SYNTHESES The effects of livestock grazing on biodiversity are multi-
trophic: a meta-analysis
Margarete A. Dettlaff,
Jessica Grenke, Tan Bao,
Isaac Peetoom Heida and
James F. Cahill Jr
Department of Biological Sciences,
University of Alberta, Edmonton,
AB T6G 2E9,Canada
*Correspondence: E-mail: ﬁlaz-
The peer review history for this arti-
cle is available at https://publons.-
Anthropogenic disturbance has generated a signiﬁcant loss of biodiversity worldwide and grazing
by domestic herbivores is a contributing disturbance. Although the effects of grazing on plants
are commonly explored, here we address the potential multi-trophic effects on animal biodiversity
(e.g. herbivores, pollinators and predators). We conducted a meta-analysis on 109 independent
studies that tested the response of animals or plants to livestock grazing relative to livestock
excluded. Across all animals, livestock exclusion increased abundance and diversity, but these
effects were greatest for trophic levels directly dependent on plants, such as herbivores and polli-
nators. Detritivores were the only trophic level whose abundance decreased with livestock exclu-
sion. We also found that the number of years since livestock was excluded inﬂuenced the
community and that the effects of grazer exclusion on animal diversity were strongest in temper-
ate climates. These ﬁndings synthesise the effects of livestock grazing beyond plants and demon-
strate the indirect impacts of livestock grazing on multiple trophic levels in the animal
community. We identiﬁed the potentially long-term impacts that livestock grazing can have on
lower trophic levels and consequences for biological conservation. We also highlight the poten-
tially inevitable cost to global biodiversity from livestock grazing that must be balanced against
Conservation, disturbance, exclosure, domestic grazers, global, graze, herbivory, trophic cascade.
Ecology Letters (2020)
Global change from human disturbance has created unprece-
dented loss of biodiversity and degradation of natural habitats
worldwide (Wake & Vredenburg 2008; Barnosky et al. 2012).
Food production can be especially impactful to biodiversity
through native land conversion into croplands or rangelands
(Barnes et al. 2009; Machovina et al. 2015). Strategies of land
sharing have been proposed to reconcile food production with
goals of conservation (Phalan et al. 2011; Tscharntke et al.
2012), but the effects of grazing by domestic herbivores on
biodiversity remain contentious (Yuan et al. 2016). Empirical
evidence has identiﬁed livestock grazing to have complex
effects on communities such as promoting invasive species
(Kimball & Schiffman 2003), controlling non-native domi-
nance (Stahlheber & D’Antonio 2013), maintaining plant
diversity via the intermediate disturbance hypothesis (Yuan
et al. 2016), and replacing the functional role of an extirpated
native grazer (Milchunas et al. 1998; Bengtsson et al. 2000;
Kohl et al. 2013). As grazing by domestic livestock currently
occupies 26% of our terrestrial land cover (UN Food & Agri-
culture Organization 2017) much of which is either protected
land or remnant natural habitat (e.g. grazing on public lands
or crown lands), there is a need to better understand the over-
all effects of livestock grazing on ecosystems in relation to
biodiversity to appropriately meet future goals of food secu-
rity and conservation.
The effects of livestock grazers have on plant communities
(Olff & Ritchie 1998; Jones 2000; Stahlheber & D’Antonio
2013; Koerner et al. 2014; T€
alle et al. 2016) and soil-biota
(Bardgett & Wardle 2003; Andriuzzi & Wall 2017; Zhou et al.
2017) are well documented, but the effects on other animal
trophic levels, especially vertebrates, has not been quantita-
tively synthesised (but see van Klink et al. 2015; Graham
et al. 2019). There have been repeated and recent calls in ecol-
ogy to include multi-trophic approaches in community analy-
ses (Fraser et al. 2015; Soliveres et al. 2016; Seibold et al.
2018). Understanding multi-trophic interactions are important
for the management of biodiversity because different organ-
isms are expected to respond differently to the same distur-
bance events (Fraser et al. 2015; Gossner et al. 2016).
Additionally, factors that alter the composition of one trophic
group can have indirect impacts on other trophic levels, e.g.
bottom-up or top-down effects (Borer et al. 2006; Beschta &
Ripple 2009; Lefcheck et al. 2015). For instance, livestock
have been identiﬁed to have multi-trophic effects on below-
ground biota through consumptions of plant biomass and
non-trophic effects, such as modifying soil characteristics and
dung deposition (Bardgett & Wardle 2003; Andriuzzi & Wall
2017). However, there are likely differences in the response of
aboveground and belowground trophic levels to disturbance
(Gossner et al. 2016). Aboveground animals can directly com-
pete with livestock grazers for similar resources, such as native
grazing ungulates (Voeten & Prins 1999; Mishra et al. 2004)
©2020 John Wiley & Sons Ltd/CNRS
Ecology Letters, (2020) doi: 10.1111/ele.13527
or herbivorous insects (Branson & Vermeire 2016; Pryke et al.
