Download Plant diversity consequences of a herbivore-driven biome

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Plant diversity consequences of a herbivore-driven biome switch
from Grassland to Nama-Karoo shrub steppe in South Africa
Michael C. Rutherford, Leslie. W. Powrie & Lara B. Husted
Rutherford, M.C. (corresponding author, [email protected]): Biodiversity Research
Division, Kirstenbosch Research Centre, South African National Biodiversity Institute, Private
Bag X7, Claremont 7735, South Africa and Department of Botany and Zoology, Stellenbosch
University, Private Bag X1, Matieland 7602, South Africa
Powrie, L.W. ([email protected]) & Husted, L.B. ([email protected]): Biodiversity
Research Division, Kirstenbosch Research Centre, South African National Biodiversity
Institute, Private Bag X7, Claremont 7735, South Africa
Abstract
Questions: How does heavy grazing change plant community structure, composition and species
richness and diversity in an ecotone between grassland and semi-arid shrub steppe-type vegetation?
Does grazing favour plants with arid affinity over those with less arid affinity? Does the grazinginduced transformation constitute a switch to the equivalent of a shrub-dominated biome?
Location: Central South Africa.
Methods: Using systematic scanning of SPOT 5 imagery and ground-truthing, a grazing treatment
area was selected that met criteria of intensity of grazing, sampling requirements, and
biogeographical position within a broad ecotonal zone. Differential vegetation responses to heavy
grazing were tested for significant differences in plant traits, vegetation structure, and species
diversity, richness and evenness. Gamma diversity was calculated for the whole study site, whereas,
independent beta diversity was calculated across the treatments assuming the additive partitioning
of diversity. In addition, the biogeographical association of grazing-induced species shifts was
determined using a range of available databases.
Results: Canopy cover and height of woody shrubs increased significantly with heavy grazing
whereas that of graminoid plants declined. The resultant species turnover was modest, apparent
extinctions of local species were minimal, species richness was maintained and species diversity was
significantly enhanced. There was a significant increase in species evenness, through possible
suppression of dominant species. Significant increases in species cover were those associated with
mainly the Nama-Karoo biome indicating that species from more arid areas are more resistant to
grazing as would be expected by the convergence model of aridity and grazing resistance.
Conclusions: The significant increase in shrub cover in heavily grazed semi-arid grassland followed
general global expectations. The study confirmed that the supposed former large shift of grassland
to shrubby Nama-Karoo in the eastern upper Karoo can indeed be readily affected by heavy grazing.
The negative connotations for biodiversity that have often been associated with intense grazing
seem, in terms of the positive responses of plant species diversity in this study, to perhaps be
exaggerated. The elevated species diversity with grazing of vegetation with a long evolutionary
grazing history in a low resource area may require a reappraisal of the application of certain grazing
hypotheses.
Keywords
Cover; Degradation; Disturbance; Encroachment; Species evenness; Grazing; Local extinction;
Species richness; Soils; Traits
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Nomenclature
http://posa.sanbi.org/searchspp.php
Abbreviations
HU: High utilization; LU: Low utilization
Running head
Diversity in a herbivore-driven biome switch
Introduction
Shrub invasion or encroachment of grassland and grassy savanna is perceived as an important
problem throughout the world (Aguiar, et al. 1996; Van Auken 2000; Watkinson & Ormerod 2001;
Liu et al. 2006). The negative impacts of shrub encroachment have included steep declines in plant
species richness (Knapp et al. 2008; Báez & Collins 2008), raised risk of animal species extinction
(Spottiswoode et al. 2009) and decreased grazing capacity which has rendered many rangelands no
longer economically viable (Smit et al. 1999). Shrub encroachment has been regarded as one of the
syndromes of dryland degradation (Scholes 2009) with the most prominent driving force identified
as chronic, high levels of herbivory by domestic animals (Van Auken 2000).
Herbivore actions have been associated with important changes in biotic assemblages,
sometimes at wide spatial scales. Thus it has been suggested that herbivore trampling and grazing
could have resulted in a biome-level change at the end of the Pleistocene (Zimov et al. 1995).
Considerable prominence has been given to the possible effects of grazing and climate in the
extensive Sahel zone of Africa (Miehe et al. 2010) with estimates of dramatic border shifts of the
Sahara Desert (Tucker et al. 1991). Equally striking descriptions of ecosystem change due to shrub
encroachment and invasion have been put forward. The shrub invasions of the North American
semi-arid grasslands in the past 150 years have resulted in changes described as ‘unparalleled’ (Van
Auken 2000). The apparent post-colonial invasion of part of the grassland biome of South Africa by
shrubby Karoo vegetation has been described by Acocks (1953) as ‘the most spectacular of all
changes in the vegetation of South Africa’. This latter view has been singularly influential in
developing national agricultural policy for arid lands (Hoffman et al. 1995) and, unsurprisingly, has
also attracted considerable scientific interest. This large area in the north-eastern upper Karoo is
interposed between the Grassland and drier Nama-Karoo Biomes (Rutherford & Westfall 1994;
Rutherford et al. 2006) and is referred to as ‘False Upper Karoo’ (Acocks 1953).
