REVIEW: The management of wild large herbivores to meet economic, conservation and environmental objectives

Abstract

Summary • Wild large herbivores provide goods and income to rural communities, have major impacts on land use and habitats of conservation importance and, in some cases, face local or global extinction. As a result, substantial effort is applied to their management across the globe. To be effective, however, management has to be science-based. We reviewed recent fundamental and applied studies of large herbivores with particular emphasis on the relationship between the spatial and temporal scales of ecosystem response, management decision and implementation. • Long-term population dynamics research has revealed fundamental differences in how sex/age classes are affected by changes in density and weather. Consequently, management must be tailored to the age and sex structure of the population, rather than to simple population counts. • Herbivory by large ungulates shapes the structure, diversity and functioning of most terrestrial ecosystems. Recent research has shown that fundamental herbivore/vegetation interactions driving landscape change are localized, often at scales of a few metres. For example, sheep and deer will selectively browse heather Calluna vulgaris at the edge of preferred grass patches in heather moorland. As heather is vulnerable to heavy defoliation, in the long term this can lead to loss of heather cover despite the average utilization rate of heather in a management area being low. Therefore, while herbivore population management requires a large-scale approach, management of herbivore impacts on vegetation may require a much more flexible and site-specific approach. • Localized impacts on vegetation have cascading effects on biodiversity, because changes in vegetation structure and composition, induced by large herbivores affect habitat suitability for many other species. As such, grazing should be considered as a tool for broader biodiversity management requiring a more sophisticated approach than just, for example, eliminating grazing from conservation areas through the use of exclosures. • Synthesis and applications. The management of wild large herbivores must consider different spatial scales, from small patches of vegetation to boundaries of an animal population. It also requires long-term planning based on a deep understanding of how population processes, such a birth rate, death rate and age structure, are affected by changes in land use and climate and how these affect localized herbivore impacts. Because wild herbivores do not observe administrative or political boundaries, adjusting their management to socio-political realities can present a challenge. Many developing countries have established co-operative management groups that allow all interested parties to be involved in the development of management plans; developed countries have a lot to learn from the developing world's example. Journal of Applied Ecology (2004) 41, 1021 –1031

Full text

Journal of Applied
Ecology
2004
41
, 1021– 1031
© 2004 British
Ecological Society
Blackwell Publishing, Ltd.Oxford, UKJPEJournal of Applied Ecology0021-8901British Ecological Society, 200412 20044161021
Review ArticleManagement of wild large herbivoresI. J. Gordon, A. J. Hester & M. Festa-Bianchet
REVIEW
The management of wild large herbivores to meet
economic, conservation and environmental objectives
IAIN J. GORDON,*† ALISON J. HESTER* and MARC O FESTA-BIANCHET‡
*
Macaulay Institute, Craigiebuckler, Aberdeen, AB15 8QH, UK;
†
CSIRO-Davies Laboratory, PMB PO Aitkenvale,
Queensland 4814, Australia; and
‡
Department of Biology, University of Sherbrooke, Sherbrooke, Canada PQ J1K 2R1
Summary
1.
Wild large herbivores provide goods and income to rural communities, have major
impacts on land use and habitats of conservation importance and, in some cases, face
local or global extinction. As a result, substantial effort is applied to their management
across the globe. To be effective, however, management has to be science-based. We
reviewed recent fundamental and applied studies of large herbivores with particular
emphasis on the relationship between the spatial and temporal scales of ecosystem
response, management decision and implementation.
2.
Long-term population dynamics research has revealed fundamental differences
in how sex/age classes are affected by changes in density and weather. Consequently,
management must be tailored to the age and sex structure of the population, rather
than to simple population counts.
3.
Herbivory by large ungulates shapes the structure, diversity and functioning of most
terrestrial ecosystems. Recent research has shown that fundamental herbivore/vegeta-
tion interactions driving landscape change are localized, often at scales of a few metres.
For example, sheep and deer will selectively browse heather
Calluna vulgaris
at the edge
of preferred grass patches in heather moorland. As heather is vulnerable to heavy
defoliation, in the long term this can lead to loss of heather cover despite the average
utilization rate of heather in a management area being low. Therefore, while herbivore
population management requires a large-scale approach, management of herbivore
impacts on vegetation may require a much more flexible and site-specific approach.
4.
Localized impacts on vegetation have cascading effects on biodiversity, because
changes in vegetation structure and composition, induced by large herbivores affect habitat
suitability for many other species. As such, grazing should be considered as a tool for
broader biodiversity management requiring a more sophisticated approach than just,
for example, eliminating grazing from conservation areas through the use of exclosures.
5.