2016), that can extend to higher trophic levels through bot-
tom-up regulation. If the abundance of certain trophic levels
is reduced, there can be loss of consumer species reliant on
these speciﬁc prey items. For instance, a reduction of herbivo-
rous insects can negatively impact bird diversity or abundance
(Wilson et al. 1999) or removal of speciﬁc plant ﬂowers can
extirpate an obligate pollinator. Grazers can also have non-
consumptive effects such as increasing plant turnover or dung
deposition that favours detritivores (Petersen et al. 2004;
Schon et al. 2011). Not all trophic levels or ecosystems are
expected to respond equally to the presence of domestic graz-
ers in an ecosystem. As such, to make proper and targeted
decisions within the realms of biological conservation and
rangeland management, there is a need to better understand
and include multi-trophic interactions when determining graz-
ing effects on ecosystems.
The effect of grazing on biodiversity patterns can depend
on climate. In areas sensitive to disturbance, even minimal
grazing can signiﬁcantly alter the abundance or diversity of
taxa within the community. For instance, ecosystems that
have high abiotic stress with extremes in precipitation or tem-
perature (e.g. the alpine or deserts) can be particularly
impacted by grazing which damages soil characteristics (e.g.
increase erosion, decrease water inﬁltration), reduces already
limited plant biomass, and decreases animal diversity (Jones
2000; Sankaran & Augustine 2004; Evju et al. 2006). How-
ever, in ecosystems with low stress, grazing can stimulate
plant growth via a compensatory response (McNaughton
et al. 1989; Ramula et al. 2019), resulting in a lower impact
from livestock on the plant community. Previous meta-analy-
ses have identiﬁed precipitation as mediating the effects of
livestock grazing on plants (Herrero-Jauregui & Oesterheld
2018) and nutrient cycling (Zhou et al. 2017), that are likely
to extend to animals in those communities. The effects of live-
stock grazing on community composition are dependent on
temperature or precipitation and thus will vary along environ-
There are many areas globally that were previously grazed
by livestock and can display biological effects of that legacy
(Valone et al. 2002). Livestock grazing can induce changes in
species composition, such as the introduction of exotics (Sink-
ins & Otﬁnowski 2012) or extirpation of species sensitive to
disturbance (Andresen et al. 1990), and these effects can per-
sist years after livestock exclusion. When grazing is frequent,
exclusion may not elicit any signiﬁcant change in composition
because the community is now adapted to grazing (Augustine
et al. 1998; Sasaki et al. 2008; Davies et al. 2016). Therefore,
with the exclusion of domestic grazers the community may
not respond linearly over time, but instead may have thresh-
olds for changes to occur in community composition (Sasaki
et al. 2008). The inclusion of climate and records of previous
grazing are required to effectively quantify the effects of graz-
ing on the trophic structure of an ecosystem.
Here we use a meta-analysis to investigate multi-trophic
impacts of domestic grazing and represent ecological commu-
nities as a food chain model. We synthesised the scientiﬁc lit-
erature on grazing to compare the effects of livestock grazing
on trophic structure. We speciﬁcally compared grazed areas to
those where grazers were excluded and categorised each of the
responding species into a trophic level (detritivores, herbi-
vores, pollinators, omnivores, predators and parasites) based
on main diet. These species were further separated into inver-
tebrates and vertebrates because there are likely different
mechanisms that each can interact with livestock. For exam-
ple, we might expect interference competition with native, ver-
tebrate herbivores and apparent competition with invertebrate
herbivores. Invertebrates also tend to have a narrower dietary
breadth relative to vertebrates. We examined a subset of the
selected studies that included measures of plants, but did not
explicitly look for studies that only tested plants. We excluded
studies that examined soil micro-organisms because there are
previous meta-analyses that explore these groups in detail
(e.g. Bardgett & Wardle 2003; Andriuzzi & Wall 2017; Zhou
et al. 2017) and our intention was to focus on mechanisms
affecting above-ground trophic levels. Using 109 published
manuscripts with relevant data, we compared the abundance
and diversity within vertebrate and invertebrate trophic levels
in grazed and grazer excluded areas. We hypothesise that live-
stock exclusion will increase the abundance and diversity of
all trophic levels because of reduced competition for plant
biomass (herbivores, pollinators anddetritivores) and an
increase in prey items (omnivores and predators). We predict
the effects of livestock exclusion on plants will parallel the
effects on animals. We also predict that the effects of livestock
exclusion on plants and animals will respond nonlinearly
overtime relative to sites that continue to be grazed (i.e. time-
since-exclusion). Finally, we predict that ecosystems with at
extremes in precipitation or temperature (e.g. desert, tundra)
will be the most sensitive to the presence of livestock grazers.
We conducted a literature search using Web of Science for all
peer-reviewed journal articles between 1970 and Nov 2019,
examining the effects of livestock grazers. We selected papers
based upon both response variable suitability and study
design. A study was only included if a response variable was
included that measured some aspect of animal diversity or
abundance. We excluded if a study only included data on
plants, fungi or soil biota. When available, we also extracted
plant response data from included papers, though that was
possible in only 46% of cases.
