Download STATUS OF RARE WOODLAND PLANTS AND LICHENS 1.0

STATUS OF RARE WOODLAND PLANTS AND LICHENS
1.0
Introduction
The parlous status of woodland birds and butterflies has long been
recognised and there are examples that show that woodland plants have
also exhibited dramatic levels of decline over the past century. This study
attempts to quantify the losses of rare and threatened woodland plants and
lichens, and identify the changes to which these losses can be attributed. In
particular, it will focus on the threatened and near threatened and Priority
BAP species in the British flora. It also covers widespread species, especially
those of international significance, in Britain, where data was available.
Woodlands are a prominent part of the British landscape and are the natural
vegetation of a great deal of the land below the tree line. Woodland cover
is below 10% in most parts of the UK due to millennia of woodland clearance
and natural climate deterioration. The cover of ancient native woodland is
even lower. The expectation would therefore be that woodland would be a
premier habitat for rare plants in Britain.
Woodland is certainly an important habitat for all groups, but for different
groups the relative importance varies markedly. For vascular plants and
bryophytes, open habitats feature much more strongly than woodland
habitats for our most threatened species. For lichens and fungi, in strong
contrast, woodlands stand out as very rich habitats for rare species. The
distribution of rare species within woods is also very patchy, with average
woods lacking rare species, but with some specific woodland habitats and
types being spectacularly rich.
1
2.0
British woodlands
2.1
Types of woodland
Woodlands have been classified in various ways over the past century by
ecologists and conservationists. These vary from very broad generalisations,
often characterised by the usually dominant tree, such as Upland Oak Wood,
to detailed phytosociological classifications, such as the National Vegetation
Classification (NVC) (Rodwell, 1991) or computer-based stand classifications
(Bunce, 1982).
An intermediate type of classification is the Peterken Stand Type
classification of the over-storey layers in coppiced ancient woodlands
(Peterken, 1993). The broad generalist classifications can be quite
problematic, for instance the interest of west Highland hyper-oceanic
woodlands for epiphytic lichen appears inversely related to the dominance
of oak, making the names Upland Oak Woods or Atlantic Oak Woods,
somewhat problematic (Coppins & Coppins, 2005). Detailed floristic
classifications can emphasise widespread common features, especially in
the ground flora, uniting otherwise rather disparate woodlands.
For example, the NVC unites acid upland temperate zone woodlands, often
with oak prominent and acid boreal birch woodlands, which never have oak,
under the name Quercus petraea – Betula pubescens – Dicranum majus
Woodland (W17). The frequent ground flora species are common between
the two woodland types but the canopy composition and epiphytic lichen
floras are different.
Stand type classifications, such as Peterken’s Stand Type Classification, can
express regional canopy variation in detail but are not comprehensive.
Peterken Stand Types work well in the lowlands but there is a desperate
need for a similar system for the uplands. Unpublished work, to allow
condition assessments of woodland lichens in Scotland, by Brian and Sandy
Coppins, demonstrated that a very useful classification was possible using a
combination of tree and shrub composition, topography and land use, in
contrast to the coppice layer composition and soil characters used in
Peterken. The NVC is certainly over-used in situations where it is too
generalist for the purpose; it expresses shared common features and irons
out local distinctness (Rackham, 2003).
In this study it was found rather difficult to use any of the current widely
used classifications. Broad systems such as the woodland habitats defined
for the Biodiversity Action Plans were sometimes too vague to easily assign
species to them. Equally, full floristic or stand type classifications were too
detailed to use for general conclusions. For this report, a broad but
comprehensive climatic-based classification was created to help understand
the distribution of species of concern within Britain. This was inspired by
the climatic zonations indicated by epiphytic lichen assemblages, which
change in more marked degree than do vascular plants (Coppins & Coppins,
2002).
2
Comment [TCW1]:
Originally published in 1981.
PETERKEN, G. P. 1981
Woodland Conservation and
Management. London,
Chapman and
Hall.
Temperate woodland: native woodland with temperate broadleaved trees
such as oak and ash prominent. Within temperate woodlands the following
species assemblages were recognised. These pick out distinctive regions
within the temperate woodlands:
•
Hyper-oceanic temperate woodland species: confined to areas of
extreme and very wet oceanic climates. These constitute distinctive
assemblages of bryophytes, lichens and ferns, which occur most
extensively in the western Highlands but with outliers in the Lake
District and North Wales (Hodgetts, 1997).
•
Oceanic temperate woodland species: species that extend into, or are
confined to, sunnier southern oceanic areas in the south west. There is
a rich assemblage of southern Atlantic, or more widely distributed
oceanic lichens, a distribution not reflected strongly in other groups of
woodland species.
•
Sub-oceanic temperate woodland species: species confined to drier
eastern areas. These include a sizeable lichen flora, now confined to
clean air areas in Scotland, and a small group of vascular plants, mainly
found in the east of central England.
•
Temperate woodland species: species that are generally found across
the country. This group includes most vascular plants as well as
significant numbers of mosses and lichens.
•
Temperate – boreal woodland species: a number of species occur in
both the temperate and boreal woods, including species with hyperoceanic, sub-oceanic and general temperate distributions.
Temperate woodlands are developed below 200 – 300m in the Scottish
Highlands. The altitudinal limit is unclear in the south, as virtually no
woodland survives beyond the limit of oak here, one of the species that best
defines this climatic zone. Oak woods reach to just over 400m in the Lake
District and Dartmoor and native beech woods grow to up to a similar
altitude in south-east Wales. Within temperate woodlands, there are strong
gradients from north to south and east to west in response to increasing
temperature and rainfall respectively.
Tree assemblages, however, have not reached equilibrium with the climate
(Svenning & Skov, 2007); the slow post-glacial recolonisation of beech and
hornbeam is particularly notable. The case of beech is especially striking;
beech can easily naturalise well north of its apparent native distribution,
and would probable grow at a greater altitude and further north than oak, if
it reached its full potential natural range (Peterken, 1996 & Maxwell, 1929).
Curiously, fossil pollen indicates that it may actually have succeeded in
spreading further north and west in the past but lost ground due to
woodland clearance (Rackham, 2003).
3
On top of natural factors of climate, recolonisation rates and soil
composition, woodland composition is also determined by changes due to
woodland exploitation and management. Small-leaved lime in particular has
lost ground in face of exploitation (Peterken, 1993 & Rackham, 2003). It has
gone from an abundant tree across the lowlands of England and Wales to
uncommon and locally extinct.
On a national scale, most vascular plant composition varies on a north to
south axis, with species numbers dropping off to the north. There are small
assemblages of species that show a west to east gradient, with a few ferns
confined to western oceanic woods and a few sub-oceanic woodland herbs
confined to the south east.
In contrast, bryophytes and lichens in temperate woodlands show a much
stronger east to west variation than vascular plants. Bryophytes and lichens
have exceptionally rich and diverse assemblages of Atlantic species,
dependent on strongly oceanic climates. These reach their apogees in very
wet hyper-oceanic temperate climates. These conditions are most strongly
developed where high mountains rise close to the western seas, as in the
western Highlands, Lake District and North Wales. For lichens, there are
also distinctive assemblages of sub-oceanic species in clean air areas along
the east coast. These are similar to the woodland lichen assemblages of
Denmark and southern Sweden.
Within all temperate woodlands, local variations between woods in
composition mainly relate to soil differences. An important axis is between
strongly acid to strongly base rich soils, and another from wet soils through
to dry soils. Woodlands on strongly acid soils are poorer in vascular plants
species than less acidic soils. The richest woods are usually large woods with
varied soil types.
A typical example would be a wood with both base rich clays and acid sands
with springlines producing areas of wet woodland. Although base rich soils
support species-rich communities, woods on chalk and limestone are rarely
the richest woods overall as they lack acid and wet soils. An additional
factor is longevity; the longer a wood has survived, the more time it has to
accumulate species - ancient woodland (Peterken, 1993). For epiphytic
species a continuity of old trees is also highly significant - old growth
woodland (Alexander et al., 2002).
Connectivity between woods also increases resilience over time (Peterken,
1993). This temporal and spatial variation is discussed further below.
Within individual woodland types, diversity is associated mainly with varied
structure producing varied light conditions. If the trees and shrubs form a
full canopy, they use so much of the light there is not enough left over to
allow smaller terrestrial or epiphytic species to flourish. There are a few
saprophytic or semi-parasitic vascular plant specialists that utilise dark
stands, especially of beech, but species diversity increases with increasing
light. As well as dappled light from incomplete canopies, temporary full
4
light from felling or coppicing and semi-permanent or permanent glades or
rides are important promoters of diversity.
Coppicing produces particularly species rich, if temporary, communities
(Rackham, 2003). Coppicing was a fundamental part of the character of
lowland woodlands for centuries before the decline of the practice. It
produces flushes of freely flowering woodland species and more weedy
species, neither of which survive well if woodland cover does not return
rapidly. In the shade phase, coppice plants survive as less freely flowering
individuals or in the seed bank.
It is difficult to see direct analogies with natural woodlands for the coppice
habitat (Rackham, 2003). It appears to be a man-made landscape, exploited
by plants evolutionarily adapted to exploit smaller canopy gaps caused by
windblow or disease. Glades that are more permanent have a different
flora, which is essentially that of our native grasslands and heaths. With the
exception of chalk and limestone grasslands, most British grassland and
heathland communities readily form from cleared and grazed woodland.
There are, however, some specialist wood edge species as well. Many
assemblages of now rare and declining epiphytic lichens appear to be
strongly associated with glades and canopy collapse deep within old growth
woodlands (Coppins & Coppins, 2005 & Sanderson, 2007a).
The issue of diversity and light is at the heart of a deep paradox within
woodland conservation. Woodlands should be among our most natural
habitats, but removing man’s management results in darker woodlands with
lower species diversity. In particular, there is the phenomenon of invasion
by native shade-bearing late succession species such as beech and holly.
These regenerate under the shade of other less shade-casting species such
as oak and hazel, which themselves require strong light to regenerate. The
result is that non-intervention woodlands in the western European lowlands,
that have been left untouched for long periods, become species-poor beech
woods (Vera, 2000). A monitored British example, The Mens in Sussex, has
shown a steady decrease in species diversity since monitoring started in
1972 (Swift & Howorth, 2006). This appears to have been caused by
increasing cover of beech and holly. Even the considerable windblow which
occurred during the 1987 storm has failed to halt this decline.
Two key species, oak and hazel, which appear to be eventually eliminated
in non-intervention stands, were however a prominent and permanent
component of natural western European woodland. It is clear that reference
non-intervention stands are failing to replicate natural conditions (Vera,
2000 & Kreuz, 2007). This issue has resulted in great controversy about the
exact drivers that maintain openness within natural woodlands.
Vera (2000) postulated that grazing and browsing animals were much more
important in driving woodland dynamics than had been previously accepted
by woodland ecologists (Peterken, 1996). Some responses to this were very
critical (Mitchell, 2005) but failed to postulate realistic alternatives (Kreuz,
2007). Svenning (2002) showed that openness was even greater in previous
5
interglacials and supported browsing animals as a major factor in these
interglacials: most floodplains were permanently clear of trees and low
productivity soils supported significant open areas of up to 30% of land
cover.
On mesic soils openness occurred but was less than 5% of the land cover.
This suggested that the loss of very large browsing in this interglacial
resulted in the grazing and browsing animals becoming a less significant
driver in natural woodlands in this period. A role for grazing and browsing in
driving natural woodland dynamics is now more generally accepted
(Peterken, pers. com. & Kreuz, 2007), but not necessarily as dominant or all
encompassing as in the more extreme interpretations of Vera’s theory.
For modern woods, the upshot is that doing nothing certainly does not
produce species rich, diverse woodlands, in the short or medium term.
Reference non-intervention stands are an important scientific objective
(Peterken, 1996) but do not conserve rich woodland floras. Either active
management of the trees is required, or in unexploited woods, more
naturalistic grazing regimes (Hodder et al., 2005) are required to prevent
late-succession tree species totally dominating.
Boreal woodland: boreal woodland must once have been widespread at
high altitudes in Britain, probably above 300 – 400m in England and Wales,
but falling above 200 – 300m in the southern Highlands, and to sea level in
the north west. The upper limit of frequent oak is probably the most
efficient definition of the beginning of zone, but the boundary is very
vague. This type of woodland has effectively been lost south of the Scottish
Highlands. On gentle slopes in many areas to the south, blanket bog has
spread to below the upper climate limit of oak, and the loss of boreal
woodland is probably natural in these circumstances. On steeper slopes, the
disappearance is likely to be due to deforestation. Boreal woodland,
however, is still widespread in the Highlands.
There will have been, and is, no sharp boundary between temperate and
boreal woodland; gradual transitions are the rule, especially on more fertile
soils. On the dominant acid soils, the presence of pine is an obvious
characteristic, and the native pine woods are deservedly famous (Steven &
Carlisle, 1959). They are not, however, the totality of boreal woodland.
They are replaced by other sorts of boreal woodland on less acidic soils and
to the north of the current native distribution of pine.
Even the current dominance of pine in some woods, may not have been long
lived. Pollen diagrams from Glen Affric, show a switch from pine-birch
woods with alternating phases of canopy collapse and regeneration to pure
pinewood in about the 18th century (Wolf & Tipping, 2003 & Shaw & Tipping,
2003). Other pollen diagrams show pine dominance for a much longer time,
particularly in the east. However, in Abernethy Forest, although always
subordinate, birch has declined in recent centuries to give the modern pure
pine stands (O’Sullivan, 1975). Associated species include silver birch on
lower ground, mountain downy birch Betula pubescence ssp tortuosa at
6
higher altitudes, and universally in the west (Worrell & Malcom, 1998),
rowan, aspen, juniper and alder. The surviving native pine woods all contain
old growth stands and appear to have been within of unclosed grazing land
for millennia.
Various birch dominated woods are associated with pine woods, or dominate
in areas without pine. Some may be woods from which pine was lost in the
past, but many are on more fertile soils, where spruce would normally
replace pines in continental boreal woods. In Scotland the place of spruce
appears to be taken by mixed boreal broad-leaved woodlands dominated by
silver or mountain downy birch which, when in good condition, can also
contain rowan, alder, hazel, aspen, birch cherry, mountain goat willow and
juniper.
These are less well known than the pine woods but can be significant for old
growth dependent mosses and lichens. Examples include silver birch woods
with aspen clones rich in rare species, and most of the other trees
mentioned above within farmland enclosing the more fertile soil adjacent to
the pine woods in Speyside. These are relatively low-altitude woodlands but
higher altitude mountain downy birch woods occur on fertile soils on
unenclosed hill land, typically with associated huge ancient alders, often
pollards, (McVean, 1956a & 1956b & Stiven & Hole, 2004), along with
mountain goat willows Salix caprea spp sphacelata, rowan and aspen.
The best-known example is on the Creag Meagaidh National Nature Reserve.
An exiting recent development (MacDonald, 2005), which accords with the
author’s personal observations, is that many more high-altitude birch woods
are at or near the natural tree line (in the sense of coherent woodland with
trees over 3m in height) than was appreciated previously.
Boreal woodlands in the Highlands have greatly reduced in area since their
greatest extend about 5,000 years ago. Much of the loss since then,
especially in the west, has been to the apparently natural spread of blanket
bog (Davies, 2003a), with the onset of wetter climates. In western Glen
Affric, Davies (2003b) shows large-scale natural decline of woodland with
peat spread, especially for pine that occupied acid gentler slopes, and
longer-term survival of boreal birch woods on slopes. These survived
millennia of human settlement, although were finally lost in the medieval
period. In the much drier Caenlochan, in Angus, pollen diagrams presented
by Huntley (1981) show high altitude birch-hazel-alder woods as opened up
and effected by grazing after 2,750 years before the present, but surviving
until as late as 200 years before the present, when they disappeared
entirely.
In the Ben Alder and Loch Treig area, an early map, the Blaeu Atlas of
Scotland, 1654, appears to show extensive open woodland in the area
(Smout et al., 2005). Some of these still survive (McVean, 1956a & 1956b)
and have been examined by the author. They proved to be the remains of
high-altitude mountain downy birch woods with rowan, mountain goat
7
willow, aspen and alder on peat-free slopes, and a single surviving pine on
the edge of the encroaching blanket bog.
As for the native pine woods, only one mapped wood, south of Loch Leven in
Lochaber, appears to have disappeared entirely in the early modern period
(Smout et al., 2005). In recent centuries losses appear to have been much
more severe for boreal birch woods on richer or damper soils than the
native pine woods on acid soils. Until recently, almost all surviving boreal
woods were heavily grazed and had little regeneration. The surviving woods,
however, have been grazed for over two millennia by domestic stock, with
native deer and wild cattle present prior to this, so they must have
regenerated under grazing pressure many times.
Man has certainly played a large part in deforesting many areas, especially
in the east and on more fertile soils, but Fenton (1998, 2004 & 2008) has
challenged the significance of this. The core of his argument is that the
densities of red deer quoted low enough for adequate tree regeneration (as
low as four-eight deer per square kilometre) are considerably below the
natural carrying capacity of the hills for red deer, which is about 20 deer
per square kilometre in the worst sort of ground. This is a result of the mild
oceanic climate, which also contributes to wide-scale natural acidification
and nutrient impoverishing of the soils.
Given this, the conclusion that moorland is likely to be as natural as
woodland and that much deforestation would have occurred naturally,
would seem difficult to refute. His position is extreme, but in the far north
west, at least very credible. In less extreme areas, the facts point to natural
woodland that was more strongly affected by grazing than lowland
woodland, with a great deal of natural openness. This was proposed by
Peterken (1996) before grazing and browsing was accepted as having any
impact on lowland natural woodlands. Natural boreal woodland in oceanic
Britain may have been a very distinctive, extreme type, with the less
extreme winters and severely leached soils allowing natural grazing impact
to be a much more significant driver of woodland dynamics, and fire much
less prominent. Given this, the modern fashion for total grazing exclusion as
a tool in boreal woodland restoration is questionable (Dennis, 1998), if
natural or semi-natural native woodland is the objective.
2.2
Longevity and size
2.2.1 Habitat longevity
Longevity in woodland habitats has been demonstrated to be a significant
factor in the richness of woodland floras (Peterken, 1993). Staring from bare
ground, time is required for woodland ecosystems to fully assemble. On a
very long timescale the tree and woodland herb floras appear still not to be
fully at climatic equilibrium since the return of trees to northern Europe
(Skov & Svenning, 2004 & Svenning & Skov, 2007). In Britain most concerns,
however, are on a shorter timescale and concern recovery of woodland
recolonising cleared land (Peterken, 1993 & Rackham, 2003); and conserving
those woods that have survived as woodland since medieval or at least early
8
modern periods. There is a rather complex nomenclature involved in
describing longevity in woodland habitats with very different meanings
(Alexander et al., 2002), which are listed below:
Virgin woodland: woodland never disturbed by man. It is often pointed
out that there is none of this in Britain, but equally there is little of it
anywhere else. Much supposed virgin woodland in other contexts was
exploited by aboriginal inhabitants in ways not immediately recognisable to
Europeans. Often used as a euphemism for old growth woodland
Primary – secondary woodland: in a British context, woodland that has
never been cleared since woodlands returned, versus woodland that has
been cleared but then recolonised. Outside Europe, this normally would
also be old growth or virgin woodland, but in Britain the term is used for
managed young growth woodland. It is not really proven that this is of much
significance over long timescales. Small leaved lime has colonised Iron Age
hill fort banks (Rackham, 2003); species rich woodlands can be found
developed over Roman farmland on the chalk in southern England
(Colebourn, 1983); and very rich woodland of medieval origin has been
studied in Oxfordshire (Day, 1993).
On the other hand, there are a few species, which appear largely confined
to primary woodland, but there always are exceptions (Spencer, 1990).
There is certainly a lot of primary woodland in Britain, but it is difficult to
distinguish from very old secondary woodland. At shorter timescales, there
are much clearer differences, which are accommodated in the ancient and
recent woodland split.
Ancient – recent woodland: this is currently the most frequently used
category in Britain. This distinguishes between older woodlands, pre-dating
1600AD (ancient woodland) and those post-dating this (recent woodland)
(Peterken, 1993 & Rackham, 2003). All recent woodland is secondary
woodland but ancient woodland can be either primary or secondary. The
date is largely one of convenience. Secondary woodlands predating 1600AD
are much more difficult to identify as maps are lacking and documentation
spasmodic. Secondary woodlands of late medieval origins are also rare,
while woodlands of post 1600AD origin are widespread. Ancient Woodland
Inventories produced even younger dates (Spencer, & Kirby, 1992), with the
earliest available accurate county map used, usually dating from the mid-tolate 18th century, but as late as the 1840s in some northern English counties.
Differences between ancient and recent woodland are usually very
pronounced, with a large suite of plant species proving to be slow colonisers
over this timescale. This has led to the compiling of lists of ancient
woodland indicator species (Peterken, 1974 & Rose, 1999). Colonisation
does occur, even if slow (Brunet & Von Oheimb, 1998, Bossuyt et al., 1999 &
Verheyen & Hermy, 2001) and can proceed at different rates in different
habitats. For example, recent woodlands on floodplains in Belgium are
9
colonised much faster by woodland species than woodland off floodplains
(Verheyen et al., 2003).
