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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 1996 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 Contents SUMMARY ..............................................................................................................................5 1. INTRODUCTION........................................................................................................7 1.1 Background ........................................................................................................7 1.2 Methodology and structure ................................................................................7 1.3 Development of workable definitions................................................................8 1.4 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 1.5 Scientific background ......................................................................................21 1.5.1 Genetics................................................................................................22 1.5.2 Ecology ................................................................................................24 1.6 Background to the main legislation concerning translocations………………26 2. TRANSLOCATIONS OF SPECIES NATIVE TO THE UK ................................27 2.1 Introduction......................................................................................................27 2.1.1 Background ..........................................................................................27 2.1.2 Types of translocations of native species.............................................28 2.2 Impacts of translocations of native species and factors affecting their success ..........................................................................................................................28 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 2.3 Summary of guidelines ....................................................................................33 2.3.1 UK Guidelines .....................................................................................33 2.3.2 International Guidelines.......................................................................43 2.4 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 2.5 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 1 Review of information, policy and legislation on species translocations 2.6 3. 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 2.7 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 2.8 Summary conclusions concerning the translocation of species native to the UK ........................................................................................................................104 INTRODUCTION OF SPECIES NOT NATIVE TO THE UK..........................106 3.1 Introduction...................................................................................................106 3.1.1 Background ........................................................................................106 3.1.2 Types of introduction.........................................................................106 3.2 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 3.3 Summary of guidelines ..................................................................................109 3.3.1 UK Guidelines ...............................................................................................110 3.3.2 International Guidelines.................................................................................112 3.4 Summary of legislation ..................................................................................117 3.4.2 Import, keeping, release and control of alien species - UK legislation ... ............................................................................................................118 3.4.3 Releases into protected areas .............................................................123 3.5 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 3.7 Wildfowl and game stocking .........................................................................133 3.7.1 Background ........................................................................................133 3.7.2 Impacts of introductions ....................................................................135 3.7.3 Discussion ..........................................................................................137 3.8 Amenity and ornamental planting, stocking or collections ...........................138 3.8.1 Background ........................................................................................138 3.8.2 Effects of introductions......................................................................141 3.9 Pets and domestic animals .............................................................................145 3.9.1 Background ........................................................................................145 3.9.2 Effects of introductions......................................................................145 2 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 3.11 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 ........................................................................................................................173 INTRODUCTION OF GENETICALLY MODIFIED ORGANISMS...............175 4.1 Introduction....................................................................................................175 4.1.1. Background ........................................................................................175 4.1.2 Genetic modification..........................................................................176 4.1.3 Types and uses of GMOs......................................................................177 4.2 Summary of guidelines and policies ..............................................................178 4.2.1 UK Guidelines ...................................................................................179 4.2.2 International Guidelines.....................................................................179 4.3 Summary of legislation ..................................................................................181 4.3.1 International Conventions and European Legislation........................181 4.3.2 UK Legislation - Environmental Protection Act 1990 ......................181 4.4 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 4.5 Discussion ......................................................................................................198 4.5.1 Does genetic modification pose different risks to conventional breeding?........................................................................................................198 4.5.2 Risk assessment for GMO release .....................................................199 4.5.3 Involvement of nature conservation bodies in GMO assessment......200 3.10 4. 3 Review of information, policy and legislation on species translocations 4.6 5. 6. 7. 8. Summary conclusions concerning introduction of genetically modified organisms .......................................................................................................201 TRANSLOCATION OF SPECIES ASSEMBLAGES .........................................203 5.1 Introduction....................................................................................................203 5.2 Summary of guidelines and policies ..............................................................204 5.3 Summary of legislation ..................................................................................205 5.4 Case studies....................................................................................................206 5.5 Consequences of community translocation ...................................................215 5.5.1 Effects of species translocation..........................................................215 5.5.2 Effects particular to translocations of species assemblages..............216 5.6 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 5.7 Summary conclusions concerning translocation of species assemblages......236 GENERAL CONCLUSIONS OF THE REVIEW................................................238 6.1 Translocations and their effects on biodiversity ............................................238 6.2 Costs of translocations ...................................................................................241 6.3 Monitoring and databases ..............................................................................242 6.4 Assessments of risks and of benefits - formats for guidelines.......................243 6.5 Co-ordination among organisations...............................................................244 6.6 A European framework of legislation............................................................246 6.7 Education and legislation...............................................................................247 6.8 Summary general conclusions of the review .................................................247 LIST OF CONTACTS AND SOURCES OF INFORMATION..........................249 BIBLIOGRAPHY ....................................................................................................253 4 Review of information, policy and legislation on species translocations SUMMARY y 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). y 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. y Non-conservation translocations for commercial, amenity and aesthetic reasons are carried out using native and non-native species, and genetically modified organisms. y Precise definitions of the different types of translocation are given, as well as other important terms used in the review. y 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. y 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. y 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. y 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 studies. y 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. y 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. y Some areas require a revision or coordination of the approach to the regulatory process: GMO releases, translocation of species assemblages, release of non-native species. 