We used the following search terms in two searches to cap-
ture all studies that compared areas exposed to livestock graz-
ers to areas that were ungrazed: Search 1) graz* OR livestock
AND exclosure* OR exclusion OR exclude* OR ungrazed
OR retire* OR fallow* OR fence* OR paddock*; Search 2)
grazing intensity OR grazing gradient OR stocking rate OR
rotation* grazing. The search included all studies globally on
Web of Science. We did not include any studies that tested
marine grazing, such as molluscs; nor did we include the
effects of livestock grazing on aquatic systems. Using these
search terms and criteria we identiﬁed 3489 published manu-
scripts. These manuscripts were individually screened for rele-
vancy using three criteria 1) a test of livestock grazing relative
©2020 John Wiley & Sons Ltd/CNRS
2A. F. Filazzola et al. Review and Syntheses
to a livestock excluded area, 2) a measured response species
being an animal and 3) there was empirical data that could be
extracted or obtained from the author. We removed dupli-
cates and any study that did not ﬁt these criteria, such as
reviews, policy documents, and manuscripts on grazer health
(e.g. cattle antibiotics, sheep shearing methods). We removed
studies that provided indirect estimates of grazers being
absent, such as biomass of plants, evidence of defoliation or
distance to water source. We were able to identify 109 studies
that matched these requirements. An individual not involved
with the initial screening validated a small subset of studies to
ensure that our criteria for inclusion and exclusion were effec-
tive (Cote et al. 2013).
The 109 selected studies were reviewed to determine the
reported species or functional groups, the response variable
measured (e.g. abundance or diversity) and species of the live-
stock grazer. We obtained details about the experimental
design from each study including the GPS coordinates, the
time-since-exclosures were established and the number of sites
used. There were few studies that tested domestic grazers
other than sheep or cattle in isolation. Therefore, we reﬁned
the studies to include livestock grazing by cattle and sheep
either in isolation, together, or in the presence of other
domesticated species (e.g. horses, goats and donkeys). Each
responding species or functional group was categorised using
available literature into its respective vertebrate class and
trophic level. We separated by vertebrate class because there
may be signiﬁcant differences in the mechanisms that grazing
affects species within each of these groups. In instances where
the description of the examined taxa was too coarse to deter-
mine trophic level (e.g. coleoptera, birds), we excluded these
comparisons from further analysis. A list of the most common
taxa found in each trophic level can be found in Table 1. The
response variables of each study were summarised into abun-
dance or diversity. To prevent pseudo-replication, where
authors reported multiple responses that included the same
sampled individuals, we selected the coarsest taxonomical
response estimate that retained information about trophic
position. For example, if a study included abundance of all
butterﬂies, abundance of polyphagous butterﬂies and abun-
dance of specialist butterﬂies, we only included abundance of
all butterﬂies. A full list of the original reported response vari-
ables and our categorisation of the response variables can be
found in Appendix A. We were able to identify 109 published
manuscripts that included 205 independent pair-wise compar-
isons of livestock grazed and livestock excluded areas. A com-
parison was treated as independent if it was from a unique
study, tested a speciﬁc trophic level, and measured either
abundance or diversity. These comparisons were categorised
into tests of plant abundance (n=40), plant diversity
(n=14), animal abundance (n=122) and animal diversity
(n=29). A workﬂow and rationale for manuscripts that were
excluded can be found in Appendix B. For each of the 205
comparisons, we extracted the means, standard deviations and
number of replicates. In studies where the raw data were pro-
vided, these statistics were calculated.
We conducted a meta-analysis to compare the effects of graz-
ing on different trophic levels using data that were extracted
from relevant studies. We followed the methodology described
by Koricheva et al. (2013) that provides a clear workﬂow
including data aggregation, calculating effect sizes, and con-
ducting statistical models. Each of the 205 comparisons was
categorised based on vertebrate class, trophic level and
response variable (i.e. abundance or diversity). To contrast
livestock grazing and livestock excluded areas, we calculated
the log-transformed ratio of means using eqn 1 (Lajeunesse
2011) as an effect size for each unique comparison (function
escalc, package metafor) (Viechtbauer 2010). We used this log-
transformed response ratio because the value is symmetrical
around zero (Viechtbauer 2010) and is commonly used in
ecology (Lajeunesse 2011).
Positive values of LRR indicate livestock exclusion increases
biodiversity relative to grazed areas. Negative values of LRR
indicate livestock grazing increases biodiversity relative to an
area where livestock grazers were excluded. We recognise an
alternative interpretation exists, such that positive values of
LRR would indicate grazing increases biodiversity and nega-
tive values indicate grazer exclusion increases biodiversity.
The only difference between these is whether an individual
view grazing or its exclusion as the active treatment. Most of
our studies (>90%) previously had livestock grazers that were
excluded rather than sites where grazing never occurred.
Therefore, we chose to frame the control as grazing and the
treatment as grazers excluded.