Acocks’ claim of a large-scale change in the vegetation of the north-eastern upper Karoo is
contentious (Hoffman et al. 1999) and has divergent levels of support. Past invasion of the grassland
by Karoo shrubs has been supported by analysis of carbon isotope composition of the soil (Bond et
al. 1994). It has further been supported by reduced livestock carrying capacities, as inferred from
lower livestock numbers in recent decades (Dean and Macdonald 1994). But, this reduction may
alternatively be ascribed to increased commercialism of farming, greater conservation awareness
and the introduction of a state-sponsored stock reduction scheme (Hoffman & Ashwell 2001). The
evidence is even less clear from the continuous pollen sequence, earlier travellers’ records and
repeat photography (Hoffman & Cowling 1990; Hoffman et al. 1995; Hoffman et al. 1999). The role
of overgrazing in this area has been bolstered by a number of local narratives on increases in shrubs
with heavy grazing, but these effects have seldom been linked to stocking density thresholds (Roux
& Vorster 1983; Roux & Theron 1987). Several studies in the region have emphasized the importance
of variation in rainfall over that of grazing in controlling growth form composition (Roux 1966;
Novellie & Bezuidenhout 1994; O’Connor & Roux 1995; Palmer et al. 1999; Beukes & Cowling 2000;
Meadows 2003), although the grazing pressures involved have seldom been indicated and may well
have been light. The role of heavy grazing may consequently have been down-played here, and for
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that matter elsewhere (Hein 2006). A state and transition model, that was developed for this region
(Milton & Hoffman 1994; Cingolani et al. 2005), indicates a number of putative threshold points that
facilitate an increase of shrubs through grazing.
In semiarid shortgrass steppe of the North American Great Plains, species diversity and
richness were found to decline with increasing grazing intensity (Milchunas et al. 1988; Milchunas et
al. 1998), a response also found for species richness in grassland in southern Australia (Dorrough et
al. 2007). Very few studies in Nama-Karoo or adjacent grassland have addressed the effect of grazing
on plant species richness or diversity. A decline in species richness with grazing in Nama-Karoo has
been mooted (Kraaij & Milton 2006) and was clearly demonstrated under extreme conditions of
grazing and trampling near to livestock water drinking points (Todd 2006).This appears to be in
accordance with the grazing reversal hypothesis (May et al. 2009). The hypothesis, which evolved
from the Dynamic Equilibrium Model (DEM) (Huston 1979), expects a decline in species diversity
with grazing of vegetation with a long evolutionary history of grazing in a low resource (e.g. semiarid) area (Michulnas et al. 1988; and see Proulx & Mazumder 1998). However, the wider supporting
evidence for the hypothesis is scant for such grazed systems in the southern hemisphere and even
further afield little change in species diversity and richness has been indicated (Cingolani et al. 2005).
By contrast, heavy grazing has significantly increased species richness in shortgrass steppe in
Mongolia (Cheng et al. 2011) and in the soil seed banks of Nama-Karoo vegetation in Namibia
(Dreber & Esler 2011). Evidence either way is often wanting; a recent review of the effects of grazing
on biodiversity in South African grassland (O’Connor et al. 2010) indicated a lack of studies where
total species richness was sampled.
A subsidiary tenet of the DEM or Generalized DEM (Milchunas et al. 1988) has been more
explicitly recognized as the convergence model of drought (aridity) and grazing resistance (Quiroga
et al. 2010). The convergence model hypothesizes that aridity and grazing are convergent selective
pressures on plants, each selecting concurrently for higher drought and grazing resistances.
Given the importance of grazing-induced encroachment of grassland by shrubs as well as the
apparently uncertain understanding of the attendant role of intensive levels of grazing, we
investigated the effects of heavy grazing on vegetation structure and diversity in the broad ecotonal
False Upper Karoo. We aimed to test the following main hypotheses:
1) Differential effects of heavy grazing transform systems dominated by perennial grasses into
low-cover, shrub-dominated systems according to the state and transition model for the
grassy eastern Karoo (Milton & Hoffman 1994; Cingolani et al. 2005); the broad-scale
corollary of this hypothesis is the grazing-induced transformation of the Grassland Biome
into the equivalent of the steppe-like Nama-Karoo Biome.
2) Heavy grazing leads to a decline in plant species diversity as predicted by the grazing reversal
hypothesis for vegetation with a long evolutionary history of grazing in a semi-arid, low
resource area (Milchunas et al. 1988; May et al. 2009).