5 Review of information, policy and legislation on species translocations y Areas requiring further research are highlighted, and these illustrate a need for more coordinated and structured monitoring and databasing of current and future translocations. y 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. y 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. 6 Review of information, policy and legislation on species translocations 1. INTRODUCTION 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 lacking. 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 knowledge. 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 7 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. 8 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. 9 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 present. 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 conspecifics. 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 Organism 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. Genotype 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. 10 Review of information, policy and legislation on species translocations Population 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 translocations. Race 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). 11 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 Extinct 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. Native 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 species. A native population is one which occurs within the particular area. Synonym: Indigenous. Non-native 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. Naturalised A non-native species or race which, after escape or release, has become established in the wild in self-maintained populations. Feral 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 organisms. Endemic 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 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 13 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 disperse. A wild animal, plant or other organism is one occurring in the wild. Introduction 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. Restocking 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. 14 Review of information, policy and legislation on species translocations Relocation 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 supplementation. Synonym: Rescue translocation. Establishment 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. 1.4 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 15 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 1988). 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. 16 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 areas; d) the biodiversity of natural and semi-natural habitats where this has been diminished over recent past decades. Biodiversity Challenge objectives (p8) 1 2 3 4 5 6 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'. 17 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 1993). 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 1 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. 18 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). y y y y y 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 19 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 classification. 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 Reserves. 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 biodiversity. 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. 2 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. 20 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 Region 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. 3 4 5 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. 21 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 species. 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. 22 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 23 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 24 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 25 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). 1.6 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 review'). 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 parties. 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. 26 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. TRANSLOCATIONS OF SPECIES NATIVE TO THE UK 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 27 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 Gamebirds Falconry Bee-keeping 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. 28 Review of information, policy and legislation on species translocations 2.2.1 Ecological impacts Deleterious 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 species. 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 loss. Beneficial 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 29 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. 30 Review of information, policy and legislation on species translocations 2.2.3 Genetic impacts Deleterious 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. Beneficial 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. 31 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. 32 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. 33 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: y y y Reintroductions, relocations and supplementations, when carried out for reasons of nature conservation, can have a valid role and should be encouraged where appropriate. 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. Reintroductions Assessment of a proposed reintroduction should consider the following. y y y y y y y y y 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 corrected? 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 following. y y 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 ; 34 Review of information, policy and legislation on species translocations y 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: y y y 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: y y y 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 Relocation y y y 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 35 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). Reintroduction Each of the guidelines on reintroductions and single-species relocations was also seen in the later NCC (1990) document, with one important exception. y All reintroduced species should have protected status, at least in the early stages of the reintroduction. Re-enforcement Extra guidelines were suggested for re-enforcements, which were not all stated explicitly in the NCC (1990) document. y y y y 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 y 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 y 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. 36 Review of information, policy and legislation on species translocations 4) Society for the Promotion of Nature Reserves (1970). A policy on introductions to nature reserves 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. y y y y 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. y y y 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. y y 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. y y 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. 37 Review of information, policy and legislation on species translocations 7) Botanical Society of the British Isles (1991).Guidelines for the transfer of rare vascular plants 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 y y 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 y y y y y y y 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 site. 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, y insect reintroduction should be particularly of nationally threatened species. 38 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: y y y y y y y y 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, y 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: y y 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. y y 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. 