Synthesis and applications
. The management of wild large herbivores must consider
different spatial scales, from small patches of vegetation to boundaries of an animal
population. It also requires long-term planning based on a deep understanding of how
population processes, such a birth rate, death rate and age structure, are affected by
changes in land use and climate and how these affect localized herbivore impacts.
Because wild herbivores do not observe administrative or political boundaries, adjusting
their management to socio-political realities can present a challenge. Many developing
countries have established co-operative management groups that allow all interested
parties to be involved in the development of management plans; developed countries
have a lot to learn from the developing world’s example.
Key-words
:bighorn sheep, conservation, habitat management, impala, population
dynamics, red deer, saiga antelope, ungulates
Journal of Applied Ecology
(2004)
41
, 1021–1031
Correspondence: I. J. Gordon, CSIRO-Davies Laboratory, PMB PO Aitkenvale, Qld 4814, Australia (fax + 61 74753 8600; e-mail
iain.gordon@csiro.au).

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A. J. Hester &
M. Festa-Bianchet
© 2004 British
Ecological Society,
Journal of Applied
Ecology
,
41
,
1021–1031
Introduction
In many parts of the world, populations of wild large
herbivores provide a substantial resource supplying
local and regional communities with goods and eco-
nomic income (Conover 1997; Barnes, Schier & van
Rooy 1999; Loibooki
et al
. 2002; Ogutu 2002). They
also have a major impact on land use and habitats of
conservation importance (Hobbs 1996; Kirby 2001).
Some species are the targets of policies to conserve
dwindling populations (Stanley Price 1989; IUCN
2002) while others are increasing in number and need
to be controlled, for example deer in UK woodlands
(Putman & Moore 1998). In this review, we highlight
how recent ecological research has investigated the
relationships between large herbivores and their re-
sources, yielding insights into the dynamics of the
herbivores and the vegetation upon which they subsist.
We emphasize the role of spatial and temporal variation
in herbivore and vegetation abundance, giving examples
of how this information could provide guidance to
those responsible for devising and implementing the
conservation, harvesting and culling of large herbivores
across the globe.
Historically, the vast populations of large herbivores
that roamed the plains of Africa, the steppes of Asia
and the prairies of America appeared to offer a vast,
bountiful resource for humans to exploit (Roosevelt
1910). However, overexploitation, predation, disease
and changes in climate and land use have reduced many
large herbivore species to levels at which they now need
to be actively conserved (Beard 1988; Teer
et al
. 1996;
Danz 1997). On the other hand, in the developed world
some species have benefited from climate and land-use
changes, reduced human off-take and the removal of
predators, and now require management to ensure that
their numbers do not affect other land-use objectives,
including agriculture, forestry and habitat requirements
for other species (Gill 1990; McShea, Underwood &
Rappole 1997).
For well over half a century, one of the prime justifica-
tions of applied ecological research has been to provide
objective information for ecosystem managers who
wish to meet environmental and economic objectives
(Elton 1924; Sheail 1985, 1987). For example, since the
mid-1950s information derived from large game counts
has been used to set harvest quotas for many ungulates
in Africa, North America and Europe, either for sport
hunting (Caughley & Sinclair 1994) or to reduce
impacts on, for example, commercial timber stocks
(Putman & Moore 1998; Terry, Mclellan & Watts
2000). While this long history of linking ecological
research with management advice has been valuable to
both parties (Sheail 1987; Sutherland 2000), ecologists
are now expected to link their science more closely with
the needs of the public if scarce public funds are to be
channelled into research rather than competing uses
such as education, health, transport and the military
(Dale
et al
. 2000). Furthermore, natural resource
managers are seeking increasingly sophisticated advice,
for example the escalating costs of both surveying
and culling large herbivores means that more precise
information is needed for managers regarding how
many animals to cull or which sections of the population
(e.g. age/sex classes and geographical location) are
damaging natural resources (Gill 1992; Georgiadis,
Hack & Turpin 2003).
Recently, both the research and management com-
munities have been questioned about the extent to
which advances in the understanding of the ecology of
natural resources have been used to guide management
(Stinchcombe
et al
. 2002). In an effort to address this
issue we present recent developments in large herbivore
ecology that are most likely to guide the development
of management planning to meet economic, conservation
and environmental objectives. For these developments
to be taken up by managers, ecologists must collaborate
with researchers in the humanities to develop method-
ologies by which ecological science-based management
is adopted by managers and policy makers.
Why does large herbivore management matter?
