• communal land;
  • conservation;
  • insects;
  • lizards;
  • monitoring;
  • rapid biodiversity assessment


  1. Top of page
  2. Summary
  3. Introduction
  4. Study area
  5. Methods
  6. Results
  7. Discussion
  8. Acknowledgements
  9. References

1. Although it is widely assumed that protected areas conserve species that would not survive elsewhere, this assumption is seldom tested. The aim of the study was to determine the respective roles of a nature reserve and commercial and subsistence rangeland in preserving terrestrial arthropods and reptiles in xeric succulent thicket in the Eastern Cape, South Africa.

2. Faunal diversity on a nature reserve (the Great Fish River Reserve Complex) was compared with a historically overgrazed commercial farm, an intensively managed, conservatively stocked commercial farm and a communal grazing area. Terrestrial arthropods and reptiles were caught in paired pitfall traps inside and outside the nature reserve, specimens being sorted into recognizable taxonomic units (RTU). The taxa occurring in each land management unit were compared using RTU diversity, a hierarchical richness index (HRI), community similarity and uniqueness.

3. The nature reserve contained more RTU and a greater HRI than adjacent land for most arthropod groups, and also supported more unique taxa than the other study localities.

4. Snakes and lizards, in contrast, were almost twice as abundant in the communal grazing area as elsewhere, although reptile species diversity was similar at all localities.

5. Each individual locality contained fewer than 62% of the total number of arthropod RTU and 55% of the total reptile RTU; the communal grazing area contained only 37% of the total number of arthropod RTU collected in the study area.

6. Nineteen (73%) of the ant, 18 (69%) of the weevil, 60 (70%) of the spider and 12 (60%) of the grasshopper RTU and 10 (66%) of the reptiles were shared by the three land uses. Six new weevil species and probably several new arachnid species were collected, but all the new weevils were confined to the nature reserve.

7. The communal grazing area differed most from the nature reserve in richness and community composition, followed by the conservatively stocked commercial farm. The historically overgrazed commercial farm was most similar to the nature reserve.

8. The communal grazing area was characterized by xeric-adapted reptiles and predatory arthropods whose ranges are centred in the semi-arid parts of South Africa. In contrast, the nature reserve and commercial farms supported more mesic-adapted reptiles and herbivorous arthropods.

9.Synthesis and applications. The data show how protected areas are key to conserving those species that decrease under heavily grazed and disturbed conditions. However, they also illustrate that diverse land-use mosaics promote gamma diversity in the xeric succulent thicket of South Africa. Conservation policies that include protected space in the wider environment and conserve habitat diversity are likely to both promote regional richness and support scarce species.


  1. Top of page
  2. Summary
  3. Introduction
  4. Study area
  5. Methods
  6. Results
  7. Discussion
  8. Acknowledgements
  9. References

Land degradation and the corresponding changes in landscape and plant diversity are likely to reduce faunal diversity and abundance. This is of particular concern in ecosystems with a low resilience, such as the xeric succulent thicket of south-eastern South Africa (La Cock, Palmer & Everard 1990; Kerley, Boshoff & Knight 1999). In such situations, the existence of protected areas is often defended on the grounds that they conserve species that are under threat elsewhere. This assumption is seldom tested, and thus it is important to evaluate which types of organisms are dependent on protected areas for their survival, and which are adequately conserved outside protected areas. It is also necessary to establish which taxonomic groups are useful indicators of the efficacy of protected areas in xeric succulent thicket (cf. McGeogh, van Rensburg & Botes 2002).

Invertebrates play important roles in altering the structure and fertility of soils, pollinating flowering plants, feeding predators (Janzen 1987) and cycling nutrients (Seastedt & Crossley 1984; Greenslade 1992). Arthropods are potential indicators of subtle habitat change because they respond to the environment at a finer scale than larger organisms and require smaller habitat patches than larger animals for survival (Wiens & Milne 1989). They are the most abundant consumers in African savannas and, in some instances, have a greater biomass than vertebrates (Gandar 1982).

The impact of land use on reptiles may not follow the trends observed for plants and arthropods. Reptiles may be disadvantaged by dense vegetation and shade and are usually not directly dependent on plants for their survival. For example, a higher diversity of lizards occurs in deserts and semi-deserts than in high-rainfall areas (Pianka 1986), a trend opposite to the diversity patterns of plants, birds and arthropods (Rosenzweig & Abramsky 1993). Degraded sites can thus be expected to harbour more species and higher numbers of reptiles than sites where the vegetation and soils are well preserved.

Focusing on the xeric succulent thicket, in south-eastern South Africa, the objectives of this study were to:

  • 1
    compare the number of recognizable taxonomic units (RTU) and the hierarchical richness indexes (HRI) of reptiles and arthropods in a nature reserve with those of adjacent commercial and communal rangeland;
  • 2
    compare arthropod community similarity across sites with different land uses;
  • 3
    determine the number of unique and shared reptile and arthropod RTU between the land management units that were recognizable;
  • 4
    identify taxonomic groups that could act as useful indicators to monitor the effect of land management on biodiversity;
  • 5
    use this information to recommend conservation strategies for terrestrial reptiles and arthropods.

Study area

  1. Top of page
  2. Summary
  3. Introduction
  4. Study area
  5. Methods
  6. Results
  7. Discussion
  8. Acknowledgements
  9. References

The Great Fish River Nature Reserve Complex is situated in the Great Fish River valley, between Grahamstown and Alice in the Eastern Cape Province, South Africa (33°11′S, 26°38′E). The vegetation is variously termed valley bushveld (Acocks 1988) or xeric succulent thicket (La Cock, Palmer & Everard 1990; Low & Rebelo 1996). The nature reserve was designated in 1976 and enlarged in 1986. Prior to proclamation it was commercial farmland. The mean annual rainfall is 410–500 mm per annum, depending on elevation, and the vegetation consists of dense shrubs of medium height interspersed with either grasses or dwarf shrubs, depending on local rainfall and management history (La Cock 1992). The bedrock comprises Ecca shales and the derived soils are shallow and clay-rich with a high moisture-holding capacity. The run-off is high, except under bush clumps, where soil porosity is higher because of zoogenic activity and litter (Palmer & Avis 1994). The vegetation has an extremely low resilience to overgrazing compared with other savannas, and might be a relic of former, higher moisture regimes. The Albany centre of endemism (Van Wyk & Smith 2001) owes its status largely to the high proportion of endemic plants, notably geophytes and succulents, contained in succulent thicket. Concern has been raised over the level of transformation of xeric succulent thicket (La Cock, Palmer & Everard 1990).

