Biol. Rev. (2012), 87, pp. 72–94.
doi: 10.1111/j.1469-185X.2011.00185.x
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Carbon storage in terrestrial ecosystems: do
browsing and grazing herbivores matter?
Andrew J. Tanentzap∗ and David A. Coomes
Forest Ecology and Conservation Group, Department of Plant Sciences, University of Cambridge, Cambridge CB2 3EA, UK
ABSTRACT
Large mammalian herbivores manifest a strong top-down control on ecosystems that can transform entire landscapes,
but their impacts have not been reviewed in the context of terrestrial carbon storage. Here, we evaluate the effects of
plant biomass consumption by large mammalian herbivores (>10 kg adult biomass), and the responses of ecosystems
to these herbivores, on carbon stocks in temperate and tropical regions, and the Arctic. We calculate the difference
in carbon stocks resulting from herbivore exclusion using the results of 108 studies from 52 vegetation types. Our
estimates suggest that herbivores can reduce terrestrial above- and below-ground carbon stocks across vegetation types
but reductions in carbon stocks may approach zero given sufficient periods of time for systems to respond to herbivory
(i.e. decades). We estimate that if all large herbivores were removed from the vegetation types sampled in our review,
increases in terrestrial carbon stocks would be up to three orders of magnitude less than many of the natural and
human-influenced sources of carbon emissions. However, we lack estimates for the effects of herbivores on below-ground
biomass and soil carbon levels in many regions, including those with high herbivore densities, and upwards revisions of
our estimates may be necessary. Our results provide a starting point for a discussion on the magnitude of the effects
of herbivory on the global carbon cycle, particularly given that large herbivores are common in many ecosystems. We
suggest that herbivore removal might represent an important strategy towards increasing terrestrial carbon stocks at
local and regional scales within specific vegetation types, since humans influence populations of most large mammals.
Key words: carbon sinks, disturbance, global, herbivory, productivity, primary consumers.
CONTENTS
I. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
II. Potential effects of herbivores on carbon stocks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
(1) Slow and non-linear responses of above-ground carbon stocks in woody vegetation . . . . . . . . . . . . . . . . . . .
(2) Effects of consumption by livestock on above-ground carbon stocks in pastures . . . . . . . . . . . . . . . . . . . . . . .
(3) Ecosystem responses to herbivores and implications for below-ground carbon stocks . . . . . . . . . . . . . . . . . .
(4) Confounding influences on the responses of terrestrial carbon stocks to herbivores . . . . . . . . . . . . . . . . . . . .
III. Estimating the effects of herbivores on carbon stocks using exclosure studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
(1) Temperate forests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
(2) Tropical forests . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
(3) Temperate grasslands and shrublands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
(4) Tropical grasslands and shrublands . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
(5) Artificial pastures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
(6) Other vegetation types: wetlands and Arctic tundra . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
IV. Synthesizing the annual effects of herbivory on carbon stocks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
(1) Effects of herbivores on terrestrial carbon stocks per unit land area . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
(2) Extrapolating the effects of herbivores on terrestrial carbon stocks to landscape
and global scales . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
(3) Temporal effects of herbivores on terrestrial carbon stocks . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
(4) Influence of herbivore body size and feeding strategy on terrestrial carbon stocks . . . . . . . . . . . . . . . . . . . . .
(5) Priorities for future research . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
* Address for correspondence (Tel: +01223 764736; Fax: +01223 333593; E-mail: [email protected]) .
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
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Carbon storage in terrestrial ecosystems: do browsing and grazing herbivores matter?
V.
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VII.
VIII.
IX.
Implications of herbivore removal for terrestrial carbon cycling and conservation . . . . . . . . . . . . . . . . . . . . . . . . .
Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Supporting Information . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
I. INTRODUCTION
Consumption of vegetation by mammalian herbivores
influences terrestrial carbon cycling, yet ecologists have
only recently begun to quantify this process (e.g. Feeley
& Terborgh, 2005; Derner, Boutton & Briske, 2006; Mills &
Cowling, 2006; Gill, 2007; Holdo et al., 2009). An unresolved
question is whether the influences of herbivory on carbon
cycling lead to significant changes in carbon stocks relative
to other factors commonly considered in the global carbon
cycle, such as deforestation. The ubiquity of herbivores, and
the capacity of some populations to demonstrate irruptive
population growth that can have correspondingly large
impacts on vegetation (Leopold, Sowls & Spencer, 1947;
Caughley, 1970; Sinclair, 1979; Forsyth & Caley, 2006),
make clarifying the linkages between herbivory and carbon
stocks an important pre-requisite for understanding the
global carbon cycle.
Herbivores are a top-down control over landscapelevel vegetation patterns (Bond, 2005), which can lead to
rapid ecosystem state changes (Holling, 1973; May, 1977;
Crawley, 1983; Huntly, 1991; Hobbs, 1996) that alter
carbon stocks (Grace, 2004). For example, large grazers
maintain grasslands or shrublands in regions where soils and
climates favour forests (Bond, 2005), and these systems have
lower above-, and potentially below-ground, carbon stocks
than forests (Grace, 2004). By contrast, ‘natural’ grasslands
and savannas are often fire-regenerated systems, in which
herbivores reduce fire frequency by reducing standing litter
that acts as a fuel source (Briggs et al., 2005; Hill et al.,
2005; Holdo et al., 2009). Declines in the incidence of
fire allow woody vegetation to encroach, and can reduce
the mortality of any established trees (Briggs et al., 2005),
increasing carbon stocks in both plant biomass and soil
organic matter (Holdo et al., 2009; though increases in
soil carbon may depend on grass persistence, Hill et al.,
2005; Coetsee, Bond & February, 2010). Primarily, the
effects of herbivores on carbon stocks in stable, established
plant communities may be small and depend on primary
productivity, herbivore consumption rates, evolutionary
histories, and climatic conditions (Milchunas & Laurenroth,
1993). Herbivory reduces growth, survival, and fitness of most
plants that are grazed or browsed, reducing above-ground
carbon stocks. In some very limited instances, low levels
of herbivory may directly benefit plants that are subjected
to herbivory by stimulating regrowth that produces greater
amounts of biomass than initially removed by herbivores
(Dyer, 1980; Owen & Wiegert, 1981; Bergman, 2002;
Rooke, 2003). Changes in species composition are also
influenced by other effects of herbivory at the ecosystem
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level, including alteration of the physical structure of habitats
and modification of macronutrient cycling (Bazely & Jefferies,
1985; Pastor et al., 1993; Bardgett & Wardle, 2003; Wolf,
Cooper & Hobbs, 2007). Variation in the responses of
plants and ecosystems to herbivores, and selective feeding by
herbivores, make it difficult to predict the community-level
effects of herbivory (Crawley, 1983; Huntly, 1991).
Human exploitation of large mammals should control
vegetation patterns and thereby influence terrestrial carbon
cycling. For example, excessive hunting by European settlers
led to the near extinctions of bison (Bison bison) in North
America and elephant (Loxodonta africana), white rhinoceros
(Ceratotherium simum), and hippopotamus (Hippopotamus
amphibious) in Africa, increasing shrub encroachment in
these regions (Owen-Smith, 1987). Reductions in elephant
populations have in fact been suggested as a possible
factor contributing to recent increases in biomass and
carbon sequestration in African tropical forests (Lewis et al.,
2009). Pre-historic and early-historic human exploitation
of deer (Cervidae), auroch (Bos taurus primigenius), tarpan
(Equus ferus ferus), and wisent (Bison bonasus) populations
may have also reduced herbivory, and consequently the
extent of open woody-grassland mosaics in Europe (Vera,
2002); in Australia, the role of Quaternary megafauna in
maintaining savanna conditions was likely influenced by
colonizing human populations (Burney & Flannery, 2005;
Miller et al., 2005). By contrast, human removal of carnivores
may be a far more important factor presently affecting
many vegetation communities by reducing predation of
large herbivores; e.g. extirpation of wolves (Canis lupus),
cougars (Puma concolor) and bears (Ursus spp.) in western North
America has contributed to increased deer populations and
thus reduced the abundance of woody plants (Beschta &
Ripple, 2009). Many animal distributions, and consequently
their impacts, are also a function of human modifications
to the distribution of forage and water across landscapes,
e.g. forest fragmentation increases favourable edge habitat
for white-tailed deer (Odocoileus virginianus) in eastern North
America (Côté et al., 2004) and elephants gravitate towards
artificial watering holes in Africa (Shannon et al., 2008).
Human-introduced diseases also strongly regulate herbivore
populations and consequently vegetation communities, e.g.
rinderpest virus has been linked to declines in populations of
large mammalian herbivores in Africa, resulting in increased
shrub establishment and cover (Prins & van der Jeugd, 1993).
Collectively, changes in many local or regional herbivore
populations can have a large impact on global vegetation.
Livestock grazing is perhaps the most widespread form of
herbivore management, and regional increases in animal
densities to meet increasing consumption (Asner et al., 2004;
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
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Green et al., 2005) can significantly alter vegetation across the
28–35 million km2 globally managed as permanent pastures
(FAO, 2008; Ramankutty et al., 2008). Although the role
humans play in manipulating herbivory is considered when
managing ecosystems (Hobbs & Norton, 1996), the means
by which herbivores may alter levels of terrestrial carbon
cycling is often overlooked.
In this paper, we quantify how large mammalian herbivores (>10 kg adult biomass) influence terrestrial carbon
stocks by reviewing studies where herbivores have been
excluded from browsing and grazing. Understanding how
herbivores influence terrestrial carbon stocks may be critical
for predicting the effects of anthropogenic activities on carbon cycling. We use the herbivore exclosure studies in our
review to estimate the impacts of herbivores on above- and
below-ground carbon stocks. We extrapolate our estimates
to the landscape and global level, and place our findings in
the context of anthropogenic influences on the global carbon
cycle, including deforestation, fire, and fossil fuel emissions.
Given that existing data are far from ideal, the purpose of
this exercise is not to generate precise estimates of how the
intensity of herbivory can affect terrestrial carbon stocks.
Rather, we wish to stimulate discussion as to the magnitude
of these changes and identify where gaps in our knowledge
currently exist. In order to be able to develop predictive
models of how carbon stocks might change over time in
response to herbivores or to different densities of herbivores,
we test whether the effects of herbivores on carbon stocks
vary over time and with characteristics of the herbivores
themselves. Lastly, we discuss how changes in herbivore
distributions and populations may affect our estimates and
the implications of herbivory for management practices that
aim to increase terrestrial carbon stocks. We acknowledge
the potential importance of small mammal and invertebrate
herbivores in reducing above-ground carbon stocks, particularly since their consumption of net primary productivity
(NPP) during population outbreaks may equal that of large
herbivores (Brown & Heske, 1990; Kessing, 2000; Howe
et al., 2006; Lovett et al., 2006; Peltzer et al., 2010). However,
we restrict our study to large mammalian herbivores since
populations of these species may consistently consume higher
quantities of NPP (Belovsky, 1997) and therefore have larger
effects on terrestrial carbon stocks.
II. POTENTIAL EFFECTS OF HERBIVORES
ON CARBON STOCKS
Herbivores influence multiple carbon stocks and fluxes in terrestrial ecosystems (Fig. 1). Ideally, all the processes involved
in carbon cycling need to be quantified in order to predict the
effects of herbivores on carbon sequestration—the overall
process by which carbon is removed from the atmosphere
and stored in alternative pools of matter. While many of
the ways in which herbivores interact with carbon cycles are
still to be entirely resolved, making our conclusions tentative,
we emphasize that this should not lead to an automatic
Andrew J. Tanentzap and David A. Coomes
Fig. 1. Effects of herbivores on carbon cycling in terrestrial
ecosystems. Lines denote fluxes that are predicted to either
increase (black lines) or decrease (grey lines) in response to
herbivory. A carbon stock represents the quantity of carbon
held within a pool of matter at a given time, while carbon fluxes
transfer carbon from one pool to another. The flow of carbon
through the Earth’s atmosphere, and aquatic and terrestrial
ecosystems comprises the global carbon cycle.
acceptance of the hypothesis that herbivores do not affect
carbon stocks. Rather, we develop a conceptual framework
in which to predict the impacts of herbivores on terrestrial
carbon stocks over decadal and century timescales by drawing out generalities from studies of the effects of herbivores
on carbon cycling at the community or ecosystem level,
and highlighting the circumstances under which variation
in responses might arise. A naïve hypothesis is that herbivores reduce terrestrial carbon stocks, because they consume
above-ground biomass, which in turn reduces below-ground
biomass (given that above- and below-ground biomass are
closely correlated; Litton, Raich & Ryan, 2007). Removal of
below-ground biomass should also reduce soil carbon levels
(e.g. Hungate et al., 1997; Litton et al., 2007). However, the
influences of herbivores on carbon cycling are often more
complicated than portrayed by this naïve hypothesis, because
as ecosystems respond to herbivores, other processes that are
indirectly related to herbivory will concurrently change. For
example, reductions in herbivory lead to the invasion of
woody vegetation onto temperate peatlands, and whilst this
increases above-ground carbon stocks, it also lowers the
water table and increases peat aeration, leading to the mobilization and emission of large pools of otherwise stable and
recalcitrant carbon (Ise et al., 2008). Other indirect influences
of herbivores include fertilization and trampling, and taken
together with the evolved responses of plants themselves to
herbivores, these indirect effects can offset losses of carbon
assumed to occur by our naïve hypothesis. We review the
direct and indirect effects of herbivores on above-ground
carbon stocks in woody vegetation and pastures, before
considering their impacts on below-ground carbon stocks.
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
Carbon storage in terrestrial ecosystems: do browsing and grazing herbivores matter?
(1) Slow and non-linear responses of above-ground
carbon stocks in woody vegetation
Herbivores have a direct effect on carbon stocks in woodlands
by consuming biomass and litter (Fig. 1). In forests and
tall shrublands (canopy height 2–8 m), several orders of
magnitude more carbon is stored within adult trees that lie
beyond the reach of herbivores than within high densities of
seedlings and saplings (Coomes et al., 2002; Luyssaert et al.,
2008). Reductions in the numbers of small trees due to
herbivores subsequently require long periods of time (e.g.
decades or centuries) to permeate through the forest size
structure and significantly affect above-ground carbon stocks
(Frelich & Lorimer, 1985; McInnes et al., 1992; Bellingham
& Allan, 2003; Mountford & Peterken, 2003). Debarking
by deer and uprooting by elephants are two mechanisms
by which herbivores can directly and rapidly reduce woody
biomass but these behaviours rarely lead to significantly
greater mortality in large-sized trees [e.g. >10 cm diameter
at breast height (dbh); Höft & Höft (1995); Akashi &
Nakashizuka (1999); Calenge et al. (2002)]. Soil carbon stocks
will also decline with reductions in the input of litterfall
associated with reductions in tree density and basal area
(e.g. Shutou & Nakane, 2004; Jonard, Misson & Ponette,
2006; Kunhamu, Kumar & Viswanath, 2009), but since
the majority of carbon in forest soils may be stored within
recalcitrant pools of organic matter, soil carbon stocks may
be much more closely related to microbial decomposition
rates than above-ground biomass (Jandl et al., 2007). Relative
changes in carbon stocks within forest soils may also be
smaller than changes in above-ground biomass (Barford et al.,
2001; Schlesinger & Lichter, 2001; Houghton & Goodale,
2004; Woodbury, Heath & Smith, 2006), and therefore, the
absolute effects of herbivores on soil carbon stocks may be
small.
Above-ground carbon stocks may be unaffected by
herbivores over a large range of herbivore densities but
decline rapidly at high density. Consider a patch model
of forest dynamics where a forest is comprised of stands
of different ages since disturbance, and where each stand
functions as an even-aged population (Watt, 1947; Coomes
& Allen, 2007). We predict a small effect of herbivores
on carbon stocks in the older patches because most of the
carbon is stored within large trees above the browse layer.
However, in younger patches, we predict that herbivores
will interact with competitive thinning. As a result, the basal
area and biomass of individual trees will increase with little
accompanying decline in tree density until space becomes
limiting, at which point, further increases in basal area will
lead to declines in tree density. Removal of a large number
of seedlings or saplings by herbivores may not prevent tree
regeneration, but simply delay individuals from reaching
larger sizes and a state of resource competition. Browsetolerant and less herbivore-palatable species may also replace
species most affected by herbivores with little eventual effect
on total basal area and biomass (e.g. Weber, Rigling &
Bugmann, 2008), although the species that are most resilient
to herbivory are often slow growing so herbivores might
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delay the accumulation of carbon stocks (Coley, Bryant &
Chapin, 1985; Herms & Mattson, 1992; Bee, Kunstler &
Coomes, 2007). Since only one of many seedlings is required
to replace an adult tree, herbivory needs to be consistently
sustained at very high levels to ensure that no seedlings and
saplings escape from the height tier in which browsing occurs.
Thus, herbivores can have little or no effect on above-ground
carbon stocks in forests.
(2) Effects of consumption by livestock
on above-ground carbon stocks in pastures
Pastures are well adapted to mammalian herbivores because
the meristems of many grassland species are situated at the
base of plants, allowing rapid recovery following removal of
foliage (Hawkes & Sullivan, 2001). However, there has been
considerable debate as to whether individual herbaceous and
woody plants can re-grow to an equal or greater biomass
(termed ‘overcompensation’) than their pre-defoliation levels
(Belsky, 1986; Paige & Whitham, 1987; Belsky et al., 1993;
Haukioja & Koricheva, 2000), and whether these individuallevel responses extend across entire communities (Belsky,
1987). Plants that are adapted to high-intensity grazing
regimes may be more capable of recovering rapidly from
herbivory because of grazing-associated increases in soil
nitrogen availability that promote growth (Holland et al.,
1992), herbivore consumption of litterfall that shades young
plants (Belsky, 1986), and/or reductions in the competitive
ability of neighbours (Belsky et al., 1993). One limitation in
applying theories of overcompensatory growth to carbon
stock calculations is that the process has largely been
demonstrated within controlled laboratory conditions or for
agricultural systems, which are less influenced by variation
in light, water, and nutrient availability than natural systems
(Belsky, 1987; Leriche et al., 2001, c.f. McNaughton, 1985;
Frank, Kuns & Guido, 2002).
Pastures are often found in regions that would have
once supported woodland, and hence, reducing rates of
biomass consumption by livestock may result in woody
encroachment (Westoby, Walker & Noy-Meir, 1989; Friedel,
1991; Laycock, 1991; Asner et al., 2004). Reversion to
woodland increases standing biomass and above-ground
carbon stocks and succession of pastures to woodland may
lead to greater gains in carbon stocks than those associated
with the removal of herbivores from native grassland
vegetation. Abandonment of woodland-pasture systems in
Europe and pasture-cropland in the eastern U.S. has led
to subsequent afforestation and woody plant encroachment
that has increased terrestrial carbon sequestration in aboveground vegetation (Houghton, Hackler & Lawrence, 1999;
Caspersen et al., 2000; Pacala et al., 2001; Saikku, Rautiainen
& Kauppi, 2008; Rhemtulla, Mladenoff & Clayton, 2009).