First, large herbivores have high economic value; they
are often an important source of revenue through sport
hunting (Williamson & Doster 1981; van der Waal &
Dekker 2000; Leader-Williams, Smith & Walpole
2001) and ecotourism (Barnes, Schier & van Rooy
1999; Ogutu 2002). They can also be major pests to
agriculture, forestry and conservation areas, and they
may present serious traffic hazards (Ratcliffe 1987;
McShea, Underwood & Rappole 1997; Ramsay 1997;
Malo, Suárez & Díez 2004). For example, at present
a trophy value of approximately US$10500 is placed
on a male elephant
Loxodonta africana
in Zimbabwe
(http://safariconsultants.com/zornframespage.htm), with
much of this revenue returning to natural resource manage-
ment organizations and local communities (Murphree
2001). In Scotland, the cull of more than 70000 red
deer
Cervus elaphus
per year generates more than £5
million per annum, and 300 permanent and 450 part-
time jobs for the rural economy (Reynolds & Staines
1997). The latter value is likely to be substantially
higher when ancillary activities such as accommodation,
transport, craft and food and drink purchases are taken
into account (K. Thomson, W. Slea & D. Macmillan,
personal communication). As long ago as 1975, deer
hunters in the USA were estimated to spend more than
$1 billion annually on pursuing their sport (Williamson
& Doster 1981) and this is likely to be substantially
higher today.
Secondly, some large mammalian herbivores are
priorities for conservation because their populations are
critically low as a consequence of habitat loss, persecution
and overhunting. Of the approximately 175 species of
ungulates in the world, 84 are listed as critically endan-
gered, endangered or vulnerable in the 2002
Red Data
Book
of the International Union for the Conservation

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of Nature (IUCN) (IUCN 2002; http://www.redlist.org/).
Large herbivores are often used as flagship species for
conservation management planning because of their
high public profile (Stanley Price 1989; Bowen-Jones &
Entwistle 2002) and because they are keystone species
in many ecosystems (Danell
et al
. in press).
Finally, across the globe, large areas are grazed by
wild herbivores that drive the structure, composition
and functioning of these ecosystems (Miles 1985; Mar-
tin 1993; Thompson, Hester & Usher 1995; Pickup,
Bastin & Chewings 1998; Wallis de Vries, Bakker & van
Wieren 1998). High densities of large herbivores can
impact upon the agricultural, conservation and envi-
ronmental value of the landscape (McShea, Under-
wood & Rappole 1997).
Successful large herbivore management, be it driven
by economic goals or by the desire to conserve and
expand specific habitats or species, requires a clear
understanding of the processes involved in plant–
herbivore interactions and their consequences for the
dynamics of both plants and herbivores; in this context
applied ecological research is crucial. The following
sections address key areas of importance to managers
to demonstrate the value of ecological research in
understanding the population dynamics of large
herbivores and their impacts on natural resources.
Density-dependent and density-independent
drivers of large herbivore population dynamics
An understanding of what factors cause animal popula-
tions to increase, decrease or remain stable is fundamental
to the provision of advice on how to manage them.
While it is advocated that species restoration and con-
servation plans should be based upon sound studies of
the species’ population ecology and habitat requirements
(Bodmer, Fang & Ibanez 1988; Stanley Price 1989),
there are still many cases where the fundamental data
required to inform the process are lacking (Stinchcombe
et al
. 2002) or where developments in theoretical ecology
are not incorporated into management plans.
Long-term ecological studies of population dynamics
in large mammalian herbivores provide a detailed under-
standing of the effects of intrinsic and extrinsic factors
in determining population size and composition
(Saether 1997; Gaillard, Festa-Bianchet & Yoccoz 1998;
Gaillard
et al
. 2000). These fundamental studies have
focused on the relationships between population density,
weather and individual survival rates of different sex/
age classes (Gaillard, Festa-Bianchet & Yoccoz 1998;
Gaillard
et al
. 2000). Density-independent effects
(Milner-Gulland 1997; Smith & Anderson 1998; Coulson
et al
. 2001) also impact upon large mammal popula-
tions. There are often interactions between density-
dependent and density-independent effects, because
malnourished animals are more likely to succumb to
severe climatic events at high than at low population
densities (e.g. roe deer
Capreolus capreolus
, Gaillard
et al
. 1997; moose
Alces alces
, Crete & Courtois 1997;
bighorn sheep
Ovis canadensis
, Portier
et al
. 1998;
alpine ibex
Capra ibex
, Jacobson
et al
. 2004). More
importantly, it has become evident that the impacts of
both density-dependent and density-independent effects
vary substantially according to a population’s sex/age
structure. This is because the survival of different sex/ age
classes is not equally affected by resource abundance
and inclement weather. In general, adults are relatively
impervious to density and weather effects, while
juveniles (and possibly senescent individuals) are
highly susceptible to both (Gaillard, Festa-Bianchet &
Yoccoz 1998; Gaillard
et al
. 2000; Coulson
et al
. 2001).