Three localities in the nature reserve (NR) were compared to two adjacent commercial farms sites (CF1 and CF2) and an adjacent communal grazing area (CGA) (Fig. 1). Localities were selected to represent a broad and representative spectrum of land management styles, and were situated along a gradient of both disturbance and stocking rates.


Figure 1. A map of the study area, depicting the sampling localities. CF1 = the historically overgrazed commercial farm Hermanuskraal; NR1 = its corresponding site in the nature reserve. CF2 = conservatively stocked, intensively managed commercial farm, Bucklands; NR2 = its corresponding site in the nature reserve. CGA = communal grazing area; NR3 = its corresponding site in the nature reserve.

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The commercial farm Hermanuskraal (CF1) has a history of overgrazing. Between about 1820 and 1880 it was part of a military outpost (Coetzee 1994) and probably subjected to intensive continuous grazing by cattle and horses. Subsequently, it became commercial farmland and until 1950–60 was heavily stocked with sheep Ovis aries Linn. and domestic goats Capra hircus Linn. The conventional dogma of the era dictated that dense vegetation harbouring ticks (Ixodidae) had to be cleared, and that the ‘open’ form of xeric succulent thicket was more productive than the dense form. In this instance the impact of the change was particularly severe because of the small size of the property (about 100 ha). By 1960 the impacts of heavy grazing necessitated a lighter stocking rate. Even then the farm was subject to bouts of up to four times the recommended stocking rate (A. Davenport, Hermanuskraal, personal communication).

The commercial farm Bucklands (CF2) has a history of successful commercial farming and at the time of the study was conservatively stocked and intensively managed for commercial livestock production. Livestock (cattle and Angora goats) were rotated according to a fixed pattern, following a short-duration grazing system. The farm contained many camps arranged in the shape of wagon-wheel (the Savory method), with long rest periods interrupted by short bouts of intensive grazing. The property has been referred to as a ‘benchmark for vegetation condition assessment in the region’ (R. Dempsey, Department of Agriculture, Grahamstown, personal communication).

A communal grazing area (CGA) at Tyefu Location on the east side of the Great Fish River was very heavily stocked, typical of an open-access system without formal property rights. The area had a high human population density of about 70 people km−2 with more than 50% unemployment (Ainslie, Fox & Fabricius 1994). Human demand for wood for cooking, heating and building was high (Ainslie, Fox & Fabricius 1994). Thus, the four study localities could be considered along a gradient of stocking rates and degrees of disturbance by humans and livestock, in the order from high to low: CGA, CF1, NR, CF2.


  1. Top of page
  2. Summary
  3. Introduction
  4. Study area
  5. Methods
  6. Results
  7. Discussion
  8. Acknowledgements
  9. References

The faunal assemblages of the nature reserve were compared to that of two commercial farms, CF1 and CF2, and a communal grazing area (CGA), using matched sites inside and outside the nature reserve. Paired site I consisted of the commercial farm Hermanuskraal (CF1) and a corresponding locality in the nature reserve (NR1). Paired site II consisted of a locality on the commercial farm Bucklands (CF2) and a locality in the nature reserve (NR2). Paired site III comprised a locality in the communal grazing area (CGA) near the village of Ndwayana and a locality in part of the nature reserve (NR3) known as Selbourne (Fig. 1).

Paired localities were 50–100 m apart and had the same elevations, slopes, geological substrata and aspects. Reptiles and arthropods were collected in pitfall traps. A trap array comprised a 20-cm high and 10-m long drift fence with a pitfall trap, consisting of a 25-litre plastic bucket at one end and two funnel traps, one on each lateral side of the drift fence, at the other. Drift fences were buried about 5 cm deep to prevent arthropods from escaping under the fence. Four trap arrays on the same side of the boundary fence formed a trap site, while eight trap arrays, positioned on opposite sides of the boundary fence, together formed a paired site. Trap arrays belonging to the same trap site were positioned 150 m apart, while trap sites belonging to the same paired site were at least 200 m apart, i.e. at least 100 m from the boundary fence. Trap arrays were arranged in pairs of two across the boundary fence (Fig. 2), each pair being positioned at the same slope, aspect and elevation. Traps were opened for 6 days per month between October (early spring) and March (late summer) of the following year. Inspection of traps was synchronized as closely as practically possible.


Figure 2. A diagram of a paired trap site. A trap array comprised a 20-cm high and 10-m long drift fence with a pitfall trap, consisting of a 25-litre plastic bucket at one end and two funnel traps at the other end, one on each lateral side of the drift fence. Four trap arrays on the same side of the boundary fence formed a trap site, while eight trap arrays, positioned on opposite sides of the boundary fence, together formed a paired trap site. The drawing is not to scale.

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Reptiles were identified to species or subspecies level. All arthropods were classified to order or family and then categorized using rapid biodiversity assessment (RBA; Beattie, Majer & Oliver 1993) into a reference collection of numbered specimens distinguished by morphological characteristics, and so labelled as recognizable taxonomic units (RTU; Beattie, Majer & Oliver 1993). To test the validity of using RTU, spiders, scorpions and weevils were subsequently identified to species level by specialized taxonomists.