Although land-use changes and deforestation often lead to
changes in herbivore densities themselves, these relationships
are rarely considered in carbon flux estimates (Houghton
et al., 1983; McGuire et al., 2001). This poses the question to
what extent can increases in carbon stocks be derived from
the removal of herbivores from pastures?
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
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Increased carbon stocks in above-ground vegetation
associated with pasture abandonment may be confounded
by the role of fire, which can often be a larger consumer
of primary production and above-ground biomass than
herbivores (Bond & Keeley, 2005). Burning of biomass
across tropical savanna is thought to emit approximately
1.65 Pg carbon (Pg C) year−1 compared to 0.19 and 1.26 Pg
C year−1 emitted from biomass burning in temperate and
tropical forests, respectively (Mouillot et al., 2006). However
not all carbon is completely oxidized by fires. Recent evidence
has suggested that charcoal derived from high-intensity fires
may comprise highly recalcitrant carbon that can be stored
in soils for hundreds or thousands of years, far exceeding
the lifetime of most grassland or savanna species (Lehmann,
Gaunt & Rondon, 2006), and sequester 0.05–0.27 Pg C
year−1 (Forbes, Raison & Skjemstad, 2006).
(3) Ecosystem responses to herbivores
and implications for below-ground carbon stocks
The majority of carbon in terrestrial ecosystems is stored
below ground (Houghton, 2007), so understanding how
herbivores affect fluxes of carbon from plants into belowground systems, and from below-ground systems to the
atmosphere and hydrosphere, is crucial to predicting the
size of terrestrial carbon stocks. Carbon enters soils from
the products of photosynthesis (including herbivore excrement), and leaves through respiration, and erosion or
leaching of soil organic matter. Consumption of aboveground biomass by herbivores alters the amount of carbon
within soils across entire ecosystems by: (1) reducing plant
litter quantity and quality; (2) altering rates of soil respiration, and (3) changing mineral nutrient cycling through waste
products. Here, we review the processes through which herbivores alter soil carbon stocks and address the difficulties
in generalizing how below-ground carbon stocks respond to
herbivory.
Plant litter is the primary input of carbon to soils
(Houghton, 2007), so changes in the quantity and quality of
litter arising from herbivory can have significant influences
on below-ground carbon stocks. Herbivores can reduce litter
quantity, and consequently soil carbon stocks, by decreasing
plant cover and shifting community composition towards
species that produce less litterfall (e.g. Pastor et al., 1993;
Wardle et al., 2001; Fornara & Du Toit, 2008). However,
herbivores can have variable effects on litter quality, and
therefore on soil carbon. Soil carbon stocks can increase
if plant composition shifts towards species with litter of
low quality to detritivores. Plants that are less-preferred
by herbivores often produce lignified leaves that are high
in secondary metabolites and decompose more slowly
(Horner, Gosz & Cates, 1988; Wardle, Bonner & Baker,
2002; Cornelissen et al., 2004). By contrast, carbon can
be allocated to regrowth rather than the production of
secondary compounds, particularly on nutrient-rich soils
(Bryant & Reichart, 1992; Herms & Mattson, 1992),
and this can increase foliar nutrient concentrations and
rates of litter decomposition (Olofsson & Oksanen, 2002;
Andrew J. Tanentzap and David A. Coomes
Chapman et al., 2003). Litter quality can also be altered by
structural changes that increase plant tolerance to herbivory;
many grassland species increase the ratio of blade to
sheath biomass when heavily grazed (McNaughton, 1984;
Coughenour, McNaughton & Wallace, 1985; Jaramillo &
Detling, 1988), leading to better quality litter that is rapidly
decomposed (Semmartin & Ghersa, 2006). The responses
of plants to herbivory also differ between evergreen and
deciduous species, with deciduous shrubs in boreal forests
and tropical savannas allocating carbon towards regrowth,
whilst evergreen shrubs in these habitats increase foliar
carbohydrate or secondary metabolite concentrations (Stock,
Roux & Heyden, 1993; Tolvanen & Laine, 1997). Soil
nutrient levels are likely to be an over-riding control over
decomposition rates and litter quality, and increases in soil
nitrogen concentrations due to herbivore excrement may
enhance the decomposition of poor-quality litter (Olofsson
& Oksanen, 2002; Fornara & Du Toit, 2008). For all these
reasons, it is difficult to predict the effects of herbivores on
litter quality and soil carbon stocks.
Herbivores influence soil respiration, and the activity and
biomass of soil microbes, within ecosystems over both short(e.g. daily) and long-term intervals (e.g. decades). Field studies
have suggested that the direction of the relationship between
above-ground herbivory and both microbial biomass and
respiration varies among habitats (Kieft, 1994; Wardle
et al., 2001; Virtanen, Salminen, & Strömmer, 2008). One
explanation for this variation over the short term is that
greater root exudation, a mechanism for transferring carbon
from individual plants into soil that increases soil microbial
activity and biomass (Bardgett, Wardle & Yeates, 1998), is
common among species adapted to high levels of herbivory
(Frank & Groffman, 1998; Ayres et al., 2004; Frost &
Hunter, 2004) and depends on above-ground carbon storage
(Holland, Cheng & Crossley, 1996). Over longer periods, soil
fertility may play an important role in mediating shifts in plant
community composition due to herbivory, which determine
the quality of carbon inputs to soils, and consequently,
the direction of the responses of microbial communities
(Sankaran & Augustine, 2004). Herbivores also influence
rates of soil microbial activity and respiration through
altering soil temperature and moisture content (Bardgett
& Wardle, 2003; Gornall et al., 2007), and this effect varies
with air temperature (Sjögersten, van der Wal & Woodin,
2008), seasonal precipitation (Classen et al., 2007), and the
seasonal timing of herbivory relative to plant production
(Stark & Grellmann, 2002). Finally, herbivores can alter soil
respiration by accelerating soil erosion and leaching through
reducing plant and standing litter cover, which increases
the exposure of the soil surface to precipitation (Wood &
Blackburn, 1981). Soil erosion can be increased further
by soil compaction, which reduces porosity and increases
surface runoff of rainwater (Wood & Blackburn, 1981).
However, losses of carbon through erosion and leaching are
typically marginal (e.g. 0.01 t C ha−1 year−1 ; Meeuwig,
1965), except when considered over extremely long time
periods (e.g. millennia; Trumbore, 2006) or in riparian plant
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
Carbon storage in terrestrial ecosystems: do browsing and grazing herbivores matter?
communities subject to fluvial erosion (Beschta & Ripple,
2008).
Herbivore excrement influences changes in nutrient
cycling and microbial activity (Bardgett & Wardle, 2003),
but excrement itself returns little carbon to soils compared
with the biomass consumed by herbivores (Pastor et al.,
1993; Butler & Kielland, 2008; Fornara & Du Toit,
2008). For nitrogen cycling, excrement has inconsistent
effects because of spatial variation in herbivore densities
and their consumption relative to site productivity (Singer
& Schoenecker, 2003). For example, the stimulation of
microbial activity by herbivore waste products accelerates
nitrogen cycling in some systems (McNaughton et al., 1989;
McNaughton, Banyikwa & McNaughton, 1997). Bison and
elk (Cervus canadensis) in Yellowstone National Park increase
mineralization of soil nitrogen (Hamilton & Frank, 2001),
leading to greater above-ground productivity and biomass
of grasslands (Frank & McNaughton, 1993). Nitrogen
deposition from livestock waste products across the Río
de la Plata grasslands of South America increases nitrogen
volatilization and leaching from soil, leading to slower litter
decomposition and a decline in soil organic carbon (Piñeiro,
Paruelo & Oesterheld, 2006). By contrast, the predominant
effect of moose (Alces alces) in boreal forests is to shift plant
community composition towards unpalatable conifers that
return low amounts of nitrogen to soil via litter (Pastor et al.,
1993). This effect is far larger than the stimulation of nitrogen
mineralization arising from excrement (Pastor et al., 1993). A
positive feedback cycle suppresses the recovery of browsed,
palatable species, resulting in the dominance of conifers with
long-lived, nitrogen-poor foliage and slowly decomposing
litter (Pastor et al., 1993).
(4) Confounding influences on the responses
of terrestrial carbon stocks to herbivores
Carbon stocks may covary with other ecological processes,
in addition to herbivory. For example, enhanced growth
and recovery of woody species following herbivore removal
might be difficult to separate from the effects of atmospheric
nitrogen deposition in some systems. Nitrogen deposition
increases carbon sequestration in temperate and boreal
forests independent of herbivore removal (Magnani et al.,
2007), and large-scale plot networks have attributed increases
in carbon stocks in Europe and the United States to this
process (de Vries et al., 2006; Thomas et al., 2009). However,
other studies have found little evidence for increased
carbon sequestration associated with atmospheric nitrogen
deposition (e.g. Nadelhoffer, Emmett & Gundersen, 1999;
McMahon, Parker & Miller, 2010), and this may arise
in forests that are not nitrogen-limited and/or because of
increased soil acidity that offsets fertilization benefits (de
Vries, 2009). In tundra and peatlands, herbivore removal
may increase moss cover (Olofsson et al., 2001; Ward et al.,
2007; Susiluoto et al., 2008), but this process may be offset
by nitrogen deposition that shifts vegetation communities
towards vascular plants, reducing peat formation and carbon
sequestration (Berendse et al., 2001). Human disturbances
77
such as fire suppression can also account for increased woody
plant abundance following herbivore removal (e.g. Nowacki
& Abrams, 2008). Increases in forest cover may also be
partly determined by the extent of historical felling in order
to clear land for livestock (Hanley et al., 2008), and thus reflect
land-use changes associated with herbivory rather than the
effects of herbivores themselves. Finally, trophic cascades
initiated by high levels of herbivory alter nutrient cycling
and hence continue to affect vegetation composition even
after herbivores are reduced or removed (Bazely & Jefferies,
1985; Pastor et al., 1993).
In summary, herbivores should lead to small declines in terrestrial carbon stocks per unit land area, but many exceptions
exist to this hypothesis. For example, reductions in aboveground carbon stocks will be small in forests and pastures,
except at very high herbivore densities or where pastures
were derived from woodland. The direction of the responses
of below-ground carbon stocks are also difficult to predict
since they vary with the characteristics of both vegetation and
herbivores, and with different abiotic factors. Additionally,
other ecological processes covary with the effects of herbivores on carbon stocks, and so herbivore removal might not
lead to predicted increases in carbon stocks.
III. ESTIMATING THE EFFECTS
OF HERBIVORES ON CARBON STOCKS
USING EXCLOSURE STUDIES
We review 108 studies that compare vegetation changes
following the exclusion of herbivores from 52 different vegetation types within six over-arching vegetation classifications
(Table 1): (1) cool temperate and boreal forest; (2) tropical
and warm temperate forest; (3) temperate grasslands and
shrublands; (4) tropical and subtropical grasslands and shrublands; (5) artificial pastures; and (6) other vegetation types:
wetlands and Arctic tundra. We include extensive pastures within grasslands and shrublands. Extensive pastures
(= rangelands) have never been sown or ploughed and are
usually regarded as natural or semi-natural vegetation. These
contrast with artificial pastures (= intensively managed systems) that occur where vegetation productivity has been
improved through the replacement of native vegetation with
high-productivity grassland and/or application of fertilizers.
A human population density of 20 individuals km−2 has been
used as a threshold for broadly distinguishing the intensity of
land cultivation (Kruska et al., 2003). Under this definition,
extensive pastures occur where human population densities
are <20 individuals km−2 compared with artificial pastures that support, or are derived from, human populations
comprising >20 individuals km−2 (Reid, Galvin & Kruska,
2008).
We estimate the effects of herbivores on annual aboveground, and in some cases, above- and below-ground carbon
stocks for each study as the difference in carbon stocks
between plots with and without herbivores divided by the
study duration. Many of these studies were not originally
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
Andrew J. Tanentzap and David A. Coomes
78
Table 1. Effects of large mammalian herbivores on terrestrial carbon stocks. Mean estimates of above- (AGB) and below-ground
(BGB) carbon stocks in plant biomass and soils (1 m depth), above- and below-ground net primary productivity (NPP), and area
affected by herbivores, were derived from the literature (refer to footnotes). We used the mean change in above- and below-ground
carbon stocks in plant biomass and soils upon herbivore exclusion from each study reviewed in Tables S1 and S2 (see online
supporting information) to estimate the annual effect of herbivore removal on carbon stocks across the global extent of eight
vegetation types ± one standard deviation. Standard deviations in each vegetation type were calculated as the square-root of the
sum of squared standard deviations for the effects of herbivores in each carbon pool. For grasslands and shrublands, we separately
estimated effects due to primarily undomesticated herbivores versus livestock in extensive pastures
C stocks (t C ha−1 )a
NPP (t C ha−1 year−1 )a
Vegetation type
AGB
BGB
Litter
Soil
AGB
BGB
Area (106 km2 )b
Global C (Mt C year−1 )c
Boreal forests
Temperate forests
Tropical forests
Temperate grasslands and
shrublands
30.5
105.0
152.4
5.8
11.0
28.5
42.0
6.8
15.9
13.2
5.8
2.1
295.5
121.5
122.5
167.5
1.2
4.8
7.0
0.7
0.8
3.0
5.5
1.3
6.63
6.21
8.94
0.04 ± 0.04
0.04 ± 0.06
0.02 ± 0.01
6.30
3.92
0.03 ± 0.04
0.01 ± 0.02
103.5
Undomesticated herbivores
Livestock
2.7
2.7
Undomesticated herbivores
Livestock
0.4
0.5
12.3
4.5
11.54
4.50
3.17
9.15
8.67
Tropical grasslands and
shrublands
Tundra
Wetlands
Artificial (intensive)
pastures
20.0
1.3
8.5
3.6
2.0
3.5
43.0
12.6
124.3
348
643.0
0.10 ± 0.03
0.09 ± 0.11
0.01 ± < 0.001
0.08 ± 0.10
0.09 ± 0.05
a All above- and below-ground data from Saugier et
al. (2001), except for wetland AGB stocks (WBGU, 1988) and NPP (Atjay et al., 1979) and
tundra soils (Ping et al., 2008), and NPP of artificial pastures (Ellis & Ramankutty, 2008). Estimates from Saugier et al. (2001) contrast with
those of Huston & Wolverton (2009), which reported similar NPP of AGB in temperate and tropical forests: 4.7 and 5.3 t C ha−1 year−1 ,
respectively. Soil carbon stocks represent average of WBGU (1988) and Carter & Scholes (2000), except for wetlands (only WBGU, 1988).
Litter values derived from means of Matthews (1997) and Atjay et al. (1979), except for wetlands and temperate and tropical grasslands and
shrublands (Atjay et al., 1979). For artificial pastures, we used a map of global above- and below-ground carbon stocks (Kapos et al., 2008)
and superimposed this upon 11 regions with human population densities ≥20 individuals km−2 derived by Ellis & Ramankutty (2008). We
then averaged the mean carbon stock per region after weighting values by the proportion of area within each region occupied by pastures
(Ellis & Ramankutty, 2008). BGB stocks from Saugier et al. (2001) may be underestimated (Robinson, 2007).
b
After excluding anthropogenic land uses (Ellis & Ramankutty, 2008), except for wetlands (mean of Lehner & Döll, 2004). For grasslands
and shrublands, we separately report the area covered by natural land cover (undomesticated herbivores) and extensive pastures (livestock),
estimated from total area of pastures in areas with human population densities <20 individuals km−2 using values from Ellis & Ramankutty
(2008). Artificial (intensive) pastures were estimated as the total area of pastures in all vegetation types where human population densities
≥20 individuals km−2 (Ellis & Ramankutty, 2008).
c
Estimated by multiplying mean change in carbon stocks upon herbivore exclusion (t C ha−1 year−1 ; Fig. 2), summed across carbon pools,
by the area of each vegetation type (ha).
intended to be used for carbon estimation, and we estimated
carbon stocks from measurements of plant biomass and litter, assuming a 50% carbon content, consistent with other
recent analyses (e.g. Lewis et al., 2009). Where studies measured the impacts of herbivores on vegetation other than
in the form of biomass, e.g. basal area and/or density of
stems, we applied published regression equations to estimate above-ground biomass and then carbon stocks (see
online supporting information: (1) Supporting methods, for
examples). If studies reported only one component of a
system, e.g. saplings, we produced estimates for only these
components. We also included studies that measured soil
carbon, and where this was reported as a percentage rather
than mass, we used measurements of bulk density reported
in the study and the study sampling depth to convert the
per cent carbon concentration of soils to a mass per unit
land area. The effects of herbivores on biomass, litter, and
soil carbon are jointly reviewed throughout the text, and
summarized in Tables S1 and S2 (see online supporting
information).
(1) Temperate forests
Mature cool temperate forests store 153–642 tonnes carbon
(t C) ha−1 in above- and below-ground living and dead
biomass (Keith, Mackey & Lindenmayer, 2009) and cover at
least 15.9 × 106 km2 (Schmitt et al., 2009). The diversity of
native large ungulate herbivores within temperate forests and
woodlands is 7 and 14 species, respectively, and dominated
by Cervidae (deer), which are the most widely distributed
family of large forest herbivores (Fritz & Loison, 2006).
Deer substantially alter the composition of temperate
forests in North America, Europe, Asia, and the southern
hemisphere (Coomes et al., 2003; Côté et al., 2004; Dolman
& Wäber, 2008; Takatsuki, 2009). Exclusion of white-tailed
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
Carbon storage in terrestrial ecosystems: do browsing and grazing herbivores matter?
deer (Odocoileus virginianus, 40 deer km−2 ) from browsing in
a mixed hardwood forest in central Canada resulted in a
fivefold difference in the density of small trees (<7.5 cm dbh)
between browsed and unbrowsed plots after 25 years (Koh,
2002). Multiplying the mean densities of trees inside and
outside of exclosures by their respective mean diameters,
the above-ground biomass of small trees was greater within
exclosures by 1.25 t C ha−1 (estimating biomass from the
regression equation for all Canadian hardwood species;
Ung, Bernier & Guo, 2008). This suggested a difference
in carbon stocks of 1.25 t C ha−1 / 25 years, which is
small in comparison to the above-ground carbon stocks of
this vegetation type (105 t C ha−1 , Table 1). By contrast,
exclusion of deer, feral goats (Capra aegagrus hircus), and cattle
(Bos taurus) from New Zealand forests for between 4 and
43 years produced results relative to outside of exclosures
ranging between a reduction in sapling density of 5.2
saplings m−2 and an increase of 0.9 m−2 (Wright, 2009),
equivalent to a difference of 0.01 t C ha−1 , assuming a
mean above-ground sapling biomass of 100 g (reported for
Agathis australis, Silvester & Orchard, 1999). The difficulties in
estimating long-term changes in above-ground carbon stocks
from these effects are that changes occur slowly, deer may
simply be removing understorey plants in the same way as
would occur under resource competition (e.g. Miyaki & Kaji,
2009), such that differences in sapling densities may represent
relatively ephemeral reductions in carbon stocks, and that
over time, unpalatable species may replace palatable species
with little eventual effect on above-ground carbon stocks (see
Section II.1; Coomes et al., 2003; Côté et al., 2004; Takatsuki,
2009).
