The exponential growth of the human population is threatening all levels of biodiversity, including large carnivores, with 77% of species’ populations experiencing continuing declines (Estes et al., Reference Estes, Terborgh, Brashares, Power, Berger and Bond2011; Ripple et al., Reference Ripple, Estes, Beschta, Wilmers, Ritchie and Hebblewhite2014). Large carnivores are particularly at risk of extinction because of their small population sizes, slow reproductive rates, complex social structures and requirement for large and contiguous habitats with sufficient prey (Cardillo et al., Reference Cardillo, Purvis, Sechrest, Gittleman, Bielby and Mace2004; Ripple et al., Reference Ripple, Estes, Beschta, Wilmers, Ritchie and Hebblewhite2014). These characteristics, and their vulnerability to negative interactions with humans, have driven declines of some of the most wide-ranging carnivores (Cardillo et al., Reference Cardillo, Purvis, Sechrest, Gittleman, Bielby and Mace2004; Ray et al., Reference Ray, Hunter and Zigouris2005; Swanepoel et al., Reference Swanepoel, Lindsey, Somers, Hoven and Dalerum2013; Wolf & Ripple, Reference Wolf and Ripple2017).
Although the leopard is considered the most adaptable large carnivore (Ripple et al., Reference Ripple, Estes, Beschta, Wilmers, Ritchie and Hebblewhite2014), range declines of 63–75% globally, 48–67% in Africa and 28–51% in southern Africa indicate that leopard populations are not as resilient to anthropogenic influences as previously believed (Jacobson et al., Reference Jacobson, Gerngross, Lemeris, Schoonover, Anco and Breitenmoser-Würsten2016). Major anthropogenic threats to leopards include ongoing habitat loss and fragmentation (Harcourt et al., Reference Harcourt, Parks and Woodroffe2001; Crooks, Reference Crooks2002; Swanepoel et al., Reference Swanepoel, Somers and Dalerum2015), depletion of prey resources (Woodroffe, Reference Woodroffe2000; Karanth & Chellam, Reference Karanth and Chellam2009), unsustainable hunting levels (Woodroffe, Reference Woodroffe2000; Harcourt et al., Reference Harcourt, Parks and Woodroffe2001) and direct persecution by people (Harcourt et al., Reference Harcourt, Parks and Woodroffe2001; Treves & Karanth, Reference Treves and Karanth2003; Treves et al., Reference Treves, Naughton-Treves, Harper, Mladenoff, Rose, Sickley and Wydeven2004; Treves & Naughton-Treves, Reference Treves and Naughton-Treves2005; McManus et al., Reference McManus, Dickman, Gaynor, Smuts and Macdonald2014; Swanepoel et al., Reference Swanepoel, Somers and Dalerum2015). Increasing human population density and loss of habitat increase the likelihood of resource competition with people, often resulting in human–carnivore conflict (Woodroffe, Reference Woodroffe2000; Cardillo et al., Reference Cardillo, Purvis, Sechrest, Gittleman, Bielby and Mace2004). Persecution often includes indiscriminate use of lethal methods to manage livestock depredation by carnivores (e.g. snares, poisoned carcasses, gin traps (leg-hold traps), gun traps, live trapping) and targeted (often retaliatory) hunting (McManus et al., Reference McManus, Dickman, Gaynor, Smuts and Macdonald2014).
Estimating a species’ density across various types of land cover (i.e. habitat and vegetation type) and land uses (i.e. residential/urban land, cultivated land, commercial and recreational land and areas protected for conservation) can help determine how population numbers are affected by landscape features. Density estimates are not equally robust, and under- or overestimating populations can have substantial implications for conservation management and policy (Foster & Harmsen, Reference Foster and Harmsen2012; Hayward et al., Reference Hayward, Boitani, Burrows, Funston, Karanth and MacKenzie2015). For example, Soisalo & Cavalcanti (Reference Soisalo and Cavalcanti2006) demonstrated how overestimating jaguar Panthera onca populations by five individuals per 100 km2 inflated the overall population estimate by 7,000 individuals across their 140,000 km2 study region. Density estimation also depends upon data quality, with a large enough sample size and capture probability for the chosen analysis method (Foster & Harmsen, Reference Foster and Harmsen2012).
Conservationists and ecologists constantly seek to improve the methodology for estimating animal population abundances and densities (Griffiths & van Schaik, Reference Griffiths and van Schaik1993; Karanth & Nichols, Reference Karanth and Nichols1998; Karanth & Nichols, Reference Karanth and Nichols2000). The use of camera traps and photo capture–recapture analysis is a common and effective non-invasive practice for obtaining data on wildlife population dynamics, particularly of rare or elusive species (Foster & Harmsen, Reference Foster and Harmsen2012). Leopards, like many other large felids, are challenging to monitor because of their large home ranges, low population density and primarily solitary and elusive nature (Karanth & Nichols, Reference Karanth and Nichols2000; Treves & Karanth, Reference Treves and Karanth2003). Global population estimates across the leopard's vast geographical distribution do not account for the health and sustainability of smaller, isolated regional metapopulations across a variety of habitat types, countries and levels of human landscape modification and pressure.