One-hundred and sixty-four reptiles and 8500 arthropods (excluding ants) were caught during 864 trap days. Ants were often encountered in large numbers and were not counted individually. Instead, a representative specimen of each ant RTU at each respective trap was collected on each trap day. Calculations for ants were thus based on daily frequencies of occurrence rather than numerical abundance. A limited number of higher taxonomic groups (orders and families) of arthropods were selected before the data were analysed, and these were chosen to represent a range of body sizes and feeding strategies. The selection of a taxonomic group depended on: (i) the number of possible finer taxonomic units within the taxon; (ii) the number of individual specimens of the taxon collected; and (iii) the contribution of the taxon towards reflecting a range of body sizes and feeding strategies in the study area. Using these criteria, eight taxonomic groups were selected: grasshoppers (Orthoptera: numerous families); crickets (Orthoptera: Gryllidae); spiders (Araneae: numerous families); scorpions (Scorpiones: Scorpionidae and Buthidae); whipscorpions (Amblypygi); solifuges (Solifugae); ants (Hymenoptera: Formicidae); and weevils (Coleoptera: Curculionidae). Scorpions, whipscorpions and solifuges were later lumped into a single group called large arachnids.

At the same paired localities where faunal data were collected, a survey of the structure and species composition of the vegetation was conducted. Only one commercial farm (CF1) was included in the vegetation survey. The percentage cover of the most palatable shrub species, Portulacaria afra Jacq. and Grewia robusta Burch., was estimated in a sample of bush clumps (dense clumps of woody vegetation dispersed in a matrix of either grass or dwarf shrubs) at each locality, as an index of range condition. An index of bush clump area was obtained by placing a bush clump in an imagined rectangle then calculating the area of the rectangle. The line intercept method was used to determine the clump to interclump ratios in 10 transects of 500 m at each locality. The plant species in each clump were recorded.

data analysis

The hierarchical richness index (HRI; French 1994) was used to compare diversity between trap sites belonging to a paired site. This index was selected because: (i) it incorporates both taxonomic diversity and abundance in a single measure of richness that is less confusing than other indexes; and (ii) it ranks sites according to easily definable, objective criteria.

HRI values were calculated for each taxonomic group at each trap array, as follows. First, all RTU of a particular taxonomic group collected at a trap array were sorted in descending order of abundance. The most abundant RTU was then allocated a weight or hierarchical index value (i) of 1, the second most abundant given an i-value of 2 and so on. The hierarchical value of each RTU was then multiplied by the number of individuals of that RTU to produce a score (s). Next, s was summed for all RTU at each locality, to produce HRI = Σ(si) (French 1994). HRI values were calculated separately for each of the four trap arrays at each locality, and means and standard errors were calculated for the trap site.

To determine the degree of similarity between trap sites, Sorenson's index of community similarity (CC; Sorenson 1948) was calculated at each paired site for each arthropod taxonomic group. CC values were not calculated for reptiles due to small sample sizes. The formula used was CC = 2s/(a + b), where s is the number of RTU that are shared by both trap sites at a paired site, a the number of RTU in trap site a, and b the number of RTU in site b. For each trap site a uniqueness index, U = Ur/Rtot, was calculated, where Ur is the number of RTU that are unique to the trap site and Rtot the total number of RTU recorded at a paired trap site. To determine whether the U-values were significantly greater than zero, 90% confidence intervals were calculated as ±z(1−α/2k)(√U(1 − U)/n), where U is the proportion of RTU unique to the nature reserve, k the number of simultaneous estimates being made, and n the total number of RTU for the paired trap site. The purpose of using a Bonferroni normal statistic (1 – α/2k) instead of the standard normal statistic (1 – α/2) is to increase the confidence interval to compensate for the fact that more than one simultaneous estimate is involved (Neu, Byers & Peek 1974).

Assessment of which taxonomic groups were potentially good biodiversity indicators was based on the frequency with which their CC, HRI or U-values differed substantially between trap sites, as this would indicate sensitivity to disturbance and intensity of herbivory. Substantially different was defined as either: (i) HRI values with non-overlapping standard errors; (ii) CC values of 0·75 and smaller; or (iii) U-values that were significantly greater than zero, according to their Bonferroni Z-statistic at the 90% confidence level. A sensitivity score, calculated as the sum of the frequencies with which a group showed a substantial difference for each respective index at each respective paired site, was calculated for each group.


  1. Top of page
  2. Summary
  3. Introduction
  4. Study area
  5. Methods
  6. Results
  7. Discussion
  8. Acknowledgements
  9. References

key vegetation differences between sites

There were key differences between the vegetation in the nature reserve, on adjacent commercial farms and in the communal area. The bush clumps in the nature reserve were significantly larger than those on commercial farms and much larger than those in the communal area, corresponding to higher clump to interclump ratios; the mean percentage cover of the palatable P. afra was higher in the nature reserve than on the commercial farm, compared with zero cover in the communal area. The mean number of woody plant species per bush clump was higher in the nature reserve than on the commercial farm, which harboured a greater diversity of species than the communal area (Table 1).

Table 1.  Comparison between vegetation in the nature reserve and on unconserved land. NR = nature reserve sites; CF = commercial farm site; CGA = communal grazing area. A Wilcoxon test was used in all instances. *** P ≤ 0·01 ** P ≤ 0·05, and * P ≤ 0·1; ± standard error
  • Coarse estimates, based on interviews with managers and agricultural extension officers, and inspection of game and livestock census data. Stocking rates vary from year to year, depending on rainfall.

  • NA, not applicable.

Mean percentage aerial cover of P. afra5038·4 ± 6·5 1·7 ± 0·7***9·915·3 ± 5·1 0·0 ± 0·0***9·9
Mean bush clump size (m2)50 8·5 ± 1·5 5·2 ± 1·0**2·110·4 ± 1·2 1·7 ± 0·2***6·9
Mean clump to total transect ratio100·62 ± 0·020·45 ± 0·03***6·00·50 ± 0·030·12 ± 0·01***6·1
Mean number of non-grass plant species per bush clump50 9·2 ± 0·5 8·2 ± 0·4*1·411·4 ± 0·7 3·8 ± 0·1***9·9
Grazing intensity (ha/large animal unit)NA177–8NA 174–6NA 

overall comparison between the nature reserve and unprotected land

The nature reserve contained the highest arthropod diversity among localities and had more RTU (156 compared with 139) and a higher abundance of arthropods than the three adjacent unprotected localities combined (Table 2). The nature reserve also had larger mean HRI values than adjacent land in most arthropod taxonomic groups (with the exception of ants and grasshoppers at paired site I; Table 3).