LRR ¼ln Xlivestock excluded
To compare the effects of livestock exclusion on different
trophic levels, we ﬁt meta-analytic random and mixed models
with the effect sizes determined from each study. To separate
the direct effects of grazing on plants from the indirect effects
on the animal community, we ran separate models for plants
and animals. We identiﬁed 54 unique effect sizes that compared
plant abundance or diversity in livestock grazed and livestock
excluded areas. We used random-effects models with the effect
sizes as the response variable and corresponding sampling
Table 1 Dominant taxa observed in each of the trophic levels separated
by vertebrate class
Trophic level Common taxa
Detritivores Collembola, Coleoptera, Diptera, Formicidae, Isoptera,
Herbivores Coleoptera, Hemiptera, Gastropoda
Parasitic Acari, Culicadae, Nematoda, Thysanoptera
Pollinator Coleoptera, Diptera, Hymenoptera, Lepidoptera
Predator Araneae, Coleoptera
Herbivores Aves, Lagomorpha, Macropodidae, Rodenta, Equidae,
Omnivores Aves, Squamata, Testudines
Predator Amphibians, Aves, Dasyuromorphia, Carnivora
©2020 John Wiley & Sons Ltd/CNRS
Review and Syntheses Livestock can reduce animal biodiversity 3
variances as the error term that includes inverse-variance
weights. We ran separate random-effects models for plant
abundance (40 effect sizes) and diversity (14 effect sizes). Ran-
dom-effects models were used because they account for differ-
ences in study methodologies and assumes the selected
comparisons are a random subset of a large population of stud-
ies conducting similar comparisons (Viechtbauer 2010). To
compare the effects of grazing on animal trophic level, we ran
random-effect models for animal abundance (122 effect sizes)
and diversity (29 effect sizes). To determine if the response to
grazing varied by trophic level, we also ran mixed-effects mod-
els with trophic level and vertebrate class of the species exam-
ined treated as a ﬁxed effect. We tested the effects of grazing
across trophic levels by comparing the effects of livestock exclu-
sion on plants and animals. We ﬁt mixed-effects models with
animal abundance and diversity as the response variables and
plant abundance or diversity as the predictor variables. To
ensure there were no differences in the effects of grazing among
grazer composition (e.g. cattle, sheep, both), we also conducted
a mixed-effects model with the response ratio as the response
variable and grazer as the moderator.
We examined the potential for bottom-up effects in which
the effects of livestock exclusion on plants had indirect effects
on animal abundance or diversity. We ran mixed-effect mod-
els with the effect size of plants as the ﬁxed effect and the
effect size of animals as the response variable. Four models
were run to compare plant abundance or diversity with ani-
mal abundance or diversity, orthogonally crossed. In these
models, a linear increase would suggest effects of livestock
exclusion on plants are paralleled in animals, no effect would
suggest effects on animals are independent of effects on
plants, and a linear decrease would suggest animals respond
inversely to changes in plants from livestock grazing.
We tested whether the effects of livestock exclusion on
plants and animals increased over time by comparing the
effect size against years since grazers were excluded. We ran
mixed-effects models with time-since-exclusion as the ﬁxed
effect and the effect sizes of livestock exclusion on plant abun-
dance (n=34), plant diversity (n=13), animal abundance
(n=107) and animal diversity (n=26) as the responses.
We tested whether the effects of livestock exclusion on the
community was mediated by climate by comparing the effect
sizes to temperature and precipitation. We used the GPS coor-
dinates associated with each study to extract the mean annual
temperature and precipitation during the duration of the
experiment from the Climatic Research Unit (CRU) (http://
www.cru.uea.ac.uk/data). We selected mean annual tempera-
ture and precipitation because these variables are frequently
used for predicting ecosystem productivity (e.g. Del Grosso
et al. 2008) and the effects on grazing (e.g. Herrero-Jauregui
& Oesterheld 2018). These variables were downloaded to the
30-arc second (c. 1 km) spatial resolution. We ﬁt mixed-effect
models for the effects of grazing on plant abundance, plant
diversity, animal abundance and animal diversity with mean
annual precipitation and mean annual temperature as the
ﬁxed effects. We expected the effects of livestock exclusion to
decrease linearly with increasing temperature and precipitation
as abiotic stress is reduced. However, we included polynomials
because some previous research has suggested the effects of
grazing to be nonlinear with changes in climate (Herrero-Jau-
regui & Oesterheld 2018). We used a second-order polynomial
ﬁt for the analyses with time-since-exclusion and climate when
it explained greater variation (R
value increase of 0.10 or
more) and had a lower AICc than a linear model.
The above models were calculated using restricted maxi-
mum-likelihood estimations (REML) because they are unbi-
ased to estimations of population heterogeneity relative to
other methods such as maximum likelihood (Viechtbauer
2005). Conﬁdence intervals for models were calculated using
Wald-type intervals. The amount of residual heterogeneity
generated from these models was obtained iteratively through
the Q-statistics method (Hartung & Knapp 2001). To deter-
mine if there were publication biases towards certain results
and absence of heterogeneity (Egger et al. 1997), we generated
funnel plots for all the mixed-effect models. We did not ﬁnd
publication biases within our tests because our funnel plots
were observed to be symmetrical that was conﬁrmed using a
regression test recommended by Egger et al. (1997). Tests of
publication bias within models can be seen in Appendix C.
All analyses and data aggregation were conducted in R Ver-
sion 3.4.3 (R Core Team 2019) and an open workﬂow can be
found online at (https://github.com/afilazzola/grazing.meta).
Trends in grazing literature
Experiments of grazing exclusion on animals were conducted
globally in 20 countries and on every continent except Antarc-
tica (Appendix E). However, there were biases in the global
distribution of studies with most experiments conducted in
Europe, North America and Australia (72%). Using the Kop-
pen-Geiger climate classiﬁcation, we identiﬁed that studies
were distributed in a large range of climates (i.e. from deserts
to tundra), but most were conducted in Cfb –temperate
humid warm (24%), BSk –arid summer dry cold (20%) and
Dfb/Dfc –warm summer (20%). In studies that tested live-
stock grazing with non-domestic grazers co-occurring in the
same ecosystem, the response of the native grazer was infre-
quently measured (8%). Most studies focused on examining
the effects of arthropods (69%) or small mammal composition
(47%). Other commonly examined taxa included reptiles
(13%) and birds (20%). Of the studies that measured animals,
46% of the studies had also measured plants.
