Journal of Ecology 2013, 101, 1225–1236
doi: 10.1111/1365-2745.12133
The response of plant diversity to grazing varies along
an elevational gradient
James D. M. Speed1*, Gunnar Austrheim1 and Atle Mysterud2
1
Museum of Natural History and Archaeology, Norwegian University of Science and Technology, NO-7491 Trondheim,
Norway; and 2Centre for Ecological and Evolutionary Synthesis (CEES), Department of Biosciences University of
Oslo, NO-0316 Oslo, Norway
Summary
1. Species richness of plants generally decreases along elevational gradients or peaks at intermediate
elevations. Land use including grazing by wild and domestic herbivores also affects plant communities and diversity, but how grazing affects plant diversity along elevational gradients is less clear.
2. Using a field experiment along an elevational gradient in Norway, we tested whether the impact
of grazing on plant diversity interacts with elevation. Vascular plant communities were sampled following 10 years of experimental sheep grazing (decreased/ceased, maintained and increased densities
of 0, 25 and 80 sheep km 2) and compared with pre-experimental baseline data.
3. The impact of grazing on species richness (alpha diversity) was investigated at the species level,
and temporal community variance was used to investigate community stability (temporal beta diversity). Spatial community variance (spatial beta diversity) at the enclosure level was used to test
whether community homogenization differed between treatments.
4. Species richness and temporal stability varied between treatments and along the elevational gradient. Where grazing was ceased, species richness declined by up 3.7 species at low elevations and
increased by up to 3.5 species at high elevations, whilst where grazing was maintained or increased,
changes were less extreme along the elevational gradient.
5. Temporal stability of plant communities was highest at low elevations and lowest at high elevations where grazing was reduced. There were no clear differences in spatial homogenization between
grazing treatments, although spatial species turnover increased in heathlands where grazing
decreased.
6. Synthesis. This study shows that the effect of grazing on plant diversity varies with elevation,
and grazing herbivores can thus directly affect elevational patterns of plant diversity. Grazing can
buffer changes in plant communities and species richness, even in the face of other environmental
drivers such as climatic warming.
Key-words: beta diversity, buffering, climate warming, determinants of plant community diversity
and structure, herbivory, homogenization, land use, species richness, stability, turnover
Introduction
Many ecosystems are grazed by wild or domestic ungulate
herbivores, and changes in herbivore density may lead to variation in patterns of plant diversity. Increases in the intensity
of herbivory favour a limited pool of resistant and tolerant
species, whilst reductions in the intensity of herbivory should
promote species more vulnerable to grazing (such as nutrientrich herbs) and over time lead to exclusion of less competitive
species (Augustine & McNaughton 1998; Krahulec et al.
2001). Thus, plant diversity may peak at intermediate levels
*Correspondence author. E-mail: [email protected]
of herbivory and with moderate densities of herbivores which
promote the persistence of species associated with grazing
(Milchunas, Sala & Lauenroth 1988). Plant diversity does not,
however, always change in response to grazing; particularly,
when environmental variation has more impact on vegetation
composition than grazing (Stohlgren, Schell & Vanden Heuvel 1999). Furthermore, the response rates of plant communities to increased and decreased herbivory may not be
symmetric, with a slower response observed when herbivory
is decreased than increased (Olofsson 2006).
The relative importance of herbivore density as a driver of
plant diversity patterns depends on factors such as habitat
productivity and spatial scale (Olff & Ritchie 1998; Bakker
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society
1226 J. D. M. Speed, G. Austrheim & A. Mysterud
et al. 2006). First, grazed plants more easily compensate for
grazing at higher productivity than lower productivity (Huston
1994; Proulx & Mazumder 1998; Austrheim & Eriksson
2001). Productive habitats are thus expected to sustain plant
diversity at higher herbivore densities than less productive
habitats. Second, the effects of grazing on plant diversity are
expected to differ with spatial scale. For example, high herbivore densities are found to increase local scale (alpha) diversity, while species turnover (beta) and regional scale (gamma)
diversity decrease as an effect of a homogenization process
(Chaneton & Facelli 1991; Olff & Ritchie 1998). However,
the role of herbivores in affecting diversity across spatial
scales is not always clearly observed (Ravolainen et al.
2010). Herbivores respond to spatial heterogeneity in vegetation (Senft et al. 1987), but can also affect the spatial heterogeneity in vegetation through impacting plant communities at
different scales and in interaction with productivity (Austrheim & Eriksson 2001) and soil nutrient cycling (Frank 1998;
Frank & Groffman 1998).
Plant diversity is also controlled by factors other than herbivory. Notably in alpine regions, diversity either shows a
decrease along elevational gradients or a humped-back relationship with peak diversity at intermediate elevations (Rahbek 1995; Grytnes 2003; Nogues-Bravo et al. 2008). These
elevational patterns in plant diversity have been attributed to
gradients in climate and productivity (Rahbek 1995; Odland
& Birks 1999; K€orner 2007; McCain & Grytnes 2010). A
peak in richness and high levels of species turnover are
observed at the forest limit (Grytnes 2003) or at transition
zones within and between forest and alpine zones (Odland &
Birks 1999), reflecting co-occurrence of species from the subalpine and alpine species pool. Although diversity patterns
along elevational gradients are well examined (Sanders &
Rahbek 2012), and it has been recognized that anthropogenic
factors influence these diversity patterns (Nogues-Bravo et al.