The concept of ancient woodland indicator species seems to work best in
the drier parts of Britain, in areas with scattered isolated woods and with
recent woods originating on arable land. It works less well in wetter areas
of the country, areas with interconnected woodlands, and where woodland
has colonised some types of natural grassland. With numerous exceptions, it
is best to regard vascular plant woodland indicator species lists as indicators
of woodland habitat that is strongly associated with ancient woodland; but
not as direct indicators of ancient woodland.
Ancient woodlands are now a keystone feature in woodland conservation
and woodland policy in Britain (Forestry Commission, 2005 & ODPM, 2006).
They represent woodland habitats that cannot be recreated in the short
term.
Old growth – young growth: ancient woodland is determined on the
continuity of the site, not the trees themselves. The trees could be
coppiced on a five-year rotation, but the woodland would still be an ancient
woodland. This is something of an alien concept to conservationists in most
other parts of the world, where such woodland would be regarded as of
little consequence (Rhind, 2003). In most parts of the world, little disturbed
or near natural woodlands are the main conservation issue. These are often
described as old growth woodlands. These are essentially woods in which
trees live out their natural lives. In many parts of the world, old growth
stands are synonymous with virgin forest as old growth stands predates the
start of European exploitation.
They can equally be found within cultural landscapes and be woods that
were managed primarily for products other than timber, such as stock
grazing or wild game (Alexander et al., 2002). Generally the term is
confined to woods that are at least semi-natural and have been in an old
growth condition for at least several generations of trees.
Disturbed young growth woods, however, will develop into old growth
woodland if left unmanaged. In Britain, most woodland types begin to
develop strong old growth characteristics after a stand age of about 200
years, although hazel-dominated stands probably become old growth in
about 100 years.
Old growth woods are especially rich in species of lichens that are
dependent on niches found only in older trees (Coppins & Coppins, 2005 &
2006, Sanderson, in prep, and Sanderson & Wolseley, 2001). As with
vascular plants, indicator lists of lichens associated with high-quality old
growth woodland have been drawn up (Coppins & Coppins, 2002). There are
four separate indexes of ecological continuity covering different climatic
10
zones. The occurrence of the species used is directly linked to old trees and
continuity of this habitat. As a result the lichen indicator indexes are more
of a direct measure of old growth continuity than vascular plants are
indicators of ancient woodland continuity.
Although Britain may have no virgin woodland, it has a surprising large
amount of old growth woodland associated with woods that were
traditionally utilised mainly for grazing, and related habitats in parklands.
For example the New Forest, Hampshire, alone has as much old growth
woodland as the whole of northern New England and southern New
Brunswick (Flower & Tubbs, 1982 & Selva, 1994). Lowland England certainly
has considerably more old growth woodland than lowland France (Rose,
1988). Only 1% of forests in Norway are older than 160 years (WWF, 2003),
yet woodland cover in Norway is often compared favourably with the
Scottish Highlands, where well over 50% of ancient native woodlands are
probably older than this.
Much old growth within the cultural landscapes of Britain is certainly of
international significance. Unlike ancient woodland, however, there has
been no attempt to carry out an inventory of old growth woodland and
there is no explicit protection given to old growth woodland (Alexander et
al., 2002).
2.2.2 Habitat size
Large areas of woodland tend to be richer in species than smaller
woodlands. This is both because larger woods harbour more varied
individual habitats and because they are able to sustain larger and hence
more sustainable populations of individual species.
Rackham (2003), however, demonstrates that for ancient coppices in
eastern lowland England the relationship is weak for plants. These woods
have been isolated for more than a millennium and have probably long
reached equilibrium. Rackham (2003) suggests that the current fashion to
apply island biogeography theory to English woods is pushed too far. He
argues that lowland ancient coppices are suffering from far more pressing
problems than millennium-old fragmentation, such as the cessation of
coppicing and increasing deer numbers. This is in contrast to habitats such
as fens and heathlands that have been massively reduced in area in the past
200 years, and where there are serious and real fragmentation problems.
Unlike woodland vascular plants, many epiphytic lichens show the more
conventional behaviour expected of woodland species; they avoid woodland
edges and can require very large areas of old growth woodland to survive
(Sanderson, 1998a & 2010). Some very interesting results produced by Ellis
& Coppins (2007) showed that modern epiphyte diversity in individual aspen
stands within boreal woodland, was related to woodland size in the 19th
century, not to modern woodland size. Where woods have declined in area
since the 19th century, then this implies a looming extinction debt. Again
this suggests that fragmentation should be regarded as a priority issue in
11
habitats that have been reduced in size in the past few centuries. Schemes
to undo fragmentation that occurred millennia ago could be something of an
extravagance.
2.3
Historic woodland management
2.3.1 Introduction
The history of woodland management in relation to nature conservation
value has only relatively recently been studied in detail, with much
pioneering work done by Oliver Rackham (Rackham, 1980). This was largely
based in East Anglia and adjacent areas (Rackham, 1986, 1989, 1990 &
2003). Unfortunately, there was no other review of a similar level of
authority based on other areas of the country until Smout et al. (2005) on
Scotland and this lacks the depth of coverage of biodiversity issues of
Rackham’s work. This has unfortunately led to a tendency for nature
conservationists to uncritically project Rackham’s work from East Anglia on
to other areas of Britain. This is in spite of Rackham’s exultations to have
regard to the unique nature of every wood and not to dwell on
commonplace shared features (Rackham, 2005).
2.3.2 Coppicing, timber and grazing
The main historic methods of exploiting woodland are grazing (either by
domestic stock or by killing wild game), coppicing bushes or young trees for
small wood with the intention of re-cutting them or felling larger trees for
timber. A variant on coppicing in grazed woodlands is to pollard trees, thus
protecting the regrowth from browsing.
It is likely that woodland exploitation started as a chaotic combination of all
methods of exploitation, but with grazing predominant (Vera, 2000). The
New Forest was a multi-use system with all the above forms of exploitation
carried out in the same piece of land by competing, and often mutually
antagonistic, interests in the early modern period (Reeves, 2006, Roberts
2002 & Tubbs, 2001). There seems to be a strong relationship between the
density of settlement and the degree of separation between types of
woodland exploitation. Areas of the country with dense settlement, with
large areas of fertile land capable of arable production, tend to segregate
the different uses of woodland. In particular, grazing is separated from
woodland all together. This separation allowed for more efficient use of
limited land resources. In contrast, areas where arable land was very
limited are likely to have retained unspecialised multi-use woodland
systems for far longer. East Anglia, where Oliver Rackham carried out much
of his work, is very much a crowded landscape, hence it is problematic to
uncritically use his work here as a template for other parts of the country.
In well settled areas of the country, fully enclosed coppice tended to
dominate woodland use at a fairly early date. If grazed at all, this would be
carefully controlled and have limited effect on the woodland composition.
Remaining unenclosed land tended to lose woodland altogether and moor,
heath, fen and down would dominate. Where woodland survived on open
land, normally forests, it would typically be in the form of
12
compartmentalised pasture woodland - embanked coppices enclosed for
several years after cutting but later thrown open to grazing, as in Hatfield
Forest (Rackham, 1989). This did not always work well and changes to
composition due to past browsing damage to coppice can be seen in
surviving examples.
In less populous areas, separation of woodland uses tended to be much
weaker and grazing in particular was often all pervasive (Smout et al., 2005
& Winchester, 2000). Land use and woodland management was often much
more fluid compared to the extreme long-term stability shown by many
lowland enclosed coppices. The management of such complex systems was
inefficient in terms of timber and wood production and, from the mind set
of the enlightenment elite, irrational. From a nature conservation
perspective, a major impact of such inefficient systems was that they
tended to preserve old growth characteristics inherited from natural
woodland. This is a key factor in the survival of internationally important
old growth dependent epiphytic lichen floras in the west of Britain. Such
systems have declined with modernity and, where they survive, tend to be
simplified.
In the uplands, a sequence of woodland development has been shown in
several areas, where subsistence peasant farming was replaced commercial
sheep farming (Winchester, 2000). The earlier subsistence farming was
characterised by extensive unenclosed woodlands, cattle dominated
grazing, references to ‘panage’ (feeding pigs on acorns) and extensive
heather dominated moors. The switch to commercial sheep farming was
associated with the loss of unenclosed woodland, disappearance of panage
and replacement of heather moor by grass moor. Wales, the Pennines and
the southern Uplands were in the latter condition by the end of the
medieval period. The Scottish Highlands remained in the former condition
into the modern period.
2.3.3 Coppice
The role of coppice in traditional management within British woodlands is
well known and described (Rackham, 2003). The scale of coppicing,
however, is probably not widely appreciated. Many attempts at coppice
revival are rather half-hearted, with too many standards retained, the
coppice stools reduced to too low a density by over shading and the ‘coupe’
(area of coppice) size much too small. The latter is very important. Much
conservation revival coppicing is carried out in very small coupes of about
quarter a hectare or less. In contrast, much historic coppicing was carried
out by felling coupes of many hectares in extent at once. This was
especially so in areas with industrial woodland industries feeding charcoal
furnaces or cutting oak for tan bark. A picture (Fig 73) in Peterken (2007a)
dating to before 1929 in the lower Wye valley at Tintern illustrates this
well. There is a similar picture in Peterken (2007b). In the lower Wye,
entire hillsides were cut at once with virtually no standards retained.
13
The biodiversity impact of controlled grazing in coppiced woodlands is
especially obscure, as the treatment has totally disappeared. It is likely to
have favoured woodland edge and woodland grassland species.
2.3.4 Pasture woodland
The role of coppice in traditional management within British woodlands is
well-known, but the extent and variety of woodland in which grazing was
the predominant use is less widely appreciated. Interest in the nature
conservation value of the habitat, however, has increased greatly since this
was first analysed in Harding & Rose (1986). As was the case with Harding
& Rose (1986), this report is considering the wider pasture woodland
habitat, rather than a narrow view of ‘wood pasture’. The former term is
used advisedly in this report as it covers all the aspects of extensively
grazed woodland habitats. This includes grazed high forest with a woodland
ground flora, park-like (savanna) stands and associated open grasslands
(Chatters & Sanderson, 1994). Wood pasture, as strictly defined by
Rackham (2003), is confined to trees over grassland without a woodland
element in the ground flora. Rackham argues that grazing in woodland
depends on there being grassland maintained under the trees. This
definition excludes grazed woodland sites such as the New Forest, which
Rackham has accepted is not a wood pasture under his definition (Rackham,
pers. com.).
In fact, grazed woodlands historically were complex systems with competing
land users who did not always share similar aims and therefore were not
necessarily logical (Tubbs, 2001 & Smout et al., 2005). In addition, shelter
and winter browse within pasture woodlands were often (and in the New
Forest still are) more valued than the summer grazing within the woods by
graziers. In this case opening up woods to a degree that encouraged
summer use would negate their greatest value in winter. Preservation of
woodland almost entirely for its value as shelter for cattle seems to have
been a feature of well-managed late medieval to early modern Highland
estates (Watson, 1997). Where summer grazing was particularly good,
typically trees would be removed altogether, producing what were often
known as lawns in England. The whole system would be complicated by the
wood and timber rights usually being separate from grazing rights. Neither
party might be getting exactly what they wanted.
Where pasture woods were, and are, being managed sustainably, then
regeneration must occur. When it occurs it tends to occur in clumps,
eventually producing a glade and grove structure but with areas of savanna
produced by canopy collapse (Vera, 2000). There are also likely to be
permanent open areas where the grazing is most productive. A pasture
woodland consisting of nothing but savanna is a dying woodland and not a
sustainable habitat. It is therefore rather odd that in some areas healthy
regenerating pasture woodlands are termed ‘infilled wood pastures’ with
the implication that this regeneration is damaging (Stiven & Hole, 2004).
The Wood Pasture and Parkland BAP explicitly rejected the narrow
definition of wood pasture and adopted the wider pasture woodland
14
definition (Definitions Sub-group, 2001), although the misleading term
‘wood pasture’ is still used in the title.
Building on Harding & Rose (1986), the following general classification of
pasture woodland can be made with reference to the past traditional use:
Lowland forests and commons: generally grazed for most of the year,
producing wood from pollards and temporary coppicing and usually
significant amounts of timber. Found mainly on forests and chases but also
on common land in some areas. Surviving pasture woodlands near London
were dominated by pollarding, but this seems to have been unusual, and
reflects distortions cased by the huge firewood market in London. In most
woods, the lord of the soil had the right to grow timber, and lowland
pasture woodlands were a major source of timber in early modern England
(Stagg, 1989). This was the native tradition of English high forest
management, but one that relied on ‘wild regeneration’ of patchy scattered
regeneration occurring where it could, rather than dense carefully
preserved ‘natural regeneration’. Strictly, ‘natural regeneration’ is a
propaganda term from modern German foresters for the unnaturally dense
and regular regeneration aimed at in shelter wood and selection felling
systems of high forest treatment (Vera, 2000). Recent research in the
numerous documents surviving from the New Forest (Reeves, 2006) gives
vivid picture of the management of such pasture woodlands in the 17th
century. The New Forest was a heavily exploited old growth woodland that
still retained its old growth characteristics (Sanderson, 2010.).
Winter grazed woodlands: these were briefly mentioned in Harding & Rose
(1986), with reference to the old growth oak woods of Horner Combe,
Exmoor. They were actually very widespread in the uplands of Britain, and
were especially important in the western Highlands (Smout et al., 2005).
Management appears to have been much like lowland pasture woodlands,
but coppicing was easier without summer grazing and, while pollards are
present in some regions, they are absent in others. In south west England
and the Scottish Highlands, these occurred in the modern period on
enclosed commons or hill grazings. In Wales and northern England most
were in enclosed land, called the ffrith or out bye. In these the importance
of grazing can be seen on old maps, as virtually all woodland in the ffrith
and out bye fails to fit to any boundaries, whether wooded enclosures or
rough grazing. In these areas there was virtually no woodland on hill
commons, a great contrast to the Scottish Highlands. In many areas, a great
deal of this often very mixed woodland was replaced by industrial coppiced
oak plantations in the 18th and 19th centuries (Smout, 2005 & Coppins &
Coppins, 2005). Given that these areas were originally grazed, it is ironic
that many conservationsists do not support them reverting to pasture
woodlands. Interestingly, in one of the best preserved areas of former
winter grazed pasture woodland in Borrowdale, Lake District, grazing was a
significant positive factor for 80% of the internationally significant
bryophytes there (Thomason, 1995). In western Scotland, the impact of 18th
and 19th century industrial coppicing on formerly winter grazed woodlands is
variable (P. Quelsh, pers. com.). There are substantial areas of surviving
15
Comment [TCW2]: The irony
that grazing management is
being opposed when these
areas were originally grazed –
i.e. they were replanted as
coppice in between.
old growth woodland which were not affected but also very large areas of
mixed pasture woodland where coppicing for charcoal was intensified in the
18th and early 19th centuries but which remained mixed grazed woodlands.
These are now reverting to old growth woodland and are nearly as lichen
rich as the woods undamaged by coppicing. In contrast, substantial areas
were converted to pure oakwood (Tittensor, 1970) and remained intensively
managed for tanbark until the early 20th century (Smout et al., 2005).
These oakwoods are still much poorer in lichens than the other types of
woodland (Coppins & Coppins, 2005).
Summer grazed woodlands: these were not described by Harding & Rose
(1986), but are a significant feature of the Scottish Highlands. They were
the woodlands that survived within the summer grazings used by cattledominated transhumance systems in medieval and early modern periods.
Most had probably been grazed for several millennia (Huntley, 1981). Tiny
relicts survive in the Pennines (Fleming, 1997 & Sanderson, 2001a). These
woods probably included much of the boreal pine and birch woods. They
also included the more isolated temperate woodlands, particularly at the
heads of inland glens.
A well-preserved example is the ten km long pasture woodland on the north
east shore of Loch Lomond at Pollochro (Sanderson, 2005a). An interesting
reference in Smout et al. (2005) refers to tenants of Craigroyston in 1758, to
the south of Pollochro, complaining that for 11 months of the year their only
pasture was in the woodlands that were about to be enclosed and converted
to oak coppice.
A characteristic identifying feature of summer grazed upland pasture
woodlands is the presence of alder pollards (Sanderson, 1998b). In winter
grazed upland woods (and in lowland grazed woodlands), alder appears to
have been mainly coppiced as the species is unpalatable, and is able to
recover even under quite heavy, if extensive, grazing. Many high altitude
former summer grazed pasture woodlands are now in very poor condition,
having failed to regenerate since the coming of sheep farming or creation of
modern sporting deer forests (Maxwell, 1929) in the late 18th or early 19th
centuries. They are now grazed heavily throughout the year, rather than
for a few months starting from May or early June through to October at the
latest (Harvey, 2002, & Winchester, 2000).
As a result, the current very open condition of many of these woods is
probably not representative of how they looked in their prime. There are
exceptions, for example, the pasture woods at Pollochro, Loch Lomond,
have younger generations of trees succeeding the oldest generation of ash
and alder pollards. The woods appear to have regenerated at least twice in
the 19th or early 20th centuries, with the pollards dating from the 18th
century. The woods were still grazed as part of a farm when they
regenerated but presumably had low deer numbers (Sanderson, 2005a).
Currently these woods are not regenerating in spite of the withdrawal of
farm stock but they still have significant numbers of red deer.
16
Contemporaneous regeneration within such woods, as at Creag Meagaidh
NNR, has only been achieved by reducing grazing to levels well below both
the natural carrying capacity of the land (Fenton, 1998, 2004 & 2008) or the
likely levels of summer grazing when they were summer grazed pasture
woods. These woods appear to have been favoured by very low deer
numbers, but quite high but seasonal stock grazing, not an easy combination
to obtain in the modern uplands.
Deer parks: privately owned enclosed pasture woodlands housing deer
were prominent status symbols in the medieval period (Rackham, 2003). A
few have survived as semi-natural ecosystems, with many old trees,
unimproved grassland and more dense wooded patches, e.g. Dinefwr in
Wales and Widen Park, Exmoor. Most, however, have been absorbed into
landscape parks, a somewhat different status symbol developed in the early
modern period. Few of these are semi-natural systems and the majority are
artificial arboretums, although they are important reservoirs of veteran
trees. Historic landscape conservation considerations usually preclude
rewilding and regenerating these habitats using wild regeneration.
2.4
Methods of floristic analysis
2.4.1 Scope
The analysis of the rare plant flora covers Britain, i.e. England, Wales and
Scotland and includes vascular plants, bryophytes and lichens. The species
covered are those included within the most recent Red Data Books (for
vascular plants - Cheffings & Farrell, 2005; for bryophytes - Church, 2001;
for lichens - Wood & Coppins, 2003) and the latest list of BAP species
(Biodiversity Reporting and Information Group, 2007). For lichens, an
unpublished spreadsheet for the second edition for the Conservation
Evaluation was used as this was digitised already, added many new species
and made some obviously required changes.
Woodland species were defined as any species which appeared to the author
to have at least some of its British population within wooded habitats.
There are obviously borderline species, where arbitrary decisions had to be
made. Of the lower plants, epiphytes which are clearly confined to trees
outside of woodland were not included in the main figures, but were
considered separately.
2.4.2 Information
The habitat, ecology and threats to the individual species were determined
using personal experience and mainly standard accounts such as Preston et
al. (2002) and Hill et al. (1991, 1992 & 1994). The data on the National
Boidiversity Network web site www.searchnbn.net/ was especially useful for
researching species not well known to the author. The UK Biodiversity
Action Plan website was also consulted. Many other publications were
consulted, which are referenced in the text. As well as the broad climatic
classification of woodland outlined in section 2.1, attempts were made to
allocate species to other classifications. The woodland Priority BAP
Habitats (Biodiversity Reporting and Information Group, 2007) proved
17
relatively straightforward but woodland Annex 1 Habitat Directive Biotopes
(European Commission, 2003) proved difficult. The data was entered into
spreadsheets.
3.0
Rare and declining plants
3.1
Vascular plants
3.1.1 Red Data Book and Biodiversity Action Plan species – numbers and
distribution
Key to abbreviations:
IUCN category
Critically Endangered (CR)
Endangered (EN)
Vulnerable (VU)
Near Threatened (NT)
Data Deficient (DD)
Least Concern (LC)
The most recent Red Data Book (Cheffings & Farrell, 2005) is a full
assessment of the whole vascular plant flora and lists 443 species that are of
Conservation Concern or Near Threatened (NT). Of these species, 48 are
clearly woodland species, 11% of the total. Similarly, of the 171 vascular
plants listed as priority Biodiversity Action Plan (BAP) species, 29 are
woodland species, 17% of the total. This, however, includes 12 whitebeam
micro-species of rocky habitats usually within woodlands, and excluding
these brings the Red Data Book (RDB) species to 8% of the total and BAP
species down to 11%. This explains a general feeling among botanists that
open habitats harbour more rare species and are under greater threat than
woodlands. This is likely to reflect the degree of protection that has been
afforded to woodland in the past half century. Much has changed and been
lost within British woodlands but the habitat remains widespread and the
losses are small compared with habitats such as fens and heathlands. A
species that exemplifies this is oxlip Primula elatior (NT), which once
occurred in both meadows and woods in the small area of eastern England
to which it is confined, but was lost from grassland sites by the 1930s
(Preston et al., 2002). It has declined in woodlands but still survives
(Rackham, 2003).