Camera-trap surveys have become an important data collection method for population and density studies because they are non-invasive, practical and affordable (Karanth, Reference Karanth1995; Karanth & Nichols, Reference Karanth and Nichols1998; Foster & Harmsen, Reference Foster and Harmsen2012). The capture–recapture method relies on individuals being identifiable (Karanth, Reference Karanth1995; Silver et al., Reference Silver, Ostro, Marsh, Maffei, Noss and Kelly2004) and has been the predominant approach used in felid density studies.
More recently, spatial capture–recapture or spatially explicit capture–recapture methods have become popular because they provide comprehensive analyses that can incorporate environmental and anthropogenic factors when estimating animal densities. These methods enable density analyses to include spatio-temporal data from capture histories, providing direct estimates of population density that remain unbiased by edge effects (Chase Grey et al., Reference Chase Grey, Kent and Hill2013). By allowing for flexibility in individual heterogeneity with the consideration of capture probability relative to trap location, spatial covariates such as habitat, and intrinsic factors such as sex and age (Foster & Harmsen, Reference Foster and Harmsen2012; Efford & Fewster, Reference Efford and Fewster2013), spatially explicit capture–recapture models are becoming increasingly robust and comprehensive.
These statistical and methodological advances provide multiple options that can be applied to specific study questions while making population density estimates increasingly reliable. Inaccurate density estimates can lead to biased population estimates, with serious implications for conservation management (Soisalo & Cavalcanti, Reference Soisalo and Cavalcanti2006).
The estimation of a baseline population density for leopards in South Africa's Western Cape province is fundamental for understanding how this regionally important population responds to landscape-scale threats and changes. By estimating densities across landscapes of varying resource availability and threats, we can examine drivers of regional density variation and improve conservation management for disjunct populations.
This study covers c. 30,000 km2 in three areas (Langeberg, Overberg and Garden Route) in the Western Cape, South Africa (Fig. 1). The topography varies from the Cape Fold Mountain peaks with an altitude of > 1,600 m that extend 1,000 km east to west, to coastal and low-lying valleys at < 500 m altitude (Thamm & Johnson, Reference Thamm, Johnson, Johnson, Anhaeusser and Thomas2006). Biomes included in our study are Thicket, Afro-temperate forest (Forest), Sandstone fynbos (Fynbos), Nama-Karoo, Succulent-Karoo and Savanna (Mucina & Rutherford, Reference Mucina and Rutherford2006).
In the Langeberg we deployed camera traps in the greater Riversdale/Heidelberg area on the southern slopes of Langeberg Mountain Range, in the greater Greyton area on the southern slope of Riviersonderend Mountain Range and in the Robertson Wine Valley (Breede River Valley) situated between the Langeberg and Riviersonderend Mountain Ranges. In the Overberg we surveyed the greater Hermanus area to the west, the greater Cape Agulhas and Arniston areas to the south along the coast, and the greater De Hoop Nature Reserve area to the eastern extent of the surveyed region. Along the Garden Route we surveyed temperate forest along the southern slopes of the Outeniqua and Tsitsikamma Mountains from George in the west to the Bloukrans River in Plettenberg Bay in the east.
We undertook seven large-scale camera-trap surveys during June 2011–March 2015, focusing on likely presence of leopards inside and outside protected areas and across different agricultural land-use zones (livestock, crops, forestry). We used Cuddeback Attack IR (Cuddeback, Green Bay, USA) digital infrared cameras and selected camera-trap locations based on the highest likelihood of leopard activity as determined by physical evidence (scat, spoor and territorial scent and scratch markings on trees), as well as habitat type and topography (Karanth & Nichols, Reference Karanth and Nichols2000). At each camera-trap station, we positioned two cameras to capture both flanks of a passing leopard. The stations were placed 3–5 km apart to achieve even coverage of the sampling area and camera stations remained active for a minimum of 3 months to ensure maximum likelihood of leopard activity being captured without violating the closed population assumption (Karanth & Nichols, Reference Karanth and Nichols1998; Devens et al., Reference Devens, Tshabalala, McManus and Smuts2018).
Individual leopards can be identified from clear photographs of both flanks by their unique rosette markings. Identification from single flank photos is possible, but needs to be based upon the observation of characteristics such as the animal's overall size, neck girth, injuries or scars. Although such characteristics may be useful, incorrectly identified individuals could bias density estimates, resulting in overestimation of the population if flank photos of the same animal were erroneously assigned to two individuals. We therefore only utilized double flank capture events for identification.