In mature forests, high levels of deer herbivory can result
in a large reduction in standing biomass over a long time
period. Exclusion of white-tailed deer (Odocoileus virginianus)
from browsing in the same Canadian forest discussed above
resulted in a mean density of trees >7.5 cm dbh of 1 182
stems ha−1 within exclosures versus 1 390 stems ha−1
outside of exclosures after 25 years (Koh, 2002). Estimating
biomass per individual using a mean dbh of 21.4 cm within
exclosures versus 18.9 cm outside of exclosures (Koh, 2002),
and the regression equation to predict biomass for all
Canadian hardwood species (Ung et al., 2008), this difference
is equivalent to 21.2 t C ha−1 . Similarly, the total basal area of
all species in an open English woodland, dominated by Fagus
sylvatica and Quercus robur was 23.2 m2 ha−1 compared with
37.3 m2 ha−1 within a 129 year-old cattle exclosure; equating
to a difference of 2.80 t C ha−1 (Mountford & Peterken, 2003;
see Supporting Methods for detailed calculations). Large
herbivores can also alter below-ground carbon stocks within
these systems through their dung and effects on litter quantity,
but many of these changes are small compared with the size
of above-ground carbon stocks. For example, 1 moose km−2
in Alaskan boreal forest contributes 0.30 t C ha−1 as dung,
without considering the effects of consumption (Butler &
Kielland, 2008), and standing litter is on average reduced by
0.08 t C ha−1 across browsed New Zealand forests compared
to exclosures (Wardle et al., 2001). However, there may be
79
little overall effect on soil carbon cycling (Pastor et al., 1993).
Wardle et al. (2001) reported small effects of herbivores on soil
carbon stocks across paired exclosure plots in New Zealand
temperate forests, ranging from a reduction of 0.01 t C ha−1
to an increase of <0.01 t C ha−1 over 14–34 years.
Deer also affect the carbon stocks of understorey
plant communities. In Quebec, Canada, high white-tailed
deer densities (>15 deer km−2 ) convert early-successional
states in post-logging balsam fir (Abies balsamea) forests to
open, graminoid-dominated communities (Tremblay, Huot
& Potvin, 2006). Reductions in biomass for the most
abundant seedlings and herbs in these forests (Abies balsamea,
Maianthemum canadense, Cornus canadensis, Rubus spp.) three
years after logging from 0.21 t C ha−1 at 0 deer km−2 to
0.04 t C ha−1 at 56 deer km−2 were negated by an increased
biomass of Gramineae: 0.20 t C ha−1 at 0 deer km−2 to
1.01 t C ha−1 at 56 deer km−2 (Tremblay et al., 2006). Longterm carbon stocks at this site, however, likely depend on
whether trees can eventually establish and how soil carbon
pools respond. Conversely, above-ground biomass of woody
and herbaceous understorey vegetation in a closed-canopy
old-growth riparian forest in Japan dominated by dwarf
bamboo (Sasamorpha borealis) significantly increased after only
three years of sika deer (Cervus nippon) exclusion: browsed,
0.15 t C ha−1 versus exclosure, 2.56 t C ha−1 (Nomiya et al.,
2002). These diverging responses may arise because high
light availability post-logging favoured graminoid invasion
in Quebec, whilst in Japan, three years of deer exclusion
were insufficient to increase significantly understorey light
availability through reductions in canopy tree recruitment.
(2) Tropical forests
Tropical forests store approximately 111–498 t C ha−1 in
above- and below-ground living and dead biomass (Keith
et al., 2009) and cover at least 11.8 × 106 km2 (Schmitt et al.,
2009). However, herbivores are unlikely to have a large
impact on carbon stocks in these forests because many
species are either threatened, restricted in range, or frugivores
(Bodmer & Ward, 2006). The impacts of herbivores may be
particularly limited in the Neotropics, where white-tailed
deer are the only ground-dwelling species that is strictly a
browser or grazer (as compared to 15 ungulate species in the
Paleotropics; Nowak, 1999). One explanation for the paucity
of large herbivores is that much of the vegetation in tropical
forests occurs above the browse tier of ground-dwelling
mammals, requiring the evolution of arboreal life-history
strategies that invariably constrain body size (Bodmer, 1989;
Coley & Barone, 1996). The significantly greater area of
forest within the Neotropics than within the Afrotropics or
Indomalaya (FAO, 2001) also suggests that the pantropical
diversity of large forest herbivores would have been severely
influenced by the late Quaternary megafauna extinction that
disproportionately affected South America (extinction of 46
of 58 mammalian genera >40 kg versus two of 44 genera in
sub-saharan Africa; Martin, 1984) and the decline in closedhabitat browsers during the Miocene associated with climatic
change and the rise of C4 plants (Fritz & Loison, 2006).
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
80
The relative scarcity of large mammalian herbivores in the
Neotropics compared with the Paleotropics may also reflect
the lack of historically large extents of savanna in which the
majority of herbivores would have resided and from which
a small number of herbivores would have evolved with
the characteristics required to colonize forests (Cristoffer &
Peres, 2003), particularly following the decline in browsers
and rise in open-habitat ungulates during the Miocene (Fritz
& Loison, 2006). High levels of hunting also limit the impacts
of large herbivores within some tropical forests at the present
time (Wright et al., 2000; Brashares et al., 2004).
Large mammalian herbivores likely have minimal impacts
on carbon stocks across global tropical forests compared
with temperate regions due to their limited distributions and
low densities within these vegetation types, and there are
few exclosure studies documenting their impacts on forest
vegetation. Evidence from moist Panamanian forests suggest
that herbivores such as agoutis (Dasyprocta punctata), pacas
(Agouti paca), brocket deer (Mazama americana), white-tailed
deer, peccaries (Tayassu tajacu), and tapirs (Tapirus bairdii), do
not alter the short-term abundances of understorey herbs
(Royo & Carson, 2005) and effects on seedling survival may
not carry through to the adult layer (Wright et al., 2000; Paine
& Beck, 2007). However, reductions in seedling densities
can still influence above-ground carbon stocks. In African
rainforests, densities of canopy tree seedlings declined from
3 330 to 1 620 seedlings ha−1 over 25 months at elephant
densities of approximately 0.4 individuals km−2 relative to
predicted abundances in the absence of elephants (Lawes
& Chapman, 2006). Lewis et al. (2009) speculate that such
interactions, specifically declines in elephant populations
that have reduced forest disturbances, may contribute to
increases in carbon sequestration of 0.63 t C ha−1 year−1
across African tropical forests over the last four decades.
Herbivores can also affect carbon stocks by reducing growth
rates, e.g. native pigs (Sus scrofa) at densities of 27–47 pigs
km−2 (Ickes, 2001) in Malaysian lowland rainforest reduce
mean tree height growth by 3.38 cm year−1 (Ickes, Dewalt
& Appanah, 2001), and this is equivalent to a reduction of
0.2 t C ha−1 year−1 across individuals in an unexclosed stand
(see online supporting information: (1) Supporting methods
for calculations).
(3) Temperate grasslands and shrublands
Humans have heavily impacted much of the world’s
temperate grasslands and shrublands, reducing the historical
extent of this vegetation type by approximately 43%
to 20.1 × 106 km2 (White, Murray & Rohweder, 2000).
Seventy-five per cent of converted grassland has been
modified for agriculture, including for rearing livestock, and
in some regions up to 70% of the historical area of grasslands
and shrublands has been converted to cropland, e.g. North
American tallgrass prairie and the South American Cerrado
(White et al., 2000). By contrast, conversion to cropland is
limited in drier regions, and these areas may experience
relatively high levels of extensive livestock grazing, e.g.
Mongolia and southern Australia (FAO, 2008). Overall,
Andrew J. Tanentzap and David A. Coomes
temperate grasslands store between 9 and 30 t C ha−1 in
above- and below-ground biomass and ≥70 t C ha−1 in soils
(White et al., 2000). These ecosystems contain a high diversity
of large undomesticated herbivores: 22 ungulate species in
temperate grasslands (Fritz & Loison, 2006).
The U.S. has among the largest areas of extensive grazing
on arid and semi-arid land (Reid et al., 2008). In the natural
short-, mixed-, and tall-grass prairies of the central U.S.,
livestock reduce above-ground carbon stocks by <0.01 to
0.50 t C ha−1 year−1 (see online supporting information:
Table S1). Although Welker et al. (2004) reported similar
declines in above-ground carbon stocks in central U.S.
grasslands of 0.24 t C ha−1 compared to exclosures (mean
age, 60 years), soil carbon levels were significantly higher
in grazed plots by 2.22 t C ha−1 . Others have reported
similar increases in ecosystem-level carbon stocks due to
herbivory, in part because grazing can increase soil carbon
accumulation through greater annual shoot turnover and
changes to plant community composition, which exceed
grazing-induced declines in above-ground carbon (Brand &
Goetz, 1986; Reeder & Schuman, 2002). Differences in root
mass and dynamics among plants can lead to both positive
and negative effects of herbivores on ecosystem-level carbon
stocks within the same region depending on the dominant
vegetation cover (Schuman et al., 1999; Bakker et al., 2004;
Derner et al., 2006; Fig. 2). Soil carbon levels follow similar
patterns: increasing at some sites whilst decreasing at others
(see online supporting information: Table S2), and the lack
of a consistent response suggests that climate may be more
important than herbivores within North America (Jobbágy
& Jackson, 2000), or at least may interact with herbivore
distributions (Olff, Ritchie & Prins, 2002).
Much of the grasslands of central Asia are vulnerable to
erosion, deepening water tables, and reductions in vegetation
cover due to extensive livestock grazing (FAO, 2000).
Grazing by 6 sheep (Ovis aries) ha−1 in arid rangeland in Inner
Mongolia, China, resulted in an above-ground biomass of
0.02 t C ha−1 after four years compared to a biomass
of 1.21 t C ha−1 within livestock exclosures (Zhao et al.,
2005). Below-ground biomass was also significantly higher
after four years in exclosures than in grazed pastures: 1.16
and 0.11 t C ha−1 , respectively (Zhao et al., 2005), and these
results were consistent with other studies from the region at
similar herbivore densities (Tables S1 and S2).
Undomesticated native herbivores determine the size of
carbon pools in natural and semi-natural temperate grasslands by influencing plant community dynamics and ecosystem processes (Knapp et al., 1999). In a tall-grass prairie in
Kansas, grazing by high densities of bison (approximately 82
animals km−2 ) consumed 1.02 t C ha−1 of graminoid biomass
in a single growing season (Towne, Hartnett & Cochran,
2005). Less palatable forbs replaced graminoids, but forb
biomass was only 0.06 t C ha−1 greater in grazed plots than
in ungrazed exclosures (Towne et al., 2005). Grazing can
also stimulate biomass production in North American grasslands (Frank & McNaughton, 1993). In Yellowstone National
Park, grazing increased above- and below-ground biomass
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
Carbon storage in terrestrial ecosystems: do browsing and grazing herbivores matter?
Above-ground biomass
Boreal
Forests
Temperate
Tropical
Temperate (wild)
Temperate (livestock)
Grasslands
& shrublands
Tropical (wild)
Tropical (livestock)
Tundra
Wetlands
Intensive pastures
Below-ground biomass
(3)
(0)
(12)
(1)
(2)
(0)
(7)
(2)
(35)
(15)
(4)
(1)
(11)
(2)
(2)
(0)
(3)
(2)
(5)
(1)
Soil
Litter
Boreal
Forests
Temperate
Tropical
Temperate (wild)
Temperate (livestock)
Grasslands
& shrublands
Tropical (wild)
Tropical (livestock)
Tundra
Wetlands
Intensive pastures
-1
81
0
1
2
3
(2)
(2)
(3)
(3)
(0)
(0)
(4)
(4)
(28)
(28)
(0)
(0)
(7)
(7)
(1)
(1)
(2)
(2)
(3)
(3)
4
-0.2
0.0
0.2
0.4
ΔC stocks upon herbivore exclusion (t C ha
−1
0.6
–1
year )
Fig. 2. Mean effect of herbivore exclusion on carbon stocks in four pools: above- and below-ground biomass, soil, and litter ±
possible range of values. We averaged changes in above- and below-ground carbon stocks in plant biomass and soils (1 m depth)
upon herbivore exclusion for each study reviewed in Tables S1 and S2 (see online supporting information) and summed values
from studies measuring different components within the same carbon pool (e.g. above-ground tree and understorey biomass). For
grasslands and shrublands, we separately estimated effects due to primarily undomesticated herbivores (wild) versus livestock in
extensive pastures. Values in parentheses denote number of studies used to derive estimates.
after one year by 0.15 and 1.09 t C ha−1 , respectively (Frank
et al., 2002). Below-ground biomass and soil carbon were
also 0.71 and 5.6 t C ha−1 greater, respectively, in grazed
plots than in exclosures of a mean age of 35 years (Frank &
Groffman, 1998). However, standing above-ground biomass
may still be greater within exclosures (Coughenour, 1991;
Frank & Groffman, 1998), because herbivore consumption
can increase with plant biomass production (Frank, 1998).
(4) Tropical grasslands and shrublands
Tropical grasslands extend over 21.7 × 106 km2 and remain
much more intact than those in temperate regions, i.e. occurring over approximately 70% of their historical extent with
the exception of flooded grasslands and savannas, which have
been heavily modified by humans (White et al., 2000). Carbon stocks in above- and below-ground biomass range from
18 to 58 t C ha−1 , which is less than the levels stored in soils
of ≥73 t C ha−1 (White et al., 2000). These vegetation types
also contain the highest diversity of large mammalian herbivores: 52 and 33 species of ungulates in tropical grasslands
and shrublands, respectively (Fritz & Loison, 2006).
The role of undomesticated mammalian herbivores has
long been recognized in determining the structure and
composition of vegetation communities in African savannas
(McNaughton, 1985; McNaughton & Georgiadis, 1986;
Guldemond & Van Aarde, 2008; Asner et al., 2009), and these
effects may influence terrestrial carbon stocks. For example,
in fire-suppressed East African savanna, the above-ground
biomass accumulation of live woody species over three years
in plots browsed by 12.2 cattle km−2 , 1.7 elephants km−2 ,
and 159 antelope km−2 (impala, Aepyceros melampus and dikdik, Madoqua kirkii) was −0.06 t C ha−1 year−1 compared
with 0.45 t C ha−1 year−1 in exclosures (Augustine &
McNaughton, 2004). The difference was similar to the 0.34 t
C ha−1 year−1 increase in above-ground herbaceous biomass
observed in South African savanna following large herbivore
exclusion (Jacobs & Naiman, 2008).
Livestock grazing in arid tropical grasslands can drive
changes in plant communities (Sinclair & Fryxell, 1985;
Bryant et al., 1990; Ludwig & Tongway, 1995; Asner et al.,
2004), and thus is one of the principal determinants of
carbon stocks in these ecosystems (Lal, 2002, 2004). In the
Sahel region of sub-Saharan Africa, ecosystem state changes
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
82
may have largely been due to colonial settlement practices
that encouraged the abandonment of traditional pastoral
livestock movements, leading to overgrazing and declines
in the productivity of grasslands (Sinclair & Fryxell, 1985).
Under high grazing pressures, deep-rooted perennial grasses
are restricted from reaching deep sources of groundwater
since resources for root growth are allocated above ground to
compensate for herbivory (Sinclair & Fryxell, 1985). Annual
grasses quickly invade, followed by unpalatable annual
dicotyledonous herbs with lower biomass per individual than
perennial or annual grasses (Rietkerk et al., 1996). Overall, reduced vegetation establishment leads to soil erosion
and desertification (Sinclair & Fryxell, 1985). This transition towards desertification is occurring in many overgrazed
arid tropical grasslands (Asner et al., 2004), and can lead to
large reductions in above-ground plant biomass and litter
inputs (summarized in Table S1). In some cases, declines
in soil moisture use associated with the reduction of grasses
allow woody species to establish (Graz, 2008). Anthropogenic
activities also contribute to shrub encroachment in many
cases (e.g. fire suppression, Asner et al., 2004). Shrub
encroachment can increase above-ground carbon stocks but
changes in soil carbon stocks are unpredictable in shrubinvaded grasslands and may depend on climate (Jackson
et al., 2002).
Large areas of grassland and savanna in South America
have been exposed to extensive grazing only since European
colonization (Solbrig, 2006), due to the relative scarcity of
large mammalian herbivores following the late Quaternary
(Webb, 1978). As a result, the use of these areas as
rangeland can have large impacts on vegetation since many
native savanna plants have a low tolerance of herbivory
(Sarmiento, 1992). The few species of large mammalian
herbivores in South American grasslands and savanna
(primarily Lama spp.) have also largely been displaced
in many areas by more competitive domesticated sheep
(Baldi, Albon & Elston, 2001). For example, in highelevation grasslands in the Argentinean Pampas, livestock
grazing can reduce above-ground biomass and litter by
0.74–5.15 t C ha−1 , depending on the duration of herbivore
exclusion (Pucheta et al., 1998). However, soil carbon levels
demonstrate diverging responses to herbivory across South
American grasslands and shrublands, ranging from −3.81 to
0.37 t C ha−1 year−1 (Chaneton & Lavado, 1996; Abril &
Bucher, 1999). These differences may arise due to within-site
variation between herbivore-accelerated nitrogen cycling,
which constrains soil organic matter formation (Neff et al.,
2002), and herbivore-induced below-ground carbon and
nitrogen allocation in plants that increase soil carbon levels
(Piñeiro et al., 2009).
(5) Artificial pastures
Reid et al. (2008) estimate that 61.2 × 106 km2 (41%) of
global land is used for livestock grazing, with 91 and
9% of this area used for extensive and intensive grazing,
respectively. However, the global area of artificial pastures
is increasing with human population growth in developing
Andrew J. Tanentzap and David A. Coomes
countries (Asner et al., 2004; Green et al., 2005), while
extensive pastures are being abandoned in developed
countries (FAO, 2008). Global conversion of natural
vegetation types into pastures is estimated to emit 0.5 Pg C
year−1 (Houghton, 2003; Strassman, Joos & Fischer, 2008),
despite the fact that soil carbon stocks may increase (Guo
& Gifford, 2002). Additionally, methane emissions from
livestock in artificial pastures are more than double those in
extensive pastures at a global level (equivalent to 0.87 versus
0.38 Pg C year−1 ), despite the significantly smaller land area
covered by artificial systems (Steinfeld et al., 2006).