The combined total of RDB and BAP woodland species is 50. Of these, 43
are confined to temperate woodland, three are boreal species and four are
found in both temperate and boreal woods. Within the species of
temperate woodlands, the majority show no strong trends towards strongly
oceanic woodland or to sub-oceanic areas within Britain. An exception is
the single species confined to hyper-oceanic woodlands - Wilson’s filmy fern
Hymenophyllum wilsonii, a small Near Threatened fern, whose low growth
18
Comment [TCW3]: But when
compared to the other 19 broad
habs, woodland is 6th or 7th
from the top! (depending on if
NT/DD included). This needs
qualification.
form and thin foliage renders it an honorary moss. Unlike many species, this
is not an edge of range species, with Britain a core part of this fern’s world
range. The lack of hyper-oceanic vascular plants in Britain is in contrast to
the abundance of hyper-oceanic bryophytes and lichens. Woodland southern
oceanic species are also limited, with only Italian lords-and-ladies Arum
italicum ssp neglectum (NT) and bastard balm Melittis melissophyllum (VU
& BAP) assigned to this group. In contrast, there are many rare southern
oceanic species of woodland lichens.
Species with a strong eastern tendency are slightly more frequent, with five
species confined to sub-oceanic woodland. These include three species of
southern, base rich, damp, ancient woodlands-rich woods: oxlip Primula
elatior (NT), Suffolk lungwort Pulmonaria obscura (EN & BAP) and crested
cow-wheat Melampyrum cristatum (VU & BAP). The other two are rather
enigmatic and more northern species. May lily Maianthemum bifolium (VU)
is a strongly continental small lily of acidic woodlands with a tenuous
scattered distribution in the east of England, which is confused by potential
garden escapes. The species has poor flowering and seed set in Britain. The
British population could also conceivably be one with a low viability, which
is periodically naturally introduced by birds. Whorled Solomon's-seal
Polygonatum verticillatum (VU & BAP) is confined to temperate woodland in
a few northern ravine woodlands in Perthshire and Angus, but is also
widespread in boreal woodland in Scandinavia. It is a possibility that it once
occurred in base rich boreal woodland in Britain but has been lost from this
woodland type through habitat loss.
The temperate woodland species include many southern species such as
spreading bellflower Campanula patula (EN & BAP), narrow leaved
bittercress Cardamine impatiens (NT), starved wood sedge Carex
depauperata (EN & BAP), white helleborine Cephalanthera damasonium (VU
& BAP), green hound’s-tongue Cynoglossum germanicum (CR & BAP),
mezereon Daphne mezereum, pale St John's-wort Hypericum montanum,
military orchid Orchis militaris (VU), lady orchid Orchis purpurea (EN),
spiked rampion Phyteuma spicatum (EN) and downy woundwort Stachys
germanica (VU). A northern species is lady`s slipper orchid Cypripedium
calceolus (CR). Widespread species include narrow-leaved helleborine
Cephalanthera longifolia (VU & BAP), lesser butterfly orchid Platanthera
bifolia (VU & BAP) and greater butterfly orchid Platanthera chlorantha
(NT). The majority are found in base rich habitats but spreading bellflower
Campanula patula (EN & BAP), copse-bindweed Fallopia dumetorum (VU &
BAP), spiked rampion Phyteuma spicatum (EN) and Arran service-tree Sorbus
pseudofennica (VU& BAP) are species of acidic woodlands.
Four species grow in both temperate and boreal woods and three in purely
boreal woodland. These species are, not surprisingly, mainly northern
species, but with juniper Juniperus communis (BAP) a more widespread
species. Juniper, however, is only really a woodland species in the north.
Coralroot orchid Corallorhiza trifida (VU), northern hawk`s-beard Crepis
mollis (EN & BAP) and small cow-wheat Melampyrum sylvaticum (EN & BAP)
are species of broadleaved stands found in both temperate and boreal
19
woodland. Twinflower Linnaea borealis (BAP), one-flowered wintergreen
Moneses uniflora (EN & BAP) and intermediate wintergreen Pyrola media
(VU) are more strongly confined to boreal woodlands, especially pine woods
and old pine plantations. The Linnaea and Pyrola can occur in moorland and
the latter is more frequent in this habitat. Coralroot orchid Corallorhiza
trifida (VU) is a saprophyte associated with trees and bushes in damp
situations and uses the under shrub creeping willow Salix repens in open
dune slack habitats as well as wet woodlands. It is also a good colonist of
secondary wet woodlands.
VASCULAR PLANTS – SPECIES OF CONSERVATION INTEREST
Distribution within climatic regions
Habitat
No. of
vascular
plants
Temperate & temperate – boreal
Temperate, hyper-oceanic
Temperate, oceanic
Temperate, suboceanic
Boreal
% vascular
plants
39
1
2
5
3
78%
2%
4%
10%
6%
Extinctions
A single species is recorded in the RDB as extinct, the enigmatic ghost
orchid Epipogium aphyllum (EX). This saprophytic orchid has a history of
fleeting appearances and may spend much of its life underground. It was
reported in 1986 and hence declared extinct in the 2005 RDB, but has since
reappeared at a single site in Herefordshire in 2009.
3.1.2 Red Data Book and BAP Species – habitats and threats
For the temperate woodland species, the strongest linking factor for many
species is a requirement for well lit conditions on wood edges, glades and in
the early regrowth of coppiced and felled stands. Species strongly
associated with the early regowth of coppice and similar habitats include
narrow leaved bittercress Cardamine impatiens (NT), starved wood sedge
Carex depauperata (EN & BAP), green hound’s-tongue Cynoglossum
germanicum (CR & BAP), lady orchid Orchis purpurea (EN) and spiked
rampion Phyteuma spicatum (EN). These species have been affected by the
reducing of intensity of management within lowland woodlands, especially
the cessation of commercial coppicing of ancient woodlands over much of
the country. Some wood edge species appear to have been especially
vulnerable, particularly spreading bellflower Campanula patula (EN & BAP)
and crested cow-wheat Melampyrum cristatum (VU & BAP), with other
declining species include pale St John's-wort Hypericum montanum (NT) and
fly orchid Ophrys insectifera (VU & BAP). As well as being negatively
affected by a reduction in temporary open space within woodlands, these
edge species have probably also been reduced by the linked problems of
intensification and abandonment of adjacent open land. Finally, there is a
small group of species that appear to be mainly associated with long term or
permanent glades, particularly the orchids narrow-leaved helleborine
Cephalanthera longifolia (VU & BAP) and red helleborine Cephalanthera
rubra (CR & BAP). Many sites for the former in Hampshire and Scotland are
20
in woodland that was historically grazed. The edge and glade species may
have benefitted from controlled or seasonal grazing within pasture
woodlands in the past. The early succession species downy woundwort
Stachys germanica (VU) was strongly associated with the mainly
compartmentalised pasture woodland of Wychwood Forest and was recorded
from the coppices themselves before enclosure (Marren, 1988). Since
enclosure in the 1850s, it has only been recorded from scrub wood edges
and lanes. It may have been associated with grazing damage in
compartmentalised pasture woodland systems, when the separation of
grazing and coppicing failed. Cynoglossum germanicum (CR & BAP) was also
recorded from Wychwood Forest and may also have occupied this niche.
The many micro-species of whitebeam recorded in the RDB are a specialist
type of woodland glade species, being associated with permanent glades
formed by base rich rock outcrops within woodlands.
Other species are less strongly associated with very open stages of woodland
development and survive in reduced populations in shady woodland but
were much more abundant in regularly coppiced woodland in the past.
These include oxlip Primula elatior (NT), Suffolk lungwort Pulmonaria
obscura (EN & BAP), lesser butterfly orchid Platanthera bifolia (VU & BAP)
and greater butterfly orchid Platanthera chlorantha (NT). The very rare
northern whorled Solomon's-seal Polygonatum verticillatum (VU & BAP) is
also thought to be suffering from increased shade from ivy in formerly
managed ravine woodlands. Shade from invasive shrubs is not mentioned
much for vascular plants, compared to bryophytes or lichens. Invasion by
the invasive exotic rhododendron will certainly have negatively affected
some sites for the hyperoceanic Wilson’s filmy-fern Hymenophyllum wilsonii
(NT).
A very different group are parasitic or partly parasitic species found in
deeply shaded woods including the orchids white helleborine Cephalanthera
damasonium VU & BAP), ghost orchid Epipogium aphyllum (EX) and bird'snest orchid Neottia nidus-avis (NT), and the unrelated yellow bird's-nest
Monotropa hypopitys (EN & BAP). Causes of declines in these species are
not clear and the occurrence of some is spasmodic in any one site, This is
especially so of the very rare ghost orchid Epipogium aphyllum.
The seven temperate – boreal and boreal species mostly occur in woodlands
that were once, or are still grazed, and several also occur in alternative
open habitats. Unbalanced grazing is likely to be a case of losses, with too
much grazing causing losses in the past but increasingly too little grazing
likely to increase threats from shading and course vegetation. Northern
hawk`s-beard Crepis mollis appears a classic example, a species of herbrich grassland and pasture woodlands in Northern England and Southern
Scotland, and now squeezed between intensification and abandonment.
21
Small cow-wheat Melampyrum sylvaticum (EN & BAP) is described by
Dalrymple (2006) as sensitive to grazing, although it is also described as
surviving light grazing. The main threat, however, appears to be the small
area of surviving base rich boreal woodland, which could have caused
genetic losses to isolated populations.
The pine specialists, twinflower Linnaea borealis and one-flowered
wintergreen Moneses uniflora, in the past showed an ability to colonise pine
plantations, but this mature pine plantation habitat has been threatened by
clear felling and then replacement by faster growing non-native conifers.
The level of protection is lower for such pine plantations than for ancient
native pinewoods.
For some species, especially the Near Threatened species, the recorded
declines may represent reduced recording effort, as in the hyperoceanic
Wilson’s filmy-fern Hymenophyllum wilsonii (NT), and be less serious than
indicated (Preston et al., 2002). In another instance, Italian lords-and-ladies
Arum italicum ssp neglectum, the Near Threatened status may relate to
recording confusion. The introduced taxon Arum italicum ssp italicum has
been more readily recognised in recent years, whereas previously these
were mis-recorded as the native subspecies, giving a false impression of a
decline (Preston et al., 2002 & French et al., 1999).
General threats recorded for temperate woodland species mainly relate to
past woodland clearance for agriculture and conversion to conifer
plantations. For example, one of only three known sites for Suffolk
lungwort Pulmonaria obscura was lost to conifer planting (Preston et al.,
2002). These are now much reduced threats, especially for ancient
woodland, due to national policies promoting the conservation and
restoration of ancient woodland within forestry (Forestry Commission, 2005)
and in the planning system (ODPM, 2006).
3.1.3 International responsibility
An innovation in the latest RDB (Cheffings & Farrell, 2005) was an indication
of the vascular plant species that may be international responsibility
species. Like the RDB and BAP lists, open ground species predominate in
this list. Beyond the endemic whitebeams, there is also almost no overlap
with the RDB and BAP lists, with only the hyperoceanic Wilson’s filmy-fern
Hymenophyllum wilsonii (NT), listed as an RDB woodland species which has
a British population that is definitely of international importance. No
woodland BAP species are listed.
Other than endemic whitebeam microspecies, only 12 woodland vascular
plant species are listed as possibly, potentially or definitely of international
responsibility. The five definite species given were all ferns: hay-scented
buckler-fern Dryopteris aemula, scaly male-fern Dryopteris affinis,
Tunbridge filmy-fern Hymenophyllum tunbrigense, Wilson’s filmy-fern
Hymenophyllum wilsonii (NT) and intermediate polypody Polypodium
interjectum. Several of these, Dryopteris aemula, Hymenophyllum
tunbrigense and Hymenophyllum wilsonii, show strongly oceanic
22
Comment [TCW4]:
Re-worded
distributions in Europe and are mainly found in temperate hyper-oceanic
woodland and temperate oceanic woodland in Britain. The poverty of
international responsibility woodland vascular plants in Britain contrasts
with the bryophyte and lichen woodland floras.
3.1.4 Influences on the general flora of woodlands
The most detailed examination of long term change in British woodlands was
produced by a 2001 resurvey of the vascular plants in 103 plots originally
recorded in 1971 (Kirby et al., 2005). This has produced qualitative results
that confirm many of the factors discussed above. These include:
Changes to the tree and shrub species
•
Oak lost stems in the lowest size classes but gained in the larger ones.
•
Young holly showed a marked increase, confirming anecdotal accounts
from lichenologists.
•
Mean basal area of trees and shrubs increased both for individual plots
and across most sites.
•
Species richness amongst saplings (25-130cm high) decreased, but small
increases in frequency were shown by some shade tolerant species
including yew, beech and holly.
•
Seedling (< 25 cm high) frequency declined for most species, but holly
showed a notable increase.
•
Open habitats (rides, glades etc) and some wet habitats (ditches, boggy
patches) became less common.
•
Grazing signs increased in the lowlands, mainly of deer other than red
deer
Ground flora changes
•
Overall species richness declined markedly
•
‘Woodland specialists’ were more likely than other species to show
decreases in frequency.
•
Open habitats declined overall
•
Increases in tree basal area were associated with species richness
declines; other variables relating to disturbance (1987 storm damage,
grazing, open habitats) were associated with increased richness.
•
Stress-tolerant species scores declined and were negatively associated
with changes in open habitats.
This confirms the ageing of stands with increasing shade and loss of open
habitats, and resulting declines in species diversity, including many
woodland specialists. This is combined in the lowlands with increasing deer
23
grazing and browsing in areas previously without deer. Interesting features
are the failure of oak regeneration and increases in beech and holly
regeneration, which reflects the expectations from non-intervention plots
(Vera, 2000, & Swift & Howorth, 2006). Canopy disruption and maintaining
of open habitats, however, were associated with increased richness.
3.2
Bryophytes
3.2.1 Red Data Book and BAP Species – numbers and distribution
The current Red Data Book (Church et al., 2001), unlike the vascular plant
and lichen RDBs, did not assess the whole bryophyte flora, but only those
found in 15 or less ten km national grid squares. An assessment of the
whole flora would be likely to add more species but it is not clear that this
would change the balance between woodland and non-woodland species. It
is possible, however, that proportionately more hyper-oceanic woodland
species would have been included.
Compared to vascular plants, coverage and data on declines is not as
detailed, simple because there are fewer bryologists recording. Overall,
however, knowledge is far more detailed than for lichens or fungi. The
bryophyte flora of the Atlantic coast is particularly well studied, for
example, Averis (1991) surveyed an extraordinary 448 woods in the Scottish
Highlands.
Church et al. (2001) gives useful summaries of the occurrence of threatened
species (excluding Near Threatened species) by habitat. This described
‘Trees and hedgerows’ and ‘Woods’ separately but the basis of this
separation is not clearly described. Purely woodland epiphytes such as
Zygodon forsteri are described under ‘Trees and hedgerows’. Other than
this, the section on woodlands emphasises that, “although woods are the
richest and most important habitats for bryophytes in many areas, there are
few threatened woodland species”. In particular, the richness of the hyperoceanic woodlands is emphasised; the bryophyte diversity of these woods is
comparable with some tropical forests (Hodgetts, 1993). Here, however,
many characteristic species are frequent, occur in more than 15 ten km
national grid squares and were not assessed in the RDB.
The analysis for this report found 23 RDB species that could be described as
woodland species on a wide basis (but excluding epiphytes of open habitats
and some species only occasionally found in woodlands), out of 208, 11% of
the total. Similarly, of the 102 bryophytes listed as priority BAP species, 17
are woodland species, 17% of the total. These are remarkably similar to the
proportions for RDB and BAP vascular plants.
The combined total of RDB and BAP woodland species was estimated for this
report as 27, considerably lower than for vascular plants or lichens. Of
these, 23 are confined to temperate woodland, three are boreal species and
one has been found in both temperate and boreal woods. Within the
species of temperate woodlands, compared to vascular plants there are far
more hyperoceanic species, with eight species (30%) confined to highly
24
oceanic woodlands. These include the mosses Daltonia splachnoides (VU &
BAP), found in a few sites in western Scotland, and Sematophyllum
demisum (EN), of small rocks in North Welsh woods, and the liverworts
Acrobolbus wilsonii (NT & BAP), Radula carringtonii (VU & BAP) and Radula
voluta (NT). In contrast to lichens, however, there are no wider woodland
southern oceanic species included in the RDB.
Temperate woodland is the main habitat for rare bryophytes, with 14
species (52%) recorded from this habitat type. These include a disparate
range of species, including the mosses Zygodon forsteri (EN & BAP) of
wound tracks on beech trees in old growth pasture woodlands in southern
England, Anomodon longifolius (VU & BAP) on base rich rock in ravines,
Atrichum angustatum (EN & BAP) on rides in the Weald, Fissidens exiguus
(NT) on wet rocks in streams in woods, Orthotrichum obtusifolium (EN &
BAP), a pollution sensitive pioneer epiphyte, and Rhytidiadelphus
subpinnatus (EN & BAP) on banks in upland woodlands. Liverworts include
Pallavicinia lyellii (VU & BAP), found in a scatter of wet woodlands.
In contrast to lichens, sub-oceanic species are even less significant than for
vascular plants, with one species confined to eastern temperate woodlands,
Anomodon attenuatus (EN). This is currently confined to basic rocks in one
ravine woodland in Angus. In addition, the pollution sensitive pioneer
epiphyte Orthotrichum speciosum (NT) is recorded from both sub-oceanic
temperate woodlands and boreal woods in eastern Scotland.
Similar to vascular plants, but unlike lichens, the number of species
confined to boreal woodland is not high, with only three (11%). These are
confined to Scotland. They include the striking green shield-moss
Buxbaumia viridis, consisting of swollen fruits emerging from fallen logs in
old growth boreal woodlands. In addition, the moss Orthotrichum
gymnostomum (EX & BAP), which was judged Extinct in the RDB, has been
rediscovered in Deeside, as part of the rich epiphytic flora of Aspen,
currently being explored in the eastern Highlands (Legg, 2004). Finally, the
liverwort Lophozia longiflora is also a dead wood specialist, but has only
ever been recorded from one native pinewood in Speyside.
BRYOPHYTES – SPECIES OF CONSERVATION INTEREST
Distribution within climatic regions
Habitat
Temperate
Temperate, hyper-oceanic
Temperate, oceanic
Temperate, sub-oceanic & temperate – boreal, sub-oceanic
Boreal
No of
bryophytes
14
8
0
2
3
% of
bryophytes
52%
30%
0%
7%
11%
Non woodland species
Six epiphytic species were judged not to be associated with woodland.
These included two species that occurred on riverside trees Cryphaea
lamyana (VU) and Myrinia pulvinata (NT) and four species that occurred on
25
trees beyond woodland, including two Orthotrichum species (Orthotrichum
pumilum CR & BAP & Orthotrichum pallens EN & BAP). The genus
Orthotrichum includes many pollution sensitive pioneer twig epiphytes,
which have declined drastically in the last 100 years. With cleaner air, they
are showing the first signs of a major recovery (BA ref, 2008).
Extinctions
Only two definitely woodland species are recorded as extinct - Weissia
mittenii (EX) and Fossombronia crozalsii (EX). Both were woodland ride
species. The latter had only ever been recorded from woodland rides but
the Weissia was also recorded from arable fields. A further species Neckera
pennata has only a single 19th century record from lowland Forfar from nonwoodland trees. It is, however, a Vulnerable RDB species in Sweden where
it is regarded as a key indicator species of woodland sites of high value
(Nitare, 2000). It may have once occurred in sub-oceanic woodland in the
east of Britain.
3.2.2 Red Data Book and BAP species – habitats and threats
Hyper-oceanic species
The resource of bryophyte rich woods in western Scotland, the Lake District
and North Wales is immense but, unlike lichens, has been well explored
(Averis, 1991 & Hodgetts, 1997). The number of important sites, however,
means that coverage by SSSIs or SACs is hardly comprehensive, with many
important sites in Scotland not notified (Hodgetts, 1997 & May, 2002).
Conservation of the resource will clearly have to rely on wider forestry and
agricultural policy in the core areas of interest. This is reflected in
Plantlife’s listing of much of the west coast of Scotland as an Important
Plant Area.
The rare hyper-oceanic species are mainly species of rock and less
frequently trees, within ravines, or very humid coastal woods in scattered
locations along the west coast. The majority of the RDB and BAP species
are found in less accessible sections of the woodlands in which they grow,
which insulates them from many threats. This is in contrast to rare lichens,
which tend to favour better lit hyper-oceanic woodlands. There is also no
strong association with old growth stands (Edwards, 1986) but rich Atlantic
bryophyte floras are strongly associated with ancient woodland with a long
continuity of woodland cover (Ratcliffe, 1986). The most serious and
insidious threat is invasion by the exotic evergreen shrub rhododendron
Rhododendron ponticum which can invade even the most inaccessible area.
Rothero (2005) emphasised the absolute seriousness of this threat and the
general ineffectualness of the response to date. Beech is also a problem
(Rothero, 2005), it is locally non-native but is well adapted to the hyperoceanic climate, and can regenerate more effectively than oak. The shade
cast by this species is much deeper than the hazel and ash it displaces in
ravines, shading out rich hyper-oceanic communities at several important
sites. Conifer plantations in adjacent accessible land have also caused
damaging increases in shade. In some sites, upstream pollution could also
be a potentially serious threat.
26
An emerging threat is ravine running, a mainly organised sport, which
involves taking groups from outdoor centres down ravines as an adventure.