Grazing effects on multiple trophic levels
Across all studies, the exclusion of livestock increased plant
abundance (mean effect SE =0.43 0.09; t
=96.1%) but had no effect on plant diversity
(mean effect SE =0.11 0.06; t
=67.1%). We found no relevant heterogeneity between
studies in both the meta-analysis on plant abundance and
plant diversity (Appendix E).
Across all animals, livestock exclusion increased abundance
(mean effect SE =0.21 0.07; t
=92.8) and diversity (mean effect SE =0.18 0.05;
=3.59, p<0.001; I
=72.4). The inclusion of trophic level
(e.g. invertebrate herbivore, invertebrate pollinator) explained a
©2020 John Wiley & Sons Ltd/CNRS
4A. F. Filazzola et al. Review and Syntheses
considerable amount of variability for mixed effects models of
=3.0; P=0.003) and diversity
=3.63; P<0.008). Livestock exclusion signiﬁ-
cantly increased the abundance of vertebrate herbivores and
vertebrate predators (Fig. 1). Livestock exclusion also signiﬁ-
cantly increased the diversity of invertebrate pollinators and
invertebrate herbivores (Fig. 1). The abundance of invertebrate
detritivores was the only animal group that signiﬁcantly
declined with grazer exclusion (Fig. 1). The diversity of omni-
vore and predator trophic levels were not impacted by grazing
for either vertebrate class (Fig. 1). We found no relevant
heterogeneity between studies in both the meta-analysis on ani-
mal abundance and animal diversity (Appendix E). The differ-
ent types of grazer assemblages (e.g. cattle, sheep, cattle and
sheep) caused no signiﬁcant difference in the effects of grazing
on plant abundance (Q
=3.99, P=0.41), plant diversity
=2.93, P=0.57), animal abundance (Q
P=0.42) or animal diversity (Q
We found evidence of potential bottom-up effects of grazing
as livestock exclusion effects on plant diversity were correlated
with exclusion effects on animal abundance (t
P=0.003; Fig. 2). This suggests livestock caused decreases or
increases in plant diversity result in the same response for
animal abundance (Fig. 2). We did not observe a pattern
when comparing plant abundance –animal abundance
=1.39, P=0.17), plant abundance –animal diversity
=0.02, P=0.98), or plant diversity –animal diversity
=0.08, P=0.94). However, comparisons with animal
diversity had relatively small sample sizes (n=8;
Effect of time-since-exclusion
The effects of grazing on animal metrics persisted for years
after exclusion of grazers from control sites (Fig. 3). However,
time-since-exclusion had no effect on plant abundance
=0.60, P=0.44) or plant diversity (t
Fig. 3). With longer time-since-exclusion, animal abundance
increased when compared to a site that was grazed (mean
effect SE =1.81 0.68; t
=2.66, P=0.008; Fig. 3). The
effect of increasing time-since-exclusion on animal diversity
was nonlinear with signiﬁcantly greater diversity relative to
grazed sites at intermediate timeframes (between 5 and
30 years), but not signiﬁcantly different at shorter or longer
=2.00, P=0.057; Fig. 3).
Figure 1 Effect sizes (LRR) of livestock exclusion on different trophic levels separated by vertebrate class for abundance (left) and diversity (right) of
respective taxa. Mean effect sizes that are signiﬁcantly different from zero are denote by asterisk (*** <0.001, ** <0.01, * <0.05). The number of
comparisons per trophic level (i.e. sample size) is in parentheses beneath the effect size. Positive values indicate grazing sites have higher measured values
and negative values indicate grazing sites have lower measured values. Conﬁdence intervals, test statistics and p-values can be found in Appendix F.
©2020 John Wiley & Sons Ltd/CNRS
Review and Syntheses Livestock can reduce animal biodiversity 5
Figure 3 Effect over time of excluding livestock grazers relative to a grazed site on plant/animal abundance and diversity. Each point represents mean effect
size calculated from a study using the log-response ratio. Solid line represents mean model ﬁt for meta-regression of time since grazed on the effect of
grazing for animal diversity and shaded areas are 95% conﬁdence intervals. Values above the dashed line represent livestock excluded sites have higher
animal abundance or diversity relative to sites that were continuously grazed. The taxa surveyed within each study are represented by circles for
invertebrates and triangles for vertebrate.
–0.4 –0.2 0.0 0.2
Effect of livestock exclusion on plant diversity (LRR)
Effect of livestock exclusion on animal abundance (LRR)
Figure 2 Effect of livestock exclusion on plant diversity was found to increase linearly with the effects on animal abundance (t
=3.63, P=0.003). The
taxa surveyed within each study are represented by circles for invertebrates and triangles for vertebrate. Models that compared the other combinations of
animal abundance, animal diversity, plant abundance and plant diversity were not signiﬁcant and are presented in Appendix G.
