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1. Up to 73% of the world’s rangelands are degraded, and increasing demand for meat in developing countries and a growing human population are likely to exert even greater pressures on rangelands in the next 20–50 years. Restoration of rangeland grazing potential and resilience is therefore important, particularly in the face of climate change.
2. We investigated the influence of past stocking rates (from 1910 to 1987), rainfall, and current grazing regimes (from 1988 to 2008) on plant assemblages, grazing potential, and diversity of palatable species in southern Karoo rangelands, South Africa.
3. We used herbivore exclusion experiments to test whether resting rangeland for 20 years enables recovery of plant assemblages (where seed sources are present within 50 m), regardless of previous grazing history. Mean annual rainfall over this period was 15% higher than the mean annual rainfall for the preceding 80 years and included two exceptionally wet years.
4. While rainfall was a primary driver of total vegetation cover, grazing history explained differences in plant species composition: plots with shared historical grazing intensity were more similar than plots with the same grazing regimes between 1988 and 2008.
5. In historically heavily-grazed exclusion plots, cover of the palatable species Tripteris sinuata (formerly Osteospermum sinuatum) returned to levels comparable to that in both exclusion and lightly-grazed plots with a moderate grazing history. Five palatable species (Pteronia empetrifolia, Tetragonia spicata, Berkheya spinosa, Hereroa latipetala and Ruschia spinosa) failed to re-establish, however, despite the presence of seed-producing plants nearby. Furthermore, only cover of P. empetrifolia increased significantly in historically moderately-grazed plots. Cover of unpalatable plants (e.g. Pteronia pallens) increased in all plots over time.
6.Synthesis and applications. These findings suggest that present species composition of arid shrublands reflects historical management at time scales greater than 20 years. Despite high rainfall enabling the return of grazing potential through recovery of a single forage species, rest alone did not ensure the return of all palatable species, with implications for rangeland resilience. Restoring the full suite of palatable species over management timeframes will require more complex interventions such as reseeding or selective clearing.
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- Materials and methods
As much as 40% of the world’s terrestrial area now functions as cropland, pasture or rangeland (Foley et al. 2005), and it is estimated that up to 73% of the world’s rangelands may be degraded (Lund 2007). In the context of a human population that is increasing at a rate of nearly 225 000 per day (Population Reference Bureau 2008), coupled with higher per capita consumption patterns in developed and emerging economies, human society needs rangelands to be highly productive.
The causes of degradation and decline of palatable cover in arid rangelands have been debated for over two decades (see review in Vetter 2005). Much of the argument over rangeland models has hinged on whether abiotic factors are more important than biotic factors as determinants of species abundance and composition (Ellis & Swift 1988; Illius & O’Connor 1999). Regardless of the underlying causes, however, the ability of rangelands to recover is still poorly understood. While there has been increasing global interest in the restoration of degraded rangelands (e.g. Aronson et al. 1993; Holmgren & Scheffer 2001; Curtin 2002), there are few long-term studies. Briske, Fuhlendorf & Smeins (2003) reported on 11 recently published (i.e. post 1995) investigations on the effects of grazing or grazing exclusion on rangelands, of which only six had datasets longer than 20 years, and only one of these (O’Connor & Roux 1995) had been conducted outside the USA. Since then, only three additional herbivore exclusion studies longer than two decades have been published, one in the USA (Valone & Sauter 2005), one in Mongolia (Sasaki et al. 2009), and one in the Sahel (Miehe et al. 2010).
A primary goal of rangeland restoration is the return of palatable cover, with a concomitant increase in grazing carrying capacity. A key question for restoration efforts concerns the role of palatable cover in restoring productivity: how complete a suite of (historically present) palatable species is necessary for restoration to be considered successful? In other words, what levels of functional redundancy (the ‘degree to which organisms have evolved to do similar things’; Rosenfeld 2002, p. 157) should be present in a restored system? An ecosystem is considered restored when it ‘contains sufficient biotic and abiotic resources to continue its development without further assistance or subsidy’ and ‘will demonstrate resilience to normal ranges of environmental stress and disturbance’ (Society for Ecological Restoration International Science & Policy Working Group 2004, p. 3), but as with many such definitions, the devil lies in the details.