This has introduced extreme trampling and disturbance to a habitat that
was barely visited in previous decades. This has become a matter of great
concern to bryologists.
Inappropriate grazing is often mentioned as a threat, but the less accessible
nature of many of the richest sites suggests that overgrazing is probably not
normally a current threat. In contrast, the total removal of grazing to
rapidly encourage tree regeneration by fencing can cause potentially serious
problems to species growing on smaller rocks. On more fertile sites
unrestricted bramble growth can smother boulders up to 2m in height
(Rothero, 2005), a considerable threat to slow colonising species.
One RDB moss, Sematophyllum demisum (EN), in particular, is vulnerable to
this problem. It is confined to north Wales, where it grows on low slabs and
small rocks, which are set in the woodland floor, rather than large ravine
rock faces. Here it is very easily overgrown and populations have declined
where grazing has been removed from protected sites (Bryan Edwards, pers.
com.). In addition, Thomason (1995), suggested that while heavy grazing
damages the general Atlantic woodland bryophyte flora, 80% of the
internationally important flora was at least partly grazing-dependent in
Borrowdale, Lake District.
Some of the very rare Atlantic ravine liverworts are curiously restricted in
distribution, for example, the European endemic Lejeunea mandonii is
absent from many apparently suitable sites. This is partly a self-selecting
feature; were these species not unaccountably restricted, they would not
be in the RDB or the BAP.
Temperate woodland species
The 14 species of general temperate woodland include a remarkable seven
species of woodland rides. Some species are largely confined to this
habitat, such as Atrichum angustatum (EN & BAP), with most of its records
from rides in Wealden woods, plus a few old records from heathlands. It is
now very rare, with increased shading as coppicing declines the probable
cause of decline. One liverwort Fossombronia crozalsii was only ever
recorded from two unsurfaced rides in conifer plantations in north Wiltshire
and Berkshire. It has not been refound since 1972, with track surfacing and
hence a lack of periodic rutting and also shading from maturing conifer
crops the possible causes of the species disappearance. Most species,
however, also occur, or occurred, more frequently in other types of
disturbed land. Examples include the narrow European endemic, Weissia
multicapsularis (EN & BAP), apparently extinct outside of Britain. This
survives in Cornwall where it typically grows on banks and track sides on sea
cliffs and inland field banks, but has also been recorded from wayside
banks, woodland rides, banks in old quarries and fallow fields in the past.
This species had a concentration of records from the Weald, where it was
likely to have occurred in rides, but has not been seen here since the 19th
27
century. Weissia squarrosa (EN & BAP) has been recorded from fields,
beside ditches and pools as well as woodland rides. The liverwort Lophozia
capitata (VU & BAP) is mainly a species of heathland, occasionally being
recorded from rides in woods.
The loss of bryophyte diversity on rides does appear to be a significant
issue. Another example is the complete disappearance of the widespread
moss Calliergonella lindbergii from Dorset, where it had previously been
recorded sparsely across the county on tracks and woodland rides, but has
not been seen since 1977 (Hill & Edwards, 2003). Similarly, the diversity of
ride flora of the Sussex Weald also has declined since the 19th century (Rose
et al., 1991). The main drivers appear to be surfacing of the most regularly
used rides, removing well lit rutted ground, and increased shade on less
used unsurfaced rides. Rutting and disturbance along rides in woodland that
favours this suite of species is now regarded as bad forestry practice.
The other species are found in several other habitats. Three are found in
less oceanic ravine woodlands or on rocks by woodland streams: Anomodon
longifolius (VU & BAP), Fissidens exiguus (NT) and Jungermannia leiantha
(CR). The latter liverwort is an interesting example of a species lost from
many sites in the 19th century, where it is impossible to say why this
happened, and hence work out exactly what the threats are.
Other species occupy a diverse range of habitats. Zygodon forsteri (EN &
BAP) is a southern species of wound tracks and root knotholes on veteran
beech trees, which is possibly the only old growth dependent species of
temperate bryophyte in the RDB or BAP lists. It is now known from three
sites, Epping Forest, Burnham Beeches and the New Forest. It has shifting
meta-populations with long occupied main colonies on longer lasting wound
tracks and more ephemeral occurrences in smaller knothole habitats. It has
small vulnerable populations dependent on new well-developed wound
tracks developing before existing main colony trees collapse. Like many
lichens, it probably needs large areas of old growth woodland to survive.
Several previously unknown meta-populations have recently been found in
the much larger old growth woodlands of the New Forest, where five
separate populations are now known, as opposed to the two known a few
decades ago. In contrast, it has declined in the smaller Epping Forest. As
well as continuity of old trees, shading by holly may also be a threat,
especially in pasture woodlands that are no longer grazed.
Pallavicinia lyellii (VU & BAP) is a puzzlingly dispersed species of the bases
of trees in wet woodland, the sides of tussocks in Molinia mires and wet
shaded sandstone rocks. This is a long persistent plant of scattered
localities occupying a narrow niche in stable conditions (Turner, 2004). This
species has declined in many of its wet woodland sites. The reasons are not
clear but Sanderson (2005b) found that the species was responding well to
small canopy clearance resulting from holly and rhododendron control in a
grazed alder wood at Cadnam Bog in the New Forest. Turner (2003) had
28
predicted that this opening up could have been harmful. Increased shading
and reductions in light levels and lack of occasional disturbance may be
issues in its wet woodland sites. Many former and current sites were on
commons and are or were once grazed, although it is unlikely to be
favoured by heavy grazing pressures.
Orthotrichum obtusifolium (EN & BAP) is a pollution sensitive pioneer
epiphyte, specialising on elm, threatened by both air pollution and elm
disease and surviving recently only in clean air areas in north east Scotland.
Several species of Orthotrichum declined drastically in response to sulphur
dioxide pollution. Most are not particularly woodland species and some
were only recorded on open grown trees. Several may now be recovering
with clearer air, but Orthotrichum obtusifolium faces extra problems as an
elm specialist.
Finally, Rhytidiadelphus subpinnatus (EN & BAP) is another puzzlingly
dispersed species, this time of acid banks in upland temperate woodlands in
Wales and northern England. Many past records were from the north of
England from areas where acidifying air pollution has reduced the
occurrence of all Rhytidiadelphus species.
Temperate, sub-oceanic and boreal woodland species
Temperate, sub-oceanic and boreal species include another rare ravine
species Anomodon attenuatus (EN) and further Orthotrichum species
Orthotrichum speciosum (NT) and Orthotrichum gymnostomum (EX & BAP).
The latter is a recently rediscovered aspen specialist (Legg, 2004). This
habitat is also rich in rare lichens and is discussed further in the lichen
section. Finally two boreal species - the moss Buxbaumia viridis (EN &
BAP) and the liverwort Lophozia longiflora (DD & BAP) - are species of fallen
dead wood and are likely to be old growth dependent species. The former
occurs ephemerally on large fallen logs as they decay, is dependent on a
continued supply of falling large trees and probably needs large areas of
habitat. Little is known about the liverwort, but it has been recorded twice
from dead wood in Speyside in native pinewoods.
Rothero (2006) considers that some level of grazing, provided it is not so
great as to cause erosion, is generally good for bryophytes. In boreal
woodlands, interesting bryophytes growing on low rocks and rotting logs
require some grazing to prevent these habitats being overwhelmed by a
coarse herb layer or dwarf shrubs. With reference to the boreal woodlands
of the Cairngorms National Park, he expresses concern that the enthusiasm
for regenerating the woods may cause the loss of bryophytes of open
habitats, including open areas of scree, low crags, small areas of mire and
fallen logs. These are richer in bryophytes, including uncommon species,
than closed canopy woodland, which is dominated by a few robust species.
3.2.3 International responsibility
International responsibility species were not listed in the last bryophyte RDB
(Church et al., 2001), but Britain is among the richest countries for
bryophytes in Europe and of significance at a world level (Ratcliffe, 1968,
29
Hodgetts, 1993 & Rothero, 2005). Rothero (2005) records that Britain has
65% of the European bryophyte flora and may have an astonishing 5% of the
global flora. Clearly there is a need for a systematic listing of international
responsibility species as has now been done for vascular plants and lichens.
It is clear from existing work, however, that Britain has a high number of
international responsibility species, and that most are to be found in the
highly oceanic Atlantic west. Hyper-oceanic woodlands are not the only
habitats of importance in this area; low montane heaths with liverwort mats
are also very important and are much more threatened. The Atlantic
temperate rain forests of the west coast, however, are among the most
important habitats for plant conservation in Britain, although this is poorly
expressed in the RDB and BAP lists. They are representative of a rare
climatic zone, the European Temperate Rain Forest biome. Other areas of
temperate rain forest are all rated as critical for global conservation by
WWF but not those in Europe, presumably because our woodlands are seminatural (Rhind, 2003). There is little evidence, however, to suggest that
little disturbed semi-natural hyper-oceanic woods in Europe are any poorer
in important species than ‘virgin’ woodlands in other areas of temperate
rainforest.
3.3
Lichens
3.3.1 Red Data Book and BAP species – numbers and distribution
Data
The level of taxonomic and distribution knowledge for lichens is even lower
than for bryophytes. Lichenology is still in an exploratory phase, with new
species to Britain and science regularly found. Important sites can still be
found, or have only ever had one or two visits. This is especially so in the
Scottish Highlands, but even in constantly visited sites, such as the New
Forest, new species of conservation interest are regularly found (Sanderson,
2010.). A measure of the lack of knowledge is that there were 272 species
in the Data Deficient category in the draft of the new Conservation
Assessment used for this report. However, in the 50 years since lichenology
was revived in Britain, sufficient knowledge was amassed to allow a
comprehensive conservation evaluation of the whole lichen flora (Woods &
Coppins, 2003). Since then knowledge has increased exponentially with a
large scale database project run by the BLS and a new Conservation
Evaluation in draft.
Unlike other fungal groups, however, lichens have the advantage of being
perennial and visible all year round, allowing for a more conventional
approach to the assessment of conservation interest and site value than is
possible for fungi.
Although lichens act like autonomous green plants, they are not plants and
many show a far finer niche specialisation than plants. This leads to large
numbers of localised species and much more markedly different
assemblages in different woodland biomes. Lichens also tend to occupy
30
habitats that are marginal for plants, including bryophytes. One of these
habitats are the drier better lit and more stressed epiphytic habitats, so it is
not surprising that woodland is particularly significant for lichens.
The draft of a new conservation evaluation of British lichens, which will
replace Woods & Coppins (2003) when finalised, listed 382 species of
Conservation Concern or as Near Threatened. Of these, 149 are found in
woodlands (39%). Adding epiphytes not normally found in woodlands
increases the total of tree associated species to 166 (43%). Similarly, of the
136 lichens listed as priority BAP species, 81 are woodland species, 60% of
the total. Adding epiphytes not normally found in woodlands increases the
total to 91 tree associated BAP species (67%). The totals and proportions of
woodland species are much higher than for vascular plants or bryophytes. In
addition, the BAP list concentrates much more on woodland species for
lichens than for vascular plants or bryophytes.
Totals
The combined total of RDB and BAP woodland species on the draft used was
155. Of these, 109 are largely confined to temperate woodland, 32 are
boreal woodland species and 14 are found in both temperate and boreal
woods. Boreal woodlands have a far larger number and percentage of rare
lichens than mosses or vascular plants. The temperate woodland species
are also more strongly differentiated between different climatic zones. As
well as 38 (25%) generally distributed species, 35 (23%) are largely confined
to hyper-oceanic temperate woodland, 33 (21%), are somewhat more widely
distributed, and usually more southern, oceanic species and 17 (11%) are
sub-oceanic species mainly confined to the east of Scotland and north east
England. The two groups of oceanic species account for 68 species (44%).
Habitat
Temperate & temperate – boreal
Temperate, hyper-oceanic & temperate – boreal, hyper-oceanic
Temperate, oceanic & temperate, oceanic (southern)
Temperate, sub-oceanic & temperate – boreal, sub-oceanic
Boreal
No of
lichens
38
35
33
17
32
% of total
25%
23%
21%
11%
21%
Hyper-oceanic temperate woodland
The majority of rare species of hyper-oceanic temperate woodland have
distributions centred on western Scotland. Many are confined to western
Scotland in Britain, especially smooth bark specialist species such as
Arthonia ilicinella (NT), Arthothelium dictyosporum (NT & BAP),
Arthothelium macounii (VU& BAP), Eopyrenula septemseptata (NT), Graphis
alboscripta (NT), Pyrenula dermatodes (CR & BAP) and Ramonia azorica
(EN). Graphis alboscripta (NT) is an apparent endemic confined to west
Scotland, while Ramonia azorica (EN) is otherwise only known from the
Azores. There are even rare fungal parasites of hyperoceanic species such
as Arthonia cohabitans (VU & BAP), an obligate parasite of the very rare
Arthothelium macounii (VU& BAP), and Opegrapha brevis (NT), an obligate
parasite of more widespread Thelotrema petractoides. Species of rougher
bark on larger trees tend not to be totally confined to western Scotland, but
31
Lecanora cinereofusca (VU & BAP), Leptogium hibernicum (NT & BAP) and
Polychidium dendriscum (VU& BAP) are not currently known in Britain
beyond Scotland. Many of the latter group of species still have occasional
or rare occurrences in the Lake District, North Wales and south west
England. These include several larger showy species, including Gomphillus
calycioides (NT & BAP), Leptogium brebissonii (NT & BAP), Leptogium
cochleatum VU & BAP), Pseudocyphellaria intricata NT & BAP),
Pseudocyphellaria lacerata VU & BAP) Pseudocyphellaria norvegica (BAP)
and Sticta canariensis independent green morph form (VU & BAP). Despite
largely remaining widespread in the Highlands, southwards most are now
very rare, declining and in some cases extinct. Similarly, whilst some
smooth bark specialists, such as Mycomicrothelia atlantica (NT) Pyrenula
hibernica (VU & BAP) and Pyrenula microtheca (NT), occur as extreme
rarities southwards, they retain larger populations in the Highlands,
A few southern species are also included with this group, including Graphina
pauciloculata (VU & BAP), a partial parasite of the Nationally Scarce lichen
Graphina ruiziana. There are also three species occurring on sheltered rock
in woodland - Arthonia atlantica (NT & BAP), Parmotrema robustum (CR &
BAP) and Porina effilata (CR & BAP). Finally, the severely threatened
Bryoria smithii (CR & BAP) has only recently been known from high altitude
oakwoods on Dartmoor.
Temperate oceanic woodland
The species grouped as temperate oceanic species are less confined to very
heavy rainfall or coastal areas and most are more southern in distribution.
Many get as far east as the New Forest and a few to Sussex along the south
coast. Species which are widespread in both the north and south or more
northern are Heterodermia japonica (NT), Megalospora tuberculosa (NT &
BAP), Fuscopannaria sampaiana (NT & BAP), Parmeliella testacea (NT &
BAP), Parmotrema arnoldii (NT) and Wadeana minuta (NT & BAP). The
southern oceanic species are typically absent from the west Highlands or are
rare there and often confined to the sunniest areas. The latter include
Agonimia octospora (NT), Arthonia invadens (NT & BAP), Mycoporum
lacteum (NT), Micarea pycnidiophora (NT), Porina hibernica (NT & BAP),
Porina rosei (NT), Rinodina isidioides (NT & BAP) and Wadeana
dendrographa (NT & BAP). Other species are absent from the west
Highlands all together and include Arthonia anglica (EN & BAP), Arthonia
astroidestera (NT), Blarneya hibernica (NT & BAP), Enterographa sorediata
(NT & BAP), Melaspilea lentiginosa (NT & BAP), Parmelinopsis horrescens
(NT & BAP), Parmelinopsis minarum (VU & BAP), Phaeographis lyellii (NT &
BAP) and Ramonia nigra (CR & BAP). Two extremely southern species are
Cryptolechia carneolutea (VU & BAP) and Opegrapha prosodea (NT & BAP),
which are confined to the far south of England and Wales.
Temperate woodland
The species found in temperate woodland but not showing strong oceanic or
sub-oceanic tendencies are a mixed group. All show large gaps in their
distributions centred on central England, representing areas of past severe
air pollution. The group includes some little known species, such as
32
Comment [TCW5]: Did Neil
add ‘stage’? By morph he
means ‘form’. Suggest change.
Arthopyrenia atractospora (NT), Chaenothecopsis caespitosa (NT) and
Chaenothecopsis savonica (NT), with a few widely scattered records, but
also includes more distinctive groups of species. These include rare species
that are widely distributed in unpolluted areas of England and Scotland,
including Bacidia circumspecta (VU & BAP), Bacidia incompta (VU & BAP),
Biatoridium monasteriense (EN & BAP), Pertusaria velata (VU & BAP),
Phlyctis agelaea (NT) and Schismatomma graphidioides (VU & BAP). More
southern species with distributions including the Welsh Marches and
southern England are Caloplaca herbidella (VU & BAP), Caloplaca lucifuga
(VU & BAP), Lecanora quercicola (NT & BAP) and Lecanora sublivescens (NT
& BAP). A group of very southern species mainly found in southern England
but showing no oceanic tendencies in their distribution, includes Collema
fragrans (EN & BAP), Enterographa elaborata (CR & BAP), Megalaria laureri
(EN & BAP) and Pyrenula nitida (VU & BAP). A final group of Collema
species, Collema fasciculare (NT & BAP), Collema nigrescens (NT) and
Collema occultatum (NT), are highly pollution sensitive species that were
once widespread across Britain but have retreated to the west and north
since the early 19th century.
As would be expected, the species classified as species of temperate –
boreal woodlands are mainly northern species in Britain such as Leptogium
saturninum (VU & BAP), Biatoridium delitescens (VU), Chaenothecopsis
vainioana (NT) and Ptychographa xylographoides (NT). Also included in this
group are two species, Caloplaca flavorubescens (EN & BAP) and Gyalecta
ulmi (EN & BAP), which were once more frequent in the south on elm, but
whose northern populations have survived better, the latter on rocks.
Temperate sub-oceanic woodland
The final group of temperate woodland lichen species are confined to suboceanic areas along the east cost, which have maintained clean air and have
surviving woodlands. The richest area is in the eastern Highlands from
Perthshire eastwards and north to the lowlands and glens around the Moray
Firth. Some species also survive sparingly in the eastern Scottish Borders
and North East England. The occurrence of some lichens with this suboceanic distribution, in the rain shadow area of Welsh Marches, suggests
that many would once have extended further into eastern England. The
flora is similar to that found in broadleaved woods in southern Scandinavia.
Most rare species are small crust forming species but the larger
Fuscopannaria ignobilis (VU & BAP) is more charismatic, while the dogtooth
Peltigera lepidophora (CR & BAP) is known from rocks in one Angus ravine
woodland. Smaller species include Bacidia subincompta (VU & BAP), Buellia
violaceofusca (NT & BAP) and the pin heads Calicium adspersum (CR & BAP),
Chaenotheca chlorella (NT), Chaenotheca gracilenta (EN & BAP) and
Chaenotheca laevigata (EN & BAP). There is an overlap with boreal
woodland, with Biatora efflorescens (NT), Catinaria neuschildii (VU),
Sclerophora pallida (VU & BAP) and Sclerophora peronella (NT) found on
broadleaved trees in both mixed temperate woodland and boreal
woodlands.
33
Boreal woodlands
The large number of rare lichens confined to boreal native pinewoods and
mixed birch woods is a great contrast to the much smaller numbers of rare
vascular plants and bryophytes species confined to this habitat. The lichen
flora of these woods has only been seriously explored in the last few
decades and was almost totally unknown before 1974 (Coppins & Coppins,
2006), but has proved to be rich and significant. Floristic variation within
the boreal woodland lichen flora, as in temperate woodlands, is large east
to west. Rare species characteristic of rich old boreal woodlands across this
climatic gradient are Alectoria sarmentosa ssp. sarmentosa (NT) on pine and
birch and Elixia flexella (NT) on pine lignum. Most rare species, however,
can be grouped into either western or eastern assemblages. The western
group typically occur as far east as the Glen Affric area and central
Perthshire, while eastern species are missing from the pine woods in the
hyper-oceanic west coast (Coppins & Coppins, 2006). This leads to the very
richest woods being found between the extremes, including the Black Wood
of Rannoch, Affric, Strathfarrar and Guisachan. Lichens of western and
central pine woods are Buellia arnoldii (NT) and Buellia sanguinolenta (NT)
on broadleaved trees and Calicium parvum (NT), Melaspilea lentiginosula
(NT) and Micarea elachista (EN) on pine. In addition, a few very rare hyperoceanic lichens occur on broadleaved trees only in the western woods. A
striking example is the remarkable hyper-oceanic pine wood at Barrisdale,
where rowans support Pyrenula dermatodes (CR & BAP) and Ramonia
azorica (EN), two strongly temperate hyper-oceanic species at their
northern most occurrences.
The central and eastern element includes many specialist of pine lignum,
including Chaenotheca xyloxena (VU), Chaenothecopsis pusiola (NT),
Cladonia botrytes (CR & BAP), Cladonia cenotea (NT), Cyphelium tigillare
(NT), Hypocenomyce anthracophila (EN), Pycnora xanthococca (VU) and
Xerotrema megalospora (NT). Bryoria furcellata (VU & BAP) occurs on pine
bark and lignum and occasionally birch. Few rare species are confined to
pine bark in these woods, with Hypogymnia farinacea (NT) an example.