39 Review of information, policy and legislation on species translocations This lead to a general policy. y y y 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. y y 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. y y y y y y y y y y y y y y 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: y y assisting natural dispersal; spreading the risk of local extinctions. 40 Review of information, policy and legislation on species translocations Criteria for project acceptability are very similar to the BBCS (1995) guidelines, requiring: y y y y y y y y y y y y 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; monitoring; a programme plan; contingency plans against failure or effects on the donor population; y 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 y '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. 41 Review of information, policy and legislation on species translocations This also accepts translocations without a comprehensive knowledge of the species' distribution. 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) guidelines. The RSPB see reintroductions as a potentially valuable tool for: y y y y 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. y y y y y y y y The species became extinct largely through human activities. Suitable habitats are available and there are not likely to be serious effects on other species. 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. y 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. 42 Review of information, policy and legislation on species translocations y y 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. y y y y y 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: y y 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; 43 Review of information, policy and legislation on species translocations y artificially high rates of immigration are required to maintain outbreeding between isolated populations on biogeographical islands. Re-enforcement should only be considered where: y y y 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. y y 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. y y y y y 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 y y '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'. 44 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 y y 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 y 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 y y y y y 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. 45 Review of information, policy and legislation on species translocations ii. Previous reintroductions. y Previous reintroductions of the same or similar species should be used to inform the reintroduction protocol. iii. Choice of release site. y 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 y y 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 y y y y y y 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 y 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 y y 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. 46 Review of information, policy and legislation on species translocations y y 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 y y y y y y y y y y y 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 duration. 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, timing). 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: y y y y y y y y y y post release monitoring - direct (e.g. tagging, telemetry) or indirect (e.g. spoor, informants); 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 necessary; 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. 47 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. y y 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. y 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 Introductions. 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: y translocation of Atlantic salmon poses genetic, ecological, disease and parasite risks to the wild Atlantic salmon. A number of recommendations were made. y y 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. 48 Review of information, policy and legislation on species translocations y y 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, disease. 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: y y wild flower seeds should never be scattered in the countryside (i.e. outside of gardens); 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 diversity This report, produced by Plantlife and endorsed by EN, produced no guidelines, but it encouraged: y y y 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: y 'indigenous seeds, where practical of British origin.' The sowing of certain grass species is mandatory and y 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 form The countryside stewardship guidelines for recreating grassland on arable land stated that: y where possible this should be by natural regeneration; 49 Review of information, policy and legislation on species translocations y y 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: y y 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: y y 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 Kingdom This advises on species to be used in the MAFF Farm Woodland Scheme. It recommends that: y y 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. 50 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. 51 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. 52 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 53 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 54 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 Appendices. 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. 55 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 bureaucracy. 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 approach. 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 y y y y y y 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.). 56 Review of information, policy and legislation on species translocations y y y y Will new species be introduced? 1. Will the species have a deleterious effect on existing flora or fauna? 2. 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. 3. 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). 4. 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? 1. 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. 2. Is there likely to be increased predation on invertebrates? 3. Is there likely to be competition for food with other species? 4. 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 57 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 58 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 necessary. 2.5.2 Reintroductions of species extinct in the UK 2.5.2.1 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 59 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. 60 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. 2.5.2.2 Impacts of reintroductions of species extinct in the UK and factors affecting their success Ecological impacts Deleterious 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. 61 Review of information, policy and legislation on species translocations Beneficial 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 62 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 stock. 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 1994b). 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 63 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 2.5.3.1 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 64 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 Ecology. 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. 65 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 Zoo). 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. 66 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. 2.5.3.2 Impacts of reintroductions of species regionally or locally extinct in the UK and factors affecting their success Ecological impacts Deleterious 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). Beneficial 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. 67 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 68 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. 69 Review of information, policy and legislation on species translocations 2.5.4 Re-enforcement of existing populations 2.5.4.1 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 report). 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 70 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 (Kew). 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. 2.5.4.2 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 71 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 minimised. 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 72 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. 73 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 2.5.6.1 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 1994). Relocations to rescue badgers are considered in section 2.6.6 'Releases for animal welfare'. 2.5.6.2 Impacts of rescue translocations and factors affecting their success 74 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 knowledge. 75 Review of information, policy and legislation on species translocations 2.5.7 Conservation seed mixes and plantings 2.5.7.1 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 76 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. 