in Ecology &
Evolution, 2: 146-147.
Williams, C.S. 1995. Conserving Europes bees: why all the buzz? Trends in Ecology & Evolution, 10: 309-310.
Williams, G., Holmes, J. & Kirby, J. 1995. Action plans for United Kingdom and European rare, threatened and
internationally important birds. Ibis, 137, S209-S213.
Williamson, M. 1989. Mathematical modes of invasions. In: Biological invasions: a global perspective, ed by
J.A. Drake, H.A. Mooney, F. di Castri, R.H. Groves, F.J. Krugef, M. Rejmánek & M. Williamson, 329350. Chichester, John Wiley & Sons.
Williamson, M. 1992. Environmental risk from the release of genetically modified organisms (GMOs) - the
need for molecular ecology. Molecular Ecology, 1: 3-8.
Williamson, M. 1993a. Invaders, weeds and the risk from genetically manipulated organisms. Experentia, 49:
Williamson, M. 1993b. Risks from the release of GMOs - ecological and evolutionary considerations.
Evolutionary Trends in Plants, 1: 5-9.
Williamson, M. 1994. Community response to transgenic plant release - predictions from British-experience of
invasive plants and feral crop plants. Molecular Ecology, 3: 75-79.
Williamson, M. & Fitter, A. 1995a. The varying success of invaders. Ecology, in press.
Williamson, M. & Fitter, A. 1995b. The characters of successful invaders. Biological Conservation, in press.
Wilson, C.G. 1989. Post-dispersal seed predation of an exotic weed, Mimosa pigra L., in the Northern Territory.
Australian Journal of Ecology, 14: 235-240.
Wilson, E.O. 1986. The current state of biological diversity. In: Biodiversity, ed by E.O. Wilson, 3-18.
Washington, National Academy Press.
Wilson, G.B., Whittigton, W.J. & Humphries, R.N. 1995. Biological flora of the British Isles, Potentilla
rupestris. Journal of Ecology, 83, 335-343.
Review of information, policy and legislation on species translocations
Wilson, J.B. 1989. Relations between native and exotic plant guilds in the Upper Clutha, New Zealand. Journal
of Ecology, 77: 223-235.
Wilson, J.B., Rapson, G.L., Sykes, M.T., Watkins, A.J. & Williams, P.A. 1992. Distributions and climatic
correlations of some exotic species along roadsides in South Island, New Zealand. Journal of
Biogeography, 19: 183-194.
Wilson, P. 1993. Plant translocations - against. BSBI News, 63: 20.
Winder, F.L.R. & Robertson, H.J. 1993. Progress reports on monitoring of grassland transplant sites.
Brampton Meadow 1987-1991. Peterborough, English Nature.
Windsor, M.L. & Hutchinson, P. 1990. The potential interactions between salmon aquaculture and the
wildstocks - a review. Fisheries Research, 10: 163-176.
Wingate, P.J. 1991. United States view and regulations on fish introductions. Canadian Journal of Fisheries and
Aquatic Sciences, 48: 167-170.
Witkowski, E.T.F. 1994. Growth of seedlings of the invasives, Acacia saligna and Acacia cyclops, in relation to
soil phosphorus. Australian Journal of Ecology, 19: 290-296.
Wolff, P.L. & Seal, U.S. 1993. Implications of infectious-disease for captive propagation and reintroduction of
threatened species. Journal of Zoo and Wildlife Medicine, 24: 229-230.
Wood, G.W., Mengak, M.T. & Murphy, M. 1987. Ecological importance of feral ungulates at Shackleford
Banks, North Carolina. American Midland Naturalist, 118: 236-244.
Woodroffe, G.L., Lawton, J.H. & Davidson, W.L. 1990. The impact of feral mink Mustela vison on water voles
Arvicola terrestris in the North Yorkshire Moors National Park. Biological Conservation, 51: 49-62.
Woodruff, D.S. & Gall, G.A.E. 1992. Genetics and conservation. Agriculture, Ecosystems and Environment.,
42: 53-73.
Woods, R.G. 1994. British wild species and varieties in gardens. In: The Common Ground of Wild and
Cultivated plants, ed by A.R. Perry & R. Gwynn Ellis, 7-16. Cardiff, National Museum of Wales.
Woolf, N.B. 1986. New hope for exotic species. Bioscience, 36: 594-597.
Worldwide Fund for Nature. 1994. WWF position paper. The conference of the parties to the Convention on
Biological Diversity. First meeting, Nassau, The Bahamas 1994.
Worldwide Fund for Nature. 1995. Biosafety and use of genetically egineered organisms. WWF Position
Worthington, T.R. & Helliwell, D.R. 1987. Transferance of semi-natural grassland and marshland onto newly
created landfill. Biological Conservation, 41: 301-313.
Wright, F.J., Galbraith, C.A. & Bendall, R., eds. 1993. Action for biodiversity in the UK. Peterborough, Joint
Nature Conservation Committee.
Yalden, D.W. 1986. Opportunities for reintroducing British mammals. Mammal Review, 16: 53-63.
Yaninek, J.S. 1988. Continental dispersal of the cassava green mite, an exotic pest in Africa, and implications
for biological control. Experimental and Applied Acarology, 4: 211-224.
Yensen, E. & Quinney, D.L. 1992. Can Townsend ground squirrels survive on a diet of exotic annuals. Great
Basin Naturalist, 52: 269-277.
Yolande, J. 1993. A bibliographical study of gene flow between crops and wild relatives in Switerland. In: Gene
transfer: are wild species in danger? ed. by J. Yolande, K. Ammann & F. Pythoud, 14-18. Bern,
Federal Office of Environment, Forests and Landscape.
Zimen, E. 1976. Wolf reintroduction: suitable areas and techniques. In: Reintroductions: techniques and ethics,
ed by L. Boitani, 151-161. Rome, World Wildlife Fund.
Zimmermann, H.G. 1991. Biological-control of mesquite, Prosopis spp, in South Africa. Agriculture,
Ecosystems and Environment, 37: 175-186.