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A review of potential biodiversity
indicators for application in British
forests
R. FERRIS1 AND J.W. HUMPHREY2
1
2
Forestry Commission Research Agency, Alice Holt Lodge, Wrecclesham, Farnham, Surrey, GU10 4LH,
England
Forestry Commission Research Agency, Northern Research Station, Roslin, Midlothian, EH25 9SY, Scotland
Summary
A review of potential indicators of biodiversity in British forests is presented, with a focus on the
usefulness of selected biotic parameters as surrogate measures of different aspects of biodiversity in
managed forests. To be effective in this respect, indicators must satisfy a number of criteria. They
must be readily quantifiable, easily assessed in the field, repeatable and subject to minimal observer
bias, cost effective, and ecologically meaningful (i.e. close association with, and identification of, the
conditions and responses of other species). It is suggested that a combination of structural
(physiognomy of stands and associated structures) and compositional indicators (indicator species or
species groups) is selected which is appropriate to the aims of management and to the particular
forest type in question. A useful approach is to identify two to three key compositional indicators,
shown to be functionally linked to a broad range of other species, such as the extent and species
composition of the broadleaved component in conifer forests; and two to three key structural
indicators, which act as surrogates for general species richness or diversity, such as the quantity and
quality of deadwood.
Introduction
Since the United Nations Conference on
Environment and Development (Rio Earth
Summit, 1992) biodiversity and sustainable
management have become key issues in forest
policy and management. In Europe, a number of
Action Plans for the sustainable management of
forests have been produced, e.g. Sustainable
Forestry: The UK Programme (Anon, 1994) and
The UK Forestry Standard (Forestry Commission, 1998); Action Plan for Biological Diversity
and Sustainable Forestry (The National Board of
© Institute of Chartered Foresters, 1999
Forestry, Sweden, 1996); Criteria and Indicators
for Sustainable Forest Management in Finland
(Ministry of Agriculture and Forestry, Finland,
1997). Within these, there is almost invariably a
component that is concerned with evaluation of
methodologies for rapid and effective biodiversity assessment. As researchers and forest managers become steadily more aware of the
complexity of this task, the need to identify biodiversity ‘indicators’ has become a research priority in recent years (Noss, 1990; Ratcliffe,
1993).
An indicator may be defined as a characteristic
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which, when measured repeatedly, demonstrates
ecological trends, and a measure of the current
state or quality of an area. Simberloff (1998)
distinguishes between two main uses of indicators:
1 where the presence and fluctuations of the indicator are believed to reflect those of other
species in the community;
2 where the presence and fluctuations of the indicator are believed to reflect chemical/physical
changes in the environment.
This review concentrates on the first of these, i.e.
biotic indicators which can be used as surrogate
measures of other components of biodiversity,
and which may provide a short-cut in survey or
monitoring programmes. Implicit within this, is
the need for measurement to avoid complexity
unacceptable to forest managers, who will have
to use the indicators to detect the effects of
management practices on biodiversity (Boyle and
Sayer, 1995). Several attempts have been made to
define criteria for the selection of appropriate
indicators for biodiversity in boreal and temperate forests, within the context of sustainable
management. Key initiatives include the ‘Montreal Process’ (Canadian Forest Service, 1995)
incorporating mainly non-European countries,
and the ‘Helsinki Process’ within Europe (Ministry of Agriculture and Forestry Finland, 1994).
Most of the biodiversity indicators which have
been proposed through these initiatives are
designed to be used as monitoring tools for gross
changes in the forest area, or for assessing loss of
ecosystem functions (Table 1). There is also a
need for indicators that can be used by forest
managers to assess biodiversity at the stand- (here
used to describe a sub-division of a forest), forestor landscape-scale.
The objective of the review is to suggest possible groups of indicators for application in British
forests which can be used to assess the effects of
changing management practices on biodiversity.
The focus is primarily at the individual forest and
stand scales rather than taking a regional perspective. Forest and woodland types under consideration include semi-natural woodlands, as
classified using The Forestry Authority Forestry
Practice Guides, 1–8 (The Forestry Authority,
1994), plus plantation forests, both broadleaved
and coniferous.
Defining biodiversity in forest ecosystems
Three key components of biodiversity can be
recognized, and provide a framework for this
review (Schulze and Mooney, 1994):
• composition,
• structure (e.g. physiognomy of forest stands
and associated habitats), and
• function (processes, e.g. nutrient cycling).
However, the main focus is on indicators relating to structure and composition as these are
more amenable to measurement by forest managers. Structural or compositional elements may
also be surrogate functional indicators, e.g. deadwood (a structural indicator) may be a good indicator of decomposition processes.