This species and Alectoria sarmentosa ssp. sarmentosa (NT) are the only
rare boreal woodland specialists with recent records from northern England
in the Pennines. Here they are confined to rocks, where they could be relic
boreal woodland species (Gilbert, 2000). Very recent exploration of old
aspen stands in the eastern Highlands has added a rich assemblage of rare
boreal species, centred on Speyside, but also found in Deeside(see Royal
Botanic Garden Edinburgh website http://rbg-web2.rbge.org.uk/lichen/).
This has resulted in an avalanche of new species to Britain, with several now
added to the RDB or listed as BAP species: Arthonia patellulata (DD & BAP),
Bacidia igniarii (VU), Bacidia vermifera (EN), Caloplaca ahtii (DD & BAP),
Diplotomma pharcidium (DD & BAP) and Lecanora populicola (NT). Also
discovered was Candelariella superdistans (DD & BAP), a lichen parasitic on
Lecanora populicola (NT). Juniper stands can also support the striking
yellow lichen Vulpicida pinastri (NT & BAP).
34
Finally, only two rare species appear to have been recorded from pine
plantations in eastern Scotland - Lecidea antiloga (VU) and Ochrolechia
arborea (NT). The former has also been recorded from native woodland in
Speyside and the latter is a recently identified species.
Non woodland epiphytes
A total of 16 species of epiphytic lichens are included in either the RDB, at
Near Threatened status or higher, or are BAP species but were not
considered to be woodland species. Of these, eight showed general
temperate distributions, seven were temperate oceanic species and one was
temperate sub-oceanic. All are strongly associated with field trees in
cultural landscapes. There will be some overlap with sites rich in rare
woodland lichens. This is particularly so where non-woodland epiphytes
occur on more open trees in parklands that also include denser and or more
sheltered stands of veteran trees. Many of the rare field tree species were
strongly associated with old elm trees.
Extinctions
Fifteen woodland species are listed as extinct, i.e. not seen for 50 years or
more. A further species, the southern oceanic Pseudocyphellaria aurata (CR
& BAP), is extinct in woodland but still occurs in a single coastal grassland in
the Isles of Scilly. For micro-lichens, extinction may not be forever, as
lichenology is still in an exploration phase. An example is the New Forest
endemic Bacidia subturgidula (CR & BAP), a specialist of lignum exposed on
standing live old hollies. This species was collected twice in the 19th
century but was not seen again until 2003, when it was rediscovered in the
New Forest. It is now known from a second site in the New Forest
(Sanderson & Cross, 2003; Cross et al., 2006). The lack of oceanic species
in this list of extinct species, compared to their abundance in the list of rare
species, is a striking feature. The one extinct species of hyper-oceanic
woodland, Arthothelium spectabile (Ex) is a small species that might yet
turn up. The species assigned to temperate woodland include four species
which were in all probability confined to field trees, leaving a couple of
dead wood pin head Calicium species that were probably mainly sub-oceanic
in their distribution. There are, however, proportionately more species of
sub-oceanic temperate and boreal woodland that are likely to be have been
woodland species than in the other groups. All this reflects the current
extent and importance of oceanic woodlands and greater historic losses and
pressure on woodland in the east.
LICHENS – SPECIES OF CONSERVATION INTEREST
Distribution within climatic regions
Extinctions
Temperate
Temperate, hyper-oceanic
Temperate, oceanic
Extinct
6
1
1
35
Extant
38
35
33
% Extinct
16%
3%
3%
Temperate, sub-oceanic
Boreal
Totals
4
4
16
17
32
154
24%
13%
10%
3.3.2 Red Data Book and BAP species – habitats and threats
General factors – woodland structure and scale
In spite of the strong distinctions between woodland lichen floras of
different climatic zones, there are considerable common factors which
apply to all habitats. In particular, most of the rare woodland species
covered here are old growth dependent to some degree, either being most
frequent in old growth woodlands or confined to such woodlands (Coppins &
Coppins, 2005 & 2006; Coppins et al., 2002; Rose, 1992 & 1993; Sanderson,
1996 & 2010.). Some lichens are specifically associated with very old trees.
For most species, however, the association with old growth woodland is a
combination of a requirement for niches best developed in post mature
(trees beyond their economic usefulness but not necessarily that old) or
ancient trees, slow growth rates requiring habitat stability and poor
colonisation ability requiring long times for colonisation of new trees. The
resulting combination is best described as continuity. As a result, although
stands rich in rare lichens will have very old trees present, the rare lichens
are often not found on the largest trees. The presence of very old trees is a
measure of stand continuity, rather than necessarily the only habitat of such
lichens (Gustafsson et al., 1992). An extreme example of this are
undisturbed Atlantic hazel woods; here rare species are confined to old
uncoppiced hazel stands, which can be identified by the presence of old or
dead hazel stems on the hazel bushes. The rare species, however, are
mainly found on younger stems and require a succession of nearby younger
stems to survive, which is lost in coppiced hazel stands (Coppins et al.,
2002).
Lichens in general tend to have much stronger niche specificity than green
plants and rare old growth dependent species are among the fussier lichens.
Niche diversity is greatest in old growth woods with varied structure and
composition. Structural variation, especially in patchy woods with a glade
and grove type structure, produces old leaning and twisted trees with varied
bark water regimes. Uniformly dense woods with predominantly straight
trees are much poorer habitats for lichens. In addition large diameter
standing dead wood and, to a much lower degree, fallen deadwood can also
be a significant habitat, and is strongly associated with old growth
woodland. Variations within and between trees and shrubs in bark pH are
also very important in producing niche diversity and are best developed in
stands with a range of both species and age classes. Above this is the
general liking of lichens for light; few lichens survive in heavily shaded
conditions. To complicate the issue further, woodland lichens are mainly
defined by their requirements for sheltered conditions, with higher humidity
found than outside woodlands. Typically, long exposure to strong summer
sunshine is avoided. Species more resistant to desiccation are likely to be
widespread also on trees growing outside of woods. Across the sharp
oceanic climatic gradient in Britain, this can lead to species which are
36
confined to very sheltered woodland conditions in the dry and sunny east
being much more tolerant of open conditions in the damp and cloudy west.
This requirement for both good, if indirect, illumination and shelter is
critical and means that classic lightly grazed high forest non-intervention
stands, even if old growth, are not high quality lichen habitats. Rapid infill
of glades by tree regeneration found in such stands produces conditions that
are too dark for rich epiphytic lichen floras to survive. Lichen-rich
woodlands require some degree of regeneration failure, in addition to welldeveloped old growth conditions. The primary mechanism for this is
localised high browsing pressures maintaining open old growth stands. This
type of woodland structure supports notable concentrations of rare and
declining lichens from habitats as disparate as New Forest beech – oak –
holly stands (Sanderson, 2007a) and Highland native pinewoods (Coppins &
Coppins, 2006). Obviously, such regeneration failures will needed to be
localised, over time and spatially, or the woodland will eventually be lost.
Historically, such habitat has mainly been provided by extensive grazed old
growth woodlands on unenclosed grazings - pasture woodland. It could also
occur in the larger and more semi-natural deer parks (Rose, 1992). Such
habitat must have also occurred in the wild wood, if these lichen
assemblages are not totally anthropogenic. The processes postulated by
Vera (2000), where grazing and browsing by wild animals partly drove the
structure of the wild wood by maintaining temporary glades, which were
later infilled, would certain provide this habitat.
Scale of habitat is also important, and some rare lichens can occur at very
low densities, even in apparently suitable habitat. An example is the
veteran beech specialist lichen Megalaria laureri (EN & BAP); this is
currently known from less than 30 trees in the New Forest, an old growth
complex covering 3,000ha and with approximately a quarter of a million
post mature and ancient trees (Sanderson, 1999 & 2001b). Even if the real
population is four times as abundant as the known trees, this is still a
frequency of tree occupancy of about 0.05%. Very large areas of old growth
woodland are needed to support some very specialist species. Sanderson
(2010.) recorded that 56 RDB species (CR to NT) and 30 BAP species had
been recorded from the 3,000ha of old growth woodland in the New Forest
between 1967 and 2007. Ellis & Coppins (2007) found that the diversity of
woodland specialist micro-lichens on discrete aspen stands in Scotland was
positively correlated with the existence of large areas of woodland in the
mid 19th century, not current day woodland size. In this case, stand-scale
epiphyte assemblies were coupled to the dynamics of the wider woodland
ecosystem, but with a significant lag in the response of epiphyte species
richness to habitat spatial structure. Even the aspen specialist Arthonia
patellulata (DD & BAP), which is not associated with old growth aspen
stands, is still associated with unfragmented woodland landscapes (see
project account on Royal Botanic Garden Edinburgh website http://rbgweb2.rbge.org.uk/lichen/) .
In summary, rich assemblages of rare lichens, from southern beech woods to
Highland native pinewoods, are strongly associated with large varied
37
structured pasture woodlands, preferably with a glade and grove structure,
providing areas of sheltered but well illuminated veteran trees (Sanderson &
Wolseley, 2001). Also significant are more relic situations, such as in old
parks or small areas of retained veteran trees, as on the boundaries of
oceanic woodlands in Exmoor (Wolseley & O’Dare, 1989).
Woodland structure – threats
Threats include felling of old growth woodlands, not necessarily for
woodland clearance but also as a part of normal woodland management. At
first sight, it would seem astonishing that this is a current threat, but
existing old growth woodland has no specific protection in Britain
(Alexander et al., 2002) and has poor recognition. Old growth is often
thought to be a feature of unmanaged foreign ‘virgin woodlands’ and the
concept of managed old growth woodland in Britain has not been wide
appreciated. The issue is particularly problematic in the uplands, where
veteran trees are smaller and less spectacular, and the age of the stands is
easily overlooked (Quelch, 2001). Old hazel stands rich in rare lichens have
proved particularly vulnerable (Coppins et al., 2002). Conservation
coppicing of probably never previously coppiced stands of hazel has been
carried out in western Scotland with the best of intentions, but with
devastating consequences to internationally rare lichen assemblages.
In the historic past, large scale losses of old growth woodland occurred as
unenclosed woodlands were enclosed in the 18th and 19th century.
Conversion to oak plantation was as devastating to lichen diversity as
clearance to open land. Many thousands of hectares of such woodland were
lost to plantations of native trees in the 19th century and yet more to
plantations of exotic conifers in the 20th century. In the former case,
however, recovery is possible; significant partial recolonisation by old
growth dependent species has been observed in 18th century oak plantations
in the New Forest. These were abandoned back to pasture woodland in the
19th century (Sanderson, 1996 & 2010.). Within 200 -300 years, good
recovery has occurred in all communities except those of dry bark on
veteran oaks and rain track and wound track assemblages of veteran beech.
In western Britain, the numerous 19th industrial oak plantations are
potential sites for such recolonisation, aided by the ability of some rare
leafy species to survive as relic populations on rocks.
Maintaining the quality of old growth woodland is an even more serious
current threat to lichen diversity. This includes the long appreciated threat
in parts of the uplands of centuries of high grazing levels, preventing any
regeneration and leading to the eventual loss of the woodlands. Lichen
diversity declines long before the total loss of trees; pasture woodlands
consisting of nothing but savanna like stands lose critical humidity in all but
the most sheltered topographies. Loss of woodland area, or tree density, is
also likely to induce an extinction debt, even in reduced, but still sheltered
and lichen rich stands (Ellis & Coppins, 2007).
A less appreciated but growing threat is the complete removal of grazing
from previously grazed woodlands as an understandable response to
38
overgrazing. As many woods rich in lichens were traditionally grazed, and
unenclosed woods may have been grazed continuously since trees returned
after the last ice age, this is a drastic action. Although removing or
drastically reducing grazing pressure will usually allow woodland to
regenerate and be saved from any actual or perceived threat, it also has the
potential to devastate rich epiphytic lichen floras by increasing shade
(Coppins & Coppins, 2005; Sanderson, 1998a and 2010.; Sanderson &
Wolseley, 2001). Declines can be rapid on fertile soils, and the expanding
canopy of existing shrub layers can be as problematic as new regeneration
(Coppins & Coppins, 1998). In less fertile sites, damaging shade can take
longer to develop. Increased cover by shade bearing late succession
species, both native species such as holly, ivy and regionally beech, and
introduced species such as sycamore and regionally beech, are particular
problematic with reduced grazing pressure. The toxic exotic evergreen
Rhododendron ponticum is a very serious problem in the west. It is not
controlled directly by grazing but, in woods managed by stock grazing, the
graziers have a considerable incentive to control seedlings, which has
preserved many important woods (B. J. Coppins pers. com.). Woods not
grazed by stock have been noticeably more vulnerable to rhododendron
invasion in some areas.
A further issue is that the total removal of grazing often fails to regenerate
a similar mixed composition of the existing pasture woodland, with dense
birch regeneration produced from the New Forest to Scotland. Sanderson
(2004) noted that total removal of grazing on New Forest floodplains tended
to produce species-poor birch dominated woodland, while secondary pasture
woodland developed under continuous grazing pressure, though thorn scrub
was mixed and dominated by ash. Bakker et al. (2004) give a detailed
description of the regeneration processes involved in an ash rich grazed old
growth floodplain woodland in the New Forest. An appreciation of the need
for balanced grazing regimes within anciently grazed woodlands, not the
precipitous removal of all grazing, is required. Research into appropriate
grazing regimes is desperately required. These are likely to involve
variations in grazing pressure over time, with periods of maintenance
grazing that suppresses much of the regeneration, alternating with periods
of lower grazing pressure allowing more regeneration. Fewer deer and
sheep and more cattle and ponies may also be beneficial.
General factors – air pollution
Many lichens are very sensitive to air pollution, with past sulphur dioxide
pollution especially damaging (Hawksworth & Rose, 1970). Only areas that
remained clean through the last two centuries harbour rich epiphytic floras.
As well as direct poisoning by sulphur dioxide, milder but still damaging
acidification has selectively removed acid sensitive species in high rainfall
areas further from pollution centres (Farmer et al., 1992). The result is
that large areas of the country have been effectively sterilised. Much of
England, from the Thames valley to north east England, has lost almost all
slow colonising old growth dependent lichens, as have south east Wales and
the central belt of Scotland. This acidifying pollution has now greatly
declined and large scale recolonisation by rapid colonising species is
39
underway, but these do not include any rare species. Old growth dependent
species have been lost from these areas for the foreseeable future. Pristine
clean air areas survived in the Scottish Highlands and the far south west of
Wales and England. Wide zones of relatively clean air survived along the
south coast of England, north Wales and the northern Lake District. Here,
some exceptionally sensitive species, such as the attractive leafy
Pseudocyphellaria species, have declined and are threatened with
extinction. This is resulting in oceanic species such as Fuscopannaria
sampaiana (NT & BAP) and Parmeliella testacea (NT& BAP) beginning to
assume distributions similar to hyper-oceanic species, as the western
Scottish populations become more significant. These species are not yet
obviously recovering. The areas of relatively clean air, however, still have
sites rich in less extremely sensitive rare lichen species.
The threat from sulphur-based pollution has greatly receded, but nitrogen
pollution is now an increasing threat. As far as lichens are concerned the
main threat is from ammonia from agriculture; nitrous oxides are apparently
not directly assimilated by lichens (van Herk, 1999). Effects have been
demonstrated by nitrous oxides from traffic, but only at the exceptionally
high levels found in inner London (Davies et al., 2007). Ammonia from
point sources from agriculture has only a short distance effect, but in areas
of highly intensive agriculture the sources combine to produce regions with
much altered epiphytic floras (van Herk, 1999). Similar effects have been
demonstrated from Britain (Wolesey et al., 2006). As far as rare epiphytic
lichens are concerned, the main threats are to field tree specialists, but
woodland lichens are also threatened in parkland and woodland sites in
zones of intensive agriculture.
General factors – elm disease
Dutch elm disease has had a devastating effect on a group of specialist
species, which were largely confined to old elm trees (Edwards, 2005).
Many, but not all, are wound track specialists. Some, such as Anaptychia
ciliaris ssp. ciliaris (VU & BAP) and Caloplaca luteoalba (VU & BAP), were
confined to field elm species. Others, including Bacidia incompta (VU &
BAP), span the field tree and woodland habitat, while Bacidia circumspecta
(VU & BAP) is confined to woodland habitats. The degree of dependence on
elm varies, with field tree species such as Anaptychia ciliaris ssp. ciliaris
(VU & BAP) and Caloplaca luteoalba (VU & BAP) and field tree occurrences
of Bacidia incompta (VU & BAP) confined largely to elm. These species in
this habitat are unlikely to survive. In woodlands, most elm specialists have
alternative substrates with wound tracks on beech trees and ash or hollow
holly, maple, sycamore and ash, but these only support a fraction of the
former populations of these lichens and may not represent sustainable
populations. A few, such as Bacidia incompta (VU & BAP), populations on
wound tracks and hollow trees in the old growth pasture woodlands of the
New Forest are still large enough to be likely to be sustainable.
Hyper-oceanic temperate woodland – habitats and& threats
The western Scottish stronghold of this habitat has many large lichen-rich
woodlands with a history of pastoral use, which are surviving old growth or
40
recovering old growth woodlands. These are one of the most important
habitats for lichen diversity in Europe and the temperate northern
hemisphere. Coppins & Coppins (2005) describe the habitat in Scotland.
The rare species among the distinctive species defining the assemblage are
largely confined to two habitats - smooth bark on hazel and sometimes
rowan and rarely holly (Graphidion – undescribed hyper-oceanic
communities) and base rich mossy bark (Lobarion) on many species of trees,
but particularly ash, hazel and elm. Other habitats, as on acid bark
(Parmelietum laevigatae) on birch, alder, willow and oak, support
characteristic hyper-oceanic lichens, but have fewer rare and threatened
species. Coppins & Coppins (2005) point out the lack of distinctiveness of
the flora of oak in this area and the greater importance of species such as
hazel and ash. As they discuss, this makes the commonly use name ‘Atlantic
Oak Woods’ for hyper-oceanic temperate woodlands rather problematic. A
better and more accurate name would be ‘Temperate Rainforest’ (Rhind,
2003). This problem has had negative practical consequences in SSSI and
SAC notification, where plantations of oak have been preferentially notified
over true native woodlands.
The resource of old lichen-rich woods in western Scotland is immense, but is
under explored. It is likely that many woods of international importance
have not yet been visited and many others are little known. As a result
statutory notification as SSSI or SAC does not reflect the distribution of the
interest, nor are consultations for forestry or regeneration works likely to
take lichen interest fully into account, if the interest is unknown.
Plantlife’s recent listing of the most of the west coast of Scotland as an
Important Plant Area may help alleviate this problem.
The flora reappears with a reduced diversity in north Wales and the Lake
District but the hyper-oceanic element here has probably been greatly
reduced in historic times by heavy exploitation of hazel by coppicing
(Sanderson, 2006a), which was far more intense than in the west Highlands
(Gilbert, 1984; Coppins et al., 2002; Smout et al., 2005). In addition, the
species of base-rich bark have lost ground here to acidification from acid
rain. Acid rain has also damaged lichen floras in the far south west of the
west Highlands, but most areas were never affected.
Rhododendron invasion is one of the most serious issues in hyper-oceanic
woodlands. Much is being done to tackle it in many sites, but this needs to
be continued and accelerated. Beech here, although well adapted to the
climate, is a recent introduction and the local lichen flora appears poorly
adapted to its hard bark. In addition it casts a very deep shade which is a
significant threat, not only to hazel specialist flora, but the hazel itself.
The non-native sycamore and Norway maple can however have rich Lobarion
communities, with rare species. These maples can deeply shade ungrazed
woods but, unlike the grazing resistant beech, are controlled in pasture
woodlands, where they are not a serious lichen conservation problem. All
these tree species are European natives whose absence is due to
happenstance in post glacial tree dispersal; automatic eradication is
41
probably not advisable. The damage to native biodiversity should be
critically assessed first.
Regeneration failure was, and is, a major issue in many woods, especially
woods within deer forests. It is interesting to note, however, that many
woods within crofting common grazings have regenerated in the presence of
grazing without special assistance (Noble, 1997). Regeneration is now
being addressed in many woods, but often this is not carried out in a way
that shows much respect to the long tradition of sustainable woodland
grazing that has formed these woods and benefited their important
bryophyte and lichen floras (Rothero, 2005; Coppins & Coppins, 2005). This
is emerging as a major threat to the hyper-oceanic flora, an astonishing fact
given that much of the grazing exclusion is publicly funded and is meant to
be a conservation measure. Until recently, a significant problem appears to
have been the split between forestry and agro-environmental grant aid.
Denser and better preserved pastoral woodlands were deemed to be
woodlands, within which there was limited encouragement to graze and
there are specific grants for excluded grazing. Only open degraded pastoral
woodlands are included in some Scottish definitions of wood pasture, which
define wood pasture as being trees over grassland (Stiven & Hall, 2004) or as
stands with less than 20% tree cover (Smout et al., 2005). These were more
likely to be included within agro-environmental schemes and with
sustainable grazing encouraged. These habitats, however, are much poorer
habitats for woodland lichens than denser pastoral woodlands. There is,
more recently, increasing acceptance of the positive role of grazing in
woodland conservation with initiatives such as the very positive Woodland
Grazing Toolkit developed for Scotland (Sumsion & Pollock, 2005).