©2020 John Wiley & Sons Ltd/CNRS
6A. F. Filazzola et al. Review and Syntheses
Local climate effects on grazing
The effect of livestock exclusion on plant abundance was
independent of mean annual temperature (t
P=0.32), total annual precipitation (t
=1.14, P=0.25) and
their interaction (t
=0.72, P=0.47). The effect of livestock
exclusion on plant diversity also did not change with tempera-
=0.23, P=0.81), precipitation (z
P=0.13) or their interaction (z
Livestock exclusion increased animal abundance more in
warmer climates (mean effect SE =1.50 0.76;
=1.97, P=0.051), but had no effect with precipitation
=1.60, P=0.11) or any interaction (t
P=0.15). The effect of livestock exclusion on animal diversity
had a unimodal relationship with mean annual temperature
=2.41, P=0.025; Fig. 4), where exclusion increased ani-
mal diversity the most at average annual temperatures
between 5 and 15 °C. Precipitation did not mediate the effect
of livestock grazing on animal diversity (t
and there was no interaction with temperature (t
Livestock grazing can be a signiﬁcant driver of global biodi-
versity patterns by affecting multiple trophic levels (Fig. 1).
Across all animals, we found excluding livestock increased
animal abundance and diversity that was likely driven by
strong effects on trophic levels directly dependent on plants
(i.e. herbivores and pollinators; Fig. 2). Detritivores were the
only trophic level that decreased in abundance with livestock
exclusion. We found evidence that the number of years since
livestock were excluded affected the animal community
(Fig. 3), indicating that our study sites were able to recover
from livestock grazing over time. We also observed the effects
of livestock exclusion to be most pronounced in relatively
mild climates, with decreasing effects at temperature extremes
(Fig. 4). Our results suggest that domestic grazers can inﬂu-
ence the trophic structure of ecological communities, and
these effects can persist after their removal.
Grazing effects on trophic levels
We found evidence that livestock exclusion increased the
abundance of plants and over time, plant diversity. Livestock
grazers were expected to decrease the abundance of the plant
community because indigenous grazers are rarely present at
the same density as livestock domestic grazers (Steuter &
Hidinger 1999; du Toit et al. 2017). However, the relationship
between plant diversity and grazing remains more mixed. For
instance, stocking rate has been identiﬁed to decrease the
number of plant species in a community, but this relationship
dissolves when using more complex measures of diversity,
such as those that include relative plant abundances (Herrero-
Jauregui & Oesterheld 2018). The effects of herbivores on
plant diversity are also largely dependent on the impact for
the dominant plant species, which if palatable tend to increase
biodiversity or if herbivore resistant decrease biodiversity
(Koerner et al. 2018). Other meta-analyses have supported the
idea that the effect of grazing on the plant community is
highly species speciﬁc, beneﬁting certain native species and
grasses, but negatively affecting exotics and ﬂowering forbs
(Stahlheber & D’Antonio 2013; T€
alle et al. 2016). Although
not tested here, there are likely signiﬁcant changes in the com-
position of plants occurring with increases in plant diversity
worthy of further investigation.
0 5 10 15 20
Mean annual temperature (°C)
Effect of livestock exclusion (LRR)
50 100 150
Mean annual precipitation (mm)
Figure 4 Relative difference between the diversity of animals on livestock grazed and livestock excluded sites was related to mean annual temperature, but
not mean annual precipitation. Each point represents mean effect size calculated from a study using the log-response ratio. Solid line represents mean
model ﬁt for meta-regression of temperature on the effect of grazing for animal diversity and shaded areas are 95% conﬁdence intervals. Values below the
dashed line represent livestock excluded sites have higher animal diversity relative to grazed sites. The taxa surveyed within each study are represented by
circles for invertebrates and triangles for vertebrate.
©2020 John Wiley & Sons Ltd/CNRS
Review and Syntheses Livestock can reduce animal biodiversity 7
As we predicted, domestic grazers impacted the trophic
levels directly reliant on the plant community (Fig. 1). We
observed a decline in vertebrate herbivore abundance and
invertebrate herbivore diversity when livestock grazers were
present, likely because of competition for food resources. The
vertebrate herbivores that were most commonly tested (89%)
for effects of grazing exclusion were rodents and lagomorphs
(e.g. rabbits, hares and pikas). Small mammals are particu-
larly affected by livestock grazing because trampling can dam-
age burrows (Schmidt et al. 2005; Horncastle et al. 2019) or
reduce vegetation cover that results in increased predation risk
from raptors (Torre et al. 2007) or other visual predators.
Grazing impacts on small mammals can have effects that
extend beyond a decline in their population. For instance,
exclusion of cattle can increase small mammal populations
(Sullivan & Sullivan 2014), which can improve nitrogen min-
eralisation (van Wijnen et al. 1999) and maintain grassland
diversity by limiting shrub encroachment (Ceballos et al.
2004). However, reduced small mammal densities may not be
viewed as negative in all human contexts, and rodent control
by livestock could be a positive ecosystem service.