2008), it is not understood how herbivory may affect patterns
of plant diversity along elevational gradients. Livestock grazing is a particularly important form of land use in mountain
ecosystems (Austrheim & Eriksson 2001), and grazing by
domestic livestock has important influences on plant communities, often maintaining diversity (Krahulec et al. 2001;
DeGabriel et al. 2011), as well as having impacts on other
taxa in mountain ecosystems (Evans et al. 2006; Loe et al.
2007). However, the impact of grazing on vegetation along
elevational gradients is hard to predict as herbivores are selective at a range of spatial scales and herbivore grazing patterns
vary between seasons and years (Mobæk et al. 2009) and
with elevation, with low elevations being used earlier in the
season than higher elevations (Mysterud, Iversen & Austrheim 2007).
Diversity can be measured at a range of scales and using
an array of indices, each with distinct properties (Whittaker
1972; Tuomisto 2010a,b; Anderson et al. 2011). This study
investigates whether the impact of grazing on plant diversity
varies along an elevational gradient, using three separate measures of diversity. We investigate (i) whether species richness
(as measured by alpha diversity) is affected by grazing, and
whether this effect varies with elevation, (ii) whether temporal
variation (temporal turnover as a measure of beta diversity
Anderson et al. 2011) in plant communities is affected by
grazing, and whether this varies with elevation and (iii)
whether grazing leads to vegetation homogenization by reducing spatial variation (as a measure of beta diversity Anderson
et al. 2011) along an elevational gradient? Our approach
involved a large-scale grazing enclosure experiment in the
alpine region of Norway, with enclosures running from 1050
to 1320 m a.s.l. Grazing was manipulated over a 10-year period, with sheep densities increased, decreased and maintained
in comparison with prior to the experiment.
As the mechanistic processes that structure plant communities such as competitive exclusion, tolerance and resilience to
herbivory and shifts in species’ ranges are expected to vary
along environmental gradients, we made specific predictions
for changes in plant diversity under varied grazing regimes at
different elevational levels; these are summarized in Table 1.
We predicted that where grazer density was maintained, there
would be no, or at most minor, changes in vascular plant
diversity (this applies to all measures of diversity): this represents a grazing status quo scenario in this region with a long
history of grazing. (Hypothesis 1). Following grazing reduction, we predicted that quadrat-level species richness would
decrease at low elevations due to the competitive exclusion of
alpine species by lowland species (Milchunas, Sala & Lauenroth 1988), but increase at high elevations following upslope
expansion of lowland species’ distributions when released
from grazing (Speed et al. 2010a, 2012). Where grazer density was increased, we predicted that we would see a
decreased quadrat-level diversity at high elevations, due to
the colimitation of distribution by herbivores and abiotic conditions decreasing species richness and leading to dominance
of resistant and tolerant species. At low elevations in the
grazing-increased treatment, we predicted that there would be
no change in quadrat diversity, or a lower decrease than at
higher elevations, as plants would be able to compensate in
the more productive environment (Olff & Ritchie 1998; Proulx & Mazumder 1998; Austrheim & Eriksson 2001; Bakker
et al. 2006). Species replacement may mask changes in net
species richness. Therefore, we also investigated temporal
turnover in community composition. We hypothesized (2) that
temporal community variance would be lowest in the grazingmaintained treatment, since there is no perturbation to a
system with a long-grazing history (Milchunas, Sala & Lauenroth 1988; Austrheim 2002), and that stability (i.e. the
opposite of temporal variance) would be lower in the grazingincreased treatment than the grazing-decreased treatment since
the response of species composition to decreased grazing
intensity is observed to occur at a lower rate than to increased
grazing intensity (Olofsson 2006). We further predicted that
the turnover would be higher in grassland communities than
heathland communities due to the slower life cycles and more
resilient nature of the shrubs in the dwarf shrub heathland and
a higher selective utilization of the grasslands by the grazers.
Finally, (3) we predicted that following the grazing-homogenization hypothesis, spatial variance across the whole
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1225–1236
Response of plant diversity to grazing 1227
Table 1. Summary of predictions of change in species richness, temporal variation and change in enclosure-level spatial variation under different
grazing treatments along an elevational gradient. The symbols +,
and 0, respectively, denote predicted positive, negative and no changes in
species richness and beta diversity. Specific predictions for richness and temporal variation are presented for high and low elevations. Spatial variation predictions encompass the whole elevational gradient. References forming the basis for each prediction are shown in footnotes
Change in
Elevation
1. Species richness
Alpha diversity
Low elevation
High elevation
2. Temporal variation
Beta diversity
Low elevation
High elevation
3. Spatial variation
Beta diversity
Whole gradient
Enclosure scale
Decreased grazer density
Competitive exclusion of
subdominant species in
absence of grazing*
+
Invasion of lowland species
following grazing release‡
or 0
Perturbation to grazing
system* with a lower rate
of change for decrease
in herbivory§
or 0
Perturbation to grazing
system* with a lower rate
of change for decrease
in herbivory§
+ or
Release from grazing
homogenization¶ and
advance of subalpine
community‡
Maintained grazer
density
0
Status quo
0
Status quo
0
Status quo: no
perturbation to
grazing system*
0
Status quo: no
perturbation to
grazing system*
0
Status quo
Increased grazer density
0 or +
Plants compensate for herbivory
in productive habitats†
No compensation for herbivory
in unproductive habitats†
Perturbation to grazing system*
Perturbation to grazing system*
Grazing homogenization¶
*Milchunas, Sala & Lauenroth (1988).