It is imperative that indicators are linked to the
types and levels of biodiversity that management
aims to generate or maintain, within the particular forest or woodland type under consideration
(Newton and Humphrey, 1997). For example,
broad objectives for improving biodiversity in
plantation forests might involve creating or maintaining rare habitats or species. Also, management might aim to increase ‘naturalness’ in both
plantations and semi-natural woodlands, or try
and meet social and cultural expectations. By
implication, indicators must be tied in to management objectives, and need to meet a number of
criteria:
• they need to be easy to assess, even for nonspecialists;
• they must be repeatable (often using different
observers) and subject to minimal observer
bias;
• they must be cost-effective, generating reliable
data for acceptable costs;
• they must be ecologically meaningful, providing data which are easy to interpret.
Compositional diversity
Compositional diversity is commonly measured
by counting the number of plant or animal species
present in a given area (Groombridge, 1992), the
turnover of species among areas (Vane-Wright et
al., 1991), or relative abundance and evenness as
part of some diversity indices (Magurran, 1988).
Of 22 biodiversity indicators proposed in Global
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Table 1: Examples of biodiversity indicators used in sustainable forest management in the USA, Finland, and the
UK, grouped according to compositional, structural and functional parameters of forest ecosystems
Parameters
Composition
Country (Source)
———————————————————————————————————————–—
USA (Williams and Marcot,
Finland (Ministry of
UK: The UK Forestry
1991)
Agriculture and Forestry,
Standard (Forestry
1997)
Commission, 1998)
Genetic diversity of tree
species
Stands managed for
conservation of genetic
resources
Protection of designated sites
and sensitive areas for
threatened or rare species/
genotypes
Vegetation and habitat types
Tree species composition
(dominant species, exotics,
number of species,
proportions, key species)
Identification of semi-natural
sites and sites for new native
woodlands
Sensitive and endemic species
Protection of valuable
biotopes (e.g. ‘old-growth’
stands; undrained mires)
Inclusion of 10–20% open
space in new planting
Community diversity
Changes in number and
percentage of threatened
species
Diversification of landscape
and habitats of conifer
woodlands by . . . strategically
sited broadleaved species (5%
per unit/area)
Stand structure
Development classes of forest
stands
Management of edge
structure of stands to . . .
benefit habitat
development
Patch size, shape, habitat
edge
Proportion of natural
regeneration in relation to
total area regenerated
annually
Fragmentation
No. and volume of dead/
decaying and ancient trees
standing/fallen per unit area
Habitat linkages
No. of trees left uncut
through the production
rotation
Habitat turnover rates
Area, and change in area
subject to forest fires and
prescribed burning
Species diversity
Structure
Function
Nutrient cycling and soil
productivity
Fish habitat suitability
Human land-use trends
Natural ecosystem function
Area of forest land subject to
erosion, fertilizer, herbicide
use
Areas of the forest where
trees can be retained for the
long-term . . . formation of
dead branches and large
fallen wood characteristic of
natural forests . . . existing
veteran trees retained
Fragmentation of important
semi-natural habitats
Damage to the hydrology of
wetlands of conservation
value
Effects of grazing by wild and
domestic herbivores
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Biodiversity Strategy: Policymakers’ Guide
(WRI-IUCN-UNEP, 1992), 11 focus specifically
on measuring species richness. However, it is
exceedingly difficult to enumerate all the organisms at even small localities, let alone for tropical
areas covering tens or thousands of square kilometres (Williams and Gaston, 1994). One way to
try and overcome this is to identify indicator
species which can provide information on the
presence and status of a range of species groups
(Noss, 1998). The use of selected species groups
as indicators of overall biodiversity appears
attractive, because if suitable indicator relationships can be shown to exist, sampling for just the
selected species might greatly reduce survey costs
(Williams and Gaston, 1994). For example, surrogate taxa could be identified whose diversity is
related to that of other taxa occupying different
trophic levels (Hammond, 1994; Oliver and
Beattie, 1996). A concern with this method is that
these relationships are often poorly understood
and rarely substantiated (McKenney et al., 1994).
Areas of particularly high species richness cannot
be assumed to coincide among different groups of
organisms (Williams and Gaston, 1994). A good
example of this is the work of Saetersdal et al.
(1993), who found that deciduous woods in
western Norway that had maximal plant diversity
were not the same ones as those that had maximal
bird diversity. Oliver and Beattie (1996) and
Jonsson (1998) have conducted further studies
which have shown there can be little agreement
between different taxa.
The most useful indicators are those which
have a large number of relatively well-known
direct and indirect relationships with other
species groups, i.e. they have an important functional role in ecosystems. These are known as
keystone species, and it is argued that they should
be conserved because they have a disproportionate effect on the persistence of all other species
(Bond, 1994).