Reductions in carbon stocks due to herbivores in artificial
pastures may exceed those in extensive pastures. For
example, Allard et al. (2007) reported 0.46 t C ha−1 lower
peak total above-ground biomass in artificial European
pastures compared to extensive pastures after three years
of grazing. Bardgett, Frankland & Whittaker (1993) similarly
measured 0.58 t C ha−1 year−1 lower soil carbon levels in
artificial pastures compared to moderately-grazed extensive
grasslands. In the eastern U.S., grazing by cattle across an
intensively managed grass-pasture resulted in a reduction in
above-ground biomass compared to exclosures of 1.26 t C
ha−1 year−1 (Skinner, 2008), but soil carbon may increase
with grazing in these systems (Weinhold, Hendrickson &
Karn, 2001).
(6) Other vegetation types: wetlands and Arctic
tundra
Between 8.2 and 10.1 × 106 km2 of global land area is
covered by wetlands (Lehner & Döll, 2004), of which
4.2 × 106 km2 contain over 30 cm of peat (Parish et al.,
2008). Peatlands are particularly important because these
formations store the largest quantity of carbon per unit
land area among vegetation types: 1 375 t C ha−1 stored
in peat versus 25 and 50 t C ha−1 in vegetation and soils,
respectively (Parish et al., 2008). Junk et al. (2006) reported 35
species of large mammalian herbivores across seven globally
important wetlands, although 25 of these species occurred in
seasonally flooded or non-flooded African savanna that may
only be utilized during the dry season. In Danish salt marshes
grazed by sheep and cattle, the late-summer peak in aboveground carbon stocks was reduced by 5.24 t C ha−1 due to
differences in litter accumulation (Morris & Jensen, 1998).
Total below-ground biomass was greater within grazed plots
by 2.30 t C ha−1 , although soil carbon (0–60 cm depth)
was less than in exclosures by 0.60 t C ha−1 (Morris
& Jensen, 1998). Similarly, in peatlands on the Tibetan
Plateau, livestock grazing reduced above-ground biomass by
1.96 t C ha−1 in zones with shallow water tables after only
one year (Hirota et al., 2005). This can represent a relatively
large reduction in long-term below-ground carbon stocks if
the majority of litter in this system enters anoxic soil that
becomes peat rather than simply being oxidized.
Arctic tundra covers 5.6 × 106 km2 and stores large
amounts of carbon in soils relative to vegetation, approximately 348 t C ha−1 (Ping et al., 2008) versus 7 t C ha−1
(Saugier, Roy & Mooney, 2001), respectively. Six species
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
Carbon storage in terrestrial ecosystems: do browsing and grazing herbivores matter?
83
of large mammalian herbivores persist within this region,
all within the Caprinae or Capreolinae subfamilies, with
the exception of wood bison (Bison bison athabascae) (Nowak,
1999). Grazing by reindeer (Rangifer tarandus) may be particularly important as it can initiate compositional changes
among lichen-, moss-, and graminoid-dominated communities that can reduce above-ground carbon stocks (Olofsson
et al., 2001; Olofsson, Stark & Oksanen, 2004; van der Wal,
2006; but see Brathen et al., 2007). For example, 60 years
of little to no grazing in Finland resulted in lichen biomass
of 1.78 versus 0.25 t C ha−1 where reindeer densities were
2.3 animals km−2 (Susiluoto et al., 2008). Moss biomass also
differed between sites: ungrazed, 0.01 t C ha−1 versus grazed,
1.40 t C ha−1 (Susiluoto et al., 2008), and this effect will have
an important influence on long-term carbon accumulation
in peat. Mack et al. (2004) reported that carbon sequestration in Arctic tundra was dependent on compositional shifts
in root biomass that influence below-ground decomposition, e.g. a shift in root biomass towards surface soil layers
(0–5 cm) and the quicker decomposition rates in these layers accounted for a loss of 20 t C ha−1 in experimentally
fertilized plots after 20 years. A better understanding of the
effects of herbivores on below-ground processes is clearly
needed.
to declines of 0.19 t C ha−1 year−1 , and changes in soils
ranging from increases of 3.81 t C ha−1 year−1 to declines
of 1.46 t C ha−1 year−1 , depending on vegetation type
(Fig. 2). These changes were often small in comparison to
the annual productivity of vegetation types. For example,
NPP in temperate and tropical forests is estimated at 7.8
and 12.5 t C ha−1 year−1 , respectively, with over 55% of
biomass produced above-ground (Table 1), yet herbivores
reduce above-ground carbon stocks in these vegetation types
by a mean of 0.89 and 0.23 t C ha−1 year−1 , respectively.
The NPP of Arctic tundra, which has the lowest value among
the major vegetation types, is 0.90 t C ha−1 year−1 (44%
produced above-ground; Table 1), and even this exceeds
many of the increases in carbon stocks associated with
herbivore removal. Regions underrepresented in our review,
e.g. the Amazon basin and Northern Africa, may reflect the
fact that these areas are depauperate in naturally occurring
large herbivores (Olff et al., 2002), and so have not attracted
study. We also lacked estimates for the effects of herbivores
on below-ground biomass and soil carbon levels in many
areas with high densities of large herbivores, e.g. tropical
grasslands and shrublands, and in regions with large areas
of high pasture cover, i.e. Eastern Europe and southern
Australia (Ramankutty et al., 2008), emphasizing the need
for additional data collection (Fig. 3).
IV. SYNTHESIZING THE ANNUAL EFFECTS
OF HERBIVORY ON CARBON STOCKS
(2) Extrapolating the effects of herbivores
on terrestrial carbon stocks to landscape
and global scales
We now attempt to synthesize the effects of herbivores
on carbon stocks, in order to predict their impacts at
landscape and global scales. We first summarize how the
estimated effects of herbivores on terrestrial carbon stocks
vary among carbon pools and vegetation types using the
exclosure studies we review. We then assume that the mean
effects of herbivores in different carbon pools can be summed
and extrapolated over the extent of global vegetation types.
Ideally, we would like to develop models to predict the sizes
of these effects in areas with different times since herbivore
exclusion and levels of herbivory. We use the studies in our
review to test whether these relationships persist. Finally, we
discuss gaps in knowledge that are identified by our review of
exclosure studies and analyses, and recommend approaches
to improve understanding of the effects of herbivores on
carbon cycling.
Herbivory leads to small reductions in carbon stocks per
unit land area each year but these effects are manifested
over large areas, and thus, are comparable to the large
reductions in global carbon stocks arising from relatively
localized disturbances. For example, deforestation emits
2.4 Pg C year−1 (Houghton, 2003), and this represents a
flux of 1.10 t C ha−1 year−1 over the global extent of boreal,
temperate, and tropical forest (Ellis & Ramankutty, 2008),
which may be similar to reductions in carbon stocks arising
from herbivory in some ecosystems (Fig. 2). We therefore
asked whether removing every large mammalian herbivore
from the land they currently occupy would have a large
effect on carbon stocks relative to other fluxes in the global
carbon cycle. Obviously, removing all herbivores is neither
desirable nor feasible, but asking such a question can help us
place carbon fluxes due to herbivores in a broader context.
Our approach was to extrapolate annual rates of carbon flux
per unit area attributable to herbivory in Tables S1 and S2
across the actual land cover influenced by that particular
herbivore-vegetation interaction. For each vegetation type,
we averaged the range of effects of herbivores on carbon
stocks within four carbon pools: above-ground and belowground biomass, litter, and soil carbon. Where changes in
soil carbon stocks were measured at different depths, we
only averaged measurements in the upper 75th percentile.
We then multiplied the sum of the four pools by the total
area of each vegetation type. Pools for which estimates were
unavailable were treated as zero in summations. To derive
(1) Effects of herbivores on terrestrial carbon stocks
per unit land area
On average, exclosure studies suggest that mammalian
herbivory reduces above-ground carbon stocks, consistent
with our naïve hypothesis. However, carbon stocks increased
in many circumstances because of the influences of herbivores
on litter decomposition and nitrogen mineralization (Fig. 2).
Removing herbivores from the vegetation types covered by
this review resulted in estimated changes in above-ground
carbon stocks ranging from increases of 1.96 t C ha−1 year−1
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
84
Andrew J. Tanentzap and David A. Coomes
Fig. 3. Map of sites for studies included in Tables S1 and S2 (see online supporting information) superimposed upon global aboveand below-ground carbon stocks (Kapos et al., 2008). = sites with data from only above-ground carbon pools (either above-ground
biomass and/or litter); = sites with data from only below-ground carbon pools (either biomass and/or soil carbon); = sites with
data from both above- and below-ground carbon pools.
errors within each vegetation type, we calculated the squareroot of the sum of squared standard deviations in each carbon
pool. Standard deviations were propagated through when
estimating the sum of effects across all the vegetation types.
Although our approach assumes that scattered exclosure
studies are representative samples of widespread vegetation
types, many of the studies reviewed did attempt to establish
exclosures in areas typical of broader unexclosed sites. We
emphasize that our results should be considered preliminary
and incomplete, but nonetheless, important to consider,
particularly given the need to identify where primary data
are lacking.
Our approximate analysis indicates that removing
herbivores from all of the vegetation types for which we had
exclosure studies would increase terrestrial carbon stocks by
5.2 × 10−4 Pg C year−1 over a period of up to 129 years,
with a standard deviation of 1.8 × 10−4 Pg C year−1 . The
vast majority (97%) of this value was attributable to shortterm changes in plant biomass and litter (mean duration
of reviewed studies ± standard error: 20 ± 2 years). Across
vegetation types, the largest increases in terrestrial carbon
stocks due to herbivore removal occurred in artificial pastures
and tropical grasslands and shrublands, and there was
little difference in the potential range of effects between
woody and herbaceous vegetation types (Table 1). Our
analysis further suggests that herbivore removal would lead
to considerably smaller increases in global carbon stocks
(about one thousandth) than gains or losses in carbon
stocks arising from annual sequestration in vegetation and
soils, or emissions from fossil fuels, deforestation, or fire
(Andreae & Merlet, 2001; Houghton, 2003, 2007). Fossil
fuel emissions, in particular, may consistently exceed gains
in plant biomass, e.g. reforestation of historically cleared
forests in the southern U.S. has increased carbon stocks and
compensated for emissions associated with past deforestation
but these gains are offset by increased emissions from
regional urbanization (Zhang et al., 2008). Although changes
in above-ground carbon stocks are transient, reaching a
stable state at which reductions in carbon stocks due to
herbivory approach zero given sufficient periods of time
(i.e. centuries; Fig. 4), intervening reductions in terrestrial
carbon stocks may still accrue. Furthermore, our review
does not consider methane emissions from ruminants, which
may represent a relatively large source of carbon emissions
associated with herbivores of 0.7 Pg C year−1 (Denman et al.,
2007, calculated from CO2 -equivalent with Global Warming
Potential over 100 years = 25; Forster et al., 2007).
(3) Temporal effects of herbivores on terrestrial
carbon stocks
Many of the effects of herbivory we report are unlikely to
increase linearly on an annual basis. We tested whether
changes in carbon stocks due to herbivory changed
non-linearly over time by plotting the estimated effects
of herbivores on carbon stocks (E) for each study in
Tables S1 & S2 against the duration of these studies (t).
We fitted curves to the relationship between E and t
using robust regression with MM-estimators (lmrob function
in the robustbase package, R v.2.8, R Development Core
Team, 2008), since preliminary exploration of our dataset
with linear models indicated non-normal residuals. MMestimators are a type of maximum-likelihood estimator that
is less sensitive to deviations from the distributions assumed
under classical least-squares regression (i.e. outliers). Hence,
robust regression is an alternative method when residuals
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
–1
year )
A)
Effect on carbon stocks (t C ha
Effect on carbon stocks (t C ha
85
B)
1.0
–1
0.5
–1
–1
year )
Carbon storage in terrestrial ecosystems: do browsing and grazing herbivores matter?
0.0
-0.5
-1.0
-1.5
-2.0
0
25
50
75
100
125
Study duration (years)
0.0
-1.0
-2.0
-3.0
-4.0
0
20
40
60
80
Study duration (years)
Fig. 4. Annual effect of herbivores on terrestrial carbon stocks averaged over study duration in forests ( ) and all other vegetation
types (grasslands, shrublands, artificial pastures, wetlands, and tundra; ) as estimated from (A) above-ground plant biomass (N = 82,
Table S1), and (B) soil carbon pools (N = 47, all depths, Table S2). The non-linear regression line (grey) was fitted to all points in A
using robust regression, and there was no significant relationship in B. Dashed lines represent mean effects.
are non-normally distributed (Maronna, Martin & Yohai,
2006). We detected a very weak positive relationship between
the estimated effect of herbivores on carbon stocks and
the duration of the studies from which these effects were
estimated, such that over long periods of time (i.e. decades)
mean annual reductions in above-ground carbon stocks
due to herbivores approached zero (E = 0.03 × ln(t) − 0.14,
explained deviance = 6%; χ12 = 15.81, P < 0.001; Fig. 4A).
One explanation is that, given sufficient time for dispersal
and establishment, less palatable species may simply
replace preferred species with little loss in carbon stocks.
This relationship would also explain why some of the
largest reductions in terrestrial carbon stocks occur within
herbaceous rather than woody vegetation. Herbaceous
biomass studies, many of which were less than five years
in duration in our review, may respond immediately to
herbivore exclusion as opposed to the slow changes in woody
biomass for which the mean effects of herbivory are likely to
decrease over time. There was no relationship between the
effects of herbivores on soil carbon stocks and study duration
(Fig. 4B, explained deviance = 2%; χ12 = 1.11, P = 0.293).
Others have reported soil carbon stocks in grasslands to
increase non-linearly over time in response to different
management practices (Conant, Paustian & Elliott, 2001) and
these changes may occur slowly over decades or centuries
(Guo & Gifford, 2002).
(4) Influence of herbivore body size and feeding
strategy on terrestrial carbon stocks
We tested whether a negative relationship existed between
the total biomass of herbivores per unit land area
and their annual effects on above-ground carbon stocks,
since previous research has shown the consumption and
biomass of individual herbivores to be positively correlated
(Shipley et al., 1994; Belovsky, 1997). We used estimates of
the effects of herbivores on mean above-ground carbon
stocks from studies in Table S1, omitting those that did
not quantify herbivore densities outside exclosures and
reporting comparisons between all levels of herbivory for
studies that included multiple herbivore densities. Where
individual studies reported multiple comparisons at the
same herbivore densities and over the same time period,
we used the mean of these values in our calculations. We
also included additional studies that quantified the response
of vegetation to the manipulation of herbivore densities
rather than the complete removal of herbivores (Cao et al.,
2004; Allard et al., 2007; Dumont et al., 2007; Holland et al.,
2008; Jauregui et al., 2008). For each of the 57 studies, we
calculated the biomass of herbivores per unit land area
by multiplying herbivore densities estimated in each study
by estimates of mean herbivore body mass. Where studies
did not report herbivore body mass, values were obtained
from the following sources: Asian water buffalo (Bubalus
bubalis; McMahon & Bradshaw, 2008); domestic horse (Equus
caballus), elephant, elk, moose, reindeer, sheep, and whitetailed deer (Shipley et al., 1994); mule deer (Odocoileus hemionus)
and sika deer (Forsyth & Duncan, 2001). We used robust
regression with MM-estimators since linear models indicated
non-normal residuals.
The total biomass of herbivores at a site (M, t km−2 )
was not correlated with their effect on above-ground carbon
stocks (deviance explained = 2%, χ12 = 1.80, P = 0.183),
contrary to our predictions. The lack of a relationship may
have been due to the high amount of variation at low
herbivore biomass (<10 t km−2 ; Fig. 5). Terrestrial carbon
stocks were reduced in control plots at all levels of herbivore
biomass relative to exclosures through the loss of aboveground plant biomass, and there appeared to be a negative
trend between herbivore biomass and their reductions in
carbon stocks when herbivore biomass was ≥10 t km−2
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
Andrew J. Tanentzap and David A. Coomes
–1
–1
Effect on carbon stocks (t C ha year )
86
0.5
0.0
-0.5
-1.0
-1.5
-2.0
0
20
40
60
80
100 120 140
–2
Total herbivore biomass (t km )
Fig. 5. Effect of herbivore biomass (t km−2 ) on terrestrial aboveground carbon stocks (t C ha−1 year−1 , N = 106). Estimates
derived from studies that exceeded the mean study duration of
16.4 years are denoted as ; all others are plotted as .
(Fig. 5). Addition of mean body mass per herbivore was not a
significant factor in the model (log-transformed, χ12 = 2.18,
P = 0.140). A model of both total herbivore biomass and
study duration was however, significantly better than a model
of only herbivore biomass since the largest reductions in
carbon stocks were estimated from studies that were shorter
than the long-term mean (log-transformed, χ12 = 12.82,
P < 0.001; Fig. 5). This relationship further suggests that
temporal scale is an important consideration, irrespective of
herbivore biomass or feeding strategy. We also acknowledge
that our estimates suggest that all carbon removed by
herbivory is eventually transferred to the atmosphere, despite
the fact that carbon can remain within terrestrial systems for
centuries after animals die if decomposition occurs slowly.
However, the percentage of above-ground vegetation that
remains stored within herbivores in the form of body mass
may be minimal, e.g. <4% (Allard et al., 2007).
(5) Priorities for future research
Our review has identified priorities for future research.
Firstly, we acknowledge that studies of herbivore impacts
are likely to have been concentrated within locations where
these impacts are perceived to be high, and therefore,
deriving a sensible area over which to extrapolate our
estimates is difficult. Distributions and population sizes
of large undomesticated mammalian herbivores are also
predicted to change globally under multiple climate change
scenarios that will differentially affect geographic regions
and their biota (Post & Stenseth, 1999). For example, some
herbivore populations are predicted to respond positively
to climate change, whilst others may respond negatively or
remain unaffected (Johnston & Schmitz, 1997; Loison, Jullien
& Menaut, 1999; Post & Stenseth, 1999; Niemelä et al., 2001;
Ogutu & Owen-Smith, 2003; Garel et al., 2004; Thuiller et al.,
2006; Levinsky et al., 2007; Vors & Boyce, 2009). Where we
review multiple impacts within a vegetation type, we assumed
that the effects of herbivores on different carbon pools were
independent, and thus additive, and this may overestimate
some impacts. Combining landscape-level satellite estimates
of vegetation productivity and standing biomass with similar
methods for estimating herbivore densities may provide
an approach for predicting the impacts of herbivores on
vegetation and carbon stocks at continental and global scales.