This finding has crucial relevance to management strat-
egies, first because it underlines that population projec-
tion forecasts must take sex/age structure into account
(Gaillard, Loison & Toïgo 2003), and secondly because
it suggests possible fundamental differences in how
exploited and unexploited populations will react to
changes in density and weather. The proportion of
juveniles and yearlings (the age classes most sensitive to
both weather and density) tends to be much greater in
harvested than in unharvested populations, while the
proportion of senescent individuals (with higher
mortality and sometimes lower fecundity; Gaillard
et al
. 2000) is much lower (Langvatn & Loison 1999;
Apollino, Bassano & Mustoni 2003; Festa-Bianchet 2003;
Festa-Bianchet, Gaillard & Côté 2003). It therefore
seems reasonable to predict that the growth rate of
heavily harvested populations may vary more than that
of unharvested or lightly harvested populations in
response to changes in weather and density. This idea
merits further consideration, because much of our
current understanding of population dynamics in un-
gulates comes from long-term monitoring of unharvested
populations (Gaillard
et al
. 2000). In particular, un-
gulate populations subject to heavy harvest should show
steep declines following seasonally harsh weather
(because of the high proportion of juveniles) and pos-
sibly rapid increases following either mild weather or a
relaxation of harvest (because of the sudden influx of
young reproducing females, and the very high adult
survival given the young age structure). If those predic-
tions are correct, sport hunting, as currently practised
in many areas, may increase population variability,
the opposite of the often-stated goal of management
programmes (Fryxell
et al
. 1991; Langvatn & Loison
1999). Management programmes that target young of the
year for a substantial proportion of the harvest should
be less likely to increase the amplitude of weather-
related population fluctuations and more likely to
maintain an age structure not radically different from
that in naturally regulated populations.
Recent research on population dynamics of herbivores
has also underlined the importance of time lags in both
weather effects and density-dependence, because of a
combination of delays in the recovery of overgrazed
vegetation and the effects of changes in the age struc-
ture of the population (Saether 1997; Post & Stenseth
1998). An important applied consequence of time lags

1024
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© 2004 British
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,
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,
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in the population response of herbivores is that if
managers determine harvest quotas based on current
population estimates, they risk overharvesting declin-
ing populations and underharvesting recovering ones,
amplifying rather than dampening fluctuations in
population density. This has been reported for white-
tailed deer
Odocoileus virginianus
in Canada and
moose
Alces alces
in Scandinavia (Fryxell
et al
. 1991;
Solberg
et al
. 1999).
The saiga antelope
Saiga tartarica
L. demonstrates
the interaction between density-dependent and density-
independent effects on a species for which management
is critical for both the economy of local communities
and for saiga conservation. The saiga occupies the
semi-arid steppes of Kazakstan, Russia and Mongolia
(Bekenov, Grachev & Milner-Gulland 1998). Historic-
ally, the species was exploited for its meat in a regulated
fashion. However, with the opening of the Chinese
medicine market in the 1990s, there has been increasing
pressure on male saiga, the horns of which fetch a
high price (Chan, Maksimuk & Zhirnov 1995; Milner-
Gulland 1997), leading to a dramatic decline in numbers
(Sharp 2002). The species was listed under Appendix
II of the Convention on the International Trade in
Endangered Species of Wild Fauna and Flora (CITES)
in 1995 (Baillie & Groombridge 1996) and was assessed
as critically endangered by IUCN in 2002 (IUCN 2002;
http://www.redlist.org/). Saiga antelope populations
are characterized by large fluctuations in size, primarily
attributed to density-independent factors (Milner-
Gulland 1997; Coulson
et al
. 2001). Summer droughts
and severe winters affect birth rates of adult and
yearling females, and mortality of both young and
old animals. Recent modelling has suggested that
sustainable legal harvests of saiga can be achieved
through risk-averse management (Milner-Gulland
1997), with quotas set that account for vulnerability to
severe weather. This example demonstrates how an
increased understanding of density-independent
and density-dependent effects could be incorporated
into future models, which could then provide more
realistic predictions of population dynamics. However,
as has been demonstrated for saiga antelope over
the past 2 years, no amount of ecologically based
advice will save a species from decimation if legal
frameworks are not implemented to reduce poaching.
This requires an understanding of proximate causes
of poaching that include local poverty, lack of law
enforcement and open trade across national boarders
(Milner-Gulland
et al
. 2001; Sharp 2002). Currently,
saiga antelope cannot sustain any harvest (Milner-
Gulland
et al
. 2003).