3) Species compositional shifts due to heavy grazing favour species with an arid affinity (NamaKaroo Biome) more than those with a less arid affinity (Grassland Biome), as expected by the
convergence model of drought (aridity) and grazing resistance (Quiroga et al. 2010).
Considering the close interrelation between vegetation and soil we also explored the effects
of heavy grazing on soil characteristics. This study in addition formed part of a national pilot research
programme on understanding the relationships between plant diversity and land degradation
(Rutherford & Powrie 2010).
Methods
Study area
The region was systematically scanned using SPOT 5 satellite imagery to detect candidate sites for
heavy grazing juxtaposed with a grassland area that could serve as a suitable lightly-grazed reference
system. After field inspection of candidate sites, the area selected for study represented a clear
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fence line contrast situated at 24° 43’ S 30° 05’ E about 6 km east of Petrusville in the Northern Cape
Province (Fig. 1). Mean annual precipitation at Petrusville was 352 mm of which 74 % fell in the
summer half-year between October and March. The site was on a slightly undulating plain at an
altitude of 1250 m amsl. Geology was dolerite mixed with Ecca shale (Volksrust Formation), and soil
was generally shallow, dark reddish brown sand. There was no indication of any pre-existing
environmental gradients across the fence line.
Fig. 1. Location of the study area in the context of the False Upper Karoo
as well as the Grassland and Nama-Karoo Biomes.
The grazing contrast area comprised the grassland reference site on the ranch of De Put and
the extensive commonage (townlands) of Petrusville that was grazed communally. De Put was
stocked with cattle at a density of 0.055 LSU ha-1. A large stock unit (LSU) is equivalent to a steer of
450 kg. The commonage was grazed continuously by a range of animals including goats, donkeys,
horses, sheep and cattle and stocking density was estimated as at least three times than that on De
Put (J. van Jaarsveld pers. comm. October 2008, January 2011, local farmer) i.e. > 0.165 LSU ha-1 .
Mean stocking density for the former Philipstown District (that included Petrusville) for the period
1971-81 has been given as 0.057 LSU ha-1 (Dean & Macdonald 1994).
In this study the grassland reference side of the fence is, where appropriate, referred to as LU
(low utilization) and the communal side is termed HU (high utilization). This terminology is intended
to fit with the broader pilot study where both grazing and firewood extraction can co-occur.
Sampling
Vegetation
Sampling was carried out towards the end of the rainy season in late summer. Fourteen sample plots
of 50 m2 (10 x 5 m) were randomly placed on each side of the fence at distances varying between 20
and 30 m to avoid possible disturbance effects near the fence. The canopy cover and mean plant
height of each vascular plant species was recorded as well as the total canopy cover and mean
height of graminoid and woody plants in each plot. Species were classed according to plant traits
(Díaz et al. 2007), namely by life history (annual, perennial), growth form (forb, geophyte, graminoid,
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woody), habit (erect, prostrate) and architecture (leafy stem, rosette, stoloniferous, tussock).
Rainfall in the months prior to sampling had been above average thus engendering confidence in the
recording of ephemeral species. Plant specimens were identified by the National Herbarium of the
South African National Biodiversity Institute (SANBI). Our sampling took advantage of a ‘natural’ field
experiment in which the consequential logistic problems in avoiding pseudo-replication (Hurlbert
2004) were insurmountable (Hargrove & Pickering 1992). See Rutherford & Powrie (2010) for
discussion on issues of pseudo-replication in fence-line contrast studies.
Soils
Ten soil samples were taken to a depth of 50 mm from each side of the fence and analyzed for
texture, pH, resistance, carbon, ammonium nitrogen, calcium, magnesium, phosphorus, potassium
and sodium by the Provincial Department of Agriculture of the Western Cape (Elsenberg) according
to methods given by The Non-affiliated Soil Analysis Work Committee (1990). Physical and biological
soil crusts were noted on each sample plot and the relative cover of herbivore dung on the soil
surface was estimated per plot.
Data analysis
Plant canopy cover and soil parameter values were analyzed using two-tailed t-tests after square
root transformation of the data (Krebs 1989). Frequency of occurrence was analyzed using the Fisher
exact test. The step-up false discovery rate correction was applied to the results of multiple tests
where appropriate (García 2004). Results for plant cover and species height were compared with
those of the independent Indicator Species Analysis of Dufrêne & Legendre (1997) according to
McCune & Grace (2002). The Shannon-Wiener index of diversity was calculated as in Krebs (1989),
but using ln base e. The values of this index were also converted to their number equivalents which
expresses a species-neutral diversity (Jost 2009). This was used as the alpha diversity for LU and HU.