A high level of functional redundancy can be expected to contribute to increased resilience (the amount of disturbance a system – including its fundamental structure, process and functions – can endure before changing to a qualitatively different state; Gunderson & Holling 2002). Different species respond differently to disturbance or environmental stress, so the presence of a number of functionally similar species enables communities to maintain ecosystem function in response to perturbation (Walker, Kinzig & Langridge 1999). As ecosystems change, and dominant species decline or are lost, functionally equivalent species can be expected to substitute for them, enabling ecosystem function to be maintained (Tilman & Downing 1994; Naeem & Li 1997; Walker, Kinzig & Langridge 1999; Petchey & Gaston 2002; Memmott et al. 2004). The original definition of ecosystem resilience (i.e. as return time to an equilibrium point following a perturbation) suggests that more resilient rangelands should be restored more rapidly, and that more diverse rangelands should be more resilient.
Continuous selective grazing in natural rangelands typically affects palatable species, leading to declines (or even losses) of this functional group (O’Connor 1991; Milton 1995a). Two questions then arise – can rangelands regain their grazing productivity with rest, and is rest alone sufficient to enable the return of the full suite of palatable species. Yayneshet, Eik & Moe (2009) found rest was sufficient to restore plant species composition, biomass and cover to heavily-grazed rangelands, but a review by Curtin (2002) concluded that in general, rangelands do not recover with rest alone, and that climate is an overwhelming factor in determining restoration of rangelands. Holmgren & Scheffer (2001) argued that favourable weather conditions should be used opportunistically, along with grazing control, to bring about effective restoration of degraded rangeland. Regeneration of the desired components of rangelands has been shown to be triggered by good rainfall events in many arid and semi-arid systems (Gutiérrez et al. 1997; Holmgren et al. 2001). These ideas are consistent with the consensus that restoration attempts are most effective when managers can restore and sustain natural processes, using appropriate climatic conditions and disturbance factors that interact with grazing to elicit desired ecological effects such as shifts in vegetation composition (Westoby, Walker & Noy-Meir 1989; Curtin 2002).
In the arid rangelands of South Africa, historical livestock census data show that stock numbers peaked in the country in the 1920s and have subsequently fallen to half the peak density value (Milton 1995a). It has been argued that these livestock declines were driven by a reduction in forage availability (Dean et al. 1995), rather than by a decline in the market for wool and meat or by stock reduction policies. Despite scientific advances over the past couple of decades, our knowledge of long-term rangeland recovery in southern Africa remains limited. In 1988, a set of exclusion plots was established at our study site in the southern Karoo. Our 20 year data set incorporates plant transects from areas that experienced different grazing intensities and histories, dating back a century. Between 1988 and 2008, mean annual rainfall was 15% higher than the mean for the preceding 80 years, and included two exceptionally high rainfall years, allowing the effects of above-average rainfall on rangeland recovery to be compared with and without grazing exclusion. Although most species in our study plots are ‘medium’ (5–20 years) to ‘very long-lived’ (>80 years) (Wiegand et al. 2000), we anticipated some changes in plant species composition over this time frame.
In this study we asked, using two decades of grazing exclusion:
How have plant assemblages in areas with historically different grazing intensities changed in response to grazing exclusion?
Does rest alone lead to return of palatable plant cover (grazing productivity), regardless of whether grazing in the past was heavy or moderate?
Given above-average rainfall and a seed source within 50 m, do all species depleted by grazing return to rested rangelands within twenty years, i.e. do the rangelands exhibit resilience in timeframes of management or ownership?
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- Materials and methods
A total of 31 perennial plant species was surveyed in the permanent plots used in this study. Long-term grazing history (prior to 1987, i.e. between 1910 and 1987; hereafter referred to as ‘grazing history’) emerged as a significant factor driving the observed similarities in species composition across all grazing treatments (‘grazing treatments’ = grazing between 1987 and 2007; ANOSIM: Global R = 0·31, P < 0·01). Across all grazing history groups, however, treatments were not significantly different from each other (anosim: Global R = −0·028, P > 0·05). Inspection of the dendrogram produced from the cluster analysis confirmed these results (Fig. 1).
Figure 1. Plots that were heavily grazed in the past were more similar in species composition, regardless of grazing regime between 1988 and 2008. Heavily-grazed plots differed from those that had been moderately grazed in the past by 59·65%. Species that explained 66% of this dissimilarity are listed in Table 2.