Table 2.  Number of recognizable taxonomic units (RTU) and number of specimens (No.) trapped at three paired sites, as well as at nature reserve sites and all unconserved localities combined. Paired site I = comparison between the historically overgrazed commercial farm (CF1) and its corresponding site in the nature reserve (NR1). Paired site II = comparison between conservatively stocked commercial farm (CF2) and its corresponding site in the nature reserve. Paired site III = comparison between the communal grazing area (CGA) and its corresponding site in the nature reserve (NR3)
 Paired site IPaired site IIPaired site IIIOverall comparison
NR1CF1NR2CF2NR3CGANature reserveUnconserved land
  • *

    Daily frequency of occurrence, as opposed to number of individuals.

Crickets 3  87 2  46 3  95 3  94 2 91 3 21 3 273 4 161
Grasshoppers 8  4211  4412 14911 150 6 42 6  915 23317 203
Large arachnids10 151 9 140 7 253 7  8710236 7 9112 64011 318
Spiders45 67339 81549 62946 4473982131292782123681554
Weevils12  39 9  4517 12213  5512 87 8 4623 24819 146
Reptiles 8  29 7  20 8  19 8  15 7 25 8 5613  7312  91
Table 3.  Mean hierarchical richness indexes (HRI) and standard error (SE) for three paired trap sites. Paired site I = comparison between the historically overgrazed commercial farm (CF1) and its corresponding site in the nature reserve (NR1). Paired site II = comparison between conservatively stocked commercial farm (CF2) and its corresponding site in the nature reserve (NR2). Paired site III = comparison between the communal grazing area (CGA) and its corresponding site in the nature reserve (NR3). SE based on n= 4
 Paired site IPaired site IIPaired site III
  • *

    Calculated using daily frequencies as opposed to number of individuals.

Ants*175·534·0215·0 39·0260·5 33·9209·873·3 95·8 8·2 48·510·3
Crickets 24·0 2·6 12·8  2·1 32·5  6·0 25·5 5·0 20·0 5·1  6·8 1·3
Grasshoppers 19·8 5·4 25·0  9·1 90·0 16·8 73·010·5 35·011·0  4·5 2·2
Large arachnids109·012·8 76·8 14·6 87·5 14·5 39·514·7132·527·2 48·5 9·9
Weevils 24·8 3·8 20·2  5·1 96·5 17·1 36·510·2 59·5 3·6 22·3 6·6

Specialists identified 40 different weevil and 113 arachnid taxa. At least six of the weevils appeared to be undescribed species, with a specimen of an undescribed species of the genus Euretus, previously regarded as monotypic, being of particular interest. All new weevil taxa were collected only in the nature reserve. Some arachnids were also probably undescribed (A. Dippenaar, personal communication). The richness of large arachnids (scorpions, whipscorpions and solifugids combined) was substantially higher in the nature reserve than on unconserved properties at all paired trap sites.

All sites had a very similar species richness of reptiles (P > 0·1, χ2 = 0·57, n= 4) but almost twice as many individuals were captured in the communal grazing area than elsewhere (Table 2). None of the localities contained more than eight of the 15 reptile species collected in the study area as a whole (Table 4).

Table 4.  Lizard and snake species and number of specimens collected during the study period. Paired site I = comparison between the historically overgrazed commercial farm (CF1) and the nature reserve (NR1). Paired site II = comparison between conservatively stocked commercial farm (CF2) and the nature reserve (NR2). Paired site III = comparison between the communal grazing area (CGA) and the nature reserve (NR3). ** P ≤ 0·01; * P ≤ 0·05; (χ2 test)
 Paired site IPaired site IIPaired site III
Pedioplanis lineoocellata1713 1 112**3142·2
Mabuya capensis 3 1 1 1 2  11 
Pachydactylus maculatus 4 1 2 3 4   1 
Pachydactylus m. mariequensis 1 1 0 0 2  *4 
Gerrhosaurus typicus 0 0 0 0 0   6 
Mabuya varia 0 0 0 4 0   0 
Cordylus cordylus 0 0 0 3 1   0 
Varanus albigularis 0 0 3 0 0   0 
Agama a. atra 1 1 0 0 0   2 
Mabuya v. variegata 1 1 0 1 0   0 
Nucras taeniolata 0 0 1 1 0   0 
Acantias g. gracilicauda 0 0 1 0 0   0 
Psammophis notostictus 2 0 3 0 3   1 
Rhinotyphlops lalandei 0 2 0 1 1   0 
Prosymna s. sundevalli 0 0 3 0 0   0 
Total number of species 8 7 8 8 7   8 
Total number of specimens2920191525**5641·0
HRI6042634960 110 

The grasshopper and weevil communities on the nature reserve differed substantially from those on unconserved land, with CC values of 0·75 and 0·76, respectively. A CC of 0·75 and lower was regarded as the threshold for two sites to be classified as supporting different communities (cf. Whittaker 1972; Cowling et al. 1989). The communities of spiders, ants, crickets and large arachnids in the nature reserve and on unprotected land were broadly similar, with CC values of 0·82, 0·84, 0·86 and 0·96, respectively (Table 5).

Table 5.  Sorenson's index of community similarity (CC) for three paired trap sites, as well as all unconserved localities combined against the nature reserve (overall comparison). Paired site I = comparison between the historically overgrazed commercial farm and its corresponding site in the nature reserve. Paired site II = comparison between the conservatively stocked commercial farm and its corresponding site in the nature reserve. Paired site III = comparison between the communal grazing area and its corresponding site in the nature reserve. CC values of < 0·76 are indicative of different communities
 Paired site IPaired site IIPaired site IIIOverall comparison
Large arachnids0·710·840·820·96

Thirty-five arthropod RTU (20·1% of the total) were unique to the nature reserve. Of these, 18 were spiders, seven were weevils, six were ants, three were grasshoppers and one was a solifugid. Eighteen RTU (10·3% of the total), comprising eight spiders, five grasshoppers, three weevils, one cricket and one ant, occurred only on unprotected land. Two reptiles (one snake and one lizard) were unique to the nature reserve and two lizards occurred only elsewhere (Table 4). The taxonomic groups with the largest proportion of taxa confined to the nature reserve were weevils, ants and spiders (in that order). Grasshoppers had the largest proportion of RTU unique to unconserved properties, while the large arachnids had the highest proportion of shared RTU, i.e. RTU that occurred in the nature reserve and on unconserved land (Table 6).