2.5.7.2 Impacts of the use of conservation seed mixes and plantings and factors affecting its success 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. 77 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. Beneficial 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 78 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): y y y y 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 species; crossing between native and introduced plants, leading to erosion of native genetic variation. 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 79 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 provenance. 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 80 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. Beneficial 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: y y y y y y y y y y y y 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; 81 Review of information, policy and legislation on species translocations y y y y y y 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 pollinators; 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 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. 82 Review of information, policy and legislation on species translocations 2.6 Translocations of native species for purposes other than conservation 2.6.1 Fisheries and angling 2.6.1.1 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 83 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 parasite. 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. 84 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 1995). 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. 2.6.1.3 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 85 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 conservation. 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 them. 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 86 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. 1991). 2.6.2 Crustacea and molluscs in aquaculture 2.6.2.1 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. 2.6.2.2 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.). 87 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 2.6.3.1 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. 2.6.3.2 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 88 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 2.6.4.1 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). 2.6.4.2 Impacts of translocations and factors relating to management 89 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 2.6.5.1 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 honey-bee. 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. 2.6.5.2 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 90 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 translocations. 2.6.6. Translocations and releases for aesthetic purposes 2.6.6.1 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 91 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). 2.6.6.2 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 92 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. 2.6.6.3 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 biodiversity. 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 93 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 relatives. 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 2.6.8.1 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. 94 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 2.7.1.1 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 95 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. 2.7.1.2 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. 96 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. 2.7.1.3 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 guidelines. 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. 2.7.1.4 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 97 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. 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 98 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. 99 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. 1. 2. 3. 4. 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 programmes? 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. 2.7.1.5 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 useful. 100 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. 2.7.1.6 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 butterfly. 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. 2.7.1.7 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 101 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. 2.7.1.8 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. 2.7.2 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 102 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 2.6.1). 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 whole. 2. Conservation interests concerned with the possible impacts of translocations on biodiversity. 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 up. 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. 103 Review of information, policy and legislation on species translocations 2.8 Summary conclusions concerning the translocation of species native to the UK y Types of translocation of native species for conservation are reintroductions, reenforcements, benign introductions, relocations and seed-sowings and plantings for habitat restoration. y 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). y 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. y 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. y 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. y 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. y Provision of sufficient and appropriate habitat is usually well-studied in conservation translocations and is achieved by restoration and/or management. y 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. y 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. 104 Review of information, policy and legislation on species translocations y 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. y 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. y 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. y 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. y 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). y 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. 105 Review of information, policy and legislation on species translocations 3. INTRODUCTION OF SPECIES NOT NATIVE TO THE UK 3.1 Introduction 3.1.1 Background 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). 3.1.2 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 106 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 Forestry Crops Fur animals Accidental introductions 3.2 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 species. 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 107 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 Competition 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. Herbivory 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. Predation 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. 108 Review of information, policy and legislation on species translocations Prey 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. 3.2.3 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. 3.3 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. 109 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 y 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 races. 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 y '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. y y y y y y y 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 individuals? Will the release be monitored? Procedures y 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. 110 Review of information, policy and legislation on species translocations y y 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. y y 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. 111 Review of information, policy and legislation on species translocations Translocations of British natives y 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 reasonable'. Translocations of species not native to Britain The reasons for introductions are: y y y for general interest as an experimental activity; to enrich the British fauna, which is poor in herpetofauna; for commercial purposes. Types of introduction allowed. y y y Non-European species should not be introduced - they may compete with native species. 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 5) RSPB The RSPB simply opposes all artificial introductions of non-native bird species (G. Williams pers. comm.). 3.3.2 International Guidelines 6) IUCN. (1987). The IUCN position statement on translocation of living organisms The basis of these guidelines was that: y '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. 