Floristic indicators
Plants are often chosen as indicators owing to
their known relationship to edaphic and climatic
factors, and their role in providing habitat for
dependent fauna. Plant community composition
(e.g. age, size and growth form of trees), species
richness (numbers of species) and diversity (a
function of both numbers of species and their
relative importance) are relatively straightforward to measure, and can provide information
on the ecological status of a site (Peterken, 1981;
Ferris-Kaan and Patterson, 1992). Classification
of plant communities using the National Vegetation Classification system (Rodwell, 1991) has
been proposed as a surrogate indicator of biodiversity, and a study is currently underway to
examine woodland vegetation patterns in
England for this purpose (Hall, 1998).
Selected plants can also function as keystone
species in supporting dependent taxa. An
example of a key compositional indicator of
forest biodiversity in managed boreal spruce
forests is birch (Betula spp.) (Figure 1). Birch
trees provide a direct source of food and shelter
for invertebrates, particularly phytophagous
insects (Watt et al., 1997), and for hole-nesting
and seed-eating birds (Patterson, 1993). Birch
also has direct and indirect effects on the
development of groundflora and soil fauna communities, through shading and soil improvement.
In turn, these species groups provide food and
shelter for birds, mammals, etc. In Finland,
broadleaves such as aspen (Populus tremula L.),
goat willow (Salix caprea L.) and other rare
species (e.g. Quercus spp.; Tilia spp.) are also
considered to be key indicator species in boreal
forests (Ministry of Agriculture and Forestry,
1997). In Swedish forests, over 25 per cent of
red-listed forest species (mostly vertebrates and
invertebrates) depend on retained broadleaves
for their survival through a forest rotation
(Gustafsson, 1998).
In boreal pine (Pinus sylvestris L.) forests, the
abundance of berry-producing ericoid shrubs
such as bilberry (Vaccinium myrtillus L.) could
prove to be a useful biodiversity indicator
(Humphrey and Coombs, 1997). Many species
are dependent on bilberry as a food source. For
example, the flowers and foliage are utilized by
Lepidoptera, which in turn provide food for
woodland grouse (Baines et al., 1994). The
berries and foliage are also consumed directly by
other birds and mammals. Where the benefits to
other taxa of groundflora species such as bilberry
are well known, it is possible to build up a knowledge of the value of particular plant assemblages.
There is considerable interest, both in Britain
and Continental Europe, in the use of vascular
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Composition
food for
invertebrates
moths/life miners
COMPOSITIONAL
INDICATOR
Composition
birds (insectivores,
hole nesters, seed
eaters etc.)
Function
soil improver –
inputs into nutrient
cycling
Birch
317
Composition
epiphytic lichens/
bryophytes
Composition
mammals – deer,
badgers, weasels
Composition
ground flora (berry
producing plants
etc.)
Composition
development of soil
fauna (earthworms)
Figure 1. Birch: a key compositional indicator of biodiversity (data from Humphrey et al., 1998; Patterson, 1993).
plants (Peterken, 1974; Honnay et al., 1999) and
bryophytes (Ratcliffe, 1968) as indicators of
ancient woodland (defined for Britain, by
Peterken (1981), as sites which have been continuously wooded since at least 1600). Of all forest
types in the UK, ancient woods are generally considered to have the highest value for biodiversity
(Ratcliffe, 1993; Peterken, 1996). The existence
of ancient woodland indicators is a consequence
of habitat fragmentation and the resulting ecological isolation of the original primary woods
(Peterken, 1996). Consequently, these indicators
are more reliable in landscapes where fragmentation and isolation of ancient woodland is greatest, since they tend to have limited powers of
dispersal to, and establishment in, newly available habitat. Where the degree of ecological isolation is low, i.e. densely wooded regions, ancient
woods are harder to distinguish floristically from
recent woods.
Lichens and fungal indicators
Lichens may be useful as indicators due to their
requirement for ecological continuity; e.g. in
ancient broadleaved woodland in the UK (Rose,
1976), and in boreal coniferous forests (Tibell,
1992). Many species of epiphytic lichens show
close association with particular tree species
(Barkman, 1958), and can also act as an important food source for many orders of invertebrates
and, in turn, insectivorous birds. Pettersson et al.
(1995) found that the abundance of fruticose and
filamentous lichens in boreal Norway spruce
(Picea abies (L.) Karst.) forests was strongly correlated with both the number of species and mass
of invertebrates.