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Five species explained almost two-thirds of the differences in grazing history (Table 2). As a rule, palatable species had higher percentage cover in areas that had histories of light grazing. The four species responsible for almost 60% of the dissimilarity were all medium to long-lived or long-lived. However, only T. sinuata and B. ciliatus could be considered good discriminating species, with average percentage cover of T. sinuata being nearly 40% greater in plots of moderate grazing history compared with those of heavy grazing history (5·4% vs. 7·4%, respectively). Similarly, B. ciliatus had an average percentage cover seven times higher in plots of heavy grazing history compared to those with a history of moderate grazing (3·5% vs. 0·5%, respectively).
Table 2. Results from simper analysis, showing species that explained up to two-thirds of the dissimilarities in plant assemblages with different grazing histories. Species with high Dissimilarity/SD ratios (‘Ratio’ in this table) contributed consistently across replicates. Species for which this ratio was greater than 1·3 can be considered as good discriminating species (Clarke & Warwick 1994); these taxa are underlined in the table. ‘Short-lived’ species are those that live for less than 5 years; ‘medium-lived’ for 5–20 years; ‘long-lived’ for 20–25 years and ‘very long-lived’ for greater than 50 years
|Species||Heavy Grazing history||Moderate grazing history||Average Dissimilarity||Ratio||Contribution %||Cumulative %|
|Average % cover||Average % cover|
|Tripteris sinuata(very palatable, medium to long-lived)||5·4||7·4||11·82||1·36||19·8||19·8|
|Pteronia pallens (poisonous, very long-lived)||10·9||6·2||10·24||1·14||17·2||37·0|
|Ruschia spinosa (palatable, long-lived)||0·3||4·3||7·12||1·28||11·9||48·9|
|Drosanthemum montaguense (slightly palatable, long-lived)||3·0||1·3||5·54||1·12||9·3||58·2|
|Brownanthus ciliatus(unpalatable, medium-lived)||3·5||0·5||4·97||1·48||8·3||66·5|
Plots with a history of heavy grazing had an average similarity of 45·5% (Table 3). Pteronia pallens accounted for over half (53·3%) of the similarity between historically heavily-grazed plots, regardless of whether grazing had been excluded or not over the last 20 years. Plots with a moderate grazing history had an average similarity of 49·9%, and Tripteris sinuata was the only species that could be considered a good discriminating species, explaining 34·3% of the similarity, regardless of grazing treatment (Table 3).
Table 3. Species that explained 90% of the similarity between plots with a history of heavy grazing, regardless of grazing treatment, or moderate grazing, regardless of grazing treatment (simper analysis). Species with high Similarity/SD ratios (‘Sim/SD’ in this table) greater than 1·3 can be considered good discriminating species (Clarke & Warwick 1994), and are underlined in the table
|Group: Heavy grazing history (Average similarity: 45·48)|
|Species||Average abundance||Average similarity||Sim/SD||Contribution %||Cumulative %|
|Pteronia pallens(poisonous, very long-lived)||0·11||24·22||1·79||53·26||53·26|
|Brownanthus ciliatus (unpalatable, medium-lived)||0·03||6·68||1·14||14·69||67·95|
|Tripteris sinuata (very palatable, medium to long-lived)||0·06||5·17||0·75||11·36||79·31|
|Galenia fruticosa (slightly palatable, medium-lived)||0·03||2·85||0·66||6·26||85·58|
|Malephora lutea (unpalatable, short-lived)||0·02||2·75||0·64||6·04||91·62|
|Group: Moderate grazing history (Average similarity: 49·91)|
|Species||Average abundance||Average similarity||Sim/SD||Contribution %||Cumulative %|
|Tripteris sinuata(very palatable, medium to long-lived)||0·07||17·14||1·6||34·34||34·34|
|Pteronia pallens (poisonous, very long-lived)||0·06||14·19||1·06||28·44||62·78|
|Ruschia spinosa (palatable, long-lived)||0·04||7·63||0·89||15·29||78·08|
|Galenia fruticosa (slightly palatable, medium-lived)||0·02||3·81||0·99||7·64||85·72|
|Pteronia empetrifolia (palatable, long-lived)||0·03||3||0·53||6·02||91·74|
Species responses to changing stocking rates over time
Although all species in all treatments showed best-fit relationships with time (percentage cover vs. months) that were linear, only some of these were statistically significant after application of sequential Bonferroni tests to correct for false detection rates. Brownanthus ciliatus cover appeared to decline over time, but the decline was not statistically significant. Drosanthemum montaguense cover increased with time (Spearman Rank R = 0·378, P = 0·0017) in plots that had a history of heavy grazing in which sheep grazing was ongoing. Cover of Pteronia pallens was weakly but positively correlated with time (Spearman Rank R = 0·254; P = 0·006) in plots with a history of moderate grazing from which all grazers had been excluded for the last 20 years.