112 Review of information, policy and legislation on species translocations Intentional introduction of non-native species Such introductions: y y y y y should only be considered if clear benefits to humans or natural communities can be foreseen; 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 y y y 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 environment; 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. y The same stock should be used as that to be introduced extensively. 113 Review of information, policy and legislation on species translocations y y The organisms should be free of diseases or parasites that could spread to other species. 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 y The extensive introduction should be closely monitored and arrangements made to restrict, control, or eradicate the species if necessary. Accidental introductions y y Accidental introductions should be discouraged where possible, with emphasis on the following. 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 humans. 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 y y y 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 y As this may involve introduction of alien species, the same care should be taken as for other intentional introductions. 114 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. y y Accidental coincident translocation of other species. Harmful ecological effects of the translocated species. y 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. y y y y y y y y 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 environment. 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; predators; competitors; 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. y y y y y y 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. 115 Review of information, policy and legislation on species translocations Procedures for ongoing translocations that are part of current commercial practice. y Periodic inspection of material prior to translocation - if any pathogens or pests are discovered the translocation must be discontinued. and/or y y 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 organisms 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, y 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: y 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 Government. A number of procedures to prevent translocation are discussed in detail, which involve the following objectives. y y y y 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. 116 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 y y 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. 3.4 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 1992. 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. 117 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 prohibited 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. 118 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: y y y y y y y y y y y 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; 119 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 removed. 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: 120 Review of information, policy and legislation on species translocations y y y y y y y y y 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 121 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 122 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. 3.4.3 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. 3.5 Fish and shellfish stocking for aquaculture 3.5.1 Background 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 occurs. 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. 123 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. 3.5.2 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. 3.5.2.1 Ecological impacts Competition 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). 124 Review of information, policy and legislation on species translocations Predation 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. 125 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. 3.5.2.2 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. 3.5.2.3 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. 126 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. comm.). 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 127 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. y y y y y y y 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 introductions. The following criteria are suggested for a species that is a candidate for an introduction. y y y y 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. 128 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 species. 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 3.6.1 Background 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 129 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. 3.6.2 Impacts of introductions 3.6.2.1 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. y y y y y 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 occurring. 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. 130 Review of information, policy and legislation on species translocations y 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 agents. 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. 3.6.2.2 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 131 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. 3.6.2.3 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. 3.6.2.4 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. 3.6.3 Discussion 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. y y y 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. 132 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: y y y 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 greatly. 3.7 Wildfowl and game stocking 3.7.1 Background 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. 133 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 134 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). 3.7.2 Impacts of introductions 3.7.2.1 Ecological impacts Herbivory 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. Predation 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, 135 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. 3.7.2.2 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 136 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. 3.7.3 Discussion 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. 137 Review of information, policy and legislation on species translocations 3.8 Amenity and ornamental planting, stocking or collections 3.8.1 Background 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). 138 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 psittacosis. 139 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). 140 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. Others 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 UK. 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. 3.8.2 Effects of introductions 3.8.2.1 Ecological impacts Herbivory 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 141 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 vegetation. 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 reproduction. 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). 142 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. 3.8.2.2 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). 3.8.2.3 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 143 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. 3.8.2.4 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 144 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. 3.9 Pets and domestic animals 3.9.1 Background 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. 3.9.2 Effects of introductions 3.9.2.1 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. 145 Review of information, policy and legislation on species translocations Herbivory 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). Predation 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. 3.9.2.2 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 146 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. 3.9.3 Discussion 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 are. 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. 147 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 following. 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 1993). 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. 148 Review of information, policy and legislation on species translocations 3.10.2 Effects of introductions 3.10.2.1 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). 3.10.2.2 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 149 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. 3.11 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 4). 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. 150 Review of information, policy and legislation on species translocations 3.11.2 Effects of introductions 3.11.2.1 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. 3.11.2.2 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 barley. 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). 151 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 significant. Raybould & Gray (1993a) suggest several issues that are in need of further investigation. y y y y Which natural habitats are most invasible. Which modifications would be likely to improve crop species' performance in these habitats. 