Some species of fungi show a marked dependence on large, well-decayed logs. Specialist fungi
such as those from the families Polyporaceae,
Hymenochaetaceae and Corticiaceae may be the
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best indicators, compared with generalist species
which are able to colonize any type or quality of
dead wood (Bader et al., 1995). These specialist
fungi are indicators not only of the presence of
deadwood, but of its quality (size, state/range of
decay conditions) and of the continuity of oldgrowth conditions. In boreal Norway spruce
forests in Sweden, Bader et al. (1995) identified
several species of wood-inhabiting fungi which are
used extensively as indicators of forests of high
nature conservation value throughout Scandinavia
(Nitare and Noren, 1992). The assumption is that
different rare and endangered species often
coexist, such that a proper choice of a few indicator species is sufficient to include many other
threatened species. Jonsson (1998) found a good
correlation between deadwood, epiphytic lichens
and wood-inhabiting fungi, illustrating that the
diversity of species groups that share similar
habitat requirements can be more closely correlated.
Invertebrate indicators
Invertebrates have received considerable attention as biodiversity indicators. They are sensitive
to habitat structure and composition, and also
have great functional significance, due to their
large numbers, wide array of life history patterns,
and high inter- and intra-specific diversity
(Samways, 1993). There are specific instances
where certain insect groups have well-known
relationships with other taxa.
Butterflies have been used as the flagship species
(species regarded as valuable for aesthetic and cultural reasons, and hence generating high public
interest) for open-space conditions in British
forests (Warren and Fuller, 1993; Ferris and
Carter, in press). Many species have rather precise
foodplant requirements, e.g. the brimstone
(Gonepteryx rhamni L.), females of which have an
extraordinary ability to find even the most isolated
bushes of both purging (Rhamnus catharticus L.)
and alder buckthorn (Frangula alnus Miller), on
which they lay their eggs. Variations in populations of the wood white (Leptidea sinapis L.),
predominantly a woodland species, have been
shown to correlate with the amount of shading in
forest rides by adjacent trees; the optimum shade
conditions being in the range of 20–50 per cent
shade (Warren, 1985). These two examples serve
to demonstrate how closely butterfly populations
can be tied to very precise habitat conditions, and
hence they may be used as indicators of habitat
change or loss which may have consequences for
a range of other dependent taxa.
Carabid beetles, as primary and sometimes
secondary carnivores, integrate a substantial
amount of ecological information relating to other
parts of the biological communities to which they
belong (Day and Carthy, 1988). Studies of ground
beetle species assemblages provide indications of
habitat differences for this reason (Refseth, 1980).
Butterfield et al. (1995) found that the species
composition of carabid communities in upland
forests in northern England was closely related to
habitat type; in particular, the species composition
of clearfelled areas differed significantly from
closed canopy stands. Similarly, hoverflies (Syrphidae) may have value as potential indicators of
diversity in other arthropod groups, since they
have varying habitat requirements and functional
roles (e.g. some depend on decaying wood, some
are phytophagous, and others are predatory (Watt
et al., 1997)). Therefore, diverse syrphid communities indicate a high diversity of habitat types,
and a correspondingly greater diversity of other
species groups which utilize these habitats
(Humphrey et al., 1999).
The wealth of existing, documented information on the relationships between invertebrates
and habitat parameters (compositional and structural), means that they offer great potential as
indicators of biodiversity. In addition to being
well-studied, invertebrates may be sampled using
established, standardized methods, and expertise
is widely available.
Vertebrate indicators
Large vertebrates are often considered to be
‘umbrella’ species (Simberloff, 1998), meaning
that because they need such large tracts of habitat,
saving them will automatically save many other
species. Although this approach does not require a
demonstration of a direct link between species, it
relies upon intuition rather than research as to
whether many other species really do fall under the
umbrella. For example, large vertebrates may not
be good indicators of taxa such as insects that
might do very well in a landscape fragmented into
small patches of suitable habitat.
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A number of well-documented cases, underpinned by ecological research, serve to illustrate
the value which some vertebrates have as indicators in forest ecosystems. At the stand scale, the
presence and abundance of vertebrates may be
related to the vertical stand structure. For
example, Harrison (1962) distinguished six
mammal and bird communities in tropical
lowland rainforests, based on the canopy level and
food resources that they exploit. Similar groupings
have been recognized for less complex forest types,
such as structurally simple oak woodlands in California and Mexico (Landres and MacMahon,
1983). In these woodlands, a number of guilds of
insect-eating birds have been identified: (1) species
that primarily glean foliage; (2) species that feed
primarily from bark; either gleaning the surface or
probing beneath (e.g. woodpeckers); (3) species
which feed primarily on flying insects; and (4)
species that feed primarily on ground-dwelling
insects. Where such groupings can be recognized,
they may be used as indicators of stand structure
(or vice versa; see later section) and the presence
of associated fauna.
Woodpeckers have a particular value as indicators because they are reliant upon the presence
of deadwood; large, old, broadleaved trees in
which to excavate nest holes, and on a supply of
invertebrate food (Angelstam and Mikusinski,
1994). As important primary nest hole excavators, woodpeckers such as the black woodpecker
(Dryocopus martius) may also be viewed as
important keystone species (Johnsson, 1993) as
they promote the development of deadwood
habitat for a range of additional species.