Gamma diversity was calculated for the whole study site by pooling the observations. Independent
beta diversity across the contrast (Anderson et al. 2011) was calculated assuming the additive
partitioning of diversity (Legendre et al. 2005). Species accumulation curves were derived using
sample-based rarefaction (Mao Tau expected richness function in EstimateS 8.0) as described by
Colwell et al. (2004). Evenness was calculated as the ratio of diversity index to maximum possible
diversity index for the given number of species (Krebs 1989). The more frequent species were
classed according to their affiliation with either the Grassland or Nama-Karoo Biomes, or both, based
on distribution records from PRECIS (Gibbs Russell & Arnold 1989), ACKDAT (Rutherford et al. 2003)
and other SANBI-curated databases.
Results
Mean total canopy cover declined significantly with heavy grazing; canopy cover and height of
woody shrubs increased significantly whereas that of graminoid plants declined (Table 1). Cover of
perennial graminoids declined from 92.9% on LU to 43.5% on HU (P = 3.0 x 10-9) whereas that of
annual graminoids increased from 0.04% on LU to 9.9% on HU (P = 0.0024). Other significant
differences include an increase on HU of cover of forbs, prostrate plants and plants with leafy stem
and a decline in that of geophytes, plants of erect habit and tussock plants (Table 1). A regrouping of
the species into life-forms (Raunkiaer 1934; Rutherford & Westfall 1994) shows heavy grazing to be
associated with a substantial increase in cover of chamaephytes and therophytes and a sharp decline
in that of hemicryptophytes (Fig. 2).
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Table 1. Species diversity, cover and heights of plants
and plant guilds under low and high utilization grazing
intensities. Total number of species was 80 with 58 on
LU and 64 on HU.
Parameters
LU
Mean
Mean number of species/50 m2
Mean Shannon-Wiener Index
Mean Species-neutral diversity
Mean Evenness Index
a
Mean Canopy cover (%)
Sum of species canopy cover (%)
Life history: Annual
b
Perennials
Growth form:Forb
Geophyte
Graminoid
Herbaceous
legume
Woody
Habit:
Erect
Prostrate
Architecture: Leafy stem
Rosette
Stoloniferous
Tussock
Mean heightsa:
Graminoid
Woody (cm)
HU
SE
Mean
P
SE
21.43
1.19
3.32
0.39
84.93
97.79
0.42
97.38
0.40
1.82
92.95
0.33
1.06
0.04
0.15
0.01
1.73
2.78
0.13
2.77
0.13
0.43
2.50
0.06
24.36
1.47
4.49
0.46
56.29
78.62
11.28
67.34
1.59
0.30
53.36
0.14
0.91
0.08
0.30
0.03
1.89
1.76
3.10
2.53
0.36
0.04
1.92
0.03
0.046
*0.0047
*0.0027
*0.025
-11
*9.8 x 10
*5.7 x 10 -6
*0.0025
-8
*2.3 x 10
*0.0052
*0.0019
-12
*2.1 x 10
*0.0056
2.30
97.65
0.15
3.03
1.86
92.92
64.9
32.4
0.24
2.79
0.11
0.29
0.42
2.50
1.9
1.8
23.23
77.13
1.48
24.99
1.75
0.02
51.86
23.3
43.0
1.53
1.65
0.37
1.75
0.78
0.02
1.78
2.6
1.8
*2.2 x 10
-6
*1.9 x 10
*0.0021
-11
*1.4 x 10
0.63
0.34
-13
*4.0 x 10
*2.9 x 10 -12
-4
*3.1 x 10
-12
LU - low utilization grazing intensity. HU: high utilization grazing
intensity. P: probability value. SE: Standard error. a Mean
canopy cover (not overlapping) and mean heights. *
Significantly different at P < 0.05 after False Discovery Rate
b
correction. Aristida congesta was regarded as a short-lived
Fig. 2. Relative contribution of cover of life forms for a) low-utilization
(LU) and b) high utilization (HU) grazing intensities.
Mean species richness or density increased with heavy grazing, but this was not significant
(Table 1). However, owing to a significant increase in mean species evenness, species diversity
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indices were significantly raised on HU (Table 1). Species richness increased significantly on HU for
graminoids (P = 0.027), annuals (P = 0.010), and plants of prostrate habit (P = 3.0E-05). The change in
species composition with heavy grazing, as reflected by the Sorensen’s dissimilarity index, indicated
a species turnover of 31%. Beta diversity between LU and HU constituted only 2.4 of the 7.2 for
gamma diversity.
Of the 29 species present in at least half the sample plots on either LU or HU, the cover of 19
species was significantly different (Table 2). Almost all shrub species in this group (Aptosimum
marlothii, Chrysocoma ciliata, Eriocephalus ericoides and Pentzia incana) increased in cover on HU.