We assume that carbon stocks increase linearly over time in
our calculations of the annual effects of herbivores on carbon
stocks. However, transitions between ecosystem states may
occur suddenly and rapidly (Holling, 1973; May, 1977),
leading to punctuated changes in carbon stocks. In systems
that respond to herbivory within several years, prolonged
study periods (i.e. decades) may only reduce the mean
annual effect of herbivores on carbon stocks. For example,
if herbivores remove all the herbaceous biomass in a forest
understorey within one year, averaging this effect over five
years of study will only under-estimate impacts. Most of the
studies in our review were too short for the differences in
carbon stocks between areas with and without herbivores to
no longer accrue (Conant et al., 2001), but the duration of
studies should be considered in the future where the effects
of herbivores are non-linear over time.
The large uncertainty associated with the relatively large
amounts of carbon stored in soils is a major impediment
in deriving an ‘herbivore flux’ for the global carbon cycle.
We identified few studies of below-ground carbon stocks
and soil carbon levels, especially involving the effects of
undomesticated large herbivores on soil carbon in forests
(Fig. 2; Table S2). Additionally, soil carbon levels per unit
area can vary between herbivore treatments within studies
because of poor sampling techniques. Many studies sample
soils at a constant depth (e.g. Table S2), and if soil density
increases with herbivory due to compaction (Berg, Bradford
& Sims, 1997), this can lead to different masses of soil being
measured at different herbivore densities. Micro-variation
in soil depth and texture also suggest that levels of total
soil carbon extrapolated over entire landscapes should be
interpreted with caution (Jobbágy & Jackson, 2000).
Studies measuring short-term CO2 fluxes or 13 C:12 C
ratios from photosynthesis and soil respiration may provide
an alternative approach to pool size determinations for
estimating the effects of herbivores on terrestrial carbon
stocks. Differences between CO2 uptake by plants and
emissions by soil respiration, termed net ecosystem exchange
(NEE), have been used to estimate the effects of herbivores on
carbon sequestration in temperate grasslands (LeCain et al.,
2002; Li et al., 2005; Risch & Frank, 2006; Owensby, Ham
& Auen, 2006), temperate and tropical pastures (Wilsey
et al., 2002; Skinner, 2008), Arctic tundra (Susiluoto et al.,
2008), and European salt marshes (Morris & Jensen, 1998).
Herbivores reduce NEE through consumption of biomass
and the subsequent effects on photosynthesis and respiration
(Morris & Jensen, 1998; Skinner, 2008), but we are aware
of only the eight aforementioned studies that have measured
NEE across long-term herbivore exclosures. The direction
of the effects of herbivores on NEE may also vary seasonally
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
Carbon storage in terrestrial ecosystems: do browsing and grazing herbivores matter?
(Owensby et al., 2006), and several studies have reported no
difference in CO2 exchange between sites with and without
herbivores (LeCain et al., 2002; Wilsey et al., 2002; Risch
& Frank, 2006), making a synthesis far too preliminary
at this stage. Other studies have used 13 C labelling to
trace CO2 exchange and compare fluxes in areas where
herbivores are present versus absent (Doescher, Svejcar &
Jaindl, 1997; Kielland & Bryant, 1998; Sinton et al., 2000).
Studies that utilize carbon tracing techniques in combination
with estimates of standing carbon stocks in plant biomass or
soils may provide a more robust approach to estimating the
effects of herbivores on terrestrial carbon stocks (e.g. Frank
et al., 1995; Wilsey et al., 2002).
V. IMPLICATIONS OF HERBIVORE REMOVAL
FOR TERRESTRIAL CARBON CYCLING
AND CONSERVATION
Herbivore removal has important consequences for strategies
that aim to increase terrestrial carbon stocks at regional
scales. The United Nations Reducing Emissions from
Deforestation and Forest Degradation in Developing
Countries Programme (REDD) is one such initiative for
reducing carbon emissions by compensating landowners in
tropical countries for forest conservation. Deforestation is
considered one of the largest sources of carbon emissions
in the tropics (Houghton, 2003), and has increased in
parts of the Brazilian Amazon over the last three decades
with increasing cattle herds and the price of Brazilian beef
(Kaimowitz et al., 2004). One potential REDD mechanism
would aim to compensate landowners for the revenue that
they would forgo by conserving forest rather than producing
livestock, in order to mitigate the emissions associated
with pasture formation and livestock production (Nepstad
et al., 2007). Although we lack studies from the Amazon
in our review, carbon stocks are significantly higher within
secondary tropical forests than pastures; across abandoned
pastures in the central Amazon basin of Brazil, aboveground biomass and soil carbon accumulated in young
secondary forests (<15 years old) at a rate of 7.0 t C
ha year−1 (Feldpausch et al., 2004). Economic drivers of
land-use changes have also been important in Europe,
where increases in livestock prices have been linked to
vegetation changes, including woodland regeneration over
400 years in Scottish uplands (Hanley et al., 2008), and may
contribute to recent afforestation across Europe (Saikku et al.,
2008; Rautiainen, Saikku & Kauppi, 2010). Market-based
economic incentives to remove some grazing and browsing
animals may prove an important component of strategies for
increasing terrestrial carbon stocks, particularly in developing
countries where livestock production is increasing (FAO,
2008). In developed countries, the effects of herbivore
removal are small compared with legislated reductions in
carbon emissions, such as the 0.07 Pg C year−1 target by
2020 in the U.K., but will benefit other pre-existing measures,
87
e.g. 0.001 Pg C year−1 target reduction by 2020 through
woodland conservation schemes (CCC, 2008).
Removing large herbivores can have unintended
consequences and may be undesirable, particularly if it
conflicts with the conservation of biodiversity. For example,
herbivory is required to maintain some semi-natural
habitats in agricultural landscapes (e.g. chalk grasslands,
heathlands, moorlands) by reducing interspecific competition
among species and creating structural heterogeneity that
provides niche opportunities for many specialist species
(van Wieren, 1995). The loss of large herbivores may also
affect invertebrate communities and co-evolved interactions
among trophic levels; e.g. removal of large herbivores
in African savannas shifts tree-nesting ant communities
associated with Acacia drepanolobium from pollinator species to
ants that attract stem-boring beetles, leading to slower growth
and higher mortality of adult trees (Palmer et al., 2008). In
many systems, the removal of large herbivores also facilitates
smaller herbivores, i.e. rodents, which can potentially
maintain the effects of larger herbivores on above-ground
biomass through seed predation and grazing (Kessing, 2000;
Smit et al., 2001; Steen, Mysterud & Austrheim, 2005).
Other groups of small herbivores, e.g. rabbits, may however
prefer vegetation grazed by large herbivores (Bakker, Olff
& Gleichman, 2009). Herbivore removal may also alter
ecosystem processes. In mesic savanna, removal of white
rhinoceros increases the height of grassland vegetation,
enhancing the spread and intensity of fires (Waldram, Bond
& Stock, 2008). Livestock introduction into western U.S.
forests and the semi-arid savannas of South America are
presumed to suppress fires by consuming fuel sources, i.e. dry
herbaceous vegetation (Bucher, 1987; Belsky & Blumenthal,
1997). Similarly, the irruption of wildebeest (Connochaetes
taurinus) populations in the Serengeti following rinderpest
eradication has led to declines in fires through reduced fuel
loads, and these changes have increased tree density and
terrestrial carbon stocks (Holdo et al., 2009). Therefore, the
benefits of other ecosystem functions, such as biodiversity and
fire regimes, will have to be balanced against carbon storage.
One limitation is that the relationship between herbivory
and many ecosystem functions is poorly understood along
gradients of herbivore density (Gordon, Hester & FestaBianchet, 2004).
Improved land management may increase carbon stocks
without herbivore removal. Moderate application of nitrogen
fertilizer in grasslands, with the exception of sites on highly
organic soils (e.g. wetlands), can increase organic carbon
input to soil and net carbon sequestration (Soussana et al.,
2004). However, the carbon costs associated with fertilizer
production may negate any increases in terrestrial carbon
stocks, as can the emissions produced from agricultural lime
that is applied to neutralize soil pH after the application of
acidifying nitrogen-based fertilizers (Schnabel et al., 2001;
Rangel-Castro et al., 2004). De-intensification of pasture
systems may be another approach to increase carbon stocks
without herbivore removal, since high rates of fertilization
can increase soil carbon mineralization and result in rates of
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
88
organic matter decomposition that exceed those of organic
carbon input (Soussana et al., 2004). Litter decomposition
is also a function of the C:N ratio of biomass, and
higher nitrogen levels may lead to more degradable plant
material (Schnabel et al., 2001). Overall, reduced stocking
densities, manipulation of plant species composition, erosion
management, and enhancement of low soil fertility grazing
lands have the potential to increase soil carbon sequestration
by between 0.05 and 0.20 t C ha year−1 (Lal, 2003). Although
such management would offset reductions in carbon stocks
due to herbivores in some vegetation types (Fig. 2), in
other regions, such as the tropics, strategies to subsidize
pasture improvement may be less cost-effective at increasing
carbon stocks than efforts to reduce deforestation and lack
the corollary benefits associated with forest conservation
(Fearnside, 1995).
Andrew J. Tanentzap and David A. Coomes
carbon stocks without herbivore removal; e.g. by 0.05–0.20
t C ha year−1 in soils (Lal, 2003), and can thus be sufficient
to offset the effects of herbivores in some vegetation types.
(5) One of the greatest limitations in deriving sensible
estimates for the effects of herbivores on carbon cycling is
quantifying the area and intensity of herbivory within a given
region. We hope our review emphasises and motivates the
need for large-scale assessments of how herbivores impact
vegetation and their relation to carbon cycling.
VII. ACKNOWLEDGEMENTS
We thank D.R. Bazely, K.J. Kirby, D.M. Wright, E.R. Lines,
S.W. Husheer, and three anonymous reviewers for providing
helpful comments that improved an earlier version of the
manuscript.
VI. CONCLUSIONS
(1) The mean effect of herbivore removal on carbon stocks
per unit land area is positive but relatively small across
vegetation types and the range of possible effects overlaps
with zero for many carbon pools. However, we lacked
studies for the effects of herbivores on soil carbon stocks in
boreal regions, despite the abundance of large mammalian
herbivores in this region and the high quantities of carbon
stored below-ground.
(2) At the global scale, the effects of herbivore removal are
larger in tropical grasslands and shrublands, and pastures,
than in forests. Changes in forests proceed more slowly than
those in herbaceous vegetation, and thus, these effects may
be small over the short time periods of most studies, during
which ecosystems can respond to herbivores through altered
plant community composition with little reduction in carbon
stocks.
(3) The removal of all large mammalian herbivores from
the vegetation types we review would increase global carbon
stocks, primarily above-ground, by 5.2 × 10−4 Pg C year−1
(standard deviation = 1.8 × 10−4 Pg C year−1 ). Although
our estimate is several orders of magnitude lower than
most natural and human-mediated fluxes in the terrestrial
carbon budget, our results are preliminary and speculative
because we lack long-term datasets, especially for soils in
non-agricultural settings. We also assume that carbon stocks
increase linearly over time in response to herbivore removal,
despite trends that suggest the effects of herbivores on
carbon stocks approach zero over prolonged time periods
(i.e. decades). Irrespective of these assumptions, and the
potentially short-term and ephemeral nature of our estimate,
the effects of herbivores on terrestrial carbon stocks are
important to consider.
(4) Herbivore removal should be considered in strategies
that aim to increase terrestrial carbon stocks at local
and regional scales. However, where herbivore removal
conflicts with other ecosystem services, i.e. the conservation
of biodiversity, improved land management may increase
VIII. REFERENCES
Abril, A. & Bucher, E. H. (1999). The effects of overgrazing on soil microbial
community and fertility in the Chaco dry savannas of Argentina. Applied Soil Ecology
12, 159–167.
Akashi, N. & Nakashizuka, T. (1999). Effects of bark-stripping by sika deer (Cervus
nippon) on population dynamics of a mixed forest in Japan. Forest Ecology and
Management 113, 75–82.
Allard, V., Soussana, J., Falcimagne, R., Berbigier, P., Bonnefond, J.,
Ceschia, E., D’hour, P., Hénault, C., Laville, P., Martin, C. & Pinarèspatino, C. (2007). The role of grazing management for the net biome productivity
and greenhouse gas budget (CO2 , N2 O and CH4 ) of semi-natural grassland.
Agriculture, Ecosystems and Environment 121, 47–58.
Andreae, M. O. & Merlet, P. (2001). Emission of trace gases and aerosols from
biomass burning. Global Biogeochemical Cycles 15, 955–966.
Asner, G. P., Elmore, A. J., Olander, L. P., Martin, R. E. & Harris, A. T.
(2004). Grazing systems, ecosystem responses, and global change. Annual Review
of Environment and Resources 29, 261–299.
Asner, G. P., Levick, S. R., Kennedy-Bowdoin, T., Knapp, D. E., Emerson, R.,
Jacobson, J., Colgan, M. S. & Martin, R. E. (2009). Large-scale impacts of
herbivores on the structural diversity of African savannas. Proceedings of the National
Academy of Sciences 106, 4947–4952.
Atjay, G. L., Ketner, P. & Duvigneaud, P. (1979). Terrestrial primary production
and phytomass. In The global carbon cycle (eds B. Bolin, E. T., Degens, S. Kempe &
P. Ketner), pp. 129–181. John Wiley & Sons, Chichester.
Augustine, D. J. & McNaughton, S. J. (2004). Regulation of shrub dynamics by
native browsing ungulates on East African rangeland. Journal of Applied Ecology 41,
45–58.
Ayres, E., Heath, J., Possell, M., Black, H. I. J., Kerstiens, G. & Bardgett,
R.D. (2004). Tree physiological responses to above-ground herbivory directly
modify below-ground processes of soil carbon and nitrogen cycling. Ecology Letters 7,
469–479.
Bakker, E. S., Olff, H., Boekhoff, M., Gleichman, J. M. & Berendse, F. (2004).
Impact of herbivores on nitrogen cycling: contrasting effects of small and large
species. Oecologia 138, 91–101.
Bakker, E. S., Olff, H. & Gleichman, J. M. (2009). Contrasting effects of large
herbivore grazing on smaller herbivores. Basic and Applied Ecology 10, 141–150.
Baldi, R., Albon, S. D. & Elston, D. A. (2001). Guanacos and sheep: evidence for
continuing competition in arid Patagonia. Oecologia 129, 561–570.
Bardgett, R. D., Frankland, J. C. & Whittaker, J. B. (1993). The effects of
agricultural practices on the soil biota of some upland grasslands. Agriculture, Ecosystems
and Environment 45, 25–45.
Bardgett, R. D. & Wardle, D. A. (2003). Herbivore-mediated linkages between
aboveground and belowground communities. Ecology 84, 2258–2268.
Bardgett, R. D., Wardle, D. A. & Yeates, G. W. (1998). Linking above-ground
and below-ground interactions: how plant responses to foliar herbivory influence
soil organisms. Soil Biology and Biochemistry 30, 1867–1878.
Barford, C. C., Wofsy, S. C., Goulden, M. L., Munger, J. W., Pyle, E. H.,
Urbanski, S. P., Hutyra, L., Saleska, S. R., Fitzjarrald, D. & Moore, K.
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
Carbon storage in terrestrial ecosystems: do browsing and grazing herbivores matter?
(2001). Factors controlling long- and short-term sequestration of atmospheric CO2
in a mid-latitude forest. Science 294, 1688–1691.
Bazely, D. R. & Jefferies, R. L. (1985). Goose feces: a source of nitrogen for plant
growth in a grazed salt-marsh. Journal of Applied Ecology 22, 693–703.
Bee, J. N., Kunstler, G. & Coomes, D. A. (2007). Resistance and resilience of New
Zealand tree species to browsing. Journal of Ecology 95, 1014–1026
Bellingham, P. J. & Allan, C. N. (2003). Forest regeneration and the influences of
white-tailed deer (Odocoileus virginianus) in cool temperate New Zealand rain forests.
Forest Ecology and Management 175, 71–86.
Belovsky, G. E. (1997). Optimal foraging and community structure: the allometry of
herbivore food selection and competition. Evolutionary Ecology 11, 641–672.
Belsky, A. J. (1986). Does herbivory benefit plants? A review of the evidence. American
Naturalist 127, 870–892.
Belsky, A. J. (1987). The effects of grazing: confounding of ecosystem, community,
and organism scales. American Naturalist 129, 777–783.
Belsky, A. J. & Blumenthal, D. M. (1997). Effects of livestock grazing on stand
dynamics and soils in upland forests of the Interior West. Conservation Biology 11,
315–327.
Belsky, A. J., Carson, W. P., Jensen, C. L. & Fox, G. A. (1993). Overcompensation by plants: herbivore optimization or red herring? Evolutionary Ecology 7,
109–121.
Berendse, F., Breemen, N. V., Rydin, H., Buttler, A., Heijmans, M., Hoosbeek, M. R., Lee, J. A., Mitchell, E., Saarinen, T., Vasander, H. & Wallén,
B. (2001). Raised atmospheric CO2 levels and increased N deposition cause shifts
in plant species composition and production in Sphagnum bogs. Global Change Biology
7, 591–598.
Berg, W. A., Bradford, J. A. & Sims, P. L. (1997). Long-term soil nitrogen and
vegetation change on Sandhill Rangeland. Journal of Range Management 50,
482–486.
Bergman, M. (2002). Can saliva from moose, Alces alces, affect growth responses in
the sallow, Salix caprea? Oikos 96,164–168.
Beschta, R. L. & Ripple, W. J. (2008). Wolves, trophic cascades, and rivers in the
Olympic National Park, USA. Ecohydrology 1, 118–130.
Beschta, R. L. & Ripple, W. J. (2009). Large predators and trophic cascades in
terrestrial ecosystems of the western United States. Biological Conservation 142,
2401–2414.
Bodmer, R. & Ward, D. (2006). Frugivory in large mammalian herbivores. In Large
herbivore ecology, ecosystem dynamics and conservation. (eds K. Danell, R. Bergström,
P. Duncan & J. Pastor), pp 232–260. Cambridge University Press, Cambridge.
Bodmer, R. E. (1989). Ungulate biomass in relation to feeding strategy within
Amazonian forests. Oecologia 81, 547–550.
Bond, W. J. (2005). Large parts of the world are brown or black: a different view on
the ‘Green World’ hypothesis. Journal of Vegetation Science 16, 261–266.
Bond, W. J. & Keeley, J. E. (2005). Fire as a global ‘herbivore’: the ecology and
evolution of flammable ecosystems. Trends in Ecology & Evolution 20, 387–394.