Spatially explicit capture–recapture methods for density analyses are more comprehensive and reliable than traditional capture–recapture analyses (Borchers & Efford, Reference Borchers and Efford2008; Kalle et al., Reference Kalle, Ramesh, Qureshi and Sankar2011; Gopalaswamy et al., Reference Gopalaswamy, Royle, Hines, Singh, Jathanna, Kumar and Karanth2012; Chase Grey et al., Reference Chase Grey, Kent and Hill2013; Thapa et al., Reference Thapa, Shrestha, Karki, Thapa, Subedi and Pradhan2014) and produce lower density estimates than non-spatial methods (Obbard et al., Reference Obbard, Howe and Kyle2010; Gerber et al., Reference Gerber, Karpanty and Kelly2012; Gopalaswamy et al., Reference Gopalaswamy, Royle, Hines, Singh, Jathanna, Kumar and Karanth2012; Noss et al., Reference Noss, Gardener, Maffei, Cuéllar and Montaño2012; Braczkowski et al., Reference Braczkowski, Balme, Dickman, Fattebert, Johnson and Dickerson2016). We calculated density estimates using two spatially explicit capture–recapture methods within the programming environment R 3.4.0 (R Development Core Team, 2017): (1) the maximum likelihood based estimator programme secr 2.10.3 (Efford et al., Reference Efford, Borchers, Byron, Thomson, Cooch and Conroy2009; Efford, Reference Efford2011, Reference Efford2016), which is a more robust version of programme DENSITY (Efford, Reference Efford2011), and (2) the Bayesian estimator programme SPACECAP 1.1.0 (Gopalaswamy et al., Reference Gopalaswamy, Royle, Hines, Singh, Jathanna, Kumar and Karanth2012). Previous studies have applied these two programmes to compare density estimates (Kalle et al., Reference Kalle, Ramesh, Qureshi and Sankar2011; Thapa et al., Reference Thapa, Shrestha, Karki, Thapa, Subedi and Pradhan2014) and to compare results of spatially explicit capture–recapture with non-spatial methods (Noss et al., Reference Noss, Gardener, Maffei, Cuéllar and Montaño2012; Braczkowski et al., Reference Braczkowski, Balme, Dickman, Fattebert, Johnson and Dickerson2016). We used these programmes for comparison of two spatially explicit capture–recapture methods and to determine the most robust and inclusive density estimate range for each study region.
For the analysis with secr we categorized camera traps as count detectors, which allows repeat detections. Count data can result from devices such as automatic camera traps. Count detectors record the presence of an animal at a trap location without restricting movement and allow > 1 detection of an individual at a particular site on any occasion. The secr.fit function was run for each regional survey phase using models of time (g0~T) and behavioural response (g0~b), as well as buffers of 15, 20, 25, 30 and 35 km surrounding the minimal convex polygon around camera stations in each camera-trapping area. These buffers ensured inclusion of all leopard home ranges within reach of camera traps (Kalle et al., Reference Kalle, Ramesh, Qureshi and Sankar2011) and made density estimation more reliable by determining the point at which the density estimate stabilized (Kalle et al., Reference Kalle, Ramesh, Qureshi and Sankar2011; Chase Grey et al., Reference Chase Grey, Kent and Hill2013).
SPACECAP was used to estimate abundance and density using spatially-explicit capture–recapture models to derive spatial Bayesian estimates with trap response. A grid of equally spaced points 1 km apart was generated in QGIS 2.8 (QGIS Development Team, 2013) and clipped to the surveyed area containing the camera trap array combined with an extended surrounding area, known as the state-space, at distances of 15, 20, 25, 30 and 35 km. These points represent all potential home range centres of all leopards within the survey. The potential home range centres file included the geographical coordinates and habitat suitability of each of these points within the state-space. We determined habitat suitability using Maxent 3.3.3 (Phillips et al., Reference Phillips, Anderson and Schapire2006; Phillips et al., Reference Phillips, Dudik and Schapire2017) model output comprising environmental variables, including eight WorldClim bioclimatic variables obtained from WorldClim website (Hijmans et al., Reference Hijmans, Cameron, Parra, Jones and Jarvis2005), human footprint index (WCS & CIESIN, 2005), altitude, anthropogenic biomes (Ellis & Ramankutty, Reference Ellis and Ramankutty2008), Globcover 2009 (FAO, 2012), South African National Biodiversity Institute ecosystem status of vegetation types (Rouget et al., Reference Rouget, Reyers, Jonas, Desmet, Driver and Maze2004), and a subset of location data obtained from two leopards resident in the area that were equipped with GPS collars. We used the Natural Breaks function (Jenks, Reference Jenks1967) in ArcMap 10.3.1 to code the model output as either 0 (not suitable = 0–0.189171) or 1 (suitable = 0.189172–0.775). The habitat suitability generated in this way (Jenks method) was recorded for each point and a home range centre input file was generated for each area's buffer distances (Fig. 2).
We ran SPACECAP with a 1 km2 pixel area, and set the model definitions to ‘Trap response present’, ‘Spatial Capture–Recapture’, ‘half normal’ detection function and the capture encounters ‘Bernoulli's process’. The ‘Trap response present’ option implements a ‘trap-specific’ behavioural response in which the probability of capture at a specific trap increases (or decreases) after the initial capture. The ‘Spatial Capture–Recapture’ option runs a spatially explicit capture–recapture analysis. The Markov-Chain Monte Carlo settings varied