Community composition of breeding birds in
forests has been the subject of numerous studies,
and James and Wamer (1982) have analysed
datasets from a wide variety of North American
forests. Their research has shown that both bird
species richness and density tend to be minimal in
coniferous forests characterized by high tree
density, low canopy height, and few species of
trees. Studies of songbird communities within
conifer forests in Britain have shown relationships with successional stages present, the spatial
pattern, and the extent of non-crop broadleaves
(Bibby et al., 1989; Patterson et al., 1995). Peck
(1989) has also found that there are tree species
preferences shown by foraging birds in upland
forests. In her work, both bird density and species
319
richness were positively correlated with the
number of tree species present. Birds are a wellstudied group, with established monitoring
methods, and their apparent relationships with
forest composition and structure make them valuable as indicators of wider biodiversity in forests.
Summary
In selecting compositional indicators, it is important to consider their relationships to specific
habitat parameters or, if they are generalists, their
ability to utilize a range of conditions and habitats. Their dispersal abilities will also need to be
known, so that they can be used at the appropriate spatial scale. In addition, their position in the
foodchain should be taken into consideration
(e.g. they may provide a food resource upon
which other taxa depend). Changes in the distribution, and abundance of indicator species
should indicate impending changes in the wider
ecosystem and the implications for species groups
(Stork and Samways, 1995). It is important that
the species chosen as indicators should provide
standards of consistency and precision about
these changes, and in order to achieve this, they
should possess some, or all of the following
attributes (modified from Brown, 1991a):
• high taxonomic and ecological diversity (many
species in each habitat or system);
• close association with, and identification of,
the conditions and responses of other species;
• high ecological fidelity;
• relatively high abundance and stable population size (i.e. always present and easy to
locate in the field);
• well-known taxonomy and easy identification;
• good background information (e.g. on behaviour, ecology, biogeography);
• large random samples encompassing all species
variation are possible;
• functional importance within the ecosystem is
understood.
Structural diversity
Stand structure, particularly the complexity and
dynamics of forest canopies, is of fundamental
importance to forest dynamics, e.g. spatial heterogeneity and temporal change in understorey
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vegetation, patterns in regeneration mosaics, and
microclimatic variation (Norman and Campbell,
1989; Song et al., 1997). The breadth of ecological relationships accounted for by habitat structure may be encompassed by three components:
heterogeneity, complexity, and scale (McCoy and
Bell, 1991). Heterogeneity encompasses the variation due to the relative abundance of different
structural components, whether in the vertical or
horizontal plane. Complexity refers to the variation resulting from absolute abundance of individual structural components, and scale takes
account of the variation due to the size of the area
or volume used to measure heterogeneity and
complexity. Heterogeneity and complexity are
aspects of habitat structure which may be applied
both within and between stands. It is possible that
for some associated species, total abundance of a
particular structural feature may be a reliable
indicator; while for others, its relative abundance
may be a better predictor, e.g. in cases where more
than a single habitat type is required to meet the
needs of that species, possibly through different
stages of a life cycle.
Structural complexity translates into a complex
of niches that enhance the diversity of associated
plant, animal and microbial communities. Structural assessments can therefore play a key role in
the description of biodiversity in both undisturbed and managed forests (Schuck et al., 1994).
In some cases they have the advantage of being
easily discernible and hence rapid to undertake,
thereby offering a valuable approach to biodiversity monitoring (Williams and Gaston, 1994;
Boyle and Sayer, 1995).
Differences between successional stages
Preserving biodiversity in temperate forests
requires the representation of all successional
stages (Franklin, 1988). Therefore, it is important
that the various stages can be easily defined. Oliver
and Larson (1990) have identified a general model
of stand development after disturbance or primary
succession: (1) stand initiation – after a disturbance, new individuals and species continue to
appear for a few years; (2) stem exclusion – after
several years, new individuals do not appear and
some of the existing ones die. The surviving ones
grow larger and express differences in height and
diameter (first one species and then another may
appear to dominate the stand); (3) understorey
reinitiation – later, forest floor herbs and shrubs
and advance regeneration again appear and
survive in the understorey, although they grow
very little; (4) old growth – much later, overstorey
trees die in an irregular fashion, and some of the
understorey trees begin growing to the overstorey.
This classification scheme may not be particularly relevant for forests where gap-phase disturbance dynamics predominate. These forests
often appear as complex mosaics with unevenaged structures (Parviainen et al., 1994).