Graminoids responded variously. Of the perennial grass species, Aristida diffusa and Heteropogon
contortus decreased in cover but that of Aristida congesta (here regarded as a short-lived perennial),
Eragrostis lehmanniana and Stipagrostis obtusa increased. The significantly different annual grass
species (Enneapogon desvauxii and Tragus berteronianus) increased in cover on HU. E. desvauxii
flourished on HU and appeared locally extinct on LU. Forbs varied with an increase in cover of
Limeum sulcatum and Tribulus terrestris but apparent local extinction of Helichrysum lineare on HU.
The geophyte Trachyandra saltii decreased significantly in cover on HU. The only alien (exotic)
species (Chenopodium album) recorded was limited to HU with a very low mean cover (< 0.01%).
Table 2. Frequency, canopy cover and height of species with a frequency of 50% and greater on either low or high utilization grazing intensities
Species
Amphiglossa sp.
Aptosimum marlothii
Aristida congesta
Aristida diffusa
Chrysocoma ciliata
Commelina africana
Cyperus usitatus
Dicoma capensis
Apocynoideae sp.
Enneapogon desvauxii
Eragrostis lehmanniana
Eriocephalus ericoides
Felicia muricata
Fingerhuthia africana
Helichrysum lineare
Helichrysum lucilioides
Heteropogon contortus
Indigofera alternans
Jamesbrittenia albiflora
Kohautia cynanchica
Lessertia cf. pauciflora
Limeum sulcatum
Pentzia incana
Polygala leptophylla
Stipagrostis ciliata
Stipagrostis obtusa
Trachyandra saltii
Tragus berteronianus
Tribulus terrestris
Growth form
Woody
Woody
Graminoid
Graminoid
Woody
Forb
Graminoid
Forb
Forb
Graminoid
Graminoid
Woody
Woody
Graminoid
Forb
Woody
Graminoid
Herbaceous legume
Woody
Forb
Herbaceous legume
Forb
Woody
Woody
Graminoid
Graminoid
Geophyte
Graminoid
Forb
Frequency
LU HU P
14
13
14
14
13
7
4
6
5
0
14
8
7
7
11
13
12
14
13
13
7
0
2
12
5
4
14
1
1
14
14
11
4
10
7
9
8
7
11
14
14
0
4
0
13
6
13
6
6
6
7
13
11
13
10
13
8
11
1.00
1.00
0.22
*0.0002
0.33
1.00
0.13
0.71
0.71
*0.0001
1.00
*0.016
*0.0058
0.44
*0.0001
1.00
*0.046
1.00
*0.013
*0.013
1.00
*0.0058
*0.0001
1.00
*0.0044
0.057
1.00
*0.013
*0.0003
Canopy cover (%)
LU
HU
Mean
SE
Mean
0.790
0.257 1.048
0.122
0.052 0.375
22.490 4.448 1.576
56.000 7.799 0.143
0.165
0.107 2.520
0.094
0.070 0.074
0.037
0.032 1.471
0.006
0.006 0.121
0.018
0.026 0.017
0.000
0.000 9.784
11.429 2.231 39.857
0.340
0.245 11.036
0.009
0.012 0.000
0.371
0.471 0.050
0.002
0.001 0.000
0.121
0.055 0.805
2.475
1.763 0.056
0.222
0.051 0.117
0.346
0.183 0.073
0.105
0.044 0.025
0.030
0.022 0.020
0.000
0.000 0.021
0.019
0.023 5.550
0.029
0.007 0.070
0.089
0.097 0.274
0.016
0.013 0.080
1.811
0.724 0.256
0.001
0.001 0.024
0.003
0.005 1.176
P
SE
0.346
0.053
0.607
0.080
1.203
0.035
0.774
0.071
0.008
2.873
2.767
1.072
0.000
0.043
0.000
0.388
0.021
0.026
0.042
0.012
0.009
0.008
1.828
0.024
0.120
0.027
0.040
0.007
0.326
0.59
*3.4 x 10-4
-9
*1.3 x 10
*5.7 x 10-11
*0.044
0.71
*0.045
0.11
1.00
*0.0025
-10
*7.2 x 10
*6.3 x 10-10
0.22
0.25
*0.0025
0.068
*0.011
*0.013
*0.025
*0.010
0.51
*0.027
*0.0030
0.119
0.15
*0.037
*0.0014
*0.0054
*0.0021
Height (m)
LU
Mean SE
0.29 0.02
0.26 0.01
0.39 0.02
0.77 0.02
0.29 0.02
0.23 0.03
0.07 0.02
0.09 0.01
0.15 0.02
0.49 0.01
0.43 0.03
0.21 0.04
0.53 0.03
0.04 0.00
0.21 0.02
0.54 0.04
0.09 0.01
0.22 0.02
0.19 0.01
0.17 0.01
0.21 0.00
0.31 0.03
0.55 0.08
0.37 0.05
0.62 0.03
0.25 0.10 -
HU
Mean
0.28
0.23
0.27
0.71
0.35
0.12
0.07
0.06
0.12
0.10
0.32
0.48
0.31
0.23
0.27
0.06
0.18
0.19
0.13
0.07
0.36
0.23
0.39
0.20
0.42
0.10
0.05
P
SE
0.02
0.01
0.02
0.09
0.02
0.02
0.01
0.00
0.02
0.01
0.01
0.02
0.07
0.03
0.05
0.01
0.04
0.02
0.02
0.00
0.02
0.03
0.04
0.02
0.03
0.02
0.01
0.53
0.084
-5
*2.6 x 10
0.54
0.058
*0.011
0.93
0.063
0.23
-9
*1.0 x 10
0.24
*0.045
0.59
*0.0018
*0.041
0.32
0.81
0.11
-7
*3.6 x 10
*0.049
0.13
*0.024
*7.9 x 10-5
-
LU: low utilization grazing intensity. HU: high utilization grazing intensity. P: probability. *: Significantly different at P < 0.05; SE: Standard error.