†Olff & Ritchie (1998), Proulx & Mazumder (1998), Austrheim & Eriksson (2001), Bakker et al. (2006).
‡Speed et al. (2010a, 2012).
§Olofsson (2006).
¶Chaneton & Facelli (1991), Austrheim & Eriksson (2001), Rooney (2009).
elevational gradient would decrease where grazing intensity
was increased (Chaneton & Facelli 1991; Austrheim & Eriksson 2001; Rooney 2009). The change in spatial variation in
community composition where grazing intensity was
decreased is a more open question (Stohlgren, Schell & Vanden Heuvel 1999): however, we predicted that as the treeline
is rising (Speed et al. 2010a), and there is high plant species
turnover at the treeline (Hofgaard 1997; Camarero, Gutierrez
& Fortin 2006; Batllori et al. 2009), spatial variation would
increase where grazing decreases. This also represents a
release from grazing homogenization.
Materials and methods
To investigate whether the response of plant diversity to grazing varies with elevation, we used a sheep grazing experiment situated on a
south-facing slope (1050–1320 m) in the mountains of southern
Norway. This region has a long history of transhumance-type livestock production, where livestock (mainly sheep) are released into
alpine ranges in early summer where they freely graze until early
autumn. Nine adjoining enclosures were erected in late summer 2001,
with an average area of 0.3 km2 each spanning the elevational gradient. There were 3 replicates of 3 sheep density treatments, 0, 25 and
80 sheep km 2 within a randomized 3-block design (Fig. 1a). The
ungrazed treatment represents a decrease in sheep density, whilst the
high sheep density (relative to typical range of sheep densities in this
region) of 80 km 2 represents an increase in sheep density. The low
density of 25 sheep km 2 approximately maintains the estimated density of 10–20 sheep km 2 before the start of manipulated grazing.
Experimental grazing of sheep occurred within the enclosures each
summer from 2002 to 2011, typically between late June and early
September in common with general livestock management in the
region (details in Mobæk et al. 2012), and the sheep were free to
graze along the elevational gradient. The site spans the treeline ecotone (Moen 1999; K€orner & Paulsen 2004), being between the forest
and the alpine zone (see data in Speed et al. 2010a). The vegetation
comprises dwarf shrub and lichen heaths, snowbeds and willow-shrub
meadows. The mean annual and summer (June to August) temperatures at the site were 1.6 and 7.9°C, respectively, whilst the mean
annual precipitation was 1430 mm (data provided by the Norwegian
Meteorological Office, interpolated to the site at an elevation of
1160 m and averaged across 1957–2009).
Permanent marked vegetation quadrats (50 9 50 cm, n = 20 per
enclosure, 180 in total) were established in the enclosures. Quadrat
location was stratified by habitat and elevation (Fig. 1a) with a maximum range of 1091–1311 m; the distribution of quadrats between
vegetation types within enclosures varied with the surface area of the
vegetation types (the numbers of quadrats within treatments and vegetation types are presented in Table S1 in the Supporting Information,
see Austrheim et al. 2008 for further information). Plant communities
were examined at the vegetation type level within this study: lichen
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1225–1236
1228 J. D. M. Speed, G. Austrheim & A. Mysterud
(a)
(b)
(c)
Fig. 1. (a) Overview of experimental site
showing grazing treatments, location of
permanent quadrats and vegetation types.
Note that some permanent quadrats appear to
be in different vegetation types to how they
are classified: this is due to the large spatial
scale at which the vegetation map was made.
(b) Modelled change in species richness
plotted over the whole experimental site
based on the elevation, grazing treatment and
vegetation type of each pixel. The selected
model is shown in Table 2 and Fig. 2. (c)
Modelled community temporal variation
(temporal beta diversity) in vascular plant
composition plotted over the whole
experimental site based on the elevation,
grazing treatment and vegetation type of each
pixel. Dissimilarity is the opposite of
stability. The selected model is shown in
Table 3 and Fig. 3. In all panels, 100 m
contour lines are shown and UTM
coordinates are in zone 32V.
