Science needs for biodiversity management: An exploration through UK policy
development
S.A. Bailey*
Corporate and Forestry Support, Forestry Commission, 231 Corstorphine Road,
Edinburgh, EH12 7AT, UK. e-mail: sallie.bailey@forestry.gsi.gov.uk
* the views presented here are purely the opinions of the author and do not reflect those of
the Forestry Commission.
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Abstract
A number of commentators have recently provided evidence where policy has
misdirected resources for biological conservation to strategies that have been
proven to be ineffective in reversing biodiversity loss. Traditionally forest policy is
based on sound scientific evidence developed to answer specific management
questions. However, increasingly political pressure and climate change are now
modifying the arena in which policy is formed, forcing reliance on first principles
to guide decisions. There is concern that forest management may be misdirected
in ways that do little to enhance biodiversity or even have deleterious effects.
Examples of evidence based policy development, using on sound scientific
research, for biodiversity within British Forestry are given. I then examine case
studies of the development of future policy highlighting research gaps and
suggest how future research directions are vital to inform policy to avoid missallocation of resource.
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Introduction
Biodiversity policy development
Generally, the development of policy is driven by the need to modify current behaviour or
circumstances to respond to a perceived need, building on the conventional view that ‘policy
is merely the choice of means to serve desired ends’ (Anon, 2005). In the case of policy for
forest biodiversity, the perceived need is to address observed biodiversity decline by
modifying management behaviour. The recent escalation of policy formation in this area was
stimulated by nothing less than global agreement that biodiversity decline must be
addressed, formalised at the United Nations Conference on Environment and Development,
Rio de Janeiro, in 1992.
A recent study of the policy implemented to protect biodiversity in European agricultural
landscapes, suggests that it has been unsuccessful for the target species (Kleijn et al., 2001;
2004). Policy had responded to the decline in avian communities, in particular waders and
meadow species, by offering incentives for management agreements designed to reverse
this decline. Since 1981, farmers in the Netherlands had received grants to modify current
agricultural practices in ways considered beneficial to biodiversity, eg., by postponing
mowing and restricting the application of fertilizer. Yet, when Kleijn et al. (2001) compared
areas with and without agreements no positive effects were found on birds, including the
wader species specifically targeted in the schemes, or vascular plants. Without the
scientifically sound evaluation of the basis of such policies, and continued monitoring, that
Kleijn et al. suggest, resources may continue to be miss-directed into unsuccessful schemes.
There is a critical need to monitor end-point to ensure actions are appropriate solution to
problem initially identified.
Prior to the recognition of declining biodiversity, wildlife resource was valued quite differently
by the forest industry often whose sole driver was the production of quality timber. The main
focus of ecological research was on wildlife ‘control’, to avoid damage to commercial crops
by deer and squirrels. Species protected by law required policy to ensure timber production
was sympathetic to their requirements, and it was this that drove much species-specific
research (primarily for raptors). Moreover, management for wildlife and conservation was
considered an insular activity, and very much confined to designated reserves. This resulted
in management strategies being localised and single purpose. The research questions
required to develop the evidence base for policy could be clearly defined, and the science
sharply focussed. Research activities were dominated by field and lab experimentation and
field observation. Generally, the development of conservation policy drew heavily on the
application of the precautionary principle. For example, currently in the case of semi-natural
woodland restoration, target woodland type and species choice are determined by current
and historic species distribution and site type. Research that developed guidelines for
genetic provenance advocates the use of local stock for planting and has defined quite
precisely the limits of localness, to ensure that planting stock is adapted to the locality.
Changing scenarios for policy development
British forest policy has adapted to meet commitments made both at Rio and subsequently
(MCPFE, PEBLDS, UKBAP), plus pressures to broaden the objectives of forest
management. In 1994, Britain adopted the principles of multi-purpose and sustainable forest
management (FC, 1994). The Forestry Commission objectives now include requirements to
‘protect Britain’s forests and woodlands’ and ‘conserve and enhance biodiversity in and
around woods and forests ’ (FC, 2004). The rapid evolution of forest biodiversity policy
contrasts dramatically with the deeply entrenched production-orientated mindset pre-Rio.
The political arena has changed further through more involvement from Europe in natural
resource management. The acceptance that mean annual temperature in the UK is expected
to increase by between 3 and 6ºC by the 2080s (Hulme et al., 2002) has had profound
implications for policy development. Faced with dramatic predictions of the movement in
‘climate space’ (Harrison et al., 2001) and changes to community membership, the main
objective of policy is no longer just to safeguard biological resources, the changing nature of
ecological resources must be anticipated to ensure action is robust and sustainable.
This, together with the global recognition of the decline in biodiversity, has exacerbated the
need for rapid policy development and has led to a very real change in the way policy is
formed. The focus of science is changing with emphasis shifting to future-casting, to ensure
management techniques developed today are working towards scenarios that can be
sustained in a changed environment. Modelling techniques are common place and their
sophistication is developing rapidly, enabling the testing of scenarios of new low impact,
multi-purpose management techniques in the light of climate change. New demands from
policy makers are creating new challenges for the research community. For example;
accommodating the breadth of temporal and spatial scale covered by research topics; the
lack of understanding, yet demand for provision of interim guidance on complex
subjects/systems; plus rationalising results with issues of values, preferences, and ethics as
society becomes more interested in countryside management. The challenges for both policy
makers and scientists will be illustrated through a case study of a new area of policy, the
formation of Forest Habitat Networks.