For all treatments and all grazing histories, cover of the palatable T. sinuata increased linearly with time (Fig. 2). We therefore used a linear-mixed effects model to assess changes in percentage cover of this species with time. Model selection using AIC selected our hypothesis (‘hypothesized model’) that moderately-grazed treatments would have a slower rate of increase than historically heavily-grazed exclusion plots, but that cover in historically heavily-grazed plots that continued to be grazed by sheep had increased slower than all other treatments, as the best model (Null model AIC = 1717·9; Percentage cover of T. sinuata increasing at the same rate within grazing histories: AIC = 1713·9; Percentage cover of T. sinuata over time increasing at the same rate within grazing treatments: AIC = 1708·3; Percentage cover of T. sinuata increasing at the same rate within grazing histories and treatments: AIC = 1672·2; Hypothesized model: AIC = 1644·1).
Figure 2. Relationship between percentage cover of the palatable shrub T. sinuata with time over nearly 20 years. The relationship between time and cover revealed a significantly greater increase in cover in plots that had a history of heavy grazing, from which grazing was excluded, compared to all other plots. Plots that had a history of heavy grazing and continued grazing increased significantly slower than all other plots. The rates of increase of T. sinuata cover in plots with a history of moderate grazing were not significantly different to each other.
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A comparison of average percentage cover of Tripteris sinuata in 2008 showed that only the historically heavily-grazed site that had continued to be grazed by sheep between 1988 and 2008 had a significantly lower percentage cover (Kruskal–Wallis One-way anova Chi-square = 9·67, P = 0·046, n = 5; Multiple comparisons with t-distributions), whilst the exclusion treatment plot on the historically heavily-grazed site was not significantly different to the historically moderately-grazed sites for all current grazing treatments.
Pteronia empetrifolia, another palatable species, showed a very different response. In areas that had been heavily grazed, even in exclusion plots, P. empetrifolia failed to re-establish. However, in areas that had been moderately grazed, cover increased significantly between 1988 and 2008 (Table 4). A comparison of average percentage cover in exclusion plots between 1988 and 2008 of four other palatable species (Tetragonia spicata, Berkheya spinosa, Hereroa latipetala, Ruschia spinosa) on the farm with a history of heavy grazing showed that 20 years of rest did not result in any significant change in percentage cover, which remained close to zero for all four species. Although these palatable species did not return to plots with a history of heavy grazing, most of these species also did not exhibit change in cover between 1988 and 2008 in plots with historically moderate grazing (Table 4).
Table 4. Changes in median percentage cover of five palatable species between 1988 and 2008 in exclusion plots that had been moderately grazed in the past. Only Pteronia empetrifolia showed a significant increase in cover. In plots that had been heavily grazed in the past, mean percentage cover remained close to zero for all five species
|Species||1988: Median cover (Lower and upper quartiles)||2008: Median cover (Lower and upper quartiles)||Mann–Whitney U||P|
|Pteronia empetrifolia||0·5% (0%; 2%)||3·5% (0%; 7%)||111||P < 0·05; (n = 38)|
|Hereroa latipetala||1% (0·5%; 2%)||1% (0%; 1·5%)||188·5||P > 0·05; (n = 40)|
|Tetragonia spicata||0·05% (0%; 0·1%)||0·01% (0%; 0·2%)||212||P > 0·05; (n = 40)|
|Ruschia spinosa||6% (1%; 8%)||3·5% (1%; 6%)||242·5||P > 0·05; (n = 40)|
|Berkheya spinosa||0% (0%; 0%)||0% (0%; 0%)||240·5||P > 0·05; (n = 40)|