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1225–1236
Response of plant diversity to grazing 1229
heaths, dwarf shrub heaths, graminoid-dominated snowbeds and tall
herb meadows (characterized by the presence of erect willow-shrubs)
were investigated as these are the dominant vegetation types. The vascular plant vegetation within these quadrats was recorded in late June
to early July of 2011, 2009 and 2007, with baseline data (prior to initiation of the experimental grazing) having been recorded during the
same period in 2001. This study focuses on the data in 2011, after
10 years of experimental grazing, whilst data from 2007 and 2009 are
presented to examine whether results may be specific to the particular
year, or part of an emerging trend. Nomenclature followed Lid & Lid
(2005). Abundance of each plant species was assessed as the frequency of 12.5 9 12.5 cm subquadrats in which that species was
found within each 50 9 50 cm quadrat although data were transformed to presence and absence format for the purpose of this study
since we were interested in losses, gains and turnover of species
within the communities.
In order to investigate how the quadrat-level species richness
responds to grazing along an elevational gradient (hypothesis 1), we
calculated the change in species richness between 2001 and 2011 for
each of the 180 quadrats. This was related to elevation, vegetation
type and grazing treatment. All interactions including the three-way
interaction were included in the candidate models, along with the null
model and a model excluding grazing. Gaussian family linear models
were used (since the change in richness followed a normal distribution), and the final model was selected based on minimum Akaike
Information Criterion (AIC).
The dependence of community temporal variation between 2001
and 2011 upon grazing treatment and along the elevational gradient
(hypothesis 2) was assessed by calculating the Sørensen pairwise dissimilarity index (the binary version of the Bray Curtis index, and a
measure of the inverse of stability) for each quadrat between the
2 years (i.e. the dissimilarity between quadrat x in 2001 and quadrat x
in 2011 for each of the 180 quadrats). This index was chosen as it is
amongst the most widely used presence–absence indices and varies
with the proportion of species common to the two communities (Baselga 2010). The dissimilarity was calculated for each quadrat between
2001 and 2011 and ranges between 0 (all species are shared) and 1 (no
species are found in both communities). This value was then used as
the dependent variable in a linear model with elevation, vegetation
type, grazing treatment and all interactions as candidate independent
variables. Model selection was based upon minimum AIC using the
same candidate models as previously, again using a Gaussian family
model, since temporal turnover followed a normal distribution.
The grazing-homogenization hypothesis (3) was tested by examining whether spatial variation in community composition (measured
across each enclosure, hence along the elevational gradient) changed
over the 10-year period, and whether this change differed between
grazing treatments. We examined beta diversity as a whole using the
multiple-site Sørensen index, and in order to examine the processes
behind changes in beta diversity (replacement of species, or gains and
losses of species), we further partitioned total beta diversity between
the components of spatial turnover (Simpson beta diversity) and nestedness beta diversity (that is spatial replacement of species, and spatial
subsetting of communities respectively; Baselga 2010). Spatial beta
diversity was estimated within all quadrats of grasslands (snowbeds
and meadows) and heathlands (dwarf shrub and lichen heaths) within
each enclosure (n = 3). The vegetation types were combined into
heathlands and grasslands since we predicted shifts between these following changes in grazing regime. One-way ANOVA tests were used to
test for differences in change in beta diversity (where a negative
change equates to homogenization) between grazing treatments.
Statistical analysis was carried out in the R statistical environment (R
Development Core Team 2012) and utilized the beta diversity functions presented by Baselga (2010). Since vegetation types are expected
to exhibit different responses in diversity and stability to changes in
grazing along the elevational gradient, the degree of change in plant
communities across a landscape will depend on the distribution and
extent of the vegetation types. The selected models of change in species richness and temporal change in community composition were
therefore plotted across the whole experimental site using a vegetation
map which had earlier been developed for the site (Fig. 1a, Rekdal
2001). This allows visualization of the magnitude of change over
space. To examine whether there was nonindependence in the response
of species richness and spatial variance to grazing across the experimental site, residuals from the selected models were tested for spatial
autocorrelation using Moran’s I statistic: no spatial autocorrelation of
residuals was found in either model (for species richness I = 0.017,
P = 0.3, for temporal variation I = 0.003, P = 0.5).
Results
SPECIES RICHNESS
A total of 125 vascular plant species were recorded. In the
2001 baseline data, there was a mean of 60.9 (standard error
SE = 2.23) species per enclosure and 12.4 (SE = 0.55) species per quadrat. The quadrat-level change in species richness
between 2001 and 2011 ranged from a loss of 9 species to a
gain of 6 species, with a mean of 0.22 (SE = 0.17) species
gained. The 10-year change in species richness varied along
the elevational gradient, and there were interactions between
grazing treatment and elevation and grazing treatment and
vegetation type in determining change in species richness
(Table 2). The interaction between elevation and grazing
treatment was also included in the selected model for the
2009 data, but not the 2007 data (Table S2).
In all vegetation types, the decadal change in species richness
increased with elevation where sheep density was decreased,
whilst there was no elevational gradient in the change in species
richness where sheep grazing was maintained (Fig. 2). Species
richness decreased at low elevations (a mean loss of up to 3.7
species) and increased at high elevations (a mean grain of up to
3.5 species) where sheep density decreased (Fig. 2). Where
sheep density was maintained, there was no change in species
richness at any elevation or in any vegetation type, with the
exception of the dwarf shrub heath where there was a minor
increase with a mean change of around one species across the
elevational gradient. Where sheep density was increased, there
was a mean species loss in the tall herb meadow of around two
species and a weak positive relationship between change in species richness and elevation (Fig. 2).