Case study
Development of Forest Habitat Networks
It is generally accepted that managing for biodiversity at the landscape scale is more likely to
have benefit for most species than if reserves or patches of habitat are considered in
isolation. Research has also shown that biodiversity decreases following fragmentation of
key habitats (Spellerberg, 1995; Bailey, In Review). Reversing fragmentation by restoring
connectivity is being frequently proposed as an effective strategy to address biodiversity
decline within fragmented habitats (Kirby, 1995; Peterken, 1995; Spellerberg, 1995). Further
interest in restoring connectivity has been stimulated by the proposition that climate change
will shift the geographic range of habitats, forcing associated species to either track this shift
through dispersal or face local extinction (Berry et al., 2002). If habitats are highly
fragmented then species will not be able to adjust to the new ‘climate space’ and thus face
extinction as they become stranded in increasingly unsuitable habitat (Holt and Keitt, 2000;
Pearson and Dawson, 2003). Within the new landscape ecological paradigm, increasing
connectivity through planting new woodlands and restoring other semi-natural habitats is
inherently attractive and is a good fit with other forest policy to increase woodland cover in
the UK (UK Forest standard). There has therefore been a rapid uptake of the concept by
policy makers, not only within the UK (UKBAP) (Forestry Commission, 2000; 2001), but also
at the European level within the Habitats and Species Directive, Pan European Ecological
Network (EEC, 1992; Council of Europe, 1996).
Inherent in the concept of ‘reversing fragmentation to reverse biodiversity decline’ is the
assumption that species will use additional woodland and semi-natural habitats to recolonise
fragments of woodland and develop viable populations. Modelling studies do suggest that
increasing connectivity in fragmented landscapes will reverse biodiversity loss (e.g., Opdam
et al., 2003). Yet, there is little systematic research available on the effects of restoration
programmes designed to reverse the effects of fragmentation (Henle et al., 2004) to evaluate
whether species are capable of recovery in newly connected landscapes. In fact, the use of
corridors, established, new or otherwise has been the source of much debate in the scientific
literature for some time (e.g., Wiens, 2002; Catchpole, 2004). The role of long distance
dispersal events (LDD), although rare is also throwing into question the level of resource
focussed on increasing connectivity. It has been suggested that it is LDD that will be key in
enabling species to keep pace with shifting ranges in fragmented landscapes. These events
are particularly important for the dispersal of otherwise sedentary species associated with
interior of forests, the species often of most conservation concern. At least four species of
saproxylic insects, previously considered to have poor dispersal powers, have experienced a
rapid expansion of their range through long distance dispersal events (Alexander, 2003;
2004). However, observations of LDD events remain elusive often due to the inability to
detect such events through experimental design.
Within this and other policies intended to benefit biodiversity often the end point can become
lost or confused. Policy documents state that it is the restoration of biodiversity that they are
striving towards. Yet, biodiversity can only become a target when it is clearly defined (Failing
and Gregory, 2003). Care must be taken that species richness does not become a surrogate
for woodland biodiversity (sensu Heilmann-Clausen and Christensen, 2005). It appears that
what we are really seeking to achieve is a compliment of species that will develop into a
functioning woodland community with the full range of natural processes and associated
structural diversity represented. If connectivity is increased it is likely that new woodlands will
quite quickly develop high numbers of the ‘usual woodland suspects’ (Honnay et al., 2002),
both native and non-native that are not suppressed by fragmentation. Yet this does not seem
to be a suitable end point of fragmentation reversal. If increasing connectivity is to benefit
those species most at risk from fragmentation and therefore develop full community
membership, more complex strategies maybe required (expanding existing woodlands, even
importing specific habitats, e.g. deadwood (Honnay et al., 2002)).
Conclusions
In an ideal world policy needs to be ‘evidence-based’ and delivery needs to structured on
well-founded ‘best practice’, as it was in the past. But we have to be realistic and accept that
there is such a high level of interest in biodiversity that policy and practice must be modified.
Referring back to the example of maintaining genetic provenance, using the existing or
historic distribution to conserve our native genetic base may be acting to reduce the
robustness of woodland ecosystems to the threats posed by global climate change. Positive
management may now be necessary to help adapt woodland ecosystems to predicted
climate change. Potential changes to policy could include; the use of non-local and/or nonnative provenances of native species; planting of species not common or accepted in a
particular region; planting of non-native species and the translocation of species, populations
or vegetation assemblages to regions that are predicted to have a more suitable climate and
thus ensure their continuing survival. The forest habitat networks case study illustrates that
modelling fulfils a vital function, however to continue to inform policy the evidence base must
be continually reinforced with field-based research that can parameterise many of the current
assumptions.
Whilst the research performed at the frontiers of academic disciplines can develop theories
based on first principles and form the basis for policy development, research then needs to
move to solving the real problems of woodland management within the new political arena.
To avoid pitfalls that Kleijn et al., (2001, 2004) highlight, end-points must be clearly defined
that are both attainable and sustainable. Policy makers must also build in an empirical
approach to implementation, facilitating the informing, attribution and validation of research
models, and be ready and willing to change approach if desired outcomes are not achieved.
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