Table 6. U-values, indicating the proportion of unique recognizable taxonomic units (RTU) and shared RTU in the nature reserve and each respective unconserved locality. U= proportions of unique RTU, relative to the total number of RTU (n) at each trap locality; ±95% Cl =±95% Bonferroni-z confidence limits. Paired site I = comparison between the historically overgrazed commercial farm CF1 and its corresponding site in the nature reserve. Paired site II = comparison between conservatively stocked commercial farm CF2 and its corresponding site in the nature reserve. Paired site III = comparison between the communal grazing area and its corresponding site in the nature reserve. Overall comparison = all localities on unconserved land compared to all localities in the nature reserve (NR). *Lower 90% confidence interval greater than zero
 Paired site IPaired site II Paired site IIIOverall comparison
Un95% ClUn95% ClUn95% ClUn95% Cl
NR0·05 0·100·24 0·20*0·45 0·24*0·23 0·18*
CricketsUnconserved0·00 30·000·25 40·460·33 30·580·25 40·46
NR0·33 0·580·25 0·460·00 0·000·00 0·00
GrasshoppersUnconserved0·27110·290·20150·220·25 80·330·25200·21*
NR0·00 0·000·27 0·24*0·25 0·330·15 0·17
Large arachnidsUnconserved0·09110·180·22 90·300·00100·000·00120·00
NR0·18 0·250·22 0·300·30 0·310·08 0·17
NR0·29 0·13*0·22 0·11*0·35 0·15*0·21 0·09*
NR0·31 0·27*0·38 0·23*0·47 0·27*0·27 0·19*

site-specific comparisons

The greatest difference between the nature reserve and unconserved land in terms of arthropod and reptile RTU diversity and abundance, as well as HRI values, occurred between NR and CGA paired site III (Tables 3 and 4). The nature reserve and the historically overgrazed commercial farm (CF1) were most similar (Table 3). The large standard errors were expected because of the small number of replicates (four) at each trap locality. The most noteworthy site-specific comparisons were as follows. The NR had a greater richness (in terms of HRI) of arthropods than the CGA in all taxonomic groups (Table 3). Standard errors of HRI values did not overlap for any taxonomic group at this paired trap site. In contrast to the pattern for arthropods, the CGA showed reptile species diversity similar to the NR, but more than twice as many individuals and almost double the HRI. Reptile HRI values at this paired site differed significantly from the expected values (P < 0·01, χ2 = 14·7). Weevil richness on the conservatively grazed commercial farm (CF2) was less than half that of NR2, while crickets showed a comparable decrease in richness on the heavily grazed commercial farm (CF1) compared with NR1.

The most dissimilar arthropod communities (in terms of their Sorenson's indexes, CC) were between NR2 and CF2 (paired site II) and between NR3 and the CGA (paired site III). The communities of crickets, grasshoppers, large arachnids and weevils differed substantially between NR2 and CF2 and there were different communities of ants, grasshoppers, spiders and weevils between NR3 and CGA. In the comparison between NR1 and CF1 (paired site I) only the spider communities differed.

The taxonomic groups that were most sensitive to land use in terms of their community composition were grasshoppers, spiders and weevils. The taxonomic groups that characterized the nature reserve (in terms of uniqueness) were the spiders and weevils (in all comparisons) and the ants (with the exception of paired site I). U-values for spiders at all localities were significantly greater than zero, indicating that they are particularly sensitive to vegetation change. The large arachnids and crickets were the least locality-unique, as none of their U-values was significantly greater than zero (Table 6).

One species of reptile was collected in the CGA only, one at CF2 only and three at NR2 only (Table 4). The uniqueness of the lizard Gerrhosoarus typicus (A. Smith 1836) to the CGA is of particular significance, as this species has not previously been recorded away from the arid Karoo biome. It is classified as ‘rare’ in the South African Red Data Book (Branch 1988).

indicator groups

When HRI, CC and U are considered simultaneously, the weevils, spiders and ants were good indicators of habitat change (column labelled ‘sensitivity score’ in Table 7). These groups contained many RTU and consequently showed a large degree of niche differentiation (hence their consistently high U-values). The weevils, which are obligate herbivores, contained many RTU and therefore displayed a high degree of niche differentiation, and were exceptionally good indicators. The large arachnids and crickets, on the other hand, were poor indicators of habitat change; they contained few RTU and are not obligate herbivores. When individual methods were considered separately, uniqueness indexes were the most effective for taxonomic groups with many RTU, i.e. the ants, weevils and spiders. HRI values were most sensitive to groups that occurred at high densities (weevils, large arachnids and crickets).

Table 7.  The frequency with which an arthropod taxonomic group showed a ‘substantial’ difference (see the Methods) between two paired trap localities, according to three indexes: hierarchical richness index (HRI), Sorenson's index of community (CC), and uniqueness (U). The sensitivity score is the sum of the frequencies
 HRICCUSensitivity score
Ants113 5
Crickets210 3
Grasshoppers130 4
Large arachnids310 4
Spiders124 7


  1. Top of page
  2. Summary
  3. Introduction
  4. Study area
  5. Methods
  6. Results
  7. Discussion
  8. Acknowledgements
  9. References

the influence of land use on biodiversity


Both the most heavily grazed (CGA) and least grazed (CF2) localities contained fewer RTU and lower HRI values than corresponding sites in the nature reserve. The Intermediate disturbance hypothesis (Connell 1975; Rozenzweig & Abramsky 1993), which predicts a parabolic relationship between diversity and disturbance, is a parsimonious explanation of this pattern, wherein CF2, CF1 and CGA, respectively, represent points of low, intermediate and high disturbance. The nature reserve represented an intermediately disturbed reference point against which the impact of disturbance could be measured, as a result of its relatively low herbivore density and the variety of disturbance agents it contained in the form of herbivores of different body sizes.