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 152 Review of information, policy and legislation on species translocations and foxes Vulpes vulpes kill mink, but in insufficient numbers to act as a regulatory mechanism. 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 3.12.2.1 Ecological impacts Herbivory 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 to 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 eutrophication. 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 153 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. Predation 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. Competition 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). 154 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 determined. 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 1995). 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 155 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 3.13.2.1 Ecological impacts Predation 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 156 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. Competition 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 157 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. 3.13.2.2 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 158 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 159 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. 160 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 161 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 162 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 invaders. 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 163 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 islands. 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 164 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 aliens. Mammals 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 165 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. Birds 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 166 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). Plants 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. 1994). 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 167 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: y y 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 168 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. 169 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. 170 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. y 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). 171 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). y 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. 172 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 y 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. y 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. y 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. y 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. y 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. y 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. y 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. 173 Review of information, policy and legislation on species translocations y 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. y 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. y 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. y 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. y 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. 174 Review of information, policy and legislation on species translocations 4. INTRODUCTION OF GENETICALLY MODIFIED ORGANISMS 4.1 Introduction 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. 175 Review of information, policy and legislation on species translocations 4.1.2 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 organism. 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 Prokaryotes 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. Eukaryotes 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 176 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). Microorganisms 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) Plants 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 177 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 and 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). Animals 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). 4.2 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. 178 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: y y y y 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. 4.2.2 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: y y y y y safety; 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 y '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 governments. 179 Review of information, policy and legislation on species translocations A biosafety protocol would be based on the following. y y y y y y y y y 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: y y y y 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 recommendation. y 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 EU. 180 Review of information, policy and legislation on species translocations 4.3 Summary of legislation This section summarises UK and EC legislation and international conventions which are relevant to controlling releases of GMOs. 4.3.1 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 (OECD/GD(93)92) 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 environment. 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. 4.3.2 UK Legislation - Environmental Protection Act 1990 Background 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 181 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 environment. 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 harmless. 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. 182 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): 183 Review of information, policy and legislation on species translocations Information relating to the organisms y y y y y y 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 microorganisms), 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 organisms; 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; 184 Review of information, policy and legislation on species translocations Conditions of release y y 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 organism, The organisms and the environment y y y y y y y y 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 organisms potential effects on other organisms, likelihood of post-release shifts in biological interactions, involvement in biogeochemical processes; 185 Review of information, policy and legislation on species translocations Monitoring, etc. y y y y y y y 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.). 4.4 Environmental impacts of GMO introduction 4.4.1 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 186 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. y y y y y y y 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 harvesting). 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 volunteers. 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 material. 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: y y baculovirus (one release)- as a biopesticide; bacteria (two releases) - both with marker genes only; 187 Review of information, policy and legislation on species translocations y y y y y 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 resistance); 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: y y y y y y y y y y 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 188 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. 189 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. 4.4.3.1 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 190 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 191 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 192 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 193 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. 4.4.3.2 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. Plants 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 194 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. y y y 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 strawberries. 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. Animals 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. Microorganisms 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, 195 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. y y y 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. 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%). 196 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 arise. 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 197 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). 4.5 Discussion 4.5.1 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. y 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. 198 Review of information, policy and legislation on species translocations y y y y y y y There is a poor understanding of the extent and consequences of gene flow in microorganisms. 