The abundance of detritivores was signiﬁcantly higher at
grazed sites, which appears contradictory to a lower availabil-
ity of plant biomass (Fig. 1). Although grazing can reduce
biomass, the quality of remaining plant tissue may increase
(e.g. higher nitrogen to ﬁbre content), which can promote
arthropod abundance (Moran 2014; Coq et al. 2018). Grazers
may also be facilitating detritivores species through non-con-
sumptive effects, such as dung deposition by livestock or
greater soil turn over (Floate 2011). Additionally, detritivores
may not be impacted by some negative effects that vertebrate
herbivores experience, such as trampling, and could take
advantage of reduced competition for plant material (Heske &
Grazing had little effect on the higher trophic levels but did
reduce the abundance of vertebrate predators (Fig. 1). Verte-
brate predators could be lower in abundance because of fewer
prey items (i.e. vertebrate herbivores) and thus livestock could
have bottom-up effects on animal abundance. Management
practices can also negatively affect vertebrate predators, such
as fences or deterrent strategies (e.g. guard dogs, trapping),
that limit predation of livestock (e.g. van Bommel & Johnson
2012). The diversity of omnivores and predators were not
affected by livestock grazing potentially because there were
enough prey items in grazed areas to support some predators
or that these animals forage over an area larger than the cat-
tle enclosure. Omnivore abundance was also not affected,
potentially because of the larger dietary breadth relative to
predators. Secondary consumers are less affected by livestock
grazing relative to herbivores or pollinators, but further work
is necessary to determine the difference in response.
An increase in invertebrate herbivore and pollinator diver-
sity with livestock exclusion was also detected (Fig. 1).
Replacement of certain plant species, such as grazing promot-
ing grasses over forbs (Stahlheber & D’Antonio 2013; T€
et al. 2016), can affect herbivores or pollinators that are obli-
gates to speciﬁc species. Pollinators can be particularly
affected by livestock grazing because even though plant diver-
sity remains the same, removal of ﬂowers still reduces pollen
and nectar resource abundance (Hatﬁeld & LeBuhn 2007). For
instance, livestock grazing has been identiﬁed to alter plant-pol-
linator visitation networks (Vanbergen et al. 2014). Grazing has
been previously demonstrated in another meta-analysis to
reduce arthropod diversity more than plant diversity because of
homogenisation of the landscape (van Klink et al. 2015). We
also found evidence that changes to plant diversity because of
livestock grazing reﬂected in the abundance of animals (Fig. 2).
Shifts in the composition of plants and a lack of landscape
heterogeneity due to grazing could be driving the observed dif-
ferences in animal diversity from grazing.
Consequences of excluding grazers overtime
The effects of livestock grazing on the biodiversity of plants
and animals can persist after herds are removed from an
ecosystem (Fig. 3). We found evidence that time since grazer
exclusion increased animal abundance and had a unimodal
relationship with diversity. The intensity of livestock grazing
is expected to mediate the effect of grazing on plants, and
consequently, animal diversity (e.g. Milchunas et al. 1998;
Boerschig et al. 2013). Unfortunately, inconsistent estimates
of livestock densities (e.g. AUM, livestock/ hectare, number
of days grazed) across studies prevented us from including
such a comparison within our meta-analysis and encourage
standardisation of livestock quantities in the future. Instead,
we observed a pattern of recovery in the relatively short-term
absence of grazers (i.e. 10–20 years). Beyond that time, the
livestock exclusion did not have an effect on animal diversity
relative to sites that were grazed. A frequently recommended
approach in rangeland recovery is passive restoration where
domestic grazers are removed without other interventions (e.g.
Beschta et al. 2013), but this might not be effective if animal
diversity beneﬁts from some frequency of grazing. For
instance, the absence of grazers can promote the dominance
of grasses (Stahlheber & D’Antonio 2013; Koerner et al.
2018) that do not support specialist herbivores or pollinators.
Restoring the diversity of a previously grazed ecosystem can
require complementary management practices, such as
increasing the number of ﬂowering plants over grasses to sup-
port invertebrates (Vaudo et al. 2015). Grazer exclusion can
also affect the physical characteristics of the landscape includ-
ing soil nitrogen (van Wijnen et al. 1999; Knops et al. 2000),
soil compaction (Halde et al. 2011), and ﬁre frequencies
(Davies et al. 2009). Evaluating the effects of grazing requires
long-term experimentation, since the effects can change in sign
and magnitude over decadal time scales.
The continued effects of grazing on ecosystems after removal
are also complicated by previous strategies for managing live-
stock. For example, legacy effects persist today from high inten-
sity grazing that occurred a century ago in the United States
with cattle (Svejcar et al. 2014) and in Sweden with reindeer
(Egelkraut et al. 2018). Exclusion of livestock might not return
the ecosystem to its original composition, but rather a different
composition with higher animal abundance or diversity. In
some environments, continued grazing is a required manage-
ment technique to support native biodiversity (Diamond et al.
2012; Germano et al. 2012; Moranz et al. 2012). However, if
managed poorly, overgrazing could also cause the reverse,
©2020 John Wiley & Sons Ltd/CNRS
8A. F. Filazzola et al. Review and Syntheses
facilitating invasive species and encouraging shrub encroach-
ment (Roques et al. 2001; D’Odorico et al. 2012). Future
research should explore the frequency of disturbance that is
occurring with the livestock grazers and the impacts it has on
species provenance (i.e. native vs. non-native).