When the decadal change in species richness was plotted
across the experimental site, in the grazing-maintained treatment, species richness was predicted to mostly increase by a
small amount due to the spatial dominance of the dwarf shrub
heath vegetation type (Fig. 1b). Whilst there was a strong
decrease in species richness in the tall herb meadows under
increased densities of sheep, this vegetation was of limited
area: there was little change in species richness across most
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1225–1236
1230 J. D. M. Speed, G. Austrheim & A. Mysterud
Table 2. Akaike Information Criterion (AIC) model selection table for linear model of change in species richness. Variables in the candidate
models are denoted by a 1 in the columns for elevation (E), vegetation type (V) and grazing treatment (G) as well as the two- and three-way
interactions. AIC values and the difference from the minimum AIC are shown, ordered with the model with most support at the top
E
V
G
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
E:V
E:G
V:G
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
5
AIC
DAIC
16
10
14
6
19
11
13
17
8
25
2
805.28
806.94
807.99
808.59
809.34
809.46
809.48
809.68
810.02
811.33
818.58
0
1.66
2.71
3.3
4.06
4.17
4.2
4.4
4.74
6.05
13.29
0
Grazing
Decreased
Maintained
Increased
5
0
–5
1050 1100 1150 1200 1250 1300
1050 1100 1150 1200 1250 1300
Elevation (m)
Elevation (m)
Tall herb meadow
Change in species richness
Graminoid snowbed
Change in species richness
1
d.f.
Dwarf shrub heath
Change in species richness
Change in species richness
Lichen heath
–5
E:V:G
5
0
–5
1050 1100 1150 1200 1250 1300
5
0
–5
1050 1100 1150 1200 1250 1300
Elevation (m)
Elevation (m)
of the landscape with the exception of a slight increase in the
snowbeds at the highest elevations (Fig. 1b). In the grazingdecreased treatment, there was a clear decrease in species
richness across the enclosures at low elevations, and an
increase at high elevations, with the threshold of no change
between 1150 and 1250 m elevation depending on the vegetation type (Fig. 1b, x intercept in Fig. 2).
TEMPORAL VARIANCE IN COMMUNITY COMPOSITION
Community stability over time was assessed through its
inverse: community dissimilarity (temporal beta diversity),
Fig. 2. Change in species richness between
2001 and 2011 in 0.25 m2 quadrats plotted
against elevation for four vegetation types
and three grazing treatments. The regression
lines (standard errors shown by dashed lines)
are plotted from a single linear model
including vegetation type as a factor, but
plotted on separate panels to increase clarity.
expressed using the Sørensen pairwise dissimilarity index.
The mean community temporal dissimilarity (i.e. temporal
beta diversity) between 2001 and 2011 was 0.16 (SE =
0.008). The selected model of community dissimilarity
included vegetation type and the interaction between elevation
and grazing treatment (Table 3), although in both 2007 and
2009, the model that did not include grazing was found to be
most parsimonious in explaining dissimilarity (Table S3). The
10-year temporal change in community composition was lowest for the lichen heath (Sørensen index between 0 and 0.1)
and highest for the snowbed and meadow vegetation types
(0.1–0.3; Fig. 3). The highest dissimilarity was generally
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1225–1236
Response of plant diversity to grazing 1231
Table 3. Akaike Information Criterion (AIC) model selection table for linear model of community temporal variation. Variables in the candidate
models are denoted by a 1 in the columns for elevation (E), vegetation type (V) and grazing treatment (G) as well as the two- and three-way
interactions. AIC values and the difference from the minimum AIC are shown, ordered with the model with most support at the top
E
V
G
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
1
E:V
E:G
V:G
E:V:G
d.f.
1
1
10
6
8
13
14
16
11
19
17
25
2
1
1
1
1
1
1
1
1
1
1
1
1
1
1
0.4
Grazing
0.3
Decreased
Maintained
Increased
0.2
0.1
0.0
1050 1100 1150 1200 1250 1300
0.3
0.2
0.1
0.0
1050 1100 1150 1200 1250 1300
Elevation
Graminoid snowbed
Tall herb meadow
0.4
Community dissimilarity
0.4
Community dissimilarity
0
1.6
1.78
3
4.17
4.26
4.84
7.13
7.63
17.09
44.37
Dwarf shrub heath
Elevation
Fig. 3. Community temporal variation (temporal beta diversity) in 0.25 m2 quadrats
between plant composition in 2001 and
composition in 2011, plotted against
elevation for four vegetation types and three
grazing treatments. Dissimilarity is the
opposite of stability. The regression lines
(standard errors shown by dashed lines) are
plotted from a single linear model including
vegetation type as a factor, but plotted on
separate panels to increase clarity.