Each of the land uses made a unique contribution to the gamma diversity of the study area. In a pair-for-pair comparison, all three paired trap sites contained different communities (CC values of 0·75 and less) for at least one taxonomic group. Conversely, each trap locality contained some unique RTU, reflected by the large U-values of several taxonomic groups at every locality. For example, 25% of grasshopper RTU did not occur in the nature reserve, while 23% of all ant RTU, 27% of all weevil RTU (including the six undescribed species) and 21% of all spider RTU occurred only in the nature reserve. Of the total number of RTU collected, 89·7% occurred in the nature reserve while 79·9% occurred outside the protected area. No single unconserved locality contained more than 62% of the total number of RTU and, in the most extreme case, the communal grazing area contained only 37% of the total number of arthropod RTU collected in the study area.

Nature conservation was the land-use type making the greatest contribution to arthropod conservation in the study area. It is also the least common form of land use in the Eastern Cape (La Cock, Palmer & Everard 1990), emphasizing its future value in biodiversity conservation. It is clear that the gamma diversity of arthropods in xeric succulent thicket will decrease substantially if other land uses replace nature conservation. The corollary, on the other hand, is that the total species complement for the study area as a whole is harboured by the variety of grazing regimes.


Land use in xeric succulent thicket has less influence on terrestrial reptiles than on arthropods, except in extreme cases such as the CGA, where degradation favours xeric-adapted species. The association of the degraded CGA with xeric-adapted reptiles is ascribed to a decrease in ground cover, resembling desertification. These species are likely to have physiological and behavioural qualities suited to arid or desert conditions (Greenberg, Neary & Harris 1994), such as large clutch sizes ability to cope with harsh climatic conditions and agility to avoid predators and catch prey in open terrain. Degraded areas are widespread in xeric succulent thicket and are probably increasing (Kerley, Knight & De Kock 1995); reptiles that are associated with these areas do not require special management. It is much more important to maintain areas with high vegetation cover for conserving those naturally occurring reptile species that do not persist in degraded, xeric conditions.

Possible causes of the higher reptile abundance in the CGA might be greater prey availability, lack of predation (particularly by birds), a positive relationship between distribution range size and local abundance (Hanksi, Kouki & Halkka 1993), improved habitat quality and uncertain production because of variable rainfall, which favours lizards over homeotherms (Morton & James 1988). Greater prey availability can be ruled out as a cause, as the CGA was characterized by invertebrate-eating reptiles yet had almost 45% fewer terrestrial arthropods, both in terms of species and abundance, than its corresponding site in the NR. Reptile abundance in the CGA may be linked directly to habitat changes. The open and barren terrain of the CGA provides suitable hunting conditions for reptiles, with enough sparse vegetation for hiding and darting amongst shrubs (cf. Mushinsky 1992). In addition to improved hunting conditions, the lack of cover at the CGA provides improved opportunities for thermoregulation of reptiles, especially in winter (Mushinsky 1992). There is broad agreement that lizards are better adapted to arid conditions and unproductive environments than other vertebrates (Pianka 1986; Morton & James 1988; Morton 1993) and that food availability, temperature and competition are determinants of their distribution and diversity.

Lizard communities elsewhere have benefited from land alteration, such as clear-cutting of forests (Greenberg, Neary & Harris 1994; but see Germano & Hungerford 1981). Sand lizards in the UK were driven to extinction in habitats where management maintained dense undergrowth, probably because of egg-shading (Corbett & Tamarind 1979). The generalization that patch diversity promotes faunal diversity (Shmida & Wilson 1985) does not apply to terrestrial Squamata in xeric succulent thicket. The CGA had a significantly lower patch diversity than the NR and commercial farms (Fabricius, Palmer & Burger 2002).

land use and regional biodiversity (gamma diversity)

Indicator groups

Taxonomic groups shown to be sensitive to land use and disturbance in xeric succulent thicket had one or more common characteristics. (i) They contained many RTU; taxonomic groups with many RTU are more sensitive to habitat change, being more stenotopic due to niche differentiation (Samways 1994). (ii) They are directly dependent on perennial vegetation during their life cycles. (iii) They are incapable of aerial migration, and hence are poor dispersers (cf. Duelli et al. 1992). Other considerations that determine whether a group would be a useful indicator are yield (the abundance of individuals representing the respective lower order taxa) and effort, i.e. the cost of collecting and identifying specimens (Duelli, Obrist & Schmatz 1999). In terms of satisfying all three criteria for indicator groups, as well as being abundant and easy to identify, the weevils appear to be appropriate indicators of biodiversity change in response to land use in xeric succulent thicket.

Other authors have identified different taxonomic groups as useful indicators of biodiversity. Duelli et al. (1992) found that spiders showed the greatest variability between sites and were the poorest dispersers. Duelli, Obrist & Schmatz (1999) concluded that, on the basis of yield and effort, the order Heteroptera and certain Hymenoptera groups should be useful indicators in agricultural landscapes. In Karoo dwarf shrubland, Dean & Milton (1995) did not find weevil abundance to be significantly different between three fields of varying vegetation but found ants to be good indicators of succession and disturbance. Rivers-Moore & Samways (1996) found that weevils were eurytopic and unaffected by land use, whereas other Coleoptera were stenotopic and were affected (mainly by differences in the amount of animal dung between sites). Unlike our study area, the physical structure of the vegetation in the different sites compared by Dean & Milton (1995) was very similar, as was that of the sites compared by Rivers-Moore & Samways (1996). The value of using indicator groups is, however, not universally accepted: Ehrlich (1994), for example, questions whether enough is known about the correlation in diversity between different groups to confidently use indicator groups as surrogates for full-scale diversity assessment of insects in temperate forests. The suitability of a taxonomic group as an indicator of vegetation change varies between vegetation types; vegetation structure might be an important determinant of this (McGeogh, van Rensburg & Botes 2002).

evaluation of methods

The sampling problems associated with pitfall traps are well documented, and it is widely accepted that data collected in this way do not always reflect the real structure of invertebrate communities (Marsh 1984; Topping & Sunderland 1992). For example, the true abundance of arthropod taxa can be masked by differences in activity patterns and differing abilities to escape from traps (Topping & Sunderland 1992). However, the purpose of this study was to compare arthropod biodiversity and not to describe the composition of arthropod communities. Because the same method was used at all localities, errors resulting from different trapping success for different types of arthropods would have been constant. Volant arthropods, which might be good indicators at the biotope level, especially in aquatic systems (Clark & Samways 1996), were, however, not sampled. Because it was impossible to replicate the different land-use treatments, limited inferences about the empirical probability of species loss under different types of land use can be made. The study would have benefited from the replication of land-use treatments and more, smaller, pitfall traps.