A significant recent problem within hyper-oceanic woods is an attempt to
revive wood and timber exploitation (Coppins et al., 2002). The richest
areas for lichens are long undisturbed woodlands, often well on the way to
recovering from past (18th or 19th century) exploitation, or in some cases
probably barely exploited at all, but maintained mainly as sheltered grazing
(Watson, 1997; Smout et al., 2005). There is no evidence that significant
wood or timber exploitation does anything other than depress lichen
diversity in these woods (Gilbert, 1984). It is imperative that a balanced
approach is adopted, where old growth woodland is given protection and
forestry management is directed to more disturbed or new woods.
There has been a great deal of work started to restore native woods
damaged by conifer planting, but this rarely seems to pay much attention to
the character of the original woodland. Restoration appears often to be by
rote, with former pasture woodlands, planted with conifers, being
converted to dense birch stands or oak plantations (Sanderson, 2006b). The
potential for restoring dull industrial oak woods originating from intensive
19th century management back to more diverse native woodlands has been
raised by lichenologists (R. Woods, pers. com.; Coppins & Coppins, 2005) but
appears to have been little considered.
Oceanic temperate woodland – habitats and threats
42
This ranges from upland woodlands in England and Wales, locally with a
hyper-oceanic element, through to lowland pasture woodlands in clean air
areas in the south and west. Both the southern temperate oceanic species
and temperate species are significant components of the threatened lichen
flora. A notable feature of these southern woods is the increasing
importance of oak compared to Scottish hyper-oceanic woodlands as a
substrate for rare and threatened lichens. These occur in both mossy baserich bark communities (Lobarion) and in other communities poorly
developed in the wettest areas. The latter include mesic bark on old oak
(Pertusarietum amarae) with crust forming species such as Pertusaria velata
(VU & BAP), which lose out to competition with mosses in very wet woods.
In the drier lowlands there is also a suite of species found on parkland trees
and glade edge trees in woodland in this habitat, including Caloplaca
lucifuga (VU & BAP), Lecanora quercicola (NT & BAP) and Lecanora
sublivescens (NT & BAP). A particularly important community is found on
dry bark on ancient oaks (Lecanactidetum premneae) in both woods and
parkland. This community requires very long continuity of old oaks
(Sanderson, 1996) and is of international importance in its own right.
Smooth bark communities (Graphidion – Graphidetum scriptae) are still
important but with holly significant as well as hazel, and with southern
oceanic species appearing in place of the hyper-oceanic species. Base-rich
bark communities are also different, with oak much more significant,
especially for small crustose species, which are squeezed out by mosses in
very wet areas. Acid bark (Parmelion laevigatae community) still supports
some rare species in the west. Towards the east, standing dead wood
becomes more significant for old growth biodiversity but rare species are
nearly all sub-oceanic specialists. Wound track and rain track species also
become important, with a particularly rich beech flora in the New Forest,
the only substantial area of old growth beech woodland surviving within a
clean air area (Sanderson, 2007a & 2010.).
Old growth stands are much less widespread south of the Scottish Highlands,
reflecting more intensive woodland management in the medieval and early
modern period. There is a closer match between sites rich in rare lichens
and SSSIs than in the Highlands, but some important sites are not covered.
SAC coverage is often poor, due to the problematic translation of European
classifications used in the Habitats Directive and the exclusion of ancient
parks from the Directive.
Many threats are similar to the hyper-oceanic woods in Scotland, especially
in the uplands. Rhododendron is also a locally serious threat, although less
widespread or as rapidly colonising in drier areas. Beech is native to the
south east but an invasive non-native to the north west. In native beech
areas, clean air old growth stands are an extremely important resource for
lichen conservation. Beyond, it is normally not only a poor substrate but its
deep shade can be a threat to lichen diversity. This raises complex issues
about post glacial spread of slow colonising trees, which are not yet at
climatic equilibrium (Peterken, 1996; Svenning & Skov, 2007) and climate
change issues (Wesche, 2003). In the south, shade cast by expanding ivy
and holly is developing into a serious issue, especially in ungrazed woods.
43
This may be due to warmer winters. Even in the well grazed New Forest
woods, cutting is required to control holly (Sanderson, 1991 & 1997; Wright
& Westerhoff, 2001) and prevent losses of rare epiphytic lichens.
Grazing is also crucial and is especially problematic in Wales, where there
are no deer, so simple stock fencing can easily exclude all grazing. There
has also been a lot of encouragement for fencing off woods in agrienvironmental schemes here. As a result, undergrazing is now a greater
threat to rare woodland lichens than overgrazing in Wales (R. Woods, pers.
com.). In the lowlands, outside of the New Forest, few pasture woodlands
are now grazed, but restoration schemes are planned for some important
sites, as at Savernake Forest (Sanderson, 2007b). Many important sites are
undergrazed and deteriorating.
The New Forest, in contrast, has remained grazed, and must be one of few
areas of extensive grazing in Britain not to drastically change its mix of
grazers and browsers in the last three centuries (Tubbs, 2001). It is still
dominated by the coarse grass grazers cattle and ponies, while most upland
pasture woods have largely lost dominant cattle grazing and become
dominated by the fine grass grazers/browsers deer and sheep. It is not
clear to what degree this or climate produces the much greater resilience to
continuous grazing shown by the New Forest pasture woodlands, than that
shown by upland pasture woodlands in the last few hundred years.
Landscape parklands containing relic pasture woodlands are important in
the clean air areas in the lowlands. An example of an exceptionally
important site is Melbury Park in Dorset, which contains a rich lichen flora
characteristic of both woodland and field trees. Some deer parks such as
Whiddon Park in Dartmoor are still essentially semi-natural pasture
woodlands, but most parks are now rather artificial environments, with the
trees managed by planting rather than natural processes. Many have been
agriculturally improved and have been damaged by nutrient enrichment
from fertiliser use, high stocking levels and felling of old trees. Most of the
best parks are SSSIs, and any problems of intensive use and lack of tree
planting should be dealt with by the condition assessment process. There
are however, important non-SSSI sites and these are much more vulnerable.
As the proportion of surviving old growth woodland is lower, there is an
even greater case for planning the restoration of industrial oak plantations
to structurally varied old growth woodlands than that given for the western
Highlands (Coppins & Coppins, 2005). Large scale conversions of 18th and
19th century oak plantations to old growth pasture woodland are planned in
the New Forest to counteract past fragmentation, amounting to several
hundred hectares (Forestry Commission, 2008; Sanderson, 2007c). There is
immense scope for more of this in areas such as Dartmoor and Exmoor and
mid and north Wales, where there are extensive oak plantations associated
with relic lichen-rich woodlands.
44
Sub-oceanic temperate woodland – habitats and threats
Although this assemblage has been lost from polluted areas in eastern
England and Scotland, it is still patchily distributed in eastern clean air
areas. The assemblage is best developed in the eastern Highlands and
adjacent lowland around the Moray Firth, with some species reappearing in
the east Scottish borders, north east England, east Wales and the Welsh
Marches. Beyond this, large areas of eastern Scotland, with clean air, have
little or no ancient woodland of any sort, let alone old growth woodland.
These areas relied on timber imports from Scandinavia and failed to
conserve their woodland as a result (Smout et al., 2005). The assemblage
is absent from these areas.
Woods deeper into eastern Highland glens grade into boreal woodland and
are more likely to be intact pasture woodlands. Beyond this the lowland
woods tend to be more altered by 19th century forestry but, given the rarity
of woodlands in this area, old 19th century oak plantations, as at Dinnet oak
wood in Aberdeenshire, can still be valuable as lichen refugia. A lichen-rich
parkland survives at Drummond Park in Perthshire.
The inland pasture woodlands share the same issues as upland pasture
woodlands in the west, with a need to find ways of introducing sustainable
grazing regimes to avoid either over- or undergrazing. Invasion by shade
bearing exotics is again a problem, as at in the Lower Findhorn Woods,
where the locally exotic beech is an issue. Overall, however, the severe
level of historic deforestation beyond some highland glens is the most
glaring issue. There is a particular need to identify and protect surviving
old growth and recovering mature growth and to expand the habitat.
Practically, expansion is mainly an issue of allowing existing young growth
near surviving sites of interest to develop into old growth woodland. New
plantings could be useful by the late 22nd and the early 23rd centuries, but
are not an immediately practical response.
Boreal woodland – habitats and threats
The most obvious woodland type in this category is the well-known and
described native pinewoods. These are certainly magnificent habitats and
are one of the most significant old growth woodland resources in Britain.
They also include some of the largest old growth stands in Britain. The
habitat is also immensely important for the diversity of British lichens.
Within the pinewoods, the major lichen habitats, in rough order of priority,
are: standing or propped up dead wood, associated old broadleaved trees
and the bark of live pines (Coppins & Coppins, 2006). The specialist rare
and threatened dead wood lichens are mainly associated with well-lit dead
wood attached to live trees, or large hulks of dry standing dead trees or
propped (on other trees or their own branches) fallen trees. Shaded and
damp dead wood is not a rich habitat for rare lichens. A curious exception
to this is the small Cladonia botrytis (CR & BAP), which is strongly
associated with well-lit, old and high cut stumps (as produced by hand
cutting) in large felled areas that have not regenerated rapidly. A rare
species associated with cultural features produced by hand felling within
native pine woods and old Scots pine plantations, which are not
45
immediately regenerated, is something of a challenge to many
preconceptions of how these woods should be managed, either for
conservation or forestry.
Lichen-rich stands of native pine wood are typically old growth stands with
open canopies and much dead wood, with requirements for high light levels
crucial (Coppins & Coppins, 2006). Similar structures are also valued in
Scandinavia (Karlson et al., 1995; Nitare, 2000). In Sweden, these are
described as ‘fire-influenced coniferous natural woodlands’ and are
produced by frequent low intensity fires. Dense and well-regenerated
pinewoods were not listed as special habitats (Karlson et al., 1995).
In oceanic Scotland, grazing may have been more significant in producing
this structure than in continental Scandinavia (Fenton, 2004 & 2008), but it
is likely that both grazing and fire were involved. The role of grazing in
pinewood conservation appears quite controversial. As is clear in the
description of the habitat given by Rackham (2006) and in Smout et al.
(2005), the native pinewood habitat is the result of millennia of pasture
woodland management. Dennis (1998) argues passionately that, without
grazing by native animals such as cattle, many components of the pinewood
ecosystem will be lost. In contrast, official actions for the pinewood BAP
only mention grazing in a negative light.
The pinewood habitat is not under direct threat, as there is now plenty of
effort to ensure the woods survive and all large areas are in SSSIs and SACs.
The need now is to ensure that the quality of the habitat is not
compromised by these very efforts to conserve it.
In contrast to the well known pine woods, the other types of boreal
woodlands are little known. These are essentially mixed broadleaved boreal
woodlands that replace the pinewoods on more base-rich soils and on acid
soils beyond the native range of pine. Most are also old growth pasture
woodlands, rich in rare and declining lichens. They are birch dominated but
calling them birch woods is not terribly helpful, as it fails to distinguish
them from secondary scrub woods.
There has been no systematic study published on the lichens of this habitat,
but recent detailed studies on stands of aspen within such woods are
revealing them as an important habitat for boreal lichens (Royal Botanic
Garden Edinburgh website http://rbg-web2.rbge.org.uk/lichen/). Aspen
stands, mostly within birch-dominated pasture woodlands on better soils
and often in enclosed farmland, are rich in species new to Britain. Such
stands face the usual problems of past overgrazing and potential future
undergrazing. The former prevents regeneration; with the latter, stands are
too dark and dense to support rich epiphytic floras. Even less explored are
unenclosed higher altitude boreal woods on less acid soils (Friday, 1990).
These share many specialised species with the pinewoods, including rare
and declining species. They appear to be quite widespread in the central
Highlands north of Rannoch Moor and in the Monadhliath, where they fill in
a large gap in the pinewood distribution. They appear to be survivors of the
46
type of woodlands that Davies (2003b) detected from pollen in western Glen
Affric. The best-known example is on the Creag Meagaidh National Nature
Reserve. Most are now very damaged by long periods of high grazing
pressure and are desperately in need of action to save them, as has
occurred at Creag Meagaidh.
3.3.3 International responsibility
The conservation evaluation (Woods & Coppins, 2003) lists all species
thought to have 10% or more of their European populations in Britain ie.
International Responsibility species (IR). The revised draft used for this
report includes 180 species judged to be International Responsibility
species, of which a quick count suggests that 142 are woodland species
(79%). Combining RDB species of Near Threatened or higher threat status
with BAP species produces a total of 83 IR species which are rare or
threatened, of which 63 are judged to be woodland species (76%). British
woodlands support a large number of lichens for which Britain has an
international responsibility.
Even more interesting are the results of a breakdown of international
responsibility by woodland types for RDB and BAP species combined. This
shows that more than 50% of the species assigned to the temperate oceanic
and temperate hyper-oceanic elements are International Responsibility
species. The temperate element is intermediate, with 39% IR species and
the temperate, sub-oceanic and boreal elements have 12% and 13% IR
species respectively. Essentially many oceanic species that are threatened
in Britain are even more threatened in Europe, while more continental and
boreal species have much larger populations in Scandinavia. The combined
temperate oceanic and temperate hyper-oceanic elements together include
67% of all woodland International Responsibility species of conservation
interest.
LICHENS – INTERNATIONAL RESPONSIBILITY SPECIES
Distribution of those species of conservation interest, which are also
International Responsibility Species
Habitats/RDB & BAP spp
Temperate, oceanic
Temperate, hyper-oceanic
Temperate
Temperate, sub-oceanic
Boreal
All
IR
spp
22
20
15
2
4
63
Percentage
of
all IR spp
35%
32%
24%
3%
6%
All
spp
33
35
38
17
32
155
Percentage
IR spp in habitat
67%
57%
39%
12%
13%
41%
It should be noted that, although conserving western oceanic woodlands is
clearly an important responsibility, some individual sub-oceanic and boreal
old growth dependent species which are rare in Britain are also threatened
in Scandinavia (Nitare, 2000). All lichen-rich old growth stands are
potentially of international significance.
47
3.4
Habitats
3.4.1 BAP habitats
Given that Biodiversity Action Plan woodland habitats cover most seminatural woodland habitats, most species covered by this report will occur
within BAP habitats. Allocating species to individual BAP habitats was not
always straightforward, especially for lower plants, but most vascular plants
were relatively straightforward to assign. The plethora of woodland plans
for individual woodland habitats that cut across the habitats important for
lichens in particular was notable. A single ‘temperate rain forest’ in the
west of Britain could include habitats termed Upland Mixed Ashwoods,
Upland Oakwood, Upland Birch Wood and Wet Woodland. On top of this,
the whole wood would very likely be a pasture woodland, virtually all the
conservation interest grazing dependent, and therefore potentially covered
by the Wood Pasture BAP. In their turn, these habitats will include distantly
related habitats in terms of their overall flora, such as species-poor oak
plantations and sub-oceanic woodlands. The concentration on canopy
dominants simplifies the classification but a more integrated landscape and
land use based approach, focused on hot spots, might have in retrospect
been more sensible. It is notable that targets have latterly been set for
ancient woodland as a whole, rather than by the originally defined
individual habitats.
The Wood Pasture and Parkland plan does have more of a landscape and
land use basis and potentially includes the majority of woods of interest for
lichens. The extension of this BAP habitat into the uplands was on the face
of it sensible biologically. Much upland wood was traditionally grazed and
the majority of species of conservation interest (in most animal groups as
well as plants) within it are grazing dependent. Whether it has been
successfully applied is more questionable. In reality the Wood Pasture plan
appears to have been left with the crumbs of degraded open upland
woodlands, which are a poor habitat for pasture woodland lichens, while
lichen-rich woodlands are left in other plans. The other plans mostly still
have to show that the need for sustainable grazing to maintain lichen-rich
woodlands has been taken on board.
For vascular plants, Lowland Mixed Deciduous Woodland is a core BAP
habitat for rare plants, followed by Upland Mixed Ashwoods, reflecting the
dominance of base-demanding and lowland species. The majority of the
Upland Mixed species are also southern and lowland, as the ‘Upland’ Mixed
Ashwoods actually includes areas such as the Wye Valley and the Mendips,
which are upland only in having steep slopes and rock outcrops. Rodwell &
Dring (2001) observe that ’North Western Mixed Ashwood’ would have been
a better name. Lowland Beech and Yew Woodlands share many species with
both mixed woods mentioned above and also have a few specialist species.
Other upland BAP woodlands (Upland Oakwood, Upland Birchwoods and
Native Pine Woodlands) have small numbers of northern and boreal species.
48
Most of the latter species occur in woods that have traditionally been
grazed so are in habitats potentially covered by the Wood Pasture plan.
Quite a few lowland species also occur in pasture woodlands. Wet
Woodlands are of very limited significance, probably reflecting the fact that
many Wet Woodlands are recent woodlands replacing richer open habitats
(Sanderson, 2007d).
VASCULAR PLANTS – PRIORITY BAP HABITATS
Distribution of species of conservation interest within BAP habitats
Priority BAP habitat
Upland Mixed Ashwoods
Lowland Mixed Deciduous
Wood Pasture
Lowland Beech and Yew Woodland
Upland Oakwoods
Native Pine Woodlands
Upland Birchwoods
Wet Woodland
No of
species
19
18
12
9
6
5
5
2
Percentage of total no of species
of conservation interest
38%
36%
24%
18%
12%
10%
10%
4%
A particularly curious feature is that several characteristic boreal pinewood
species have more populations within mature Scots pine plantations than
within in native pinewoods. These are not included within the native
pinewoods, which is ironic, as exactly equivalent mature oak plantations
tend to dominate actions within the Upland Oakwood BAP.
For bryophytes, attempting to determine exactly which hyper-oceanic
species were found in either or both of Upland Mixed Ashwoods and Upland
Oakwood was not easy, but most ravine species appear likely to be primarily
species of Upland Mixed Ashwoods. Upland birch woods, which do not yet
have an official definition, were taken to be boreal woodland only, although
it is possible that acid hyper-oceanic temperate woods without oak will be
included, making the allocation of species near impossible. Pasture
woodlands with beech dominant are explicitly excluded from the Lowland
Beech and Yew Woodland BAP.
Upland Mixed Ashwoods stand out as a core habitat for rare bryophytes.
The habitat includes general temperate and sub-oceanic woodland species
as well as many hyperoceanic species. Most other plans have a reasonable
representation of rare species, with between 10 and 21%, but Wet Woodland
has only one species and Lowland Beech and Yew Woodland none.
BRYOPHYTES – PRIORITY BAP HABITATS
Distribution of species of conservation interest within BAP habitats
49
Priority BAP habitat
Upland Mixed Ashwoods
Pasture Woodland
Lowland Mixed Deciduous
Upland Oakwood
Native Pine
Upland Birch Woods
Wet Woodland
Lowland Beech and Yew Woodland
No of
species
12
6
5
5
4
3
1
0
Percentage of total no of species
of conservation interest
43%
21%
18%
18%
14%
11%
4%
As with bryophytes, partitioning oceanic lichens between Upland Mixed
Ashwoods and Upland Oakwood, or for that matter Wet Woodland, was not
easy. As with bryophytes, in Upland Birchwoods only boreal stands were
included; hyper-oceanic birchwoods were taken as being Upland Oakwood,
if without hazel or Upland Mixed Ashwoods, if they had hazel. In addition,
Atlantic hazel woods were also included in the Upland Mixed Ashwoods.
Pasture woodlands with beech dominant are explicitly excluded from the
Lowland Beech and Yew Woodland BAP.
Published habitat details confirmed the observations of Coppins & Coppins
(2006) that hazel and ash floras are considerably more significant than oak
floras in hyper-oceanic areas. Base-rich woods are richer than acidic woods
in rare lichens; Upland Mixed Ashwood leads by a long way over other BAP
habitats classified by their canopy type. Of these, Upland Oakwood and
Native Pinewood are also of great significance, followed by Lowland Mixed
Deciduous Woodland and Upland Birchwoods. Neither Wet Woodland nor
Lowland Beech and Yew Woodland are of much significance. However, if
beech-dominated pasture woodlands such as the New Forest were included
within the latter BAP, its importance would rise dramatically to 50 species
or 32% of woodland species of conservation interest.
The association of woodland lichens with traditionally grazed woodlands is
exceptionally high, with 92% of woodland species of conservation interest
found within pasture woodlands.
LICHENS – PRIORITY BAP HABITATS
Distribution of species of conservation interest within BAP habitats
Priority BAP habitat
Wood pastures
No of
species
143
50
Percentage of total no of species
of conservation interest
92%
Upland Mixed Ashwood
Upland Oakwood
Native Pinewood
Lowland Mixed Deciduous Woodland
Upland Birchwoods
Lowland Beech and Yew Woodland
Wet Woodland
78
44
44
24
21
1
1
50%
28%
28%
15%
14%
1%
1%
3.4.2 SAC Annex 1 habitats
Fully allocating woodland species of conservation interest within SAC Annex
1 habitats proved impossible in the timescale of this project. This was at
least partly due to the ill-defined and confused nature of the definitions of
some of the SAC woodland habitats in Britain. These tend to be quite short
definitions, which are anchored to a definitive phytosociological
classification. A major part of the problem is this combination of short
generalised definitions with precise, but European, phytosociological
vegetation classifications. The latter were not based on the communities
found in Britain but on related European vegetation. In contrast, the
descriptions of Scandinavian Annex 1 woodlands should be consulted as an
example of what should have been done, giving clear but widely
encompassing descriptions of ecosystems (European Commission, 2003).