Interactions of grazing and climate
The effects of livestock exclusion on biodiversity were depen-
dent on temperature, but not precipitation (Fig. 4). Previous
ecological models have described co-occurring gradients of
herbivore pressure and environmental stress as driving shifts
in the sign of interactions among species in the stress gradient
hypothesis (Bertness & Callaway 1994; Smit et al. 2009; Dan-
gles et al. 2013) or changing the complexity of trophic struc-
ture in the exploitation hypothesis (Oksanen et al. 1981;
Oksanen & Oksanen 2000). Each of these models proposed
that in high-stress environments, the effects of natural her-
bivory are intense compared to other ecosystems because
plant species in these systems have traits that are targeted at
tolerance of abiotic stress instead of disturbance (sensu Grime
1977) and natural predators are mostly absent to regulate her-
bivory (Oksanen et al. 1981). However, we found evidence
that contradicts these models, as the effects of livestock graz-
ing appeared to be neutral in high-stress environments with
cold annual temperatures. These cold ecosystems may have
less diversity relative to temperate ecosystems (Gaston 2000),
and thus livestock grazers have a larger potential to reduce
species richness in mild climates where diversity is higher. Spe-
cies at climate extremes could also be more tolerant to distur-
bance and have greater capacity to recover following grazing.
For instance, in cold ecosystems adaptations to abiotic stress
could overlap with traits that inhibit herbivory (D
ıaz et al.
2007; Fensham et al. 2011, 2014) and in relatively ‘hot’
ecosystems, animal abundances might be less affected because
of high plant productivity. This pattern is supported by the
neutral relationship that we observed between climate and
grazing effects on plant diversity. Although not tested here,
variation in soil texture or resources from grazing can also
have signiﬁcant effects on plant composition (Schrama et al.
2013) and thus animal diversity. Inclusion of soil variables
and soil biota that have been identiﬁed to vary with livestock
(Bardgett & Wardle 2003; Andriuzzi & Wall 2017; Zhou et al.
2017) could further disentangle the indirect effects of grazing
on above-ground trophic levels.
Implications for conservation
Livestock exclusion can beneﬁt the abundance and diversity of
multiple trophic levels. However, abandoning grazing in certain
environments may not result in an increase to biodiversity and
in some instances can cause further loss. For instance, we
observed grazing having a positive effect on plant diversity and
four studies within our meta-analysis where animal diversity
increased with livestock grazing, contradicting the general trend
(Ranellucci et al. 2012; Schmidt et al. 2012; Verga et al. 2012;
Tabeni et al. 2013). In all four studies, livestock grazing main-
tained grassland structure by suppressing woody encroachment,
which supports speciﬁc animal species. Although the conversion
of grasslands to shrublands has been attributed to overgrazing
(Archer et al. 1995; Van Auken 2009), continued grazing in
these systems might be required to minimise shrub cover. In
other ecosystems, such as forests, livestock production is some-
times described as causing habitat loss because the generated
rangelands do not provide the same ecosystem services or func-
tions as the previous native habitat (Machovina et al. 2015). If
there are persistent effects of grazing, restoration to previous
conditions can be impractical and instead these rangelands rep-
resent novel ecosystems with a different set of species composi-
tion and functions (Cingolani et al. 2005; Hobbs et al. 2009).
Our analysis used coarse estimates of community composition,
such as abundance and diversity that neglect species-level
changes in composition. When examined at the species-level the
effects of grazing can be signiﬁcantly magniﬁed relative to com-
munity measures (Supp & Ernest 2014). For instance, at-risk
species may be especially sensitive to livestock relative to other
species if grazing reduces the abundance of plant species that
they are dependent (Ausden et al. 2005; Schtickzelle et al. 2007).
The impacts of livestock grazing on conservation are thus
dependent on target organism (plants, primary consumers,
predators) and goals set by land managers (improving diversity
The production of livestock has increased signiﬁcantly in
spatial extent since the 1960s and is projected to continue to
expand in developing countries (Thornton 2010), potentially
threatening indigenous animal diversity on a global scale.
Future increases in climate variability is also expected to
threaten food security and increase conversion of land into
rangelands (Sloat et al. 2018). To meet this demand, livestock
grazers will continue to be placed on land shared by indige-
nous animal species, thereby potentially threatening the global
biodiversity of herbivores and pollinators. These impacts are
expected to be most pronounced in mild climates, such as
temperate ecosystems, and are likely to persist after grazers
are removed. Identifying the aspects of grazing that most
impact animal biodiversity could be used to further develop
more effective management practices. For example, some
forms of rotational grazing are effective in environments with
low abiotic stress and when precipitation less variable (Haw-
kins 2017). Techniques for mitigation will not erase all the
effects of livestock grazing and these negative cascading
effects may be an inevitable consequence that society will need
to balance with the socio-economic beneﬁts.
The authors thank Dr. Jens Roland for comments on an ear-
lier draft. The authors also thank three anonymous reviewers,
Dr. Eric Seabloom, and Dr. Jonathan Chase for their com-
ments that signiﬁcantly improved the quality of the manu-
script. This research was funded by a Killam Post-Doctoral
Fellowships, a NSERC Post-Doctoral Fellowship awarded to
A.F, and an NSERC Discovery Grant awarded to JFC.
All authors contributed to developing the purpose of the
manuscript and participated in the data extraction from
©2020 John Wiley & Sons Ltd/CNRS
Review and Syntheses Livestock can reduce animal biodiversity 9
published articles. AF analysed the data and wrote the manu-
script. AB, CB, JG, JC and MD provided comments on the
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Additional supporting information may be found online in
the Supporting Information section at the end of the article.
Editor, Eric Seabloom
Manuscript received 22 August 2019
First decision made 8 February 2020
Second decision made 26 March 2020
Third decision made 1 April 2020
Manuscript accepted 3 April 2020
©2020 John Wiley & Sons Ltd/CNRS
12 A. F. Filazzola et al. Review and Syntheses