The adoption of an RBA approach was appropriate because the study objective was to obtain relative data to compare sites, and only about 5–10% of invertebrates have scientific names (Samways 1994). The disadvantage of RBA was that the response of different taxa within a taxonomic group to different levels of grazing intensity could not be determined. Our use of RBA was validated by the specialists’ classification of arachnids, and, to lesser extent, of weevils. We recognized 86 arachnid RTU and the arachnologist identified 100 species. With weevils we were only 65% successful, identifying 26 RTU compared with the specialist's 40 species. Another weakness was that data were collected for two seasons only (spring and summer). Prolonged and more intensive trapping might also have resulted in fewer unique RTU. Factors affecting the temporal density of organisms, such as breeding season and hibernation, could result in different patterns emerging in different seasons (Coombes & Sotherton 1986; Desender & Alderweireldt 1988; Kromp & Steinberger 1992). Factors other than land use could have affected the differences between localities. These include climatic differences between sites, microtopographic differences and temporary changes in species composition.


Unconserved xeric succulent thicket plays a role in promoting gamma diversity among arthropods and reptiles, but protected areas are vital to the conservation of species that decrease under intensive herbivory and consumptive use by people (‘decreaser species’). The results reported here, while strongly supporting the notion that commercial and subsistence agriculture has a negative impact on biodiversity in xeric succulent thicket, should not be extrapolated to less fragile vegetation types.


  1. Top of page
  2. Summary
  3. Introduction
  4. Study area
  5. Methods
  6. Results
  7. Discussion
  8. Acknowledgements
  9. References

André Boshoff and Graham Kerley provided useful comments and support, while Charlie Shackleton and Cindy Kulongowski commented on the manuscript. Tony Philips and Ash Davenport permitted us to work on their properties, while the Ndwayana community leaders gave us access to the communal area. L. Antony, C. Dubase, L. Gqamane, S. Hoyi, S. Loli and C. Tyatya assisted with collecting specimens. Ansie Dippenaar and Rolf Oberprieler identified arachnid and weevils specimens, respectively. Martin Villet assisted with the identification of RTU. Brad Fike of Sam Knott NR is thanked for his help.