All but two of the above 29 species were fully identified and could be classed according to
their affiliation with either the Grassland or Nama-Karoo biomes, or with both biomes (Table 3).
There were about twice as many species affiliated with the Nama-Karoo Biome than with the
Grassland Biome. Species that increased significantly in cover on HU were split fairly evenly between
affiliation with either the Nama-Karoo or with both biomes. No species with sole Grassland affinity
increased significantly with heavy grazing. Species where cover declined significantly on HU (i.e.
significantly greater on LU) were split equally between affiliation with either Grassland or with both
biomes. No species with sole Nama-Karoo affinity decreased significantly with heavy grazing. Most
of the remaining species which showed no significant difference between LU and HU were of NamaKaroo affiliation.
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Table 3. Grazing response of speciesa according to
their biome affiliation.
Grazing response (canopy cover) Biome affiliation
Increase with heavy grazing
Nama-Karoo
Nama-Karoo and
Grassland
Aptosimum marlothii
Cyperus usitatus
Chrysocoma ciliata
Eragrostis lehmanniana
Enneapogon desvauxii
Limeum sulcatum
Eriocephalus ericoides
Tragus berteronianus
Pentzia incana
Tribulus terrestris
Grassland
Stipagrostis obtusa
Decrease with heavy grazing
No significant change
Aristida congesta
Helichrysum lineare
Aristida diffusa
Heteropogon contortus
Indigofera alternans
Jamesbrittenia albiflora
Kohautia cynanchica
Trachyandra saltii
Dicoma capensis
Felicia muricata
Commelina africana
Helichrysum lucilioides
Fingerhuthia africana
Lessertia cf. pauciflora
Polygala leptophylla
Stipagrostis ciliata
a
Species occurring in at least 50% of sample plots
on either low- or high-utilization grazing
intensities.
Concentrations of soil phosphorus, potassium and sodium increased significantly with heavy grazing
whereas those of calcium and magnesium declined (Table 4). There was no significant difference in
organic carbon and ammonium nitrogen. The mean cover of herbivore dung was 1.43% on HU which
was significantly greater (P = 3.5 x 10-6) than the 0.004% on LU. Patches of cyanobacterial soils crusts
were present to a greater or lesser degree on all plots on LU and were absent on HU where only
weak, brittle physical soil crust were noted.
9
Table 4. Soil properties under low and high utilization grazing intensities.
Parameters
LU
Mean
Clay (%)
Silt (%)
Sand (%)
pH (KCl)
Resistance (ohm)
Carbon (%)
Ammonium nitrogen (%)
HU
SE
Mean
P
SE
1.60
1.40
97.00
5.40
0.16
0.31
0.39
0.04
1.90
1.00
97.10
5.29
0.10
0.00
0.10
0.03
0.14
0.22
0.81
0.043
1924.00
0.32
0.034
159.55
0.04
0.003
1981.00
0.39
0.046
98.10
0.04
0.005
0.77
0.19
0.058
Calcium (cmol(+) kg-1)
4.11
0.14
3.58
0.07
*0.0055
Magnesium (cmol(+) kg-1)
2.82
0.09
2.50
0.07
*0.010
56.90
2.81
76.00
4.77
*0.0037
114.60
4.57
198.80
14.93
25.30
0.62
28.00
0.63
Phosphorus (citric acid mg kg -1)
Potassium (mg kg -1)
Sodium (mg kg-1)
*2.4 x 10-4
*0.0068
LU: low utilization grazing intensity. HU: high utilization grazing intensity. P: probability value. SE: Standard error. * Significantly different
at P < 0.05 after False Discovery Rate correction.