The most striking problem is with the temperate rain forests. These are one
of Europe’s most important habitats but the category intended to include
them 91A0 Old sessile oak woods with Ilex and Blechnum in the British Isles
is bizarrely confined to acid sessile oak woods. This excludes the base-rich
woods with ash and hazel dominant, which are richer in rare species, or acid
woods with birch dominant. The reference to holly (Ilex) as a defining
feature is odd as this is hardly now a conspicuous component of many hyperoceanic woodlands in Britain.
In fact, the whole Annex 1 habitat seems to be based on the BlechnoQuercetum described from the generally strongly acidic Killarney oak – holly
woods by Braun-Blanquet & Tüxen (1952) (European Commission, 1991),
making it a woefully inadequate description of the range of variation of the
temperate rain forests of Britain. The practical result of this is that many
internationally important sites are not listed as SACs for this Annex 1 habitat
as they are not acid enough or lack oak (May, 2002).
This is particularly problematic in the west Highlands, as here woods
without oak dominance are richer in rare lichens and mosses than oakdominated woods (Coppins & Coppins, 2006). As a solution to the
misdescription of the temperate rain forests biome, JNCC has notified some
of the most important ravine woodlands as Annex 1 9180 * Tilio-Acerion
forests of slopes, screes and ravines. This is problematic, however, as
although southern ravines as in the Wye Valley do represent northern
outliers of Tilio-Acerion, with species such as large leaved lime present, the
ravine woods of western Scotland are better referred to the Alno-Ulmion
(Rodwell, 1991; Rodwell et al., 2000). This is a very different if equally
important woodland type. An example of the oddities this produces is the
51
Resipole ravine in the Sunart SAC, Ardnamurchan. This is one of the most
important epiphytic lichen sites in Europe.
The Annex 1 habitat into which such ravines have been shoehorned, the
9180 * Tilio-Acerion forests of slopes, screes and ravines, however, is only
listed as a qualifying feature for the SAC, but not as a primary reason for
site selection for the SAC. Other less acid temperate rain forests, without
large ravines and with lower oak dominance but of the highest importance,
such as Ellary or Loch Melfort, Argyll, are not SACs. Extreme, but vitally
important types of temperate rain forests, such as Atlantic hazel woods
have no European recognition at all, for example the outstanding Atlantic
hazel wood at Ballachuan, Seil, Argyll (Copins et al., 2002). Far better
would have been to describe the temperate rain forest properly in the first
place. Priority status for old growth, or little disturbed, native woodland
within the rainforest could also be justified.
Other issues include:
•
The narrow definition of the 91C0 * Caledonian forest as only boreal
pine woods is a sad contrast to the inclusiveness of the equivalent
Scandinavian Annex 1 habitat 9010 * Western Taïga. Boreal mixed birch
woods have therefore been excluded from the British boreal woodland
Annex 1 habitat, resulting in important birchwoods, with aspen clones
rich in rare lichens, being excluded from immediately adjacent
pinewood SACs in Speyside.
•
91A0 Old sessile oak woods with Ilex and Blechnum in the British Isles
have been extended by JNCC to include sub-oceanic woods in north
eastern Scotland. These woods have few Atlantic characteristics
(Rodwell & Dring, 2001), but include a distinctive lichen assemblage
similar to that of Scandinavian woods, including many rare species.
These woods are a better match for the Scandinavian Annex 1 habitat
9020 * Fennoscandian hemiboreal natural old broad-leaved deciduous
forests (Quercus, Tilia, Acer, Fraxinus or Ulmus) rich in epiphytes.
•
Pasture woodlands in Scandinavia that cannot be included with other
woodland types are placed in Annex 1 habitat 9070 Fennoscandian
wooded pastures. The equally important pasture woodlands and
parklands of Britain are not included in Annex 1. As a result, lowland
pasture woodlands without beech and parklands in general have no
European recognition. These include some of the most important
lichen-rich veteran tree sites in Europe, such as Melbury Park in Dorset
and Bocconoc Park in Cornwall. Also excluded are rich pasture sites,
such as Savernake Forest, which lack native beech.
•
Finally, the most remarkable absence is the Atlantic lowland mixed
deciduous woodland, as defined by the abundance of bluebell (EndymioCarpinetum type of vegetation). As Rodwell & Dring (2001) state, “it is
much to be regretted that a separate Endymion-Carpinetum could not
be included as an unambiguous category for the distinctive oak –
52
hornbeam woodlands with [bluebell] Hyacinthoides that are confined to
the north-west lowlands of the Atlantic zone.” They also add,
“although [hornbeam] Carpinus has a limited range in the UK, there
seems no reason to regard the distribution of this tree, rather than the
general extent of mixed broadleaf woodland with Hyacinthoides, as
setting the bounds of the Carpinion with us”. For vascular plants, these
are the richest woods for rare vascular plants in Britain and bluebell
Hyacinthoides non-scripta itself is an International Responsibility
species.
The above problems are not just pedantic. Misnotified SACs will be difficult
to defended from planning applications as habitats of European importance
if it turns out that the Annex 1 habitat concerned is not actually present.
3.4.3 Relationships to open habitats
The biodiversity losses caused by abandonment and loss of grazing on low
productivity land in the lowlands is well appreciated. Within pasture
woodlands, the removal of grazing not only causes the loss of grazingdependent woodland plants and epiphytes, but also losses from associated
open communities. This problem may be starting to be significant in the
uplands. The retreat of farming in general, and in particular the exclusion
of grazing from pasture woodlands or from moorlands intended to become
woods by conservationists, may be beginning to causing serious biodiversity
losses. Holland et al. (2008) gives a chilling example where a population of
up to 60 flowering spikes of the Vulnerable Red Data Book species small
white orchid Pseudorchis albida disappeared in five years from a heath from
which the grazing had been removed within the Tyndrum Community
Woodland, Perthshire. The author has seen several examples of base-rich
flushes within open areas in now fenced off Highland pasture woodlands,
which would be regarded as SSSI quality in their own right in the lowlands,
which are disappearing under rank grass. Such communities have probably
remained open and species-rich for millennia and may have been more
natural than the woodlands regenerated in their place (Fenton, 2008).
53
4.0
CONCLUSIONS
4.1
Vascular plants
Woodlands are clearly an important part of Britain’s biodiversity, and there
are particular areas of concern. These mainly relate to wood edge plants,
woodland grassland species and early succession plants.
Britain’s richest woodlands for vascular plants, those to the south, are on
average getting darker and suffering from a change in their grazing regime
from very little browsing or grazing to uncontrolled deer grazing. The
combination of increasing shade and unbalanced grazing is unlikely to
promote a diverse flora and is certainly inimical to the main groups of plants
under threat.
The answer lies presumably within woodland policy. There is a lot of
pressure not to clear fell or create large openings in woodland. The
concept of continuous cover forestry is presented as an environmentally
friendly version of forestry and there are landscape objections to large
felling blocks. In discussing continuous cover forestry, Mason et al. (1999)
define clear felling as the cutting down of all trees on an area of more than
0.25 ha. This would exclude traditional scale coppicing seen in hot spots for
rare plants such as the Wye Valley (Peterken, 2007a & 2007b). As an
example, the Woodland Trust is recommending continuous cover forestry
methods for restoring conifer plantations to native woodland on replanted
ancient woodland sites. This is applied even where these were originally
coppices or pasture woodlands (The Woodland Trust, 2005). In both
situations, the continuous cover approach is of no obvious benefit to the
rarer plants associated with the original woodland types, but it seems to be
assumed that it must be of benefit (Sanderson, 2006a).
Any presumption against openness in our woods is not likely to benefit
declining vascular plants. Our woodlands are not mini tropical rain forests;
being regularly opened up to an extreme degree has been part of our
woodland ecosystems for millennia. The biota appears pre-adapted to this
and it is remarkable how little of the flora is well adapted to closed canopy
forest - the starting point of the theories of Vera (2000).
Forestry policy needs to counteract the pressures, deliberate or
inadvertent, to reduce openness in woodlands. Specific issues are likely to
include encouragement for large scale fellings(as coppice or shelter wood
regeneration fellings to restore habitat for early succession species), deer
control, management of woodland edges to maintain more open edge
habitat, interaction with adjacent land use and the management and
retention of woodland grasslands.
There is also a serious international anomaly with the non-inclusion of
lowland mixed bluebell woods within the Habitats Directive. Any
opportunity to achieve international recognition for the importance of these
woods should be pursued.
54
In the north and west the diversity of the vascular plant flora is reduced and
here issues are dominated by the needs of bryophytes and lichens.
Important rare vascular plant populations do occur in woodlands but these
share issues with the bryophytes and lichens.
4.2
Bryophytes
Again, rare bryophytes are concentrated in open habitats, but the Red Data
Book for bryophytes (Church et al., 2001) was not compiled in the same way
as other RDBs, as only geographically restricted species were considered and
this may have resulted in the internationally important temperate rain
forest flora being under represented. This Atlantic woodland habitat is
easily the most important for bryophytes in Britain and Europe. This
internationally important habitat deserves a great deal of attention, which
can be combined with action associated with lichens. Issues include
international recognition of importance - the woods being part of a
European cultural landscape appears to have prejudiced international
recognition (Rhind, 2003) and this needs challenging. Habitats Directive
characterisation of hyper-oceanic woodland is woefully inadequate and fails
to cover the range of variation of temperate rain forest and urgently
requires re-examination. At a site level, much more characterising of the
actual range of positive management regimes benefiting lower plants is
required. To date there appears to have been too much disregard of
cultural landscapes and the positive role of grazing. This has resulted in too
great a readiness to resort to grazing exclusion and non-cultural models of
the habitat, which ignore millennia of human involvement. If Fenton (2008)
is correct, these non-cultural models may even be ignoring natural models
of the functioning of these ecosystems, in which high grazing levels may be
a natural function of the oceanic climate. Essentially overgrazing is not the
issue - instead it is unbalanced grazing. Too many sheep or deer may be a
problem but this is not solved by undergrazing the habitat.
Outside of hyper-oceanic woodland, the bryophyte ride flora appears to be a
significant feature, which is particularly under threat. This group of species
is suffering from a decrease in openness in woodlands, similar to the
problems faced by woodland grassland vascular plants.
Another group of species showing a serious decline are Orthotrichum
mosses, which are very sensitive to air pollution but are also pioneer species
and good colonists. The loss of elm has confounded the threat to some
species. These appear to be just starting a countrywide recovery, as would
be expected of pollution sensitive pioneer species, with improving air
quality. These species require clean air policies to be maintained for this
recovery to continue.
There are a few old growth dependent bryophyte species of veteran trees
and large fallen dead wood. These share issues with numerous lichen
species.
4.3
Lichens
55
This group is different, with woodland the single most important habitat for
rare and threatened lichens. Epiphytic lichens on field and wayside trees
are a significant issue that Plantlife has concentrated on, but there are far
more rare and declining lichen species within woods that outside them. In
particular, the oceanic woods, those with hyper-oceanic and southern
oceanic lichen assemblages, are of international importance and share
issues with the Atlantic bryophyte flora. As with the bryophytes, these
include the international recognition of the importance of the habitat. At a
European level, this would include a properly delimited Habitats Directive
Annex 1 habitat for temperate rain forest. At an international level, there
needs to be recognition that the European temperate rain forest biome is a
centre of plant and lichen diversity of international significance on a par
with other temperate rain forest biomes (Rhind, 2003).
At a site level, much more characterising of the actual range of positive
management regimes benefiting lower plants is required. To date there
appears to have been too much disregard of cultural landscapes and the
positive role of grazing. In the case of lichens this is even more critical than
for bryophytes, as rich epiphytic lichen assemblages are much more old
growth dependent and are found in more accessible and better-lit sites.
The maintenance of grazing is even more significant for lichens than for
bryophytes and the need for balanced grazing must be emphasised. The
problems relating to grazing are not best characterised as overgrazing but as
imbalanced grazing, with undergrazing an increasing problem.
The majority of rare and threatened woodland lichens are old growth
dependent species. This also applies to less oceanic woodlands, which are
less internationally significant but can still be of high importance, where old
growth survives in clean air areas. Old growth is still not specifically
protected in Britain and there are serious issues with the maintainance of
habitat quality within old growth woodlands. This relates to the heart of
the Vera debate - rich epiphytic lichen floras are associated with old growth
woodland suffering canopy break up without immediate regeneration. This
habitat requires a significant grazing impact to emerge. Existing action for
some important woodland types, such as boreal native pinewoods, does not
appear to be fully taking into account the conditions required for the
survival of the majority of the rare species found within them. The role of
grazing within old growth woodlands is confused within the BAP process, and
some habitats, which include numerous traditionally grazed woodlands with
a high biodiversity value, only mention grazing in the plan as a problem.
The Wood Pasture and Parkland BAP is potentially making things worse in
some areas. It can gives the impression that grazing is only an issue with
park-like habitats, while the greatest lichen diversity, including numerous
BAP lichen species, is associated with grazed woodlands, not a savanna-like
habitats. There is a need to accept that pasture woodlands are those woods
that require grazing to maintain their biodiversity importance, irrespective
of whether this fits in with conservationists’ preconceptions or forestry
policy.
56
There is a need for the conservation of old growth dependent epiphytic
floras to be pushed up the agenda to ensure they are not being
compromised by well meaning efforts to conserve woodlands, as well as by
over exploitation. Swedish forestry practice may be a model to follow
(Karlson et al., 1995).
4.4
Issues
The following are issues which this report suggests require research and
debate:
Fragmentation and new woodland
•
Fragmentation of open habitats has been far more severe in the
lowlands than in woodland. In addition, open habitats were often very
extensive in the early modern period but woodland has largely been
highly fragmented for millennia. Perhapsrestoration of open habitats
from farmland should greatly exceed new woodland creation on such
habitats? This is not to argue against all new woodland, just that the
priorities should be sorting out management and condition within
existing ancient woodlands and expanding grazed open habitats. With
increased demands for food production, there is also a need to prioritise
the most effective habitat restorations. Open habitats can be restored
to high quality habitats much faster than woodland habitats. Open
habitat restoration, however, should have room for new pasture
woodlands, for example being created out of older recent woodlands
within heathland restoration schemes.
•
In the uplands, open semi-natural habitat is still very extensive and
there is certainly room for woodland expansion. There appears to have
been large scale fragmentation of some woodlands, especially mixed
boreal birch woodland on better soils, since the early modern period.
Again, this should not detract from sorting out management and
condition within existing ancient woodlands. On the other hand, much
native woodland restoration appears to be in the form of plantations,
often blanket planting at 2000 trees per hectare, with cultivation
(mounding). The latter has allowed planting into wet heath and flushes
which were probably naturally open habitats (Fenton, 2008). The
contrast with real native woodlands, nearly all of which were pasture
woodlands, as opposed to 19th oak plantations, is striking. Native
woodlands are highly patterned, typically with a glade and grove
structure, with open areas mainly on wet ground and the richest
grazing. This patterning is fundamental to biodiversity with these
woods, with groups as diverse as woodland epiphytes, open ground
vascular plants and rare glade and woodland edge butterflies depending
on it. It is also probably a natural feature, not an anthropogenic
artefact. Serious native woodland restoration in the uplands could be
carried out by planting initial groves, leaving large open areas for future
expansion and by allowing grazing for extended periods. In reality,
planting appears to be driven by grant aid mechanisms designed for
commercial softwood plantations, not native woodlands, and with little
57
thought given to grazing and grazing dependent rare species. A further
issue is the use of the NVC, rather than the actual existing local
woodland composition, as a guide to planting. This appears to be
ironing out local distinctness and replacing it with an idealised image of
what woods are expected to be, rather than what they were.
Woodland management
•
Within lowland ancient woodland, especially on base-rich soils, there is
clearly a decline in rare specialist species, matched by a decline in the
more widespread common species. This appears to be related to largely
increasing shade in woodlands. There is a need for research into
effective methods of management that maintain diversity within seminatural woodlands. These will probably include forestry systems that
open up woods much more than currently tends to occur, such as large
scale coppicing and more extensive shelterwood fellings in high forest.
All of these will have increasing deer numbers as a problem. For rare
species, there are likely to be those that have very special requirements
relating to lost management systems, such as compartmentalised grazed
coppice. Edge species and woodland grassland species are especially
likely to be associated with lost arts such as controlled stock grazing
within intensively managed woods, as opposed to extensive pasture
woodland management.
•
How should woodlands within intensively farmed landscapes be buffered
against adjacent intensive land use? Will planting or regenerating
recent woodlands around existing ancient woodlands help or mitigate
against woodland edge specialists? If these require low productivity
species-rich habitats, would nutrient stripping and managed grassland
on woodland margins be more effective for declining plants and
invertebrates?
•
Woodlands within the uplands have been part of extensive pastoral
farming systems for millennia. The majority of woodland plant and
animal species of conservation interest are grazing dependent and have
survived in grazed woods since trees returned after the ice age. Our
oceanic woods are likely to have been naturally and strongly influenced
by grazing before man’s arrival. Ungrazed intensively managed
woodlands are a recent early modern innovation in many parts of the
uplands. Unbalanced and inappropriate grazing has been a major
problem in the uplands but not just for woodlands; moorlands have also
been widely mismanaged. The answer is surely not to fence woods out
of extensive grazing systems, but instead to restore balanced grazing to
the whole landscape. The flora and fauna of the uplands appears to
have thrived until the early modern period within low input farming
landscapes. To quote Prof. Chris Smout, St. Andrews (Bullock, 2005):
“Why are we fixated on the Atlantic period? Our knowledge of
biodiversity depends on the work of 19th and 20th century scientists and
we only tend to value what we had left in the 20th century. Wouldn’t it
be more sensible to use the last 400 years as the model?”
58
•
Little-disturbed grazed old growth woodlands are exceptionally rich in
rare species of epiphytic lichens, especially in hyper-oceanic climates.
In the these climates, rich bryophyte floras also occur and the lower
plant assemblage is of international importance. Such woods are an
international responsibility and there should be detailed research into
how to conserve these woods and their associated floras. The
observational evidence of specialists does not appear to find that many
grant aided conservation management activities are beneficial and
some, such as total grazing exclusion, are positively damaging.
Written by Neal Sanderson, 2008. The views in this report are those
of the author and not necessarily Plantlife
59
5.0
REFERENCES
Alexander, K.N.A., Smith, M., Stiven & Sanderson, N. A. (2002) English
Nature research reports No 494. Defining ‘Old Growth’ in the UK
Context. Peterborough: English Nature.
Averis, A. B. G (1991) A Survey of the Bryophytes of 448 Woods in the
Scottish Highlands. Scottish Field Unit Report. Edinburgh: Nature
Conservancy Council.
Bakker, E. S., Olff. H., Vanderberghe, C., Maeyer, K. De, Smit, R.,
Gleichman, J, M, & Vera, F. W. M (2004) Ecological anachronisms in
the recruitment of temperate light-demanding tree species in
wooded pastures. Journal of Applied Ecology. 41: 571–582.
Biodiversity Reporting and Information Group (2007) Report on the Species
and Habitat Review, Report to the UK Biodiversity Partnership.
Peterborough: JNCC.
Blockeel, T. L. (2008) Orthotrichum consimile in Derbyshire. Field
Bryology 94:
23-26.
Bossuyt, B., Hermy M. & Deckers J. (1999) Migration of herbaceous plant
species across ancient – recent forest ecotones in central Belgium.
Journal of Ecology 87: 628-638.
Braun-Blanquet, J. & Tüxen, R (1952) Irische Pflanbzengesellschaften.
Veröffentlichungen des Geobotanischen Institutes Rübel in Zürich.
25: 224-415.
Brunet, J. & Von Oheimb, G. (1998) Migration of vascular plants to
secondary woodlands in southern Sweden. Journal of Ecology, 86,
429–438.
Bullock, D. J. (2005) Large herbivores in upland Britain: what can the past
tell us about the future? In: The Role of Large Herbivores in Shaping
the Upland Landscape of Britain. What does the science of herbivore
ecology tell us? (ed. J. Fenton) 9-16. Edinburgh: National Trust for
Scotland.
Bunce, R. G. H. (1982) A Field Key for Classifying British Woodland
Vegetation, Part 1. Institute of Terrestrial Ecology, Cambridge.
Chatters, C. & Sanderson, N. A. (1994) Grazing Lowland Pasture Woodlands.
British Wildlife 6: 78-88.
Cheffings, C. M. & Farrell, L. (2005) Species Status No. 7 The Vascular Plant
Red Data List for Great Britain. Peterborough: JNCC.
60
Church, J. M., Hodgetts, N. G., Preston, C. D., & Stewart, N. F. (2001) Red
Data Books of Britain and Ireland Mosses and Liverworts.
Peterborough: JNCC.
Colebourn, P. (1983) Hampshire’s Countryside Heritage 2: Ancient
Woodland. Winchester: Hampshire County Council.
Coppins, A. M. & Coppins, B. J. (1998) Lichen Survey of Horner Woods NNR –
1998. Unpublished Report to the National Trust.
Coppins A. M. & Coppins, B. J. (2002) Indices of Ecological Continuity for
Woodland Epiphytic Lichen Habitats in the British Isles. London:
British Lichen Society.
Coppins A. M. & Coppins, B. J. (2005) Lichens – the Biodiversity Value of
Western Woodlands. Botanical Journal of Scotland. 57: 141-153.