344.92
343.33
343.15
341.92
340.75
340.66
340.08
337.79
337.29
327.84
300.56
DAIC
0.4
Community dissimilarity
Community dissimilarity
Lichen heath
AIC
0.3
0.2
0.1
0.0
1050 1100 1150 1200 1250 1300
found for the grazing-increased treatment. Dissimilarity
tended to increase along the elevational gradient in the grazing-decreased treatment, but not where grazing was maintained or increased. At low elevations, dissimilarity was
higher in the grazing-maintained and grazing-increased treatments than where grazing was decreased, whilst at high elevations, temporal dissimilarity was similar between treatments,
but generally highest in the grazing-decreased treatment
(Fig. 3). When plotted across the whole experimental site,
lichen heaths in all grazing treatments, and all plant communities at low elevations where grazing had declined, were predicted to be the most stable communities in terms of
composition over time, whilst community composition was
most temporally variable in the grazing-increased treatment
(Fig. 1c).
0.3
0.2
0.1
0.0
1050 1100 1150 1200 1250 1300
Elevation
Elevation
SPATIAL VARIANCE IN COMMUNITY COMPOSITION
The mean enclosure-level spatial beta diversity (expressed as
Sørensen multiple-site index) in the baseline data set was
0.79 (SE = 0.019) in the grasslands and 0.80 (0.013) in the
heathlands. Spatial turnover of species constituted the main
part of this (0.68 and 0.62, SE = 0.021 and 0.021 for grasslands and heathlands, respectively) with only a small contribution of spatial nestedness (0.10 in grasslands and 0.18 in
heathlands, SE = 0.007 and 0.018, respectively). There were
only minor changes in beta diversity over 10 years with a
mean decrease in total beta diversity across grasslands of
0.007 (SE = 0.005) and of 0.016 (SE = 0.006) in heathlands.
There were no significant differences in change in total beta
diversity between grazing treatments in either the heathlands
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1225–1236
1232 J. D. M. Speed, G. Austrheim & A. Mysterud
or grasslands (analysis of variance: F2,6 = 1.02, 2.76 for
grasslands and heathlands, respectively, P > 0.05 for both,
Fig. 4a,d). However, in the grazing-decreased treatment, spatial turnover was significantly higher (Fig. 4e, F = 6.06,
P = 0.036), whilst community nestedness was significantly
lower (Fig 4f, F = 5.92, P = 0.038) than in the grazingmaintained treatment within the heathlands. Conversely, there
was a trend for an increase in turnover and a decrease in nestedness in the increased grazing treatment in the grassland
habitats (Fig 4b,c), although these differences were not significant (F = 2.07, 2.74 for turnover and nestedness, respectively, P > 0.05 for both).
processes along environmental gradients. We found that
where the density of grazing herbivores was decreased, there
was a positive relationship between change in species richness
and elevation, with a mean net loss of up to 3.7 species at
low elevations and a net gain of up to 3.5 species at high elevations. Contrastingly, where grazing was maintained or
increased, changes in species richness were modulated (closer
to zero) and did not vary along the elevational gradient. Community stability over 10 years decreased with elevation where
grazing was decreased; stability was thus lower where grazed
than ungrazed at low elevations, but tended to be lower where
ungrazed at the highest elevations. It has been recognized that
land use and other anthropogenic influences in mountains
may impact on elevational patterns of diversity (NoguesBravo et al. 2008): here, we provide evidence for a direct
impact of grazing livestock on plant diversity along an elevational gradient.
Discussion
Diversity patterns along elevational gradients are well examined but the importance of herbivory along these gradients is
less well understood. By analysing alpha diversity as well as
spatial beta diversity and temporal stability, we have shown
that the impact of grazing on plant diversity and temporal variance in community composition varies along an elevational
gradient, highlighting the perils of averaging ecological
BUFFERING BY HERBIVORY
The increase in richness and community temporal dissimilarity with elevation in the grazing-decreased treatment is as we
Grasslands
Change in index between 2001 and 2011
(a)
(b)
Total
(c)
Turnover
0.04
0.04
0.04
0.02
0.02
0.02
0.00
0.00
0.00
–0.02
–0.02
–0.02
–0.04
–0.04
–0.04
–0.06
Decreased
Increased
–0.06
Grazing treatment
Decreased
Increased
–0.06
Grazing treatment
Nestedness
Decreased
Increased
Grazing treatment
Heathlands
Change in index between 2001 and 2011
(d)
(e)
Total
Turnover
(f)
0.04
0.04
0.04
0.02
0.02
0.02
0.00
0.00
0.00
–0.02
–0.02
–0.02
–0.04
–0.04
–0.04
–0.06
Decreased
Increased
Grazing treatment
–0.06
Decreased
Increased
Grazing treatment
–0.06
Nestedness
Decreased
Increased
Grazing treatment
Fig. 4. The change in enclosure-level spatial variation (spatial beta diversity) between 2001 and 2011 in three different grazing treatments within
grassland (a–c) and heathland (d–f) quadrats. Total spatial beta diversity, spatial turnover beta diversity and spatial nestedness beta diversity are
shown. Mean and standard errors are shown, n = 3.