To date most studies on large mammalian herbivores
have concentrated on temperate species (reviewed by
Gaillard
et al
. 2000) and have focused on long-term
monitoring programmes of a few populations, mostly
in the northern hemisphere and mostly unexploited
(Clutton-Brock, Guiness & Albon 1982; Festa-Bianchet,
Gaillard & Côté 2003). These populations may behave
differently from those in other environmental contexts,
or those subject to controlled harvests. As such there
are limits to the generalizations and extrapolations that
can be made from these studies. We encourage more
research to further our understanding of the impact of
intrinsic and extrinsic processes on population dynamics
in tropical ungulates, including predation (Messier
1994) and the effects of harvests on sex/age structure
and on age-specific vital rates (Sinclair 1977; Owen-Smith
1993; Mduma, Sinclair & Hilborn 1999).
Population response to age- and sex-specific
culling
Many management schemes prescribe sex-specific levels
of culling. Ecological research has demonstrated how
culling different sexes has very different impacts on
population dynamics (Mysterud, Coulson & Stenseth
2002). A failure to take this into account can lead to
unexpected and often undesirable consequences
(Gaillard, Loison & Toïgo 2003).
Many African antelope species have been put forward
as possible candidates for sustainable wildlife harvest-
ing schemes (Darling 1960), to the extent that attempts
were made in the 1970s to domesticate some species
(e.g. Lewis 1975). The impala
Aerycerus melampus
is a
medium-sized antelope, ubiquitous in the semi-arid
bush savannas of southern Africa (Kingdon 1972;
Smithers 1983). It is a highly social species in which the
females range in medium to large groups, with each
group accompanied by an adult male (Murray 1982).
As is common in many polygynous antelope species
(Jarman 1974), the males carry horns whereas the
females are hornless. Historically, the hunting pressure
on impala was very heavily male-biased as the species
was hunted for its trophy value (Fairall 1985). It was
suggested that male-biased hunting pressure may limit
population size because female fecundity may be
reduced when trophy males are removed (Fairall 1985;
Ginsberg & Milner-Gulland 1994). This was disputed
by other authors because young males can fertilize
females in the absence of trophy males (Mysterud,
Coulson & Stenseth 2002). In saiga antelope, however,
the sex-biased harvest is thought to have led to repro-
ductive collapse (Milner-Gulland
et al
. 2003).
Modelling, using parameters derived from auteco-
logical studies of impala, suggests that, under certain
circumstances, strongly sex- and age-biased hunting
(as occurs in game ranches and under trophy hunting)
can lead to population collapse in ungulates (Fairall
1985; Mysterud, Coulson & Stenseth 2002). This may
also be true in other species where males represent a
greater economic resource than females (e.g. elephant
Loxodonta africana
, Milner-Gulland & Mace 1991;
moose
Alces alces
, Solberg
et al
. 2002; but see Laurian
et al
. 2000; saiga
Saiga tartarica
, Milner-Gulland
et al
.
2003). For example, many hunted populations of
wapiti
Cervus elaphus canadensis
have post-hunt ratios of
five or fewer males per 100 females (Bender
et al
. 2002)

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,
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and few males survive past 4 years of age (Biederbeck,
Boulay & Jackson 2001), possibly affecting the timing
of conceptions (Noyes
et al
. 1996). Most of the evidence,
however, currently suggests that extreme sex ratio biases
(less than five males per 100 females) are required to
affect population productivity.
More recently, concerns have been raised about how
sport hunting mortality leads to sex- and age-specific
mortality rates that are radically different from those in
unhunted populations. Sport harvest may have evolu-
tionary consequences such as changes in life-history
parameters, as have been recorded in commercially
exploited fish, including changes in size and age at
maturity and possibly in reproductive effort (Jennings,
Reynolds & Mills 1998; Law 2001; Harris, Wall &
Allendorf 2002; Festa-Bianchet 2003; Olsen
et al
. 2004).
Recent evidence strongly suggests that high levels of trophy
hunting, whereby males with the largest horns are targeted,
can select for small-horned males over a few generations
(Coltman
et al
. 2003). It is becoming evident that managers
should consider evolution, as well as population dynamics,
in deciding which animals to cull. Moving towards ‘evo-
lutionarily enlightened management’ (Ashley
et al
. 2003)
will be a major challenge facing the management of
large herbivores over the next few years.
The link between sex-specific culling programmes
and trophy harvest opportunities is clearly demonstrated
in red deer, where recent research has concluded that
emigration of male red deer from natal areas increased
with female density (Clutton-Brock
et al
. 2002). The
authors incorporated this information into an economic
model to assess optimal culling strategies on Scottish
estates, taking account of deer numbers on neighbouring
estates. They advocated that to maximize economic
returns from hunting stags, estate managers should
reduce female densities to around 50% of the ecological
carrying capacity. This will reduce male emigration and
possibly encourage immigration from neighbouring
populations. However, this does not account for the
fact that stags may move large distances in search of
hinds during the mating season (mating commutes;
sensu
Hogg 2000). As the mating season coincides
with the hunting season in Scotland, stags are likely to
be shot on land holdings (estates) other than those
where they spend most of their lives (Sibbald, Hooper
& Gordon 2001). In bighorn sheep, ram movements
during the rut are affected by both social rank and the
relative availability of ewes in neighbouring populations.