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 species. 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: y y 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; 199 Review of information, policy and legislation on species translocations y y 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 products; 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, 200 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 agencies. 4.6 Summary conclusions concerning introduction of genetically modified organisms y 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. y Current releases in the UK are very small-scale and experimental, with strict protocols to prevent escape. y 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. y 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. y 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. y 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. 201 Review of information, policy and legislation on species translocations y Certain genetic modifications may increase the likelihood of such effects, e.g. changed growth rate, pest resistance or disease resistance. y Unexpected or unstable effects of the modification on the phenotype may lead to unpredictable changes in the ecology of the species. y 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. y 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 greater. y 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. y 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. y 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. y 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 agencies. y 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 abroad. 202 Review of information, policy and legislation on species translocations 5. TRANSLOCATION OF SPECIES ASSEMBLAGES 5.1 Introduction 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. 203 Review of information, policy and legislation on species translocations 5.2 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. y y y y 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 conservation. 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. 204 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: y y suitable vegetation types for turfing; appropriate methods of turf lifting, storage and re-laying. 5.3 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 205 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. 5.4 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 Background 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 206 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). Translocation 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. Management 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 site. Monitoring 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). 207 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 Background 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.). Translocation 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. Management 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. Monitoring 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. 208 Review of information, policy and legislation on species translocations Middlebere Heath, Dorset Background 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. Translocation 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. Management 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. Monitoring 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 Background 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 209 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). Translocation 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. Management The donor site was traditionally lightly grazed by cattle, but the translocated turves were managed only by occasional strimming over the subsequent four years. Monitoring The donor site was surveyed and the translocated turves were monitored botanically for four years. 'Spreading' 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. Background 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). 210 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. Translocation 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 Management 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 spring. Monitoring 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'. 'Blading' 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 Background 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. 211 Review of information, policy and legislation on species translocations Translocation 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 years Management 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. Monitoring 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 Background 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. Translocation 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). Management The management of neither the donor or the receptor sites are described by Piekarczyk (1991). 212 Review of information, policy and legislation on species translocations Monitoring 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 Background 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). Translocation 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 Management of the grassland before the repairs consisted of occasional cutting. This continued uninterrupted after the repairs. Monitoring 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 Background 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 site. 213 Review of information, policy and legislation on species translocations Translocation 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 cohesion. Management 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). Monitoring 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. 'Blading' 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 Background 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. 214 Review of information, policy and legislation on species translocations Translocation 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. Management 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. Monitoring 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. 5.5 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. 5.5.1 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. 215 Review of information, policy and legislation on species translocations 5.5.2 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 translocation. 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. 5.5.2.1 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 216 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 Dongas 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). Hockley 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 217 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). Middlebere 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). 218 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. Newhall 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). Potatopot 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 219 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. Potatopot 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). Ashcott 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.). 220 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. Dongas 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 221 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. Ashington 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. 5.5.2.2 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 Dongas 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. 222 Review of information, policy and legislation on species translocations Hockley 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. Potatopot 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. Potatopot 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. 223 Review of information, policy and legislation on species translocations Ashcott 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 Dongas 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). Ashington 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). Thrislington 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. 5.5.2.3 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. 224 Review of information, policy and legislation on species translocations 5.5.2.4 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. 5.5.2.5 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 225 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. 5.5.2.6 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. 5.6 Discussion 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, 226 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 species. 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 227 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. Disturbance 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. Climate The dry weather that followed turf translocation at Middlebere was thought to have caused death of some turves. Environment 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 228 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. Management 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 vegetation. 