Brand, M. D. & Goetz, H. (1986). Vegetation of exclosures in southwestern North
Dakota. Journal of Range Management 39, 434–437.
Brashares, J. S., Arcese, P., Sam, M. K., Coppolillo, P. B., Sinclair, A. R. E.
& Balmford, A. (2004). Bushmeat hunting, wildlife declines, and fish supply in
West Africa. Science 306, 1180–1183.
Brathen, K. D., Ims, R. A., Yoccoz, N. G., Fauchald, P., Tveraa, T. &
Hausner, V. H. (2007). Induced shift in ecosystem productivity? Extensive scale
effects of abundant large herbivores. Ecosystems 10, 773–789.
Briggs, J. M., Knapp, A. K., Blair, J. M., Heisler, J. L., Hoch, G. A., Lett, M. S.
& McCarron, J. K. (2005). An ecosystem in transition: causes and consequences
of the conversion of mesic grassland to shrubland. BioScience 55, 243–254.
Brown, J. H. & Heske, E. J. (1990). Control of a desert-grassland transition by a
keystone rodent guild. Science 250, 1705–1707.
Bryant, J. P. & Reichart, P. B. (1992). Controls over secondary metabolite
production in arctic woody plants. In Arctic ecosystems in a changing climate:
an ecophysiological perspective. (eds F. S. Chapin, R. L. Jefferies, J. F. Reynolds,
G. R. Shaver & J. Svoboda), pp. 377–390. Academic Press, New York.
Bryant, N. A., Johnson, L. F., Brazel, A. J., Balling, R. C., Hutchinson, C. F.
& Beck, L. R. (1990). Measuring the effect of overgrazing in the Sonoran Desert.
Climatic Change 17, 243–264.
Bucher, E. H. (1987). Herbivory in arid and semi-arid regions of Argentina. Revista
Chilena de Historia Natural 60, 265–273.
Burney, D. A. & Flannery, T. F. (2005). Fifty millennia of catastrophic extinctions
after human contact. Trends in Ecology & Evolution 20, 395–401.
Butler, L. G. & Kielland, K. (2008). Acceleration of vegetation turnover and
element cycling by mammalian herbivory in riparian ecosystems. Journal of Ecology
96, 136–144.
Calenge, C., Maillard, D., Gaillard, J., Merlot, L. & Peltier, R. (2002).
Elephant damage to trees of wooded savanna in Zakouma National Park, Chad.
Journal of Tropical Ecology 18, 599–614.
Cao, G. M., Tang, Y. H., Mo, W. H., Wang, Y. S., Li, Y. N. & Zhao, X. Q. (2004).
Grazing intensity alters soil respiration in an alpine meadow on the Tibetan plateau.
Soil Biology and Biochemistry 36, 237–243.
89
Carter, A. J. & Scholes, R. J. (2000). Spatial global database of soil properties.
International Geosphere-Biosphere Programme Data Information Systems.
Toulouse.
Caspersen, J. P., Pacala, S. W., Jenkins, J. C., Hurtt, G. C., Moorcroft, P. R.
& Birdsey, R. A. (2000). Contributions of land-use history to carbon accumulation
in U.S. forests. Science 290, 1148–51.
Caughley, G. (1970). Eruption of ungulate populations, with emphasis on Himalayan
Thar in New Zealand. Ecology 51, 53–72.
Chapman, S. K., Hart, S. C., Cobb, N. S., Whitham, T. G. & Koch, G. W.
(2003). Insect herbivory increases litter quality and decomposition: an extension
of the acceleration hypothesis. Ecology 84, 2867–2876.
Chaneton, E. J. & Lavado, R. S. (1996). Soil nutrients and salinity after long-term
grazing exclusion in a flooding Pampa grassland. Journal of Range Management 49,
182–187.
Classen, A., Overby, S., Hart, S., Koch, G. & Whitham, T. (2007). Season
mediates herbivore effects on litter and soil microbial abundance and activity in a
semi-arid woodland. Plant and Soil 295, 217–227.
Coetsee, C., Bond, W. J. & February, E. C. (2010). Frequent fire affects soil
nitrogen and carbon in an African savanna by changing woody cover. Oecologia
162, 1027–1034.
Coley, P. D. & Barone, J. A. (1996). Herbivory and plant defenses in tropical forests.
Annual Review of Ecology and Systematics 27, 305–335.
Coley, P. D., Bryant, J. P. & Chapin, F. S. (1985). Resource availability and plant
antiherbivore defense. Science 230, 895–899.
Committee on Climate Change [CCC]. (2008). Building a low-carbon economy—the
UK’s contribution to tackling climate change. The Stationery Office, Norwich.
Conant, R. T., Paustian, K. & Elliott, E. T. (2001). Grassland management and
conversion into grassland: effects on soil carbon. Ecological Applications 11, 343–355.
Coomes, D. A. & Allen, R. B. (2007). Mortality and tree-size distributions in natural
mixed-age forests. Journal of Ecology 95, 27–40.
Coomes, D. A., Allen, R. B., Forsyth, D. M. & Lee, W. G. (2003). Factors
preventing the recovery of New Zealand forests following control of invasive
deer. Conservation Biology 17, 450–459.
Coomes, D. A., Allen, R. B., Scott, N. A., Goulding, C. & Beets, P. (2002).
Designing systems to monitor carbon stocks in forests and shrublands. Forest Ecology
and Management 164, 89–108.
Cornelissen, J. H. C., Quested, H. M., Gwynn-Jones, D., Van Logtestijn,
R. S. P., De Beus, M. A. H., Kondratchuk, A., Callaghan, T. V. & Aerts, R.
(2004). Leaf digestibility and litter decomposability are related in a wide range of
subarctic plant species and types. Functional Ecology 18, 779–786.
Côté, S. D., Rooney, T. P., Tremblay, J. P., Dussault, C. & Waller, D. M.
(2004). Ecological impacts of deer overabundance. Annual Review of Ecology, Evolution,
and Systematics 35, 113–147.
Coughenour, M. B. (1991). Biomass and nitrogen responses to grazing of upland
steppe on Yellowstone’s northern winter range. Journal of Applied Ecology 28, 71–82.
Coughenour, M. B., McNaughton, S. J. & Wallace, L. L. (1985). Shoot growth
and morphometric analyses of Serengeti graminoids. African Journal of Ecology 23,
179–94.
Crawley, M. J. (1983). Herbivory: the dynamics of animal-plant interactions. Blackwell
Scientific Publications, Oxford.
Cristoffer, C. & Peres, C. A. (2003). Elephants versus butterflies: the ecological
role of large herbivores in the evolutionary history of two tropical worlds. Journal of
Biogeography 30, 1357–1380.
De Vries, W. (2009). Assessment of the relative importance of nitrogen deposition
and climate change on the sequestration of carbon by forests in Europe: an overview
introduction. Forest Ecology and Management 258(S1), vii–x.
De Vries, W., Reinds, G. J., Gundersen, P. & Sterba, H. (2006). The impact of
nitrogen deposition on carbon sequestration in European forests and forest soils.
Global Change Biology 12, 1151–1173.
Denman, K. L., Brasseur, G., Chidthaisong, A., Ciais, P., Cox, P. M., Dickinson, R. E., Hauglustaine, D., Heinze, C., Holland, E., Jacob, D.,
Lohmann, U., Ramachandran, S., Da Silva Dias, P. L., Wofsy, S. C. &
Zhang, X. (2007). Couplings between changes in the climate system and biogeochemistry. In Climate change 2007: the physical science basis. Contribution of Working
Group I to the Fourth Assessment Report of the Intergovernmental Panel on Climate Change.
(eds S. Solomon, D. Qin, M. Manning, Z. Chen, M. Marquis, K. B. Averyt,
M. Tignor & H. L. Miller), pp. 499–588. Cambridge University Press,
Cambridge.
Derner, J., Boutton, T. & Briske, D. (2006). Grazing and ecosystem carbon
storage in the North American Great Plains. Plant and Soil 280, 77–90.
Doescher, P. S., Svejcar, T. J. & Jaindl, R. G. (1997). Gas exchange of Idaho
fescue in response to defoliation and grazing history. Journal of Range Management 50,
285–289.
Dolman, P. M. & Wäber, K. (2008). Ecosystem and competition impacts of
introduced deer. Wildlife Research 35, 202–214.
Dumont, B., Garel, J. P., Ginane, C., Decuq, F., Farruggia, A., Pradel, P.,
Rigolot, C. & Petit, M. (2007). Effect of cattle grazing a species-rich mountain
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
90
pasture under different stocking rates on the dynamics of diet selection and sward
structure. Animal 1, 1042–1052.
Dyer, M. I. (1980). Mammalian epidermal growth-factor promotes plant growth.
Proceedings of the National Academy of Sciences 77, 4836–4837.
Ellis, E. C. & Ramankutty, N. (2008). Putting people in the map: anthropogenic
biomes of the world. Frontiers in Ecology and the Environment 6, 439–447.
Fearnside, P. M. (1995). Potential impacts of climatic change on natural forests and
forestry in Brazilian Amazonia. Forest Ecology and Management 78, 51–70.
Feeley, K. J. & Terborgh, J. W. (2005). The effects of herbivore density on soil
nutrients and tree growth in tropical forest fragments. Ecology 86, 116–124.
Feldpausch, T. R., Rondon, M. A., Fernandes, E. C. M., Riha, S. J. & Wandelli, E. (2004). Carbon and nutrient accumulation in secondary forests regenerating on pastures in central Amazonia. Ecological Applications 14, S164–S176.
Food and Agriculture Organization of the United Nations [FAO]. (2000).
Terrastat: human-induced land degradation due to agricultural activities. Food and Agriculture
Organization of the United Nations, Rome.
Food and Agriculture Organization of the United Nations [FAO]. (2001).
Global forest resources assessment 2000. Food and Agriculture Organization of the United
Nations, Rome.
Food and Agriculture Organization of the United Nations [FAO].
(2008). FAOSTAT. Food and Agriculture Organization of the United Nations,
Rome. http://faostat.fao.org/
Forbes, M. S., Raison, R. J. & Skjemstad, J. O. (2006). Formation, transformation
and transport of black carbon (charcoal) in terrestrial and aquatic ecosystems. Science
of the Total Environment 370, 190–206.
Fornara, D. A. & Du Toit, J. (2008). Browsing-induced effects on leaf litter quality
and decomposition in a southern African savanna. Ecosystems 11, 238–249.
Forster, P., Ramaswamy, V., Artaxo, P., Berntsen, T., Betts, R., Fahey,
D. W., Haywood, J., Lean, J., Lowe, D. C., Myhre, G., Nganga, J., Prinn, R.,
Raga, G., Schulz, M. & Van Dorland, R. (2007). Changes in atmospheric
constituents and in radiative forcing. In Climate change 2007: the physical science basis.
Contribution of Working Group I to the Fourth Assessment Report of the Intergovernmental Panel on
Climate Change. (eds S. Solomon, D. Qin, M. Manning, Z. Chen, M. Marquis,
K. B. Averyt, M. Tignor & H. L. Miller), pp. 129–234. Cambridge University
Press, Cambridge.
Forsyth, D. M. & Caley, P. (2006). Testing the irruptive paradigm of large-herbivore
dynamics. Ecology 87, 297–303.
Forsyth, D. M. & Duncan, R. P. (2001). Propagule size and the relative success
of exotic ungulate and bird introductions to New Zealand. American Naturalist 157,
583–595.
Frank, D. A. (1998). Ungulate regulation of ecosystem processes in Yellowstone
National Park: direct and feedback effects. Wildlife Society Bulletin 26, 410–418.
Frank, D. A. & Groffman, P. M. (1998). Ungulate versus landscape control of soil C
and N processes in grasslands of Yellowstone National Park. Ecology 79, 2229–2241.
Frank, D. A., Kuns, M. M. & Guido, D. R. (2002). Consumer control of grassland
plant production. Ecology 83, 602–606.
Frank, D. A. & McNaughton, S. J. (1993). Evidence for the promotion of
aboveground grassland production by native large herbivores in Yellowstone
National Park. Oecologia 96, 157–161.
Frank, A. B., Tanaka, D. L., Hofmann, L. & Follett, R. F. (1995). Soil carbon
and nitrogen of Northern Great Plains grasslands as influenced by long-term grazing.
Journal of Range Management 48, 470–474.
Frelich, L. E. & Lorimer, G. G. (1985). Current and predicted long-term effects
of deer browsing in hemlock forests in Michigan, USA. Biological Conservation 34,
99–120.
Friedel, M. H. (1991). Range condition assessment and the concept of thresholds a
viewpoint. Journal of Range Management 44, 422–426.
Fritz, H. & Loison, A. (2006). Large herbivores across biomes. In Large herbivore ecology
ecosystem dynamics and conservation. (eds K. Danell, R. Bergström, P. Duncan &
J. Pastor), pp. 19–49. Cambridge University Press, Cambridge.
Frost, C. J. & Hunter, M. D. (2004). Insect canopy herbivory and frass deposition
affect soil nutrient dynamics and export in oak mesocosms. Ecology 85, 3335–3347.
Garel, M., Loison, A., Gaillard, J., Cugnasse, J. & Maillard, D. (2004). The
effects of a severe drought on mouflon lamb survival. Proceedings of the Royal Society of
London Series B-Biological Sciences 271, S471–S473.
Gill, R. A. (2007). Influence of 90 years of protection from grazing on plant and
soil processes in the subalpine of the Wasatch Plateau, USA. Rangeland Ecology and
Management 60, 88–98.
Gordon, I. J., Hester, A. J. & Festa-Bianchet, M. (2004). The management of
wild large herbivores to meet economic, conservation and environmental objectives.
Journal of Applied Ecology 41, 1021–1031.
Gornall, J., Jónsdóttir, I., Woodin, S. & Van Der Wal, R. (2007). Arctic mosses
govern below-ground environment and ecosystem processes. Oecologia 153, 931–941.
Grace, J. (2004). Presidential address: understanding and managing the global carbon
cycle. Journal of Ecology 92, 189–202.
Graz, F. P. (2008). The woody weed encroachment puzzle: gathering pieces.
Ecohydrology 1, 340–348.
Andrew J. Tanentzap and David A. Coomes
Green, R. E., Cornell, S. J., Scharlemann, J. P. W. & Balmford, A. (2005).
Farming and the fate of wild nature. Science 307, 550–555.
Guldemond, R. & Van Aarde, R. (2008). A meta-analysis of the impact of African
elephants on savanna vegetation. Journal of Wildlife Management 72, 892–899.
Guo, L. B. & Gifford, R. M. (2002). Soil carbon stocks and land use change: a meta
analysis. Global Change Biology 8, 345–360.
Hamilton, E. W. & Frank, D. A. (2001). Can plants stimulate soil microbes and their
own nutrient supply? Evidence from a grazing tolerant grass, Ecology 82, 2397–2402.
Hanley, N., Davies, A., Angelopoulos, K., Hamilton, A., Ross, A., Tinch, D.
& Watson, F. (2008). Economic determinants of biodiversity change over a 400-year
period in the Scottish uplands. Journal of Applied Ecology 45, 1557–1565.
Haukioja, E. & Koricheva, J. (2000). Tolerance to herbivory in woody vs.
herbivorous plants. Evolutionary Ecology 14, 551–562.
Hawkes, C. V. & Sullivan, J. J. (2001). The impact of herbivory on plants in
different resource conditions: a meta-analysis. Ecology 82, 2045–2058.
Herms, D. A. & Mattson, W. J. (1992). The dilemma of plants—to grow or defend.
Quarterly Review of Biology 67, 283–335.
Hill, M., Roxburgh, S., Carter, J. & McKeon, G. (2005). Vegetation state change
and consequent carbon dynamics in savanna woodlands of Australia in response to
grazing, drought and fire: a scenario approach using 113 years of synthetic annual
fire and grassland growth. Australian Journal of Botany 53, 715–739.
Hirota, M., Tang, Y., Hu, Q., Kato, T., Hirata, S., Mo, W., Cao, G. &
Mariko, S. (2005). The potential importance of grazing to the fluxes of carbon
dioxide and methane in an alpine wetland on the Qinghai-Tibetan Plateau.
Atmospheric Environment 39, 5255–5259.
Hobbs, N. T. (1996). Modification of ecosystems by ungulates. Journal of Wildlife
Management 60, 695–713.
Hobbs, R. J. & Norton, D. A. (1996). Towards a conceptual framework for
restoration ecology. Restoration Ecology 4, 93–110.
Höft, R. & Höft, M. (1995). The differential effects of elephants on rain forest
communities in the Shimba Hills, Kenya. Biological Conservation 73, 67–79.
Holdo, R. M., Sinclair, A. R. E., Dobson, A. P., Metzger, K. L., Bolker,
B. M., Ritchie, M. E. & Holt, R. D. (2009). A disease-mediated trophic cascade
in the Serengeti and its implications for ecosystem C. PLoS Biology 7, e1000210.
Holland, E. A., Parton, W. J., Detling, J. K. & Coppock, D. L. (1992).
Physiological responses of plant populations to herbivory and their consequences
for ecosystem nutrient flow. American Naturalist 140, 685–706.
Holland, J. N., Cheng, W. & Crossley, D. A. (1996). Herbivore-induced changes
in plant carbon allocation: assessment of below-ground C fluxes using carbon-14.
Oecologia 107, 87–94.
Holland, J. P., Waterhouse, A., Robertson, D. & Pollock, M. L. (2008). Effect
of different grazing management systems on the herbage mass and pasture height
of a Nardus stricta grassland in western Scotland, United Kingdom. Grass and Forage
Science 63, 48–59.
Holling, C. S. (1973). Resilience and stability of ecological systems. Annual Review of
Ecology and Systematics 4, 1–23.
Horner, J., Gosz, J. & Cates, R. (1988). The role of carbon-based plant secondary
metabolites in decomposition in terrestrial ecosystems. American Naturalist 132,
869–883.
Houghton, R. A. (2003). Revised estimates of the annual net flux of carbon to the
atmosphere from changes in land use 1850-2000. Tellus 55B, 378–390.
Houghton, R. A. (2007). Balancing the global carbon budget. Annual of Review Earth
Planetary Sciences 35, 313–47.
Houghton, R. A. & Goodale, C. L. (2004). Effects of land-use change on the
carbon balance of terrestrial ecosystems. In Ecosystems and land use change (eds
R. DeFries, G. Asner & R. A. Houghton), pp. 85–98. American Geophysical
Union, Washington, DC.
Houghton, R. A., Hackler, J. L. & Lawrence, K. T. (1999). The US carbon
budget: contributions from land-use change. Science 285, 574–578.