Middle-ranking rams are more likely to move to ewe
groups up to 50 km away in years when there are few
breeding opportunities in their natal population (Hogg
2000). There is clearly a need to understand the role
of both short-term movement (mating commutes)
and long-term dispersal among hunted populations
(McCullouch 1996) and between hunted populations and
protected areas (Hogg 2000), if managers are to more
effectively account for metapopulation responses to
management strategies. Large-scale male movements
during the rut also demonstrate that biological processes
often occur at a much larger scale than that affected
by individual management plans, hence there is great
value in co-operation between different resource managers
to ensure that their management targets are not jeopard-
ized by the activities of others (see below, Managing large
herbivores to meet multiple objectives: the future).
Herbivore impacts on vegetation: local to
landscape
Herbivore distribution and associated impacts on
vegetation are scale-dependent (Senft
et al
. 1987;
Bailey
et al
. 1996; Roguet, Dumont & Prache 1998;
Rietkerk
et al
. 2000). It is, therefore, fundamentally
important to understand the scales of impact driving
vegetation or landscape change in large herbivore
dominated ecosystems. For example, at the landscape
scale heavy grazing may lead to increasing dominance
of grazing-tolerant or unpreferred plant species that may
reduce diversity, whereas at local scales heavy pressure
on preferred vegetation might locally increase diversity
through the provision of new germination niches by
trampling or improved nutrient cycling (Crawley 1997).
Large herbivores generally have extensive ranges, and
therefore their management tends to be focused at a large
scale, which may not be the most appropriate scale of
management for the resources themselves (Palmer
et al
.
2003). Much work has been done to define desirable
densities of different herbivores for particular aims, and
to explore how culling or other management regimes
should be employed to achieve those aims (Welch 1984;
Beaumont
et al
. 1994). Other studies have focused on the
vegetation responses to grazing pressure (utilization
rates) rather than on management through fixed her-
bivore densities, particularly in grass–shrub systems
(Archer 1996; Armstrong
et al
. 1997). Without a full
understanding of what drives herbivores to distribute
themselves across the landscape, however, all these
‘large-scale’ approaches have their limitations.
Ecological research has shown how key resources,
such as vegetation, water and shelter, together with aspects
of herbivore sociability and gregariousness, all drive the
distribution of herbivores and thus their impacts on
resources at a range of scales (Hunter 1962; Kolasa &
Pickett 1991; Milchunas & Lauenroth 1993; Schaefer
& Messier 1995; Bailey, Dumont & Wallis de Vries
1998; Pastor
et al
. 1998; Illius & O’Connor 2000; Apps
et al
. 2001). Much of this research suggests that the
distribution of herbivores is primarily determined by
abiotic factors, such as terrain or distance to shelter/water,
and herbivore responses to vegetation heterogeneity
operate within these higher level constraints (Bailey
et al
. 1996; Tainton, Morris & Hardy 1996; Adler, Raff
& Lauenroth 2001; Landsberg
et al
. 2003). The effects
of vegetation heterogeneity on herbivore distribution
are complex. Herbivores are generally attracted to pre-
ferred vegetation, but the spatial relationship between
preferred and non-preferred vegetation is of para-
mount importance in driving the system dynamics

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I. J. Gordon,
A. J. Hester &
M. Festa-Bianchet
© 2004 British
Ecological Society,
Journal of Applied
Ecology
,
41
,
1021–1031
(Reichman, Benedix & Seastedt 1993; Archer 1996;
Hester
et al
. 1999; Illius & O’Connor 2000; Palmer
et al
. 2003). Recent research has clearly shown that dif-
ferent spatial patterns of vegetation types can change
herbivore behaviour and their concomitant impacts on
the dynamics of the vegetation itself (Clarke, Welch &
Gordon 1995; Wallis de Vries 1996; Illius & O’Connor
2000; Oom
et al
. 2002). For example, Palmer
et al
.