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1225–1236
Response of plant diversity to grazing 1233
predicted and supports that lowland species increase their
upper elevational limit when not constrained by grazing. This
is observed for the treeline mountain birch at the site (Speed
et al. 2010a), whilst plant community composition at high
elevations had also observed to become more similar to that
of lower elevations in the absence of grazers (Speed et al.
2012). In the grazing-decreased treatment, the decrease in species richness at low elevations and increase at high elevations
suggest that the elevation of peak species richness (usually
just above the forest line; Grytnes 2003) may rise in the
absence of grazers. This parallels the elevational advance of
the treeline in the absence of sheep (Speed et al. 2010a).
Interestingly, the treeline at the experimental site is at an elevation of around 1150–1200 m which corresponds to the elevation at which the change in species richness is around zero
(Fig. 1b, the x intercept in Fig. 2). This highlights the interaction between treeline position, as a key community determinant, and land use, as a key ecosystem driver in determining
community structure (Hofgaard & Wilmann 2002; Camarero,
Gutierrez & Fortin 2006; Batllori et al. 2009).
Competitive interactions decrease at higher elevations (Callaway et al. 2002), and the decrease in species richness at low
elevations where grazing was decreased is probably due to
the increasing dominance of competitive species (Krahulec
et al. 2001). Indeed, at this experimental site, the plant traits
that are associated with species showing positive responses to
increased grazing include low plant height and high root-toshoot ratios (Evju et al. 2009). Furthermore, there is some
evidence that grazing can counteract the competitive exclusion of low and prostrate species at high elevations such as
Omalotheca supina and Sagina procumbens (Austrheim et al.
2008). The interaction between grazing treatment and elevation in determining change in vascular species richness was
also seen in the 2009 data, but not the 2007 data, showing
that the role of grazing in shaping plant communities cannot
be assessed in short-term studies.
The change in species richness under grazing varied
between vegetation types. In the grazing-maintained treatment, there was no net change in species richness with the
exception of in the dwarf shrub heath vegetation, where there
was a minor increase. Where grazing was increased, there
was also a small increase in species richness in the dwarf
shrub heath vegetation, but a clear decrease in the tall herb
meadow, which is highly selected by grazing sheep (Mobæk
et al. 2009). The net loss of species from these communities
is likely to be of the highly selected herbs in the productive
vegetation (Br
athen et al. 2007; Austrheim et al. 2008).
We predicted that the community stability would be high
where grazing was maintained (status quo) and indeed stability tended to be highest across the elevational gradient in the
grazing-maintained treatment. Community turnover was not
related to grazing after 6 and 8 years of grazing (Table S3),
so whilst the findings presented for the 10-year change may
be part of an emerging trend, we cannot rule out that they are
a result of the environmental conditions between 2009 and
2011. During the experimental grazing period, growing season temperatures at the experimental site were higher and less
variable [mean 2002–2011 of 8.93°C, standard deviation (SD)
of 0.76] than immediately before (mean 1992–2001 7.75°C
SD = 1.05, data presented in Speed et al. 2011). Thus, the
plant community changes observed here are responses to
changes in both grazing and temperature. However, temporal
stability was highest in the grazing-maintained treatment
(averaged across the elevational gradient), indicating that climate change had a minimal impact where grazing was maintained. Indeed, evidence has previously been presented
supporting buffering of climate-driven vegetation changes in
plant communities by herbivores (Collins et al. 1998; Post &
Pedersen 2008; Speed et al. 2012).
As predicted, stability tended to be lower in the grazingincreased treatment than the grazing-maintained treatment.
However, stability varied with elevation and was higher in
the grazing-decreased treatment at low elevations than at the
same elevation in both the grazing-increased and grazingmaintained treatments. This implies that grazing is promoting
temporal dissimilarity through processes of both colonization
and extinction of plant species at the local scale (quadrat-level
species). Indeed, sheep have been seen to increase both local
colonization and extinction under long-term grazing (Gibson
& Brown 1991), as well as to maintain species richness by
acting as dispersal agents (Fischer, Poschlod & Beinlich
1996). Since the heathlands were more stable than grasslands,
when viewed across this landscape, temporal turnover appears
relatively low due to the broader extent of the heathlands;
however, the heathlands may respond on a longer time-scale.
HOMOGENIZATION
Homogenization of plant communities is expected to arise
from herbivory, as grazing-sensitive species are lost (Chaneton & Facelli 1991; Olff & Ritchie 1998). Long-term homogenization of mountain vegetation at a regional scale has been
attributed to both a warming climate and intensive grazing
(Britton et al. 2009; Ross et al. 2012), whilst herbivory has
been directly linked with community homogenization at the
landscape scale (Austrheim & Eriksson 2001; Rooney 2009).
In our study, we found only weak evidence in support of the
grazing-homogenization hypothesis. Enclosure-level beta
diversity (diversity across the whole elevation gradient within
enclosures) showed a slight negative change over 10 years.
Where grazing was decreased, there was an increase in spatial
turnover (species replacement), and a decrease in community
nestedness (community subsetting) compared with the grazing-maintained treatment, but only in the heathland habitats.