Discussion
Our study’s grassland reference site corresponded to the most grassy state, namely ‘perennial
grasses in matrix of shrubs’ in the state and transition model of Milton & Hoffman (1994). Indeed,
with perennial grass cover about forty times that of shrub cover on the LU site it may be argued that
this state was even more grassy than expected by the model. The heavily grazed site (HU) does not
correspond with the state and transition model’s most degraded state in which the effects of heavy
grazing are exacerbated by extended drought. It corresponds rather to intermediate states in which
palatable and unpalatable shrubs co-occur and perennial grasses are reduced but remain present.
State and transition models are considered to characterize rangelands not at equilibrium. Evidence
from arid environments suggests that these systems are well described by the non-equilibrium
paradigm (Vetter 2004, 2005) in which, for example, land degradation is not viewed as a necessary
outcome of heavy communal grazing (Ward et al. 1998). However, there is also a growing
realization that most rangelands show elements of both equilibrium and nonequilibrium dynamics
(Vetter 2004, 2005; Todd &Hoffman 2009).
Possible biome correspondence of the LU and HU sites are subject to the biome criteria used.
Taking relative cover of Raunkiaer-type plant life forms as an equivalent measure of importance to
that used to distinguish biomes by Rutherford & Westfall (1994), it is evident that with
hemicryptophytes greater than 90% (Fig. 2) LU qualified as Grassland Biome, scale considerations
aside. On HU, with cover of hemicryptophytes below 90% but with the sum of hemicryptophytes and
chamaephytes greater than 75%, Nama-Karoo Biome was indicated. Caveats of how fully
representative this area is of the biome are possibly changed soil fertility and unusually high level of
plant diversity (see below), as well as possible influence of different grazing animals species and
length of grazing history, and even potentially unrepresentative phenology (Wessels et al. 2011).
Interpretation of findings relative to the grazing reversal hypothesis and some of its predecessors
are not necessarily straight forward. Comparison of results is hindered by the operational difficulties
in determining the evolutionary history of grazing of a type of vegetation or community (Oesterheld
& Semmartin 2011). Comparisons are also made difficult where the concept of diversity is used
10
without, or with only incidental, reference to its richness or evenness components (e.g. Huston
1979; Milchunas et al. 1988, Milchunas & Lauenroth 1993; May et al. 2009) or where diversity and
richness are used (or referenced) interchangeably and species evenness is not mentioned (e.g.
Proulx & Mazumder 1998; Cingolani et al. 2005) despite the potential influence of the lastmentioned on diversity. ‘Diversity...as measured by species richness...’ (Cingolani et al. 2005) may be
widely used but remains potentially confusing especially when ‘...evenness can account for more
variation in Shannon’s diversity index (H’) than richness...’ (Wisley & Stirling 2007). This problem of
context is exemplified by the results of the present study that showed a significant increase in
species diversity and evenness but showed no significant difference in species richness.
We assume the vegetation of our study site falls into the category of systems of low resources
(productivity) and long evolutionary history of grazing. If so, the significant increase found in species
diversity with heavy grazing does not appear to support the grazing reversal hypothesis. The
significant elevation in species evenness might reduce the community dominance hierarchy and
thereby possibly weaken control of ecosystem functioning and stability in terms of the mass ratio
hypothesis (Sasaki & Lauenroth 2011). Increased species evenness through grazing in a number of
systems elsewhere has been attributed to suppression of dominant species (Schultz et al. 2011).
Competitive release may in turn allow for an increase in species number. The lack of support of our
results for the grazing reversal hypothesis is perhaps not too surprising. Other studies on grazing and
species richness have shown conflicting results to the model predictions (e.g. Adler et al. 2005) and
have indicated the possibility of negative, neutral and positive effects of grazing on small-scale
species richness (Schultz et al. 2011).
Although there was a relatively modest level of species turnover with heavy grazing, all
significant increases in species cover were those associated with either the Nama-Karoo biome or
with this biome and the Grassland Biome; none were associated with only the less-arid Grassland
Biome. This appears to show that species from more arid areas are more resistant to grazing and
tends to support the convergence model of drought (aridity) and grazing resistance (Quiroga et al.
2010).
The much larger size of the pool of species in the Grassland Biome than that in the NamaKaroo Biome (Thuiller et al. 2006) appears to go together with a mean alpha (plot) diversity in
Grassland roughly twice that of Nama-Karoo (Cowling et al. 1989). Although ecotones may
theoretically be expected to raise species richness using species pools from both sides (Risser 1995;
Kark & van Rensburg 2006), this may not apply in a very broad ecotone where species availability
may be limited by excessive distance and other factors. Although the species richness of the
reference Grassland site at 700 m2 (Fig. 3) was well below the mean given for 1000 m2 for the
Grassland Biome (Cowling et al. 1989), the corresponding species richness at 700 m2 on the heavy
grazed site was above the mean given for 1000 m2 for the Nama-Karoo Biome. That this relatively
high richness seems to derive from increases in, and addition of, karroid species (see above) and not
be due to additional floristic elements from the Grassland Biome apparently negates any benefit
from the possibly available pools in its ecotonal position. The relatively high species richness may
simply be a consequence of comparing it with a mean that includes the more arid part of the NamaKaroo. There appeared to be no species that could be classed as exclusively ecotonal.