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 229 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 230 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 5.5.2.1) 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 231 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 preservation. 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 232 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. 5.6.3 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). 233 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 234 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. 5.6.4 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. 235 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. 5.7 Summary conclusions concerning translocation of species assemblages y 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. y Turf-lifting is the most common technique used, although soil transfer is an occasional method. y Many translocated communities are of high conservation value, e.g. from SSSIs, and are usually of herbaceous or dwarf shrub vegetation. y 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. y 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. y 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). y 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. y 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 valuable. y 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. y 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. y 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. y 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. y It should be a policy to educate the public and developers: 1. about this best practice; and 2. that translocation cannot preserve a community. 237 Review of information, policy and legislation on species translocations 6. GENERAL CONCLUSIONS OF THE REVIEW 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. 6.1 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. 238 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 researched. Genetic considerations for a successful translocation are widely discussed, but poorly researched. The Plymouth Pear Pyrus cordata example in section 2.5.3.2 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 239 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 2.5.4.2 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 2.5.2.1). 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 240 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 2.5.3.1) 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. 6.2 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 241 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 costed. 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 accounting. 6.3 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 Britain. 242 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 http://ifs.plants.ox.ac.uk/wwd/wwd.htm). 6.4 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 243 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. 6.5 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 sufficient. 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 244 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. 245 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.). 6.6 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). 246 Review of information, policy and legislation on species translocations 6.7 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 include: y y 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 conservation, 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, y y y 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. y y y y y 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. 6.8 Summary general conclusions of the review y 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. y 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 247 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. y 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. y 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. y 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. y 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. y 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. y 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. y 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. y International links would also allow development of responses to common legislation in Europe and to international conventions which are concerned with translocations. y 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. 248 Review of information, policy and legislation on species translocations 7. LIST OF CONTACTS AND SOURCES OF INFORMATION 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. Name John Akeroyd Organisation 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 Richard Crawshaw IUCN Invasive Species Specialist Group National Rivers Authority (Southern Region) Clive Cummins Roger Daniels Maguelonne DejeantPons 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 Committee Joint Nature Conservation Committee Scottish Natural Heritage Oxford Forestry Institute English Nature English Nature National Rivers Authority Department of the Environment European Wildlife Division Institute of Terrestrial Ecology 249 John Cooper Ian Evans Vin Fleming Hugh Frost Tony Gent Mary Gibson Stephen Gledhill Felicity Grant Alan Gray Expertise Plant translocations and use of wildflower seed Crustacea releases SNH policy on freshwater translocations Mammal translocations EN policy on GMOs English legal system IUCN policy and research on captive breeding Badger translocation EN licensing issues under the Wildlife and Countryside Act IUCN policy and research on introductions Licensing under Section 30 of the Salmon and Freshwater Fisheries Act 1975 Herpetofauna translocations Plant translocations Council of Europe policy and legislation on translocations SNH policy on freshwater translocations 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 Division National Rivers Authority Monitoring of translocated fish English Nature EN monitoring of community translocations Joint Nature Conservation Marine introductions Committee Joint Nature Conservation Policy and research on vertebrate Committee translocations Scottish Office Environment Section 16 licensing Department 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 Committee translocations 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 restoration Institute of Terrestrial Ecology Bird and mammal reintroductions Scottish Natural Heritage SNH policy and Species Action Programme English Nature EN monitoring of community translocations English Nature EN monitoring of community translocations Ministry for Agriculture Section 16 licences of fish Fisheries and Food introductions University of Oxford Mammal reintroductions Institute of Zoology IUCN Red List categories IUCN Re-Introduction Specialist IUCN policy and implementation of Group reintroductions 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 Isles 250 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 Organisation National Rivers Authority (Southern Region) Margaret Palmer Joint Nature Conservation Committee Department of the Environment Bill Parish Toxic Substances Division Botanical Society of the British David Pearman Isles 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 Heather Robertson Art Schwartz Rowley Snazell Pritpal Soorae Brian Spencer Mark Stanley Price 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 ballast Licensing under Section 30 of the Salmon and Freshwater Fisheries Act 1975 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 translocations Institute of Terrestrial Ecology Community translocations and habitat restoration National Rivers Authority NRA research on fisheries translocations Institute of Terrestrial Ecology Research on GMOs Charles Blandford Associates Community translocations and plant reintroductions English Nature EN monitoring of community translocations Canadian Circumpolar Institute, Community translocation Canada Institute of Terrestrial Ecology Community translocations and habitat restoration IUCN Re-Introduction Specialist IUCN policy and implementation of Group reintroductions Ministry for Agriculture Shellfish translocations Fisheries and Food IUCN Re-Introduction Specialist IUCN policy and implementation of Group reintroductions 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 Committee translocations Institute of Terrestrial Ecology Butterfly reintroductions Institute of Terrestrial Ecology Community translocations and habitat restoration 251 Review of information, policy and legislation on species translocations Martin Warren Tony Watts British Butterfly Conservation Society Ministry for Agriculture Fisheries and Food Plant Variety Rights Office and Seeds Division 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 Service 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 252 Review of information, policy and legislation on species translocations 8. 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