Houghton, R. A., Hobbie, J. E., Melillo, J. M., More, B., Peterson, B. J.,
Shaver, G. R. & Woodwell, G. M. (1983). Changes in the carbon cycle in
terrestrial biota and soils between 1860 and 1980: a net release of CO2 to the
atmosphere. Ecological Monographs 53, 235–262.
Howe, H. F., Zorn-Arnold, B., Sullivan, A. & Brown, J. S. (2006). Massive and
distinctive effects of meadow voles on grassland vegetation. Ecology 87, 3007–3013.
Hungate, B. A., Holland, E. A., Jackson, R. B., Chapin, F. S., Mooney, H. A.
& Field, C. B. (1997). The fate of carbon in grasslands under carbon dioxide
enrichment. Nature 388, 576–579.
Huntly, N. (1991). Herbivores and the dynamics of communities and ecosystems.
Annual Review of Ecology and Systematics 22, 477–503.
Huston, M. A. & Wolverton, S. (2009). The global distribution of net primary
production: resolving the paradox. Ecological Monographs 79, 343–377.
Ickes, K. (2001). Hyper-abundance of native wild pigs (Sus scrofa) in a lowland
dipterocarp rain forest of peninsular Malaysia. Biotropica 33, 682–690.
Ickes, K., Dewalt, S. & Appanah, S. (2001). Effects of native pigs (Sus scrofa) on
woody understorey vegetation in a Malaysian lowland rain forest. Journal of Tropical
Ecology 17, 191–206.
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
Carbon storage in terrestrial ecosystems: do browsing and grazing herbivores matter?
Ise, T., Dunn, A. L., Wofsy, S. C. & Moorcroft, P. R. (2008). High sensitivity of
peat decomposition to climate change through water-table feedback. Nature Geoscience
1, 763–766.
Jackson, R. B., Banner, J. L., Jobbágy, E. G., Pockman, W. T. & Wall, D. H.
(2002). Ecosystem carbon loss with woody plant invasion of grasslands. Nature 418,
623–626.
Jacobs, S. & Naiman, R. (2008). Large African herbivores decrease herbaceous plant
biomass while increasing plant species richness in a semi-arid savanna toposequence.
Journal of Arid Environments 72, 891–903.
Jandl, R., Lindner, M., Vesterdal, L., Bauwens, B., Baritz, R., Hagedorn, F., Johnson, D. W., Minkkinen, K. & Byrne, K. A. (2007). How strongly
can forest management influence soil carbon sequestration? Geoderma 137, 253–268.
Jaramillo, V. & Detling, J. K. (1988). Grazing history, defoliation, and competition:
effects on shortgrass production and nitrogen accumulation. Ecology 69, 1599–1608.
Jauregui, B. M., Rosa-Garcia, R., Garcia, U., Wallisdevries, M. F.,
Osoro, K. & Celaya, R. (2008). Effects of stocking density and breed of goats
on vegetation and grasshopper occurrence in heathlands. Agriculture, Ecosystems &
Environment 123, 219–224.
Jobbágy, E. G. & Jackson, R. B. (2000). The vertical distribution of soil organic
carbon and its relation to climate and vegetation. Ecological Applications 10, 423–436.
Johnston, K. M. & Schmitz, O. J. (1997). Wildlife and climate change: assessing
the sensitivity of selected species to simulated doubling of atmospheric CO2 . Global
Change Biology 3, 531–544.
Jonard, M., Misson, L. & Ponette, Q. (2006). Long-term thinning effects on the
forest floor and the foliar nutrient status of Norway spruce stands in the Belgian
Ardennes. Canadian Journal of Forest Research 36, 2684–2695.
Junk, W., Brown, M., Campbell, I., Finlayson, M., Gopal, B., Ramberg, L.
& Warner, B. (2006). The comparative biodiversity of seven globally important
wetlands: a synthesis. Aquatic Sciences 68, 400–414.
Kaimowitz, D., Mertens, B., Wunder, S. & Pacheco, P. (2004). Hamburger
connection fuels Amazon destruction. Centre for International Forestry Research, Bogor.
Kapos, V., Ravilious, C., Campbell, A., Dickson, B., Gibbs, H., Hansen, M.,
Lysenko, I., Miles, L., Price, J., Scharlemann, J. P. W. & Trumper, K. (2008).
Carbon and biodiversity: a demonstration atlas. UNEP-WCMC, Cambridge.
Keith, H., Mackey, B. G. & Lindenmayer, D. B. (2009). Re-evaluation of forest
biomass carbon stocks and lessons from the world’s most carbon-dense forests.
Proceedings of the National Academy of Sciences 106, 11635–11640.
Keesing, F. (2000). Cryptic consumers and the ecology of an African savanna.
BioScience 50, 205–215.
Kieft, T. L. (1994). Grazing and plant-canopy effects on semiarid soil microbial
biomass and respiration. Biology and Fertility of Soils 18, 155–162.
Kielland, K. & Bryant, J. P. (1998). Moose herbivory in taiga: effects on
biogeochemistry and vegetation dynamics in primary succession. Oikos 82, 377–383.
Knapp, A. K., Blair, J. M., Briggs, J. M., Collins, S. L., Hartnett, D. C.,
Johnson, L. C. & Towne, E. G. (1999). The keystone role of bison in North
American tallgrass prairie. BioScience 49, 39–50.
Koh, S. (2002). The response of four species of spring flowering perennial herbs to grazing by whitetailed deer in southern Ontario: a multiple-scale evaluation. PhD Thesis, York University,
Toronto.
Kruska, R. L., Reid, R. S., Thornton, P. K., Henninger, N. & Kristjanson,
P. M. (2003). Mapping livestock-oriented agricultural production systems for the
developing world. Agricultural Systems 77, 39–63.
Kunhamu, T., Kumar, B. & Viswanath, S. (2009). Does thinning affect litterfall,
litter decomposition, and associated nutrient release in Acacia mangium stands of
Kerala in peninsular India? Canadian Journal of Forest Research 39, 792–801.
Lal, R. (2002). Carbon sequestration in dryland ecosystems of West Asia and North
Africa. Land Degradation and Development 13, 45–59.
Lal, R. (2003). Offsetting global CO2 emissions by restoration of degraded soils and
intensification of world agriculture and forestry. Land Degradation and Development 14,
309–322.
Lal, R. (2004). Carbon sequestration in dryland ecosystems. Environmental Management
33, 528–544.
Lawes, M. & Chapman, C. (2006). Does the herb Acanthus pubescens and/or elephants
suppress tree regeneration in disturbed Afrotropical forest? Forest Ecology and
Management 221, 278–284.
Laycock, W. A. (1991). Stable states and thresholds of range condition on North
American rangelands: a viewpoint. Journal of Range Management 44, 427–33.
Lecain, D. R., Morgan, J. A., Schuman, G. E., Reeder, J. D. & Hart, R. H.
(2002). Carbon exchange and species composition of grazed pastures and exclosures
in the shortgrass steppe of Colorado. Agriculture, Ecosystems & Environment 93, 421–435.
Lehmann, J., Gaunt, J. & Rondon, M. (2006). Bio-char sequestration in terrestrial
ecosystems: a review. Mitigation and Adaptation Strategies for Global Change 11, 403–427.
Lehner, B. & Döll, P. (2004). Development and validation of a global database of
lakes, reservoirs and wetlands. Journal of Hydrology 296, 1–22.
Leopold, A., Sowls, L. K. & Spencer, D. L. (1947). A survey of over-populated
deer ranges in the United States. Journal of Wildlife Management 11, 162–177.
91
Leriche, H., Leroux, X., Gignoux, J., Tuzet, A., Fritz, H., Abbadie, L. &
Loreau, M. (2001). Which functional processes control the short-term effect of
grazing on net primary production in grasslands? Oecologia 129, 114–124.
Levinsky, I., Skov, F., Svenning, J. & Rahbek, C. (2007). Potential impacts of
climate change on the distributions and diversity patterns of European mammals.
Biodiversity and Conservation 16, 3803–3816.
Lewis, S. L., Lopez-Gonzalez, G., Sonke, B., Affum-Baffoe, K., Baker, T. R.,
Ojo, L. O., Phillips, O. L., Reitsma, J. M., White, L., Comiskey, J. A.,
Djuikouo, M. N., Ewango, C. E. N., Feldpausch, T. R., Hamilton, A. C.,
Gloor, M., Hart, T., Hladik, A., Lloyd, J., Lovett, J. C., Makana, J.,
Malhi, Y., Mbago, F. M., Ndangalasi, H. J., Peacock, J., Peh, K. S.,
Sheil, D., Sunderland, T., Swaine, M. D., Taplin, J., Taylor, D., Thomas,
S. C., Votere, R. & Woll, H. (2009). Increasing carbon storage in intact African
tropical forests. Nature 457, 1003–1006.
Li, S., Asanuma, J., Eugster, W., Kotani, A., Liu, J. J., Urano, T., Oikawa, T.,
Davaa, G., Oyunbaatar, D. & Sugita, M. (2005). Net ecosystem carbon dioxide
exchange over grazed steppe in central Mongolia. Global Change Biology 11,
1941–1955.
Litton, C. M., Raich, J. W. & Ryan, M. G. (2007). Carbon allocation in forest
ecosystems. Global Change Biology 13, 2089–2109.
Loison, A., Jullien, J. & Menaut, P. (1999). Relationship between chamois and
isard survival and variation in global and local climate regimes: contrasting examples
from the Alps and Pyrenees. Ecological Bulletins 47, 126–136.
Lovett, G. M., Canham, C. D., Arthur, M. A., Weathers, K. C. & Fitzhugh,
R. D. (2006). Forest ecosystem responses to exotic pests and pathogens in eastern
North America. BioScience 56, 395.
Ludwig, J. A. & Tongway, D. J. (1995). Desertification in Australia: an eye to grass
roots and landscapes. Environmental Monitoring and Assessment 37, 321–327.
Luyssaert, S., Schulze, E., Borner, A., Knohl, A., Hessenmoller, D.,
Law, B. E., Ciais, P. & Grace, J. (2008). Old-growth forests as global carbon
sinks. Nature 455, 213–215.
Mack, M. C., Schuur, E. A. G., Bret-Harte, M. S., Shaver, G. R. & Chapin,
F. S. (2004). Ecosystem carbon storage in arctic tundra reduced by long-term
nutrient fertilization. Nature 431, 440–443.
Magnani, F., Mencuccini, M., Borghetti, M., Berbigier, P., Berninger, F.,
Delzon, S., Grelle, A., Hari, P., Jarvis, P. G., Kolari, P., Kowalski, A. S.,
Lankreijer, H., Law, B. E., Lindroth, A., Loustau, D., Manca, G., Moncrieff, J. B., Rayment, M., Tedeschi, V., Valentini, R. & Grace, J. (2007).
The human footprint in the carbon cycle of temperate and boreal forests. Nature
447, 849–851.
Maronna, R. A., Martin, D. J. & Yohai, V. J. (2006). Robust statistics: theory and
methods. Wiley, Chichester.
Martin, P. S. (1984). Prehistoric overkill: the global model. In Quaternary extinctions: a
prehistoric revolution. (eds P. S. Martin & R. G. Klein), pp. 354–403. University of
Arizona Press, Tuscon.
Matthews, E. (1997). Global litter production, pools, and turnover times: Estimates
from measurement data and regression models. Journal of Geophysical Research 102,
18771–800.
May, R. M. (1977). Thresholds and breakpoints in ecosystems with a multiplicity of
stable states. Nature 269, 471–477.
McGuire, A. D., Sitch, S., Clein, J. S., Dargaville, R., Esser, G., Foley, J.,
Heimann, M., Joos, F., Kaplan, J., Kicklighter, D. W., Meier, R. A.,
Melillo, J. M., Moore, B., Prentice, I. C., Ramankutty, N., Reichenau, T.,
Schloss, A., Tian, H., Williams, L. J. & Wittenberg, U. (2001). Carbon
balance of the terrestrial biosphere in the twentieth century: analyses of CO2 ,
climate and land-use effects with four process-based ecosystem models. Global
Biogeochemical Cycles 15, 183–206.
McInnes, P. F., Naiman, R. J., Pastor, J. & Cohen, Y. (1992). Effects of moose
browsing on vegetation and litter of the boreal forest, Isle Royale, Michigan, USA.
Ecology 73, 2059–2075.
McMahon, C. & Bradshaw, C. (2008). To catch a buffalo: field immobilisation of
Asian swamp buffalo using etorphine and xylazine. Australian Veterinary Journal 86,
235–241.
McMahon, S. M., Parker, G. G. & Miller, D. R. (2010). Evidence for a recent
increase in forest growth. Proceedings of the National Academy of Sciences 107, 3611–3615.
McNaughton, S. J. (1984). Grazing lawns: animals in herds, plant form, and
coevolution. American Naturalist 124, 863–886.
McNaughton, S. J. (1985). Ecology of a grazing system: the Serengeti. Ecological
Monographs 55, 259–294.
Mcnaughton, S. J. & Georgiadis, N. J. (1986). Ecology of African grazing and
browsing mammals. Annual Review of Ecology and Systematics 17, 39–65.
Mcnaughton, S. J., Oesterheld, M., Frank, D. A. & Williams, K. J. (1989).
Ecosystem-level patterns of primary productivity and herbivory in terrestrial habitats.
Nature 341, 142–144.
McNaughton, S. J., Banyikwa, F. F. & McNaughton, M. M. (1997). Promotion
of the cycling of diet-enhancing nutrients by African grazers. Science 278, 1798–1800.
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
92
Meeuwig, R. O. (1965). Effects of seeding and grazing on infiltration capacity and
soil stability of a subalpine range in central Utah. Journal of Range Management 18,
173–180.
Milchunas, D. G. & Lauenroth, W. K. (1993). Quantitative effects of grazing on
vegetation and soils over a global range of environments. Ecological Monographs 63,
327–366.
Miller, G. H., Fogel, M. L., Magee, J. W., Gagan, M. K., Clarke, S. J. &
Johnson, B. J. (2005). Ecosystem collapse in pleistocene Australia and a human
role in megafaunal extinction. Science 309, 287–290.
Mills, A. J. & Cowling, R. M. (2006). Rate of carbon sequestration at two thicket
restoration sites in the eastern Cape, South Africa. Restoration Ecology 14, 38–49.
Miyaki, M. & Kaji, K. (2009). The dynamics of forest stands affected by sika deer on
Nakanoshima Island—Change of size structure similar to the thinning effect. In Sika
Deer: biology and management of native and introduced populations. (eds D. R. McCullough,
S. Takatsuki & K. Kaji), pp. 181–191. Springer, Tokyo.
Morris, J. T. & Jensen, A. (1998). The carbon balance of grazed and nongrazed Spartina anglica saltmarshes at Skallingen, Denmark. Journal of Ecology 86,
229–242.
Mouillot, F., Narasimha, A., Balkanski, Y., Lamarque, J. & Field, C. B.
(2006). Global carbon emissions from biomass burning in the 20th century. Geophysical
Research Letters 33, L01801.
Mountford, E. P. & Peterken, G. E. (2003). Long-term change and implications
for the management of wood-pastures: experience over 40 years from Denny Wood,
New Forest. Forestry 76, 19–43.
Nadelhoffer, K. J., Emmett, B. A. & Gundersen, P. (1999). Nitrogen deposition
makes a minor contribution to carbon sequestration in temperate forests. Nature 398,
145–148.
Neff, J. C., Townsend, A. R., Gleixner, G., Lehman, S. J., Turnbull, J. &
Bowman, W. D. (2002). Variable effects of nitrogen additions on the stability
and turnover of soil carbon. Nature 419, 915–917.
Nepstad, D., Soares-Filho, B., Merry, F., Moutinho, P., Rodrigues, H. O.,
Bowman, M., Schwartzman, S., Almeida, O. & Rivero, S. (2007). The costs and
benefits of reducing carbon emissions from deforestation and forest degradation in the Brazilian
Amazon. Woods Hole Research Center, Falmouth.
Niemelä, P., Chapin, F., Danell, K. & Bryant, J. (2001). Herbivory mediated
responses of selected boreal forests to climatic change. Climatic Change 48, 427–440.
Nomiya, H., Suzuki, W., Kanzashi, T., Shibata, M., Tanaka, H. &
Nakashizuka, T. (2002). The response of forest floor vegetation and tree
regeneration to deer exclusion and disturbance in a riparian deciduous forest,
central Japan. Plant Ecology 164, 263–276.
Nowacki, G. J. & Abrams, M. D. (2008). The demise of fire and ‘‘mesophication’’ of
forests in the eastern United States. BioScience 58, 123–138.
Nowak, R. M. (1999). Walker’s mammals of the world. John Hopkins University Press,
Baltimore.
Ogutu, J. & Owen-Smith, N. (2003). ENSO, rainfall and temperature influences
on extreme population declines among African savanna ungulates. Ecology Letters 6,
412–419.
Olff, H., Ritchie, M. E. & Prins, H. H. T. (2002). Global environmental controls
of diversity in large herbivores. Nature 415, 901–904.
Olofsson, J., Kitti, H., Rautiainen, P., Stark, S. & Oksanen, L. (2001). Effects
of summer grazing by reindeer on composition of vegetation, productivity and
nitrogen cycling. Ecography 24, 13–24.
Olofsson, J. & Oksanen, L. (2002). Role of litter decomposition for the increased
primary production in areas heavily grazed by reindeer: a litterbag experiment.
Oikos 96, 507–515.
Olofsson, J., Stark, S. & Oksanen, L. (2004). Reindeer influence on ecosystem
processes in the tundra. Oikos 105, 386–96.
Owen, D. F. & Wiegert, R. G. (1981). Mutualism between grasses and grazers: an
evolutionary hypothesis. Oikos 36, 376–378.
Owensby, C. E., Ham, J. M. & Auen, L. M. (2006). Fluxes of CO2 from grazed and
ungrazed tallgrass prairie. Rangeland Ecology and Management 59, 111–127.
Owen-Smith, N. (1987). Pleistocene extinctions: the pivotal role of megaherbivores.
Paleobiology 13, 351–362.
Pacala, S. W., Hurtt, G. C., Baker, D., Peylin, P., Houghton, R. A., Birdsey, R. A., Heath, L., Sundquist, E. T., Stallard, R. F., Ciais, P., Moorcroft, P., Caspersen, J. P., Shevliakova, E., Moore, B., Kohlmaier, G.,
Holland, E., Gloor, M., Harmon, M. E., Fan, S., Sarmiento, J. L., Goodale,
C. L., Schimel, D. & Field, C. B. (2001). Consistent land- and atmosphere-based
U.S. carbon sink estimates. Science 292, 2316–2320.
Paige, K. N. & Whitham, T. G. (1987). Overcompensation in response to
mammalian herbivory: the advantage of being eaten. American Naturalist 129,
407–416.