Although profile drawings and crown projection
maps give basic information on stand structure, a
number of direct measurements can help to define
developmental stages. Parviainen et al. (1994) list
the following parameters: number of trees in
different diameter classes, tree species composition, layering, basal area and yield of individual
trees, structure and yield of the entire stand, tree
form, and mortality.
The total area and both spatial and temporal
distribution of broad structural or successional
stages may be used as an indicator of biodiversity
at the forest or landscape scale. Measuring variables such as the diameter of the median tree in a
stand; the range of sampled diameters; the basal
area of the stand; the estimated number of canopy
layers; and the number of tree species, allowed
stand structure variation to be determined for
virgin and managed forests in Finnish and Russian
Karelia (Uuttera et al., 1997). From these data, it
is possible to make predictions concerning habitat
suitability for dependent taxa, based on knowledge of their niche requirements, e.g. minimum
core area of particular stand types, proportion of
the stand occupied by a shrub layer etc. Canopy
arthropods have been widely studied in this
context, and research has been undertaken in
North America (Schowalter, 1989), central Europe
(Simandl, 1993), and Scandinavia (Pettersson et
al., 1995), as well as in Britain (Ozanne, 1996). In
studies of Scots pine stands of varying age,
Simandl (1993) found significant differences
related to stand age among most of the arthropod
taxa; medium-aged stands showed a higher arthropod density in the crown strata. In contrast, the
canopies of old-growth forests of the western Cascades in Oregon, dominated by Douglas fir
(Pseudotsuga menziesii (Mirb.) Franco), western
hemlock (Tsuga heterophylla (Raf.) Sarg.), and
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western red cedar (Thuja plicata D. Don), were
shown to have much higher species and functional
diversity compared with those of younger stands
(Schowalter, 1989). Similar results were obtained
for boreal Norway spruce forests in Sweden, with
old-growth forests having significantly greater
invertebrate density than managed forests, and
nearly five times as many invertebrates per branch
(Pettersson et al., 1995). In plantation forests of
Scots pine, Corsican pine (Pinus nigra var. maritima (Ait.) Melville), Norway spruce and Sitka
spruce in the UK, Ozanne (1996) found high densities of canopy arthropods, dominated by insects
which feed on epiphytes, e.g. Collembola and Psocoptera, and by mites (Acarina). However,
relationships between invertebrate communities
and habitat structure are complex interactions
between numerous variables; e.g. Gunnarson
(1996) found that arthropod density in Norway
spruce forests in Sweden was influenced by an
interactive effect of bird predation with vegetation
structure.
Diversity within stands
Factors such as tree density, locations, species composition, and crown size and shape are key determinants of canopy and gap structures (Song et al.,
1997). Variation in biomass, volume and diameter
in a tree population offer possibilities for greater
species richness through disturbance created by
competition and mortality (Huston, 1994). Differences in forest stand structure between forest
ownership groups in central Finland were assessed
using three main stand structure variables: the
range of tree diameter distribution, the number of
tree storeys and the number of tree species
(Maltamo et al., 1997). Buongiorno et al. (1994)
used tree diameter distribution as a descriptor of
stand structure and this in turn as a determinant
of biodiversity. In studies of Sitka spruce plantations in upland Britain, tree diameter diversity has
been shown to correlate with plant species diversity in the ground flora (Forestry Commission
Research Agency, unpublished data).
Picozzi et al. (1992) have developed a method
of stand structure description for assessing
habitat suitability for capercaillie (Tetrao urogallus), which recognizes the distinctions between
developmental stages, but is more precise and
quantitative. The method combines measures of
321
ground vegetation cover, based on proportions
and growth form of key species, bilberry and
heather (Calluna vulgaris (L.) Hull); nearest
neighbour distance between trees, branching
pattern, height and diameter at breast height
(d.b.h.). The criteria are used to assign the forest
stand to one of 28 types, of varying suitability for
capercaillie. Although applied specifically to Scots
pine forests, the method involves parameters that
are relatively easy to assess and, more importantly, are well understood by foresters.
Methods for the assessment of stand structure
need to address both its horizontal and vertical
components (see Ferris-Kaan and Patterson,
1992). Horizontal patchiness can be assessed in a
variety of ways, for example, recording the frequency, relative abundance, and spatial distribution of different vegetation communities.
Spatial heterogeneity of tree canopies may be
assessed using hemispherical photography, e.g.
across a range of stand growth stages in Scots
pine forests in France (Walter and Himmler,
1996). Vertical structure can be assessed in terms
of biomass (although this gives no information on
the distribution of plant structure), or as a series
of hierarchical strata (Brown, 1991b). When
forests have several canopy layers, more niches
are added. It has been shown in various studies
(e.g. MacArthur, 1964; Wilson, 1974; Moss,
1978) that bird species richness and/or diversity
correlates positively with some measure of foliage
height diversity (vertical structure) in a variety of
forest types. This measure of habitat complexity
has been used in the study of bird diversity (for a
review see Petty and Avery, 1990).