(2003) examined the impacts of red deer on heather
Calluna vulgaris
moorland. As predicted from previous
studies (Clarke, Welch & Gordon 1995; Hester & Baillie
1998), patterns of impact on heather were strongly linked
with its location relative to preferred grass patches
(at a 1-km
2
scale or less), demonstrating that it was
impossible to predict the severity and pattern of heather
utilization from only management-scale (> 100 km
2
)
parameters such as herbivore density and total area
of grassland (Stohlgren, Schell & van den Heuvel
1999; Ryerson & Parmenter 2001). Thus, it appears
as though the impacts of large herbivores on non-
preferred resources are most strongly driven by the
position of these resources relative to preferred res-
ources (Ball, Danell & Sunesson 2000). The success of
large-scale herbivore management to control impacts
on vegetation is unpredictable, because of the weakness
of the relationship between herbivore density and dis-
tribution of foraging impact in large, heterogeneous
areas. These studies demonstrate that, before accurate
predictions can be made about the consequences of dif-
ferent natural resource management scenarios, there
has to be a shift of focus from simple consideration of
the relative abundance of different resources and/or
species (Archer 1996), to a consideration of their spa-
tial distribution within the landscape. Refinements to
management might include ‘artificial’ manipulation of
localized vegetation composition (e.g. through targeted
grazing by domestic stock), with the aim of manipulat-
ing the distribution of other, free-ranging herbivores in
the area (Gordon 1989). For example, if herbivore use
of a vulnerable, highly preferred area of vegetation is
‘unacceptably’ high even after major reductions in
overall herbivore densities, manipulation of the vege-
tation elsewhere could alter herbivore distributions
and consequently their impact (Rea 2003). When linked
to spatially explicit process models (Boone
et al
. 2002),
new technology, such as remote sensing, GPS and
GIS, will be able to provide valuable information at the
appropriate spatial scale for informing management
decisions (Sibbald & Gordon 2001; Danks & Klein
2002; Johnson
et al
. 2002; Stalmans, Witkowski &
Balkwill 2002).
Implications for biodiversity conservation
With increasing concern for biodiversity and conserva-
tion internationally, many countries have now signed
agreements targeted at specific plant or animal species
and habitats designated as of international or national
importance (Department of the Environment 1994;
CITES, http://www.cites.org/). These obligations require
strong, underpinning ecological knowledge upon which
to devise appropriate management regimes to achieve
the agreed targets. However, in many cases these obli-
gations highlighted a widespread lack of understand-
ing of what drives the impacts of free-ranging large
herbivores on biodiversity (Wallis de Vries 1996).
Notwithstanding the fundamentally important effects
of abiotic factors, maximization of vegetation diversity
in a landscape is widely hypothesized to require inter-
mediate levels of herbivory (Grime 1973; Crawley
1997; Olff & Ritchie 1998; Ritchie & Olff 1999; Bullock
et al
. 2001), although definitions of ‘intermediate’ are
not always easy. To integrate the management of large
herbivores with biodiversity/environmental objectives,
the relationships between grazing and biodiversity
must be understood. A recurring problem, however,
has been that many experimental treatments of ecosys-
tems have tended to be either grazing ‘on’ (unknown
grazing species’ contribution to grazing pressure,
unknown herbivore density, unknown seasonality of
grazing) or grazing ‘off’ (using exclosures) (Hester
et al
. 2000). ‘Exclosure’ use as a management tool has
already highlighted the inappropriateness or short-term
nature of the ‘benefits’ of simple removal of herbivores,
rather than manipulation of their densities. But managers
are unlikely to develop management plans based on
varying wild herbivore densities until qualitative theor-
etical hypotheses about desirable herbivore impacts
can be expressed as recommendations for actual densities
under a range of different conditions. This problem is
gradually being redressed (Bullock 1996; Bullock
et al
.
2001) but generally requires complex or costly experi-
mental designs and careful measurement of all key
driving factors in different systems, before widespread
practical generalizations can be made.
One example of the widespread use of grazing
removal, to conserve or expand a vulnerable habitat, as
opposed to the manipulation of herbivore density,
is the case of woodland regeneration in Scotland.
Native woodland and scrub now covers less than 4% of
Scotland’s land area (Mackenzie 1999), yet theoretically
it could cover more than 50% (Towers
et al
. 2004).
Many woodland communities and some of their
associated plant and animal species have now been
designated for protection and/or expansion under the
UK Habitats and Species Directive. Grazing is thought
to have played a major role in woodland decline
across the whole of the UK (Birks 1988; Milne
et al
.
1998), and the impacts of both red deer and sheep in
suppressing regeneration have been clearly demon-
strated (Beaumont
et al
. 1994; Hester, Mitchell & Kirby
1996; Miller, Cummins & Hester 1998). Elsewhere in
Europe, widespread deforestation and suppression
of regeneration by large herbivores is also a major problem
(Humphrey, Gill & Claridge 1998; Hester
et al
. 2004),
with several European countries having designations
to protect and expand such habitats. The knock-on
effects of forest declines of such magnitude on other

1027
Management of
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herbivores
© 2004 British
Ecological Society,
Journal of Applied
Ecology
,
41
,
1021–1031
aspects of biodiversity are relatively poorly understood
(Hester
et al
. 2000). One exclosure study, for example,
found higher densities of invertebrate species in un-
grazed Scottish native pinewoods than in grazed
pinewoods (Baines, Sage & Baines 1994). To date,
where woodland cover is now greatly restricted, fencing
has been widely used as a ‘quick fix’ in protecting
woodland sites and encouraging regeneration. How-
ever, problems associated with fencing include deaths
of protected species such as woodland grouse (e.g.
capercaillie
Tetrao urogallus
), high cost, frequent snow
damage, adverse landscape impact and problems asso-
ciated with higher densities of wild herbivores in the
surrounding areas due to loss of land within their range
(Hester
et al
. 2000).