Coppins A. M. & Coppins, B. J. (2006) The lichens of the Scottish native
pinewoods. Forestry. 79: 249- 259.
Coppins A. M., Coppins, B. J. & Quelsh P. (2002) Atlantic hazelwoods. Some
observations on the ecology of this neglected habitat from a
lichenological perspective. British Wildlife 14: 17-26.
Davies, A. (2003a) Torran Beithe: Holocene history of a blanket peat
landscape. In: The Quaternary of Glen Affric and Kintail, Field
Guide. (ed. Tipping, R. M.) 41-48.
Davies, A. (2003b) Crarnach Mór and Caamban: woodland history and landuse in alluvial settings. In: The Quaternary of Glen Affric and
Kintail, Field Guide. (ed. Tipping, R. M.) 97-84.
Davies, L., Bates J. W., Bell, J. N. B., James, P. W. & Purvis O. W. (2007)
Diversity and sensitivity of epiphytes to oxides of nitrogen in London.
Environmental Pollution. 146: 299-310.
Day, S.P. (1993) Woodland origin and 'ancient woodland indicators': a casestudy from Sidlings Copse, Oxfordshire, UK. The Holocene 3: 45-53.
Dalrymple, S. (2006) Unravelling the causes of small cow-wheat’s rarity.
Looking to the Hills, Newsletter of the Uplands Lead Co-ordination
Network. 12: 17-19.
Definitions Sub-group (2001) Annex 1 A definition of wood pasture and
pasture woodland. In: Wood-pasture and parkland habitat action
plan: Progress report. No. 459 – English Nature Research Reports.
(ed: R. Watson) 38-40. English Nature, Peterborough.
61
Dennis, R. (1998) The Importance of Traditional Cattle for Woodland
Biodiversity in the Scottish Highlands. A Personal View. R. Dennis,
Nethybridge.
Edwards, B. (2005) Elm Lichens. Unpublished Draft. Salisbury: Plantlife
International.
Edwards, M. E. (1986) Disturbance histories of four Snowdonia woods and
their relation to Atlantic bryophyte distribution. Biological
Conservation. 37: 301-320.
Ellis C. J. & Coppins, B. J. (2007) 19th century woodland structure controls
stand-scale epiphyte diversity in present-day Scotland. Diversity and
Distributions, A Journal of Conservation Biogeography 13: 84–91.
European Commission (1991) Corine Biotopes Manual Habitats of the
European Community. Data Specifications Part 2. European
Commission.
European Commission (2003) Interpretation Manual of European Union
Habitats. Version EUR25. European Commission.
Farmer, A. M., Bates, J. B., & Bell, N. J. (1992) Ecophysiological effects of
acid rain on bryophytes and lichens. In Bryophytes and Lichens in a
Changing Environment. (eds: J W Bates & A M Farmer) 284-313.
Oxford University Press, Oxford.
Fenton, J. H. C. (1998) Unnatural environment – have campaigners for the
return of Scotland's forests got it wrong? New Scientist, 2126: 22.
Fenton, J. H. C. (2004) Wild thoughts ….a new paradigm for the uplands.
ECOS, 25: 2-5.
Fenton, J. H. C. (2008) A postulated natural origin for the open landscape of
upland Scotland. Plant Ecology & Diversity. 1: 115-127.
Fleming, A. (1997) Towards a History of Wood Pasture in Swaledale (North
Yorkshire). Landscape History 19: 57-73.
Flower, N. & Tubbs, C. R. (1982) Management Proposals for the Unenclosed
Woodlands and Woodlands of Special Importance in the Statutory
Enclosures. NCC, Lyndhurst.
Forestry Commission (2005) Keepers of time: A statement of policy for
England’s ancient and native woodland. Cambridge: Forestry
Commission, England.
Forestry Commission (2008) Crown Lands - Management Plan 2008-2013.
www.forestry.gov.uk/forestry/INFD-7A3F82>, Forestry Commission.
62
French, C., Murphy, R. & Atkinson, M. (1999) Flora of Cornwall. Camborne:
Wheal Seton Press.
Friday, (1990) A Comparative Survey of the Lichen Flora of Four Betula
Woods in North East Scotland. A report to Scottish Natural Heritage.
Gilbert, O. L. (1984) Some effects of disturbance on the lichen flora of
oceanic hazel woodland. Lichenologist 16: 21-30.
Gilbert, O. L. (2000) The New Naturalist Lichens. HarperCollins, London.
Gustafsson, L., Friskesjo, A., Ingelog ,T., Pettersson, B. & Thor, G. (1992)
Factors of importance to some lichen species of deciduous
broadleaved woods in southern Sweden - Lichenologist 24 : 255-266.
Harding, P. T. & Rose, F. (1986) Pasture - Woodlands in Lowland Britain.
ITE, Huntingdon.
Hawksworth, D. L. & Rose, F. (1970) Quantitative scale for estimating
sulphur dioxide air pollution in England and Wales using epiphytic
lichens. Nature, Lond. 227: 145-148.
Harvey, G. (2002) The Forgiveness of Nature. The Storey of Grass. London:
Vintage.
Hill, M. O. & Edwards, B. (2003) Mosses and Liverworts of Dorset.
Dorchester: Dorset Environmental Records Centre.
Hill, M. O., Preston, C. D. & Smith, A. J. E. (1991) Atlas of the Bryophytes
of Britain and Ireland Volume 1 Liverworts. Harley Books.
Hill, M. O., Preston, C. D. & Smith, A. J. E. (1992) Atlas of the Bryophytes
of Britain and Ireland Volume 2 Mosses. Harley Books.
Hill, M. O., Preston, C. D. & Smith, A. J. E. (1994) Atlas of the Bryophytes
of Britain and Ireland Volume 3 Mosses. Harley Books.
Hodgetts, N. G. (1993) Atlantic bryophytes on the western seaboard. British
Wildlife. 4: 287-295.
Hodgetts, N. G. (1997) Atlantic Bryophytes in Scotland. Botanical Journal of
Scotland. 49: 375-385.
Hodder, K H, Bullock, J M, Buckland, P C, & Kirby K J (2005) English Nature
Research Reports 648. Large Herbivores in the Wildwood and in
Modern Naturalistic Grazing Systems. Peterborough, English Nature.
Holland, J., Pollock, M. & Waterhouse, T. (2008) The consequences for
biodiversity. In: Farming’s Retreat from the Hills (eds. A. Renwick &
T. Waterhouse) 32-35.
63
Huntley, B., (1981) The past and present vegetation of the Caenlochan
National Nature Reserve, Scotland II Palaeoecological Investigations.
New Phytologist 87: 189-222.
Karlson, J., Norén, M. & Wester, J. (1995) Key Habitats in Woodland.
Jönköping, Sweden: National Board of Forestry.
Kirby K. J., Smart, S.M., Black, H.I.J., Bunce, R.G.H., Corney, P.M. and
Smithers R.J. (2005) Long term ecological change in British woodland
(1971-2001). English Nature Research Reports Number 653.
Peterborough: English Nature.
Kreuz, A. (2007) Closed forest or open woodland as natural vegetation in the
surroundings of Linearbandkeramik settlements? Veget. Hist.
Archaeobot. 17: 51–64
Legg, C. (2004) Extinct Aspen Bristle Moss rediscovered. Looking to the
Hills, Newsletter of the Uplands Lead Co-ordination Network. 12:
24.
Mason, B., Kerr, G. & Simpson, J. (1999) What is Continuous Cover Forestry?
Information Note. Edinburgh: Forestry Commission.
MacDonald, A. (2005) Further development of the treeline model. Looking
to the Hills, Newsletter of the Uplands Lead Co-ordination Network
13: 15-16.
Marren, P. R. (1988) the past and present distribution of Stachys germanica
L. in Britian. Watsonia 17: 69-68.
Maxwell, J. S. (1929) Loch Ossian Plantations. An Essay in Afforesting High
Moorland. Carrour: Maxwell.
May, R. (2002) Special Areas of Conservation (SACs) Analysis of the UK List
and Additional Sites Needed in Preparation for the Atlantic
Biogeographic Seminar 2002. London: WWF-UK.
McVean, D. N. (1956a) Ecology of Alnus glutinosa (L.) Gaertn. V notes on
some British Alder populations. J. Ecol. 44: 321 – 330.
McVean, D. N. (1956b) Ecology of Alnus glutinosa (L.) Gaertn. VI post glacial
history. J. Ecol. 44: 331 – 333.
Mitchell, F. J. G., (2005) How open were European primeval forests?
Hypothesis testing using palaeoecological data. Journal of Ecology.
93: 168–177.
ODPM (2006) Planning for Biodiversity and Geological Conservation – A
Guide to Good Practice. Office of the Deputy Prime Minister, London.
64
O’Sullivan, P. E. (1975) Vegetation history and the native pinewoods. In:
Native Pinewoods of Scotland, Proceedings of Aviemore Symposium,
1975 (ed: R. G. H. Bunce & J. N. R. Jeffers) 60-69.
Nitare, J. (2000) Signalarter. Jönköping: Skogsstyrelsen. Jönköping:
Skogsstyrelens Förlag.
Noble, R. (1997) Changes in native woodland in Assynt, Sutherland, since
1774. In: Scottish Woodland History (ed: T. C. Smout) 126 - 134.
Preston C. D., Pearman, D. A. & Dines T. D. (2002) New Atlas of the British
and Irish Flora. Oxford: Oxford University Press.
Peterken, G. F. (1974) A method for assessing woodland flora for
conservation using indicator species. Biol Con. 6: 239-245
Peterken, G. F. (1993) Woodland Conservation and Management, Second
Edition. Chapman and Hall.
Peterken, G. F. (1996) Natural Woodland Ecology and Conservation in
Northern Temperate Regions. Cambridge: Cambridge University
Press.
Peterken, G. F. (2007a) Wye Valley. London: Collins.
Peterken, G. F. (2007b) Woodlands of the Lower Wye Origins, History and
Management. Monmouth: Wye Valley AONB Unit.
Preston C. D., Pearman, D. A. & Dines T. D. (2002) New Atlas of the British
and Irish Flora. Oxford: Oxford University Press.
Quelch, P. R. (2001) An Illustrated Guide to Wood Pasture in Scotland.
Lochgilphead: Forestry Commission.
Rackham, O. (1980) Ancient Woodland, its History, Vegetation and Use in
England. London: E. Arnold.
Rackham, O. (1986) The History of the Countryside. London: J. M. Dent &
Sons Ltd.
Rackham, O. (1989) The Last Forest The Story of Hatfield Forest. London:
J. M. Dent & Sons Ltd.
Rackham, O. (1990) Trees and Woodland in the British Landscape, revised
Edition. London: J. M. Dent & Sons Ltd.
Rackham, O. (2003) Ancient Woodland, its History, Vegetation and Use in
England. New Edition. Dalbeattie: Castlepoint Press.
65
Rackham, O. (2006) The New Naturalist Library Woodlands. London:
Collins.
Ratcliffe, D. A. (1986) An ecological account of the Atlantic bryophytes in
the British Isles. New Phytologist 67: 365-439.
Reeves, R. P. (2006) Use and Abuse of a Forest Resource: New Forest
Documents 1632-1700. Lyndhurst: New Forest Ninth Centenary
Trust.
Rhind, P. (2003) Comment: global conservation and Britain’s coastal
temperate rainforest. British Wildlife 15: 97-102.
Roberts, P. (2002) Minstead Life In a 17th Century New Forest Community.
Southampton: Nova Foresta Publishing.
Rodwell, J. S. (1991) British Plant Communities. Volume 1, Woodlands and
Scrub. Cambridge: Cambridge University Press.
Rodwell, J. S. & Dring, J. C. (2001) European significance of British
woodland types. Peterborough: English Nature.
Rodwell, J. S., Dring, J. C., Averis, A. B. G., Proctor, M. C. F., Malloch, A.
J. C., Schaminée, J. N. J., & Dargie, T. C. D (2000) Review of
coverage of the National Vegetation Classification, JNCC Report, No.
302. Peterborough: JNCC.
Rose, F., Stern, R. C., Matcham H. W. & Coppins, B. J. (1991) Atlas of
Sussex Mosses Liverworts and Lichens. Brighton: Brighton Borough
Council.
Rose, F. (1988) Phytogeographical and ecological aspects of Lobarion
communities in Europe. Bot. J. Lin. Soc. 69: 69-79.
Rose, F. (1992) Temperate forest management: its effects on bryophytes
and lichen floras and habitats. In: Bryophytes and Lichens in a
Changing Environment. (eds: J W Bates & A M Farmer) 211-233.
Oxford: Oxford University Press.
Rose, F. (1993) Ancient British Woodlands and Their Epiphytes. British
Wildlife 5: 83-93.
Rose, F. (1999) Indicators of Ancient Woodland. British Wildlife 10: 241251.
Rothero, G. P. (2005) Oceanic Bryophytes in Atlantic Oakwoods. Bot. J.
Scot. 57: 135-140.
66
Rothero, G. P. (2006) Bryophytes. In: In: The Nature of the Cairngorms
Diversity in a Changing Environment. (eds: P. Shaw & D. Thompson)
195-214. Edinburgh: Scottish Natural Heritage.
Sanderson, N. A. (1991) Notes on holly cutting in the New Forest. In:
Pollard and Veteran Tree Management (ed. H. J. Read) 53-55.
Corporation of London.
Sanderson, N. A. (1996) Lichen Conservation within the New Forest Timber
Inclosures. Hampshire Wildlife Trust.
Sanderson, N. A. (1997) A Review of Holly Cutting in the New Forest.
Hampshire Wildlife Trust.
Sanderson, N. A. (1998a) New Forest Epiphytic Lichen Data Base Volume 4.
Part 3 Summary of Results. Hampshire Wildlife Trust.
Sanderson, N. A. (1998b) Glen Finglass Historic Landscape Survey, Final
Report 1998. A Botanical Survey & Assessment report to The
Woodland Trust.
Sanderson, N. A. (1999) New Forest Rare Lichen Monitoring Project.
(Catillaria laureri, Parmelia minarum and Enterographa elaborata). A
report to Hampshire Wildlife Trust.
Sanderson, N. A. (2001a) Lichen Survey and Initial Historical Landscape
Assessment of King’s Forest Gleltsdale, Cumbria. A Botanical Survey
& Assessment report to East Cumbria Countryside Project & English
Nature.
Sanderson, N. A. (2001b) Epiphytic Lichen Monitoring in the New Forest
2000. LIFE Job L33A2U. A report by Botanical Survey & Assessment
to Forest Enterprise.
Sanderson, N. A. (2004) Fencing Proposals Under LIFE 3, Ecological Report
Contract No LIFE02/NAT/UK/8544. A report to the New Forest LIFE 3
Project.
Sanderson, N. A. (2005a) Site Condition Monitoring for Lichens in Scotland,
Pollochro Woods SSSI. 1. Establishing Baseline Site Condition
Monitoring. A British Lichen Society report to Scottish Natural
Heritage.
Sanderson, N. A. (2005b) Vegetation Survey of Cadnam Bog, New Forest,
2005. A Botanical Survey & Assessment report to the National Trust.
Sanderson, N. A. (2006a) Pyrenula hibernica at Ceunant Llennyrch,
Meirionnydd. British Lichen Society Bulletin. 98: 23-27.
67
Sanderson, N. A. (2006b) Lichen Survey and Monitoring Ceunant Llennyrch,
Gwyndd. A Botanical Survey & Assessment report to CCW.
Sanderson, N. A. (2007a) Enterographa elaborata: Canopy Collapse and
Lichen Diversity in New Forest Beech Woods. British Lichen Society
Bulletin. 100: 27–30.
Sanderson, N. A. (2007b) Preliminary Lichen Survey of Savernake Forest,
2007. A report by Botanical Survey & Assessment to English Nature.
Sanderson, N. A. (2007c). New Forest Inclosure Habitats: Habitat
Fragmentation & Landscape History. Botley: Hampshire Wildlife
Trust.
Sanderson, N. A. (2007d) Hampshire Wetlands Habitats Project 2006, Survey
& Assessment. Curdridge: Hampshire Wildlife Trust.
Forest
Sanderson, N. A. (2010) Chapter 9 Lichens. In: Biodiversity in the New
(ed. A. C. Newton) 84-111. Newbury, Berkshire; Pisces Publications.
Sanderson, N. A. & Wolseley, P. (2001). Management of pasture woodlands
for lichens. In: Habitat Management for Lichens. (ed. A. Fletcher)
05-1 – 05-25. London: British Lichen Society.
Selva, S. B. (1994) Lichen diversity and stand continuity in the northern
hardwoods and spruce-fir forests of northern New England and
Western New Brunswick. The Bryologist 97: 424-429.
Shaw & Tipping, (2003) Allt an Laghair, East Glen Affric: recent woodland
history from pollen analysis. In: The Quaternary of Glen Affric and
Kintail, Field Guide. (ed. Tipping, R. M.) 105-110.
Skov, F. & Svenning, J.-C. (2004) Potential impact of climatic change on the
distribution of forest herbs in Europe. Ecography 27: 366-380.
Smout, C. T., MacDonald, A. R. & Watson, F. (2005) A History of the Native
Woodlands of Scotland 1500 – 1920. Edinburgh: Edinburgh University
Press.
Spencer, J W & Kirby, K. J. (1992) An inventory of ancient woodland for
England and Wales. Biological Conservation 62: 77-93.
Stagg, D. J. (1989) Silvicultural inclosure in the New Forest to 1780. Proc.
Hampsh. Field Club Archaeol. Soc. 45: 135-145.
Steven, H. M. & Carlisle, A. (1959) The Native Pinewoods of Scotland.
Oliver & Boyd.
68
Stiven, R. & Hole, K. (2004) Wood Pasture. Perth: Scottish Natural
Heritage.
Sumsion, L. & Pollock, M. (2005) Woodland Grazing Toolkit. Argyll: Argyll
and Bute Local Biodiversity Partnership.
Svenning, J.-C. & Skov, F. (2007) Could the tree diversity pattern in Europe
be generated by postglacial dispersal limitation? Ecology Letters 10:
453-460.
Svenning, J.-C. (2002) A review of natural vegetation openness in northwestern
Europe. Biological Conservation, 104: 133-148.
Swift, A. L. & Howorth, R. (2006) Long-Term Structural Changes, and Their
Causes, in The Mens, a Minimum Intervention Mixed Deciduous
Ancient Woodland in West Sussex, UK. Henfield: Sussex Wildlife
Trust.
The Woodland Trust (2005) The conservation and restoration of plantations
on ancient woodland sites. Grantham: The Woodland Trust.
Thomason, D (1995) Case Study. Grazing in western sessile oakwoods in the
Lake District. Biological Journal of the Linnean Society: 56: 49-51.
Tittensor, R. M. (1970) History of the Loch Lomond Oakwoods. Scottish
Forestry 24: 100-118.
Tubbs, C. R. (2001) The New Forest. New Forest Centenary Trust,
Lyndhurst.
Turner, J. (2003). Veilwort Pallavicinia lyellii 2002 Report Number 218.
London: Plantlife.
Turner, J. (2004). Species Dossier for Pallavicinia lyellii. London:
Plantlife.
van Herk, C. M. (1999) Mapping of ammonia pollution with epiphytic lichens
in the Netherlands. Lichenologist 31: 9-20.
Vera, F. W. M. (2000) Grazing Ecology and Forest History. Wallingford:
CABI Publishing.
Verheyen, K. & Hermy, M. (2001) The relative importance of dispersal
limitation of vascular plants in secondary forest succession in Muizen
Forest, Belgium. Journal of Ecology 89: 829–840.
Verheyen, K., Bossuyt, B., Honnay, O., Hermy, M. (2003) Herbaceous plant
community structure of ancient and recent forests in two contrasting
forest types. Basic Appl. Ecol. 4: 537–546.
69
Watson, F. (1997) Rights and responsibilities: wood-management as seen
through baron court records. In: Scottish Woodland History (ed: T.
C. Smout) 101 - 114.
Wesche, S. (2003) English Nature Research Reports Number 528. The
implications of climate change for the conservation of beech
woodlands and associated flora in the UK. Peterborough: English
Nature.
Winchester, A. J. L. (2000) The Harvest of the Hills, Rural Life in Northern
England and the Scottish Borders, 1400–1700. Edinburgh: Edinburgh
University Press.
Wolesey, P. A., James, P. A., Theobald, M. R. & Sutton, M. A. (2006)
Detecting Changes in epiphytic lichen communities at sites effected
by atmospheric ammonia from agricultural sources. The
Lichenologist. 38: 161-176.
Wolf & Tipping, (2003) Recent woodland history in the pinewoods of east
Glen Affric. In: The Quaternary of Glen Affric and Kintail, Field
Guide. (ed. Tipping, R. M.) 97-104.
Wolseley, P. A. & O’Dare, A. M. (1989) Exmoor Woodland Lichens Survey
1987-1988. Somerset Trust for Nature Conservation.
Woods, R.G. & Coppins, B.J. (2003) A conservation evaluation of British
lichens. London: British Lichen Society.
Worrell, R. & Malcolm, D.C. (1998) Anomalies in the Distribution of Silver
Birch (Betula pendula Roth) Populations in Scotland. Botanical
Journal of Scotland. 50: 1-10.
Wright, R. N. & Westerhoff, D. V. (2001) New Forest SAC Management Plan.
Lyndhurst: English Nature.
WWF (2003) State of Europe’s Forest Protection. Vienna: WWF.
70