Foliage height diversity has only more recently
been related to invertebrate species richness and
diversity (e.g. Stinson and Brown, 1983).
Schowalter (1989) has demonstrated that the
species composition of canopy arthropod communities changes with the successional stage in
temperate forests. In both the conifer forests of
western Oregon and the deciduous forests of
western Carolina, sucking herbivores such as
aphids dominate young forests, but arthropod
biomass of old forests is evenly split between
defoliators and their predators. At present, too
little is known about the forces shaping ecosystem
structure to explain with certainty why such striking differences occur in the anthropod communities of young and old forests. However, it seems
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likely that part of the explanation lies in the
greater number of niches resulting from the relatively high compositional and architectural diversity of old compared with young forests (Perry,
1994). Habitat variables such as temporary gaps,
permanent open space, and waterbodies all add
to this diversity.
The fine-scale structural diversity within each
stratum is less easy to measure, but may well be
an important determinant of biodiversity, e.g. for
communities of defoliating arthropods. The
precise arrangement of leaves or needles on
branches exerts an influence on the amount of
incident radiation reaching the sub-canopy layers
and forest floor and its spectral composition.
Clearly, however, direct measurements of canopy
structure are impractical for large forest
canopies. Indirect techniques, which are based
on the close relationship between radiation penetration and canopy structure, provide a useful
alternative. For a review see Norman and Campbell (1989).
Dead and dying wood
An important component of stand structure is the
abundance of dead and dying wood. This is a key
factor for biodiversity in forest ecosystems, particularly when a range of forms of deadwood are
present (Harmon et al., 1986). Snags and logs
form the base of a foodchain including microbes,
invertebrates, small mammals, and birds (Perry,
1994). Thomas (1979), in his compilation of the
wildlife of northeastern Oregon forests, found
that 178 vertebrates – 14 amphibians and reptiles, 115 birds and 49 mammals – used fallen logs
as habitats. In addition, dying and dead wood
provides excellent sites for natural regeneration,
acting as an important seedbed and providing a
well-drained site for seeds to germinate. It is also
important to aquatic organisms when it falls into
streams and rivers, and dead trees provide nesting
sites for several bird species (Ratcliffe and
Peterken, 1995). Deadwood can therefore act
as an indicator of elements of compositional
Structure
stand development/
structure
Function
(disturbance)
fuel for forest fires
Composition
wood-utilizing
fungi
Composition
wood-utilizing
invertebrates
Function
Substrate for
natural
regeneration
STRUCTURAL
INDICATOR
Deadwood
Composition
epiphytic lichens/
bryophytes
Function
slow release
fertilizer – input into
nutrient cycling
Function
energy source for
carbon cycling
Composition
cavity nesting
birds, bats
Figure 2. Deadwood: a key structural indicator of biodiversity (based on Hodge and Peterken, 1998).
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diversity, but its presence (in different forms –
size, location and decay state) also indicates the
likelihood of particular functions being met, e.g.
nutrient cycling (Figure 2).
The presence of suitable large, decaying, broadleaved trees indicates habitat suitability for holenesting birds such as woodpeckers (Angelstam and
Mikusinski, 1994; see earlier section), as well as
other taxa which utilize cavities, e.g. woodland
bats (Mayle, 1990). Species diversity of small
mammals has also been found to show a good correlation with the levels of fallen deadwood present
in both natural young and old-growth stands in
the USA (Carey and Johnson, 1995).
In addition to measurement of the quantity of
deadwood present (see Brown, 1974), it is important to have some measure of the ‘quality’ of this
resource, i.e. its decay status. This is required
because many species have rather exacting needs,
e.g. bats prefer snags with loose bark still attached,
and an essential prerequisite for any primary excavator (e.g. woodpeckers) is some degree of wood
softening by heart-rot fungi (Perry, 1994). Various
scoring systems for assessing the qualitative
characteristics of deadwood have been proposed
(e.g. Koop, 1989; Hunter, 1990), and these can be
used to provide a measure of the rate of turnover
of deadwood in the forest system.
Summary
Structural indicators can be recognized at a
number of spatial scales, from the landscape scale
down to single trees. Forest structure itself can be
broadly defined in terms of its vertical or horizontal composition, and these can be combined
to give a three-dimensional profile for any forest
stand. Indirect methods may be used to define
structure, such as visual assessments of vertical
layering and vegetation cover, but great care
needs to be exercised when interpreting results
from such approaches. Direct measurements are
possible, particularly at smaller scales, i.e. at or
below the level of the forest stand. When measuring individual components of structure, such as
snags and fallen deadwood, direct measurements
may be used to provide quantitative data, but
may require qualitative estimates to enhance the
indicator value. To be useful as indicators, structural parameters need to satisfy a number of criteria:
323
• close association with specific plant and animal
groups;
• clearly defined and relatively easy to measure;
• possible to quantify (i.e. for setting targets);
• observer error known and minimized.