Temporal turnover in the heathlands was greater at higher elevations, so the increased spatial turnover and decreased spatial
nestedness may be explained by invasion of (less rich) high
elevation communities by species not otherwise found in
heathlands. Indeed, species highly selected by sheep such as
Geranium sylvaticum and Ranunculus acris showed upslope
movements where grazing decreased, but downslope shifts
where grazing increased (Speed et al. 2012). The weak support for grazing-induced spatial homogenization suggests that
spatial homogenization may be a longer-term process than
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1225–1236
1234 J. D. M. Speed, G. Austrheim & A. Mysterud
alpha diversity scale changes. Alternatively, this could be due
to the fact that spatial variance was assessed along the environmental gradient and at a relatively small spatial scale
(0.25 m2 quadrats). At such small spatial scales, the impact of
herbivory may serve to increase heterogeneity due to the
localized nature of herbivore impacts, whilst homogenization
may occur at larger spatial scales (Olff & Ritchie 1998; Austrheim & Eriksson 2001). Furthermore, environmental variation may have greater impacts on plant community
composition than grazing (Stohlgren, Schell & Vanden Heuvel 1999) and, along environmental gradients plant communities differ in their response to herbivory, and herbivore habitat
selectivity itself varies.
HERBIVORE HABITAT SELECTIVITY AND PLANT
DIVERSITY
In our study, the grazers were free to roam within the elevational range of the enclosures and free to select between the
different vegetation types. The sheep at the site are known to
prefer the grassland habitats and show distinct patterns of elevation use within and between years (Mysterud, Iversen &
Austrheim 2007; Mobæk et al. 2009). Furthermore, habitat
selection varies with sheep density such that the preferred
habitat is less selected at higher densities (Mobæk et al.
2009). Thus, the intensity of herbivory received by the quadrats (in different vegetation types and elevations) over the 10year grazing period was not standardized, but reflects natural
preference of sheep. For example, whilst we predicted a
decrease in species richness at high elevations where grazing
was increased, and an increase at low elevations due to compensation at different productivity levels, this was not
observed. We suggest that this may be due to preferential
grazing by sheep at lower elevations over the course of the
season (Mysterud, Iversen & Austrheim 2007) and thus a relatively higher grazing intensity at low elevations compared
with higher elevations. Our study found that the most selected
vegetation type, the tall herb meadow, was also the least stable, whilst lichen heaths, the least selected vegetation, was
the most temporally stable. However, other studies that have
standardized herbivory through simulating herbivore activity
have found that the plant communities least selected by herbivores were also the most susceptible to a standardized level
of herbivory (Speed et al. 2010b).
Just as there are regions along elevational gradients where
plant species turnover peaks (Odland & Birks 1999; Grytnes
2003), there are likewise positions along gradients of grazer
density where plant species turnover peaks (Peper et al.
2011). Thus, the impact of grazing on the diversity and stability of plant communities along elevational gradients depends
on the interactions between plant communities and herbivores:
that is both on the intrinsic response of the plant community
to a certain intensity of herbivory and the spatial pattern of
grazing by the herbivore population, as well as the spatial distribution of the different plant communities. The variation in
herbivore habitat utilization and the spatial distribution of the
different plant communities that respond differently to grazing
(shown in Fig. 1b,c) highlights the need to take a wider perspective when drawing inference from experimental studies.
The findings of this study suggest that the use of grazing to
manage alpine biodiversity (DeGabriel et al. 2011) will be particularly challenging in the context of climatic change. Since
grazing interacts with elevation in determining changes in biodiversity and community composition, the same management
practice will have different outcomes at different points along
an elevational gradient. Reducing grazer density may increase
alpine species richness, but this may be driven by an elevational
advance of lowland species at the expense of a loss of the
alpine elements in the communities (Speed et al. 2012). It
should of course be remembered that reductions in herbivore
density in landscapes with long histories of grazing (as in this
study) represent perturbation to systems, just as increasing herbivore density drives changes in systems without long histories
of grazing (Milchunas, Sala & Lauenroth 1988).
Conclusions
In this study, we have shown that there is elevational variation in the impact of grazing on plant community diversity
and temporal turnover, demonstrating that the response of
plant communities to perturbations may be masked if averaged across environmental gradients. Grazing buffers changes
in diversity by preventing increases in species richness at high
elevations and preventing decreases in species richness at low
elevations. Herbivores may therefore affect elevational patterns of diversity. Whilst changes in mountain species richness have been linked to a warming climate (Pauli et al.
2012), the results of our study show that changes in mountain
plant diversity can also be caused by changing land use; thus,
climate change cannot be assumed to be the cause of vegetation change if land use change is not first ruled out.
Acknowledgements
We are grateful to the field assistants Eva Sofie Dahlø, Marianne Evju, and
Suzanne Hansen, and to the anonymous reviewers whose constructive input
greatly improved the presentation of this study. Funding was provided by the
Research Council of Norway, Environment 2015 program (Project 183268/S30)
and the Directorate for Nature Management in Norway.
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Received 31 January 2013; accepted 11 June 2013
Handling Editor: Sandra Lavorel
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1225–1236