11
Fig. 3. Species accumulation curves derived by use of sample-based
rarefaction for low-utilization (LU) and high utilization (HU) grazing intensities.
Results of this study were in broad agreement with the findings of the global analysis of
responses to grazing for the plant traits of life history, canopy height, habit, architecture and the
graminoid growth form (Díaz et al. 2007). However, this study did not find a neutral response to
grazing in the woody and forb growth forms but rather a significant increase. This may be explained
by the inclusion of many diverse ecosystem types other than grassland in the global dataset that was
analyzed.
Although cover of perennial graminoids decreased significantly with heavy grazing, the
perennial Eragrostis lehmanniana on the contrary increased markedly under the same treatment. E.
lehmanniana is regarded as a vigorous competitor in disturbed areas (Beukes & Cowling 2000) and
was the major increaser with grazing in a range of long-term field experiments elsewhere in the
Northern Cape Province (O’Connor 1985). The grass, Aristida congesta, has often been regarded as
an indicator of overgrazing, especially where it occurs as an annual (Vorster 1982). In this study area
it was judged as a short-lived perennial grass and its marked decline with heavy grazing may have
resulted from the several-fold increase in the competitive E. lehmanniana, as suggested by Beukes &
Cowling (2000) under similar circumstances. The shrub with the most marked increase in cover with
heavy grazing was not one of the unpalatable species such as Chrysocoma ciliata but the moderately
palatable Eriocephalus ericoides (Esler et al. 2006). These and other species level changes, including
the few apparent local extinctions on HU (Felicia muricata and Helichrysum lineare) or on LU
(Enneapogon desvauxii and Limeum sulcatum), suggest support for recognizing two different species
pools in low-productivity systems with long evolutionary history of grazing: 1) a grazing-resistant
pool that increases in periods of high grazing intensity, and 2) a less grazing-resistant pool that
increases in periods of lower grazing intensity (Cingolani et al. 2005). It has been suggested that
availability of pools of woody plant species capable of encroaching on grassland-type systems vary
considerably across the globe (Ward 2005).
The clear dominance of the grass Aristida diffusa on LU helped confirm the status of the
grassland reference site as a lightly grazed and not as an ‘unnatural’ ungrazed site where A. diffusa
would have been expected to be much diminished (Bosch & Gauch 1991).
The significantly higher cover of dung with heavy grazing is as expected with higher stocking
rates (du Toit et al. 2008). There appeared to be a fertilization effect of the soil with significant
increases in phosphorous (from dung, Williams & Haynes 1995) and in potassium and sodium (from
urine, Haynes & Williams 1992). Reasons for significant declines in calcium and magnesium with
heavy grazing are unclear. However, both these cations have been shown to not affect soil after
urine additions and have also been shown to decline along a degradation gradient in semi-arid
grassland (Henning & Kellner 1994) and with heavy grazing in semi-arid savanna (Moussa et al.
2009). Any elevation of nutrient resources through fertilization on HU is unlikely to allow
repositioning of this site within the grazing reversal model (Proulx & Mazumder 1998) owing to the
semi-arid conditions that continue to be limiting.
12
Conclusions
The significant increase in shrub cover in heavily grazed semi-arid grassland followed general
expectations found globally for similar systems and supported the Milton & Hoffman (1994) state
and transition model. This study confirmed that the supposed former large shift of grassland to
shrubby Nama-Karoo in the eastern upper Karoo can indeed be readily affected by heavy grazing. By
contrast, species richness did not follow the decline predicted by the grazing reversal hypothesis
(Milchunas et al. 1988; May et al. 2009) for vegetation with a long evolutionary grazing history in a
semi-arid, low resource area. The significant increase in species evenness, through possible
suppression of dominant species, resulted in a significant increase in species diversity with heavy
grazing. Our results tend to support the convergence model of drought (aridity) and grazing
resistance (Quiroga et al. 2010).
This study shows that the negative connotations for biodiversity that have often been
associated with intense grazing (Reid et al. 2010) have perhaps been exaggerated (see also Ward
2005). Although a herbivore-induced switch of biomes from Grassland to steppe-like Nama-Karoo
was indicated with markedly enhanced species diversity, the species turnover remained modest and
species richness was maintained.
Acknowledgements
We thank M. Vermeulen for permission to work on De Put; M. Mtubu of the Renosterberg
Municipality for permission to work on the Petrusville townlands; J. van Jaarsveld for local farming
information; and staff of the Rolfontein Nature Reserve for assistance in the field.
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