General recommendations and conclusions
Types of indicator
It is clear that we cannot measure everything.
Instead, we have to select a few variables that will
represent key components of forest biodiversity.
These representative elements are indicators –
variables that we choose to monitor, reflecting
what we consider to be important, based on the
best knowledge currently available, and what it is
feasible to measure.
Broadleaved tree species such as oak (Quercus
spp.) or birch (Betula spp.), and ancient woodland
indicator species, are examples of key compositional indicators. However, it is important that
the interrelationships between indicator species/
groups, their habitat requirements, and the species
groups that they are intended to indicate are properly understood. This review has highlighted
numerous studies which have examined such
relationships for both invertebrates and bird communities, suggesting that these taxa can be used as
indicators of habitat diversity at the stand and
forest scales and, in the case of birds, also at the
wider landscape scale. Relationships can be both
direct, e.g. predator–prey relationships, where
predator abundance and diversity are directly
linked to prey diversity and abundance; and indirect, where different species groups are linked
through a mutual dependence on the same habitat
or substrate, e.g. wood-utilizing fungi and epiphytic bryophytes on deadwood.
The most widely-used structural indicators are
measures of stand structure (vertical structure
and horizontal patchiness) including, specifically,
deadwood. In fact, the quantity and quality of
deadwood is considered to be a key structural
indicator of biodiversity in forest ecosystems, as
it functions as a habitat for a wide range of
species groups and has a pivotal role in ecosystem
functioning. This last point is especially relevant,
since it may be argued that the best indicators are
those which have a key role in the functioning of
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forest ecosystems and are linked directly and indirectly with a wide range of species groups.
Identification of a suite of biodiversity indicators
Some indicators are easier to measure than
others. Forest managers need indicators that are
practical and straightforward, and these should
include structural measures such as deadwood
and vertical structure, and compositional indicators such as tree species and ground flora
composition. Many of the characteristics of vegetation may be discerned by non-specialists, using
relatively simple assessment techniques. For each
major forest and woodland type, a useful
approach would be to identify indicators in the
two categories – compositional and structural:
• two to three key compositional indicators that
can be shown to be linked functionally to a
broad range of other species, e.g. the extent and
species composition of the broadleaved component in conifer forests;
• two to three key structural indicators which act
as surrogates for general species richness/diversity, e.g. the quantity and quality of deadwood.
Current research as part of a biodiversity assessment project (Forestry Commission Research
Agency, unpublished) recognizes the need for such
a concise suite of measures, for application at the
stand scale, and a similar view has been taken by
Van Den Meerschaut and Vandekerkhove (in
press) for the development of a biodiversity index
for use in Belgian forests. This index uses a scoring
system, based on a combination of key structural
indicators of biodiversity. These include features of
stand structure (canopy closure/cover, stand age,
number of storeys, spatial tree species mixture,
number of native tree species, variation in diameter classes, number of large trees: d.b.h. > 40 < 80
cm, number of very large trees: d.b.h. > 80 cm, and
incidence of naturally regenerating native tree
species); the herb layer (number of vascular plant
species, rarity, number of bryophyte species, total
vegetation cover); deadwood (number of large
trees with a d.b.h > 40 cm, variation in diameter
classes). Each of these factors has a weighted
maximum score, and they are summed to give a
comparative measure.
Such a suite of indicators needs to be developed
with a clear appreciation of the spatial scales to
which each may be reliably applied. Furthermore,
specific hypotheses, frameworks and objectives
need to be stated. For example, there needs to be
an accepted standard of what actually represents
a healthy or desirable condition, i.e. a benchmark.
Indicators also need to be selected so that if future
states/conditions are projected, they can be related
to alternative management scenarios, and an
appropriate system for monitoring put in place to
detect any changes or trends in biodiversity.
Finally, indicators need to be found which are
relatively easy to measure, allowing them to be
assessed by non-specialists, which in itself may
help to make them cost-effective, and ecologically
reliable. Failure to meet these needs may mean that
they are simply not trusted and adopted by practitioners. When benefits from the use of indicators
can be demonstrated, both in terms of acting as an
early-warning system of likely deleterious changes,
and also to show whether particular management
targets are being met; they will become an integral
part of sustainable forest management.
Acknowledgements
We would like to thank Gary Kerr, Peter Freer-Smith,
Keith Kirby and Robin Gill for helpful comments on
earlier drafts of the text.
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