Managing large herbivores to meet multiple
objectives: the future
The future management of wild large herbivores will
require ecologists to co-operate with sociologists, econ-
omists, politicians and the public. As shown by the
severe decline of the saiga (Milner-Gulland
et al
. 2003)
and the reintroduced Arabian oryx
Oryx leucoryx
(Spalton, Lawrence & Brend 1999), it is irrelevant how
much information from ecological research is provided
to policy makers for the conservation of target species
if other factors are not taken into account. For ex-
ample, if herbivore populations are being decimated
by poaching because of local poverty, lack of law en-
forcement and open illegal trade across national borders,
then these sociological issues must be addressed. If they
are not brought under control, extinction is a real
likelihood for many large herbivore species (Ludwig,
Hilborn & Walters 1993). The management of large
herbivores must undergo a sea-change, where the
ecological understanding of population dynamics and
habitat relationships is linked with socio-economic
studies that address the issues relating to human–wildlife
interactions (du Toit, Walker & Campbell 2004).
In many countries, wildlife (usually referring only to
large mammals) belongs to the state and the right to
hunt wildlife is regulated by a government, as is the
protection of wildlife within national parks and
reserves (Geist 1994). Over the past two decades, par-
ticularly in developing countries, there has been a shift
in government policy towards handing over the right to
use, if not the ownership of, wildlife outside protected
areas to local communities (Harris & Shilai 1997;
Hulme & Murphree 2001; but see Prins, Grootenhuis &
Dolan 2000). This change in approach is derived from
the philosophy that, in order to conserve wildlife out-
side protected areas, local people must derive some
benefit (usually financial) that outweighs the costs of
co-existing with that wildlife (Murphree 1998). Much
research has been concentrated on the benefits of the
community conservation approach to rural economies
and people’s attitude to wildlife (Hulme & Murphree
2001). Ecological research is now required to help local
communities determine the most effective ways of
managing wildlife in their area (Féron
et al
. 1998; De
Garine & de Garine-Wichatitsky 1999) and to provide
cost-effective means of population assessment. Costly
aerial surveys (Jolly 1969; du Toit 2002; Jachmann 2002)
are inappropriate in this context but low-cost methods
could be employed, such as dung surveys (Laing
et al
.
2003), bicycle surveys (Gaidet, Fritz & Nyahuma
2003) or by using information from culls and harvests
(e.g. age structure and fecundity rates). Over time, indi-
ces of population responses to environmental condi-
tions and management decisions should become
available. While there are arguments as to the value of
this approach to wildlife conservation (Hulme &
Murphree 2001; du Toit, Walker & Campbell 2004), the
philosophy is still valid and should be adopted in
developed countries, where local communities could be
encouraged to have a more positive attitude towards
wildlife in their local area.
The home range of wild herbivores often extends
over land held under more than one ownership, pro-
viding an additional challenge to management. To
date, the management of ungulate populations has
tended to focus on the management of single popula-
tions within defined management units, for example
estates, community lands and nature reserves/national
parks. Ecological research needs to develop a more
detailed understanding of the interactions between
subpopulations (e.g. immigration and emigration rates)
and the consequences of management of one popula-
tion on neighbouring populations, especially where the
management units have very different goals (e.g. hunt-
ing and conservation). The outcomes of this research
should be coupled with policy and management instru-
ments that facilitate the co-operative management of
large herbivore populations. Management units must
more closely reflect the biology of the populations rather
than the human-defined ownership and jurisdictional
boundaries. In many respects, the developing world is
leading the way in approaches to public involvement in
management and co-operative management, and the
developed world would do well to learn from the relative
success or otherwise of different attempts to manage
large herbivore populations more holistically, as part
of a socio-ecological system rather than in isolation.
Acknowledgements
Thanks go to Jean-Michel Gaillard, Glenn Iason, Norman
Owen Smith, Robin Pakeman and an anonymous referee
for their valuable comments on the manuscript. I.J.G.
and A.J.H. acknowledge the support of the Scottish
Executive Environment and Rural Affairs Department.
In addition, I.J.G. was supported by CSIRO.
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