  1. Top of page
  2. Summary
  3. Introduction
  4. Study area
  5. Methods
  6. Results
  7. Discussion
  8. Acknowledgements
  9. References
  • Acocks, J.P.H. (1988) Veld Types of Southern Africa. Memoirs of the Botanical Survey of South Africa No. 59. Department of Agriculture Technical Services, Pretoria, South Africa.
  • Ainslie, A., Fox, R. & Fabricius, C. (1994) Towards Policies for Feasible and Sustainable Natural Resource Use: The Mid-Fish River Zonal Study, Eastern Cape. Final Report to the LAPC Natural Resource Management Programme. Land and Agriculture Policy Centre, Johannesburg, South Africa.
  • Beattie, A.J., Majer, J.D. & Oliver, I. (1993) Rapid biodiversity assessment: a review. Rapid Biodiversity Assessment. Proceedings of the Biodiversity Assessment Workshop, 3–4 May 1993 (eds A.J.Beattie & J.D.Majer), pp. 414. Macquarie University, Sydney, Australia.
  • Branch, W.R. (1988) Field Guide to the Snakes and other Reptiles of Southern Africa. Struik, Johannesburg, South Africa.
  • Clark, T.E. & Samways, M.J. (1996) Dragonflies (Odonata) as indicators of biotope quality in the Kruger National Park, South Africa. Journal of Applied Ecology, 33, 10011012.
  • Coetzee, C.G. (1994) Forts of the Eastern Cape – Securing a Frontier, 1799–1878. C.G. Coetzee, Grahamstown, South Africa.
  • Connell, J.H. (1975) Some mechanisms producing structure in natural communities: a model and evidence from field experiments. Ecology and Evolution of Communities (eds M.L.Cody & J.Diamond), pp. 460490. Harvard University Press, Boston, MA.
  • Coombes, D.S. & Sotherton, N.W. (1986) The dispersal and distribution of polyphagous predatory Coleoptera in cereals. Annals of Applied Biology, 108, 461474.
  • Corbett, K.F. & Tamarind, D.L. (1979) Conservation of the sand lizard, Lacerta agilis, by habitat management. British Journal of Herpetology, 5, 799823.
  • Cowling, R.M., Gibbs Russel, E., Hoffman, M.T. & Hilton-Taylor, C. (1989) Patterns of plant species diversity in southern Africa. Biotic Diversity in Southern Africa (ed. B.J.Huntley), pp. 1950. Oxford University Press, Cape Town, South Africa.
  • Dean, W.R.J. & Milton, S.J. (1995) Plant and invertebrate assemblages on old fields in the arid southern Karoo, South Africa. African Journal of Ecology, 33, 113.
  • Desender, K. & Alderweireldt, M. (1988) Population dynamics of adult and larval Carabid beetles in a maize field and its boundary. Journal of Applied Entomology, 106, 1319.
  • Duelli, P., Obrist, M.K. & Schmatz, D.R. (1999) Biodiversity evaluation in agricultural landscapes: above-ground insects. Agriculture, Ecosystems and Environment, 74, 3364.
  • Duelli, P., Studer, M., Marchand, I. & Jacob, S. (1992) Population movements of arthropods between natural and cultivated areas. Biological Conservation, 54, 193207.
  • Ehrlich, P.R. (1994) Foreword: Biodiversity and ecosystem function: need we know more? Biodiversity and Ecosystem Function (eds E.-D.Schulze & H.A.Mooney), pp. ixxi. Springer Verlag, Berlin.
  • Fabricius, C., Palmer, A.R. & Burger, M. (2002) Landscape diversity in a conservation area and commercial and communal rangeland in xeric succulent thicket, South Africa. Landscape Ecology, 17, 531537.
  • French, D.D. (1994) Hierarchical richness index (HRI): a simple procedure for scoring ‘richness’, for use with grouped data. Biological Conservation, 69, 207212.
  • Gandar, M.V. (1982) Trophic ecology and plant–herbivore energetics. Ecology of Tropical Savannas (eds B.J.Huntley & B.H.Walker), pp. 514534. Springer-Verlag, Berlin, Germany.
  • Germano, D.J. & Hungerford, C.R. (1981) Reptile population changes with manipulation of Sonoran desert shrub. Great Basin Naturalist, 41, 129138.
  • Greenberg, C.H., Neary, D.G. & Harris, L.D. (1994) Effect of high-intensity wildfire and silvicultural treatments on reptile communities in sand-pine scrub. Conservation Biology, 8, 10471057.
  • Greenslade, P. (1992) Conserving invertebrate diversity in agricultural, forestry and natural ecosystems in Australia. Agriculture Ecosystems and Environment, 40, 297312.
  • Hanksi, I., Kouki, J. & Halkka, A. (1993) Three explanations for the positive relationship between distribution and abundance. Species Diversity in Ecological Communities (eds R.E.Ricklefs & D.Schluter), pp. 108116. University of Chicago Press, Chicago, IL.
  • Janzen, D.H. (1987) Insect diversity of a Costa Rican dry forest: why keep it and how? Biological Journal of the Linnaean Society, 30, 343356.
  • Kerley, G.I.H., Boshoff, A.F. & Knight, M.H. (1999) Ecosystem integrity and sustainable land-use in the thicket biome, South Africa. Ecosystem Health, 5, 104109.
  • Kerley, G.I.H., Knight, M.H. & De Kock, M. (1995) Desertification of subtropical thicket in the Eastern Cape, South Africa: are there alternatives? Environmental Monitoring and Assessment, 37, 211230.
  • Kromp, B. & Steinberger, K.-H. (1992) Grassy field margins and arthropod diversity: a case study on ground beetles and spiders in eastern Austria (Coleoptera: Carabidae; Arachnida: Aranei, Opiliones). Agriculture Ecosystems and Environment, 40, 7193.
  • La Cock, G.D. (1992) The conservation status of subtropical transitional thicket, and regeneration through seeding of shrubs in the xeric succulent thicket of the Eastern Cape. MSc Thesis. Rhodes University, Grahamstown, South Africa.
  • La Cock, G.D., Palmer, A.R. & Everard, D.A. (1990) Re-assessment of the area and conservation status of subtropical transitional thicket (valley bushveld) in the Eastern Cape, South Africa. South African Journal of Photogrammetry, Remote Sensing and Cartography, 15, 231235.
  • Low, A.B. & Rebelo, A.G. (1996) Vegetation of South Africa, Lesotho and Swaziland. Department of Environment Affairs and Tourism, Pretoria, South Africa.
  • McGeogh, M.A., Van Rensburg, B.J. & Botes, A. (2002) The verification and application of bioindicators: a case study of dung beetles in a savanna ecosystem. Journal of Applied Ecology, 39, 661672.
  • Marsh, A.C. (1984) The efficacy of pitfall traps for determining the structure of a desert ant community. Journal of the Entomological Society of South Africa, 4, 115120.
  • Morton, S.R. (1993) Determinants of diversity in animal communities of arid Australia. Species Diversity in Ecological Communities (eds R.E.Ricklefs & D.Schluter), pp. 159169. University of Chicago Press, Chicago, IL.
  • Morton, S.R. & James, C.D. (1988) The diversity and abundance of lizards in arid Australia: a new hypothesis. American Naturalist, 132, 237256.
  • Mushinsky, H.R. (1992) Natural history and abundance of southeastern five-lined skinks, Eumeces inexpectatus, on a periodically burned sandhill in Florida. Herpetologica, 48, 307312.
  • Neu, C.W., Byers, C.R. & Peek, J.M. (1974) A technique for analysis of utilization-availability data. Journal of Wildlife Management, 38, 541545.
  • Palmer, A.R. & Avis, A. (1994) The Description, Mapping and Evaluation of Recent Changes in the Contemporary Vegetation Patterns. Report to the LAPC. Roodeplaat Grassland Institute, Grahamstown, South Africa.
  • Pianka, E.R. (1986) Ecology and Natural History of Desert Lizards. Princeton University Press, Princeton, NJ.
  • Rivers-Moore, J.L. & Samways, M.J. (1996) Game and cattle trampling, and impacts of human dwellings on arthropods at a game park. Biodiversity and Conservation, 5, 15451557.
  • Rozenzweig, M.L. & Abramsky, Z. (1993) How are diversity and productivity related? Species Diversity in Ecological Communities (eds R.E.Ricklefs & D.Schluter), pp. 5265. University of Chicago Press, Chicago, IL.
  • Samways, M.J. (1994) Insect Conservation Biology. Chapman & Hall, London, UK.
  • Seastedt, T.R. & Crossley, D.A. (1984) The influence of arthropods on ecosystems. Bioscience, 34, 157161.
  • Shmida, A. & Wilson, M.V. (1985) Biological determinants of species diversity. Journal of Biogeography, 12, 120.
  • Sorenson, T. (1948) A method of establishing groups of equal amplitude in plant sociology based on similarity in species content. Biologiske Skrifter Kongelige Danske Videnskabernes Selskab, 5, 134.
  • Topping, C.J. & Sunderland, K.D. (1992) Limitations to the use of pitfall traps in ecological studies exemplified by a study of spiders in a field of winter wheat. Journal of Applied Ecology, 29, 485491.
  • Van Wyk, A.E. & Smith, G. (2001) Regions of Floristic Endemism in Southern Africa. Umdaus Press, Pretoria, South Africa.
  • Whittaker, R.H. (1972) Evolution and measurement of species diversity. Taxon, 21, 213251.
  • Wiens, J.A. & Milne, B.T. (1989) Scaling of ‘landscapes’ in landscape ecology, or, landscape ecology from a beetle's perspective. Landscape Ecology, 3, 8796.