Paine, C. E. T. & Beck, H. (2007). Seed predation by Neotropical rainforest mammals
increases diversity in seedling recruitment. Ecology 88, 3073–3087.
Palmer, T. M., Stanton, M. L., Young, T. P., Goheen, J. R., Pringle, R. M. &
Karban, R. (2008). Breakdown of an ant-plant mutualism follows the loss of large
herbivores from an African savanna. Science 319, 192–195.
Andrew J. Tanentzap and David A. Coomes
Parish, F., Sirin, A., Charman, D., Joosten, H., Minayeva, T., Silvius, M.,
Stringer, L. (eds) (2008). Assessment on peatlands, biodiversity and climate change: main
report. Global Environment Centre and Wetlands International, Kuala Lumpur.
Pastor, J., Dewey, B., Naiman, R. J., Mcinnes, P. F. & Cohen, Y. (1993). Moose
browsing and soil fertility in the boreal forests of Isle Royale National Park. Ecology
74, 467–480.
Peltzer, D. A., Allen, R. B., Lovett, G. M., Whitehead, D. & Wardle, D. A.
(2010). Effects of biological invasions on forest carbon sequestration. Global Change
Biology 16, 732–746.
Piñeiro, G., Paruelo, J. M., Jobbágy, E. G., Jackson, R. B. & Oesterheld, M.
(2009). Grazing effects on belowground C and N stocks along a network of
cattle exclosures in temperate and subtropical grasslands of South America. Global
Biogeochemical Cycles 23, GB2003.
Piñeiro, G., Paruelo, J. M. & Oesterheld, M. (2006). Potential long-term impacts
of livestock introduction on carbon and nitrogen cycling in grasslands of Southern
South America. Global Change Biology 12, 1267–1284.
Ping, C., Michaelson, G. J., Jorgenson, M. T., Kimble, J. M., Epstein, H.,
Romanovsky, V. E. & Walker, D. A. (2008). High stocks of soil organic carbon
in the North American Arctic region. Nature Geoscience 1, 615–619.
Post, E. & Stenseth, N. (1999). Climatic variability, plant phenology, and northern
ungulates. Ecology 80, 1322–1339.
Prins, H. T. & Van Der Jeugd, H. P. (1993). Herbivore population crashes and
woodland structure in East Africa. Journal of Ecology 81, 305–314.
Pucheta, E., Cabido, M., Díaz, S. & Funes, G. (1998). Floristic composition,
biomass and aboveground net plant production in grazed and protected sites in a
mountain grassland of central Argentina. Acta Oecologica 19, 97–105.
R Development Core Team. (2008). R: a language and environment for statistical computing.
R Foundation for Statistical Computing, Vienna.
Ramankutty, N., Evan, A. T., Monfred, A. C. & Foley, J. A. (2008). Farming the
planet: 1. Geographic distribution of global agricultural lands in the year 2000.
Global Biogeochemical Cycles 22, GB1003.
Rangel-Castro, J. I., Prosser, J. I., Scrimgeour, C. M., Smith, P., Ostle, N.,
Ineson, P., Meharg, A. & Killham, K. (2004). Carbon flow in an upland
grassland: effect of liming on the flux of recently photosynthesized carbon to
rhizosphere soil. Global Change Biology 10, 2100–2108.
Rautiainen, A., Saikku, L. & Kauppi, P. E. (2010). Carbon gains and recovery from
degradation of forest biomass in European Union during 1990–2005. Forest Ecology
and Management 259, 1232–1238.
Reeder, J. D. & Schuman, G. E. (2002). Influence of livestock grazing on C
sequestration in semi-arid mixed grass and short-grass rangelands. Environmental
Pollution 116, 457–463.
Reid, R., Galvin, K. & Kruska, R. (2008). Global significance of extensive grazing
lands and pastoral societies: an introduction. In Fragmentation in semi-arid and
arid landscapes. (eds K. A. Galvin, R. S. Reid, R. H. Behnke & N. T. Hobbs),
pp. 1–24. Springer, Dordrecht.
Rhemtulla, J. M., Mladenoff, D. J. & Clayton, M. K. (2009). Historical forest
baselines reveal potential for continued carbon sequestration. Proceedings of the National
Academy of Sciences 106, 6082–6087.
Rietkerk, M., Ketner, P., Stroosnijder, L. & Prins, H. H. T. (1996). Sahelian
rangeland development: a catastrophe? Journal of Range Management 4, 512–519.
Risch, A. C. & Frank, D. A. (2006). Carbon dioxide fluxes in a spatially and
temporally heterogeneous temperate grassland. Oecologia 147, 291–302.
Robinson, D. (2007). Implications of a large global root biomass for carbon sink
estimates and for soil carbon dynamics. Proceedings of the Royal Society B 274,
2753–2759.
Rooke, T. (2003). Growth responses of a woody species to clipping and goat saliva.
African Journal of Ecology 41, 324–328.
Royo, A. A. & Carson, W. P. (2005). The herb community of a tropical forest in
central Panama: dynamics and impact of mammalian herbivores. Oecologia 145,
66–75.
Saikku, L., Rautiainen, A. & Kauppi, P. E. (2008). The sustainability challenge of
meeting carbon dioxide targets in Europe by 2020. Energy Policy 36, 730–742.
Sankaran, M. & Augustine, D. J. (2004). Large herbivores suppress decomposer
abundance in a semiarid grazing ecosystem. Ecology 85, 1052–1061.
Sarmiento, G. (1992). Adaptive strategies of perennial grasses in South American
savannas. Journal of Vegetation Science 3, 325–336.
Saugier, B., Roy, J. & Mooney, H. A. (2001). Estimations of global terrestrial
productivity: converging toward a single number? In Terrestrial global productivity. (eds
J. Roy, B. Saugier & H. A. Mooney), pp. 543–558. Academic Press, San Diego.
Schlesinger, W. H. & Lichter, J. (2001). Limited carbon storage in soil and litter of
experimental forest plots under increased atmospheric CO2 . Nature 411, 466–469.
Schmitt, C. B., Burgess, N. D., Coad, L., Belokurov, A., Besançon, C.,
Boisrobert, L., Campbell, A., Fish, L., Gliddon, D., Humphries, K., Kapos,
V., Loucks, C., Lysenko, I., Miles, L., Mills, C., Minnemeyer, S., Pistorius,
T., Ravilious, C., Steininger, M. & Winkel, G. (2009). Global analysis of the
protection status of the world’s forests. Biological Conservation 142, 2122–2130.
Schnabel, R. R., Franzluebbers, A. J., Stout, W. L., Sanderson, M. A. &
Stuedemann, J. A. (2001). The effects of pasture management practices. In The
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society
Carbon storage in terrestrial ecosystems: do browsing and grazing herbivores matter?
potential of U.S. grazing lands to sequester carbon and mitigate the greenhouse effect. (eds
R. F. Follett, J. M. Kimble & R. Lal), pp. 291–322. Lewis Publishers, Boca
Raton.
Schuman, G. E., Reeder, J. D., Manley, J. T., Hart, R. H. & Manley, W. A.
(1999). Impact of grazing management on the carbon and nitrogen balance of a
mixed-grass rangeland. Ecological Applications 9, 65–71.
Semmartin, M. & Ghersa, C. M. (2006). Intra-specific changes in plant morphology,
associated with grazing, and effects on litter quality, carbon and nutrient dynamics
during decomposition. Austral Ecology 31, 99–105.
Shannon, G., Druce, D. J., Page, B. R., Eckhardt, H. C., Grant, R. &
Slotow, R. (2008). The utilization of large savanna trees by elephant in southern
Kruger National Park. Journal of Tropical Ecology 24, 281–289.
Shipley, L. A., Gross, J. E., Spalinger, D. E., Hobbs, N. T. & Wunder, B. A.
(1994). The scaling of intake rate in mammalian herbivores. American Naturalist 143,
1055–1082.
Shutou, K. & Nakane, K. (2004). Change in soil carbon cycling for stand
development of Japanese cedar (Cryptomeria japonica) plantations following clearcutting. Ecological Research 19, 233–244.
Silvester, W. B. & Orchard, T. A. (1999). The biology of kauri (Agathis australis) in
New Zealand. I. Production, biomass, carbon storage, and litter fall in four forest
remnants. New Zealand Journal of Botany 37, 553–571.
Sinclair, A. R E. (1979). The eruption of the ruminants. In Serengeti: dynamics
of an ecosystem. (eds A. R. E. Sinclair & M. Norton-Griffiths), pp. 82–103.
University of Chicago Press, Chicago.
Sinclair, A. R. E. & Fryxell, J. M. (1985). The Sahel of Africa: ecology of a disaster.
Canadian Journal of Zoology 63, 987–94.
Singer, F. J. & Schoenecker, K. A. (2003). Do ungulates accelerate or decelerate
nitrogen cycling? Forest Ecology and Management 181, 189–204.
Sinton, P., Seastedt, T., Bennett, B., Bock, C. & Bock, J. (2000). C3 and C4
species changes identified by δ 13 C values of soil organic matter in a Colorado
prairie. Western North American Naturalist 60, 445–450.
Sjögersten, S., Van Der Wal, R. & Woodin, S. J. (2008). Habitat type
determines herbivory controls over CO2 fluxes in a warmer Arctic. Ecology 89,
2103–2116.
Skinner, R. H. (2008). High biomass removal limits carbon sequestration potential of
mature temperate pastures. Journal of Environmental Quality 37, 1319–1326.
Smit, R., Bokdam, J., Den Ouden, J., Olff, H., Schot-Opschoor, H. &
Schrijvers, M. (2001). Effects of introduction and exclusion of large herbivores on
small rodent communities. Plant Ecology 155, 119–127.
Solbrig, O. T. (2006). Economic growth and environmental change. In The Cambridge
economic history of Latin America: the long twentieth century. (eds V. Bulmer-Thomas,
J. H. Coatsworth & R. C. Conde), pp. 329–376. Cambridge University Press,
Cambridge.
Soussana, J. F., Loiseau, P., Vuichar, N., Ceschia, E., Balesdent, J., Chevallier, T. & Arrouays, D. (2004). Carbon cycling and sequestration opportunities
in temperate grasslands. Soil Use and Management 20, 219–230.
Stark, S. & Grellmann, D. (2002). Soil microbial responses to herbivory in an
Arctic tundra heath at two levels of nutrient availability. Ecology 83, 2736–2745.
Steen, H., Mysterud, A. & Austrheim, G. (2005). Sheep grazing and rodent
populations: evidence of negative interactions from a landscape scale experiment.
Oecologia 143, 357–364.
Steinfeld, H., Gerber, P., Wassenaar, T., Castel, V., Rosales, M. &
De Haan, C. (2006). Livestock’s long shadow: environmental issues and options. Food
and Agriculture Organization of the United Nations, Rome.
Stock, W. D., Roux, D. & Heyden, F. (1993). Regrowth and tannin production in
woody and succulent karoo shrubs in response to simulated browsing. Oecologia 96,
562–568.
Strassmann, K., Joos, F. & Fischer, G. (2008). Simulating effects of land use
changes on carbon fluxes: past contributions to atmospheric CO2 increases and
future commitments due to losses of terrestrial sink capacity. Tellus B 60, 583–603.
Susiluoto, S., Rasilo, T., Pumpanen, J. & Berninger, F. (2008). Effects of grazing
on the vegetation structure and carbon dioxide exchange of a Fennoscandian fell
ecosystem. Arctic, Antarctic, and Alpine Research 40, 422–431.
Takatsuki, S. (2009). Effects of sika deer on vegetation in Japan: a review. Biological
Conservation 142, 1922–1929.
Thomas, R. Q., Canham, C. D., Weathers, K. C. & Goodale, C. L. (2009).
Increased tree carbon storage in response to nitrogen deposition in the US. Nature
Geoscience 3, 13–17.
Thuiller, W., Broennimann, O., Hughes, G., Alkemade, J. R. M., Midgley,
G. F. & Corsi, F. (2006). Vulnerability of African mammals to anthropogenic
climate change under conservative land transformation assumptions. Global Change
Biology 12, 424–440.
Tolvanen, A. & Laine, K. (1997). Effects of reproduction and artificial herbivory
on vegetative growth and resource levels in deciduous and evergreen dwarf shrubs.
Canadian Journal of Botany 75, 656–666.
Towne, E. G., Hartnett, D. C. & Cochran, R. C. (2005). Vegetation trends
in tallgrass prairie from bison and cattle grazing. Ecological Applications 15,
1550–1559.
93
Tremblay, J. P., Huot, J. & Potvin, F. (2006). Divergent nonlinear responses of the
boreal forest field layer along an experimental gradient of deer densities. Oecologia
150, 78–88.
Trumbore, S. (2006). Carbon respired by terrestrial ecosystems—recent progress
and challenges. Global Change Biology 12, 141–153.
Ung, C. H., Bernier, P. & Guo, X. J. (2008). Canadian national biomass equations:
new parameter estimates that include British Columbia data. Canadian Journal of
Forest Research 38, 1123–1132.
Van Der Wal, R. (2006). Do herbivores cause habitat degradation or vegetation state
transition? Evidence from the tundra. Oikos 114, 177–186.
Van Wieren, S. E. (1995). The potential role of large herbivores in nature conservation
and extensive land use in Europe. Biological Journal of the Linnean Society 56(S1), 11–23.
Vera, F. W. M. (2002). Grazing ecology and forest history. CABI Publishing, Wallingford.
Virtanen, R., Salminen, J. & Strömmer, R. (2008). Soil and decomposer responses
to grazing exclusion are weak in mountain snow-beds. Acta Oecologica 33, 207–212.
Vors, L. S. & Boyce, M. S. (2009). Global declines of caribou and reindeer. Global
Change Biology 15, 2626–2633.
Waldram, M. S., Bond, W. J. & Stock, W. D. (2008). Ecological engineering by
a mega-grazer: white rhino impacts on a South African savanna. Ecosystems 11,
101–112.
Ward, S. E., Bardgett, R. D., McNamara, N. P., Adamson, J. K & Ostle, N. J.
(2007). Long-term consequences of grazing and burning on northern peatland
carbon dynamics. Ecosystems 10, 1069–1083.
Wardle, D. A., Barker, M. B., Yeates, G. W., Bonner, K. I. & Ghani, A. (2001).
Introduced browsing mammals in New Zealand natural forests: aboveground and
belowground consequences. Ecological Monographs 71, 587–614.
Wardle, D. A., Bonner, K. I. & Baker, G. M. (2002). Linkages between plant litter
decomposition, litter quality, and vegetation responses to herbivores. Functional
Ecology 16, 585–595.
Watt, A. S. (1947). Pattern and process in the plant community. Journal of Ecology 35,
1–22.
Webb, S. D. (1978). A history of savanna vertebrates in the New World. Part II:
South America and the Great Interchange. Annual Review of Ecology and Systematics 9,
393–426.
Weber, P., Rigling, A. & Bugmann, H. (2008). Investigating the impact of grazing
on the dynamics of mixed Pinus sylvestris and Quercus pubescens stands by applying a
forest gap model. Ecological Modelling 210, 301–311.
Weinhold, B. J., Hendrickson, J. R. & Karn, J. F. (2001). Pasture management
influences on soil properties in the Northern Great Plains. Journal of Soil and Water
Conservation 56, 27–31.
Welker, J. M., Fahnestock, J. T., Povirk, K. L., Bilbrough, C. J. & Piper,
R. E. (2004). Alpine grassland CO2 exchange and nitrogen cycling: grazing history
effects, Medicine Bow Range, Wyoming, USA. Arctic Antarctic and Alpine Research 36,
11–20.
Westoby, M., Walker, B. & Noy-Meir, I. (1989). Opportunistic management for
rangelands not at equilibrium. Journal of Range Management 42, 266–74.
White, R. P., Murray, S. & Rohweder, M. (2000). Pilot analysis of global ecosystems:
grassland ecosystems. World Resources Institute, Washington, DC.
Wilsey, B. J., Parent, G., Roulet, N. T., Moore, T. R. & Potvin, C. (2002).
Tropical pasture carbon cycling: relationships between C source/sink strength,
above-ground biomass and grazing. Ecology Letters 5, 367–376.
Wissenschaftlicher
Beirat
der
Bundesregierung
Globale
Umweltveränerungen [WBGU]. (1988). Die Anrechnung biolischer Quellen und
Senken in Kyoto-Protokoll: Fortschritt oder Rückschlang für den globalen Umweltschutz
Sondergutachten 1988. Wissenschaftlicher Beirat der Bundesregierung Globale
Umweltveränerungen, Bremerhaven.
Wolf, E. C., Cooper, D. J. & Hobbs, N. T. (2007). Hydrologic regime and herbivory
stabilize an alternative state in Yellowstone National Park. Ecological Applications 17,
1572–1587.
Wood, M. K. & Blackburn, W. H. (1981). Grazing systems: their influence on
infiltration rates in the rolling plains of Texas. Journal of Range Management 34,
331–335.
Woodbury, P. B., Heath, L. S. & Smith, J. E. (2006). Land use change effects on
forest carbon cycling throughout the southern United States. Journal of Environmental
Quality 35, 1348–1363.
Wright, D. M. (2009). Impact of deer invasion on New Zealand forests. PhD Thesis,
University of Cambridge, Cambridge.
Wright, S., Zeballos, H., Dominguez, I., Gallardo, M., Moreno, M. &
Ibanez, R. (2000). Poachers alter mammal abundance, seed dispersal, and seed
predation in a Neotropical forest. Conservation Biology 14, 227–239.
Zhang, C., Tian, H., Pan, S., Liu, M., Lockaby, G., Schilling, E. &
Stanturf, J. (2008). Effects of forest regrowth and urbanization on ecosystem
carbon storage in a rural–urban gradient in the southeastern United States.
Ecosystems 11, 1211–1222.
Zhao, H., Zhao, X., Zhou, R., Zhang, T. & Drake, S. (2005). Desertification
processes due to heavy grazing in sandy rangeland, Inner Mongolia. Journal of Arid
Environments 62, 309–319.
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94
IX. SUPPORTING INFORMATION
Andrew J. Tanentzap and David A. Coomes
Table S2. Effects of herbivores on soil carbon stocks.
Supporting references.
Additional supporting information may be found in the
online version of this article.
Supporting methods.
Table S1. Effects of herbivores on terrestrial above-ground
carbon stocks through consumption of plant biomass in
temperate, tropical and Arctic regions.
Please note: Wiley-Blackwell are not responsible for the
content or functionality of any supporting materials supplied
by the authors. Any queries (other than missing material)
should be directed to the corresponding author for the
article.
(Received 30 July 2010; revised 9 April 2011; accepted 10 May 2011; published online 2 June 2011)
Biological Reviews 87 (2012) 72–94 © 2011 The Authors. Biological Reviews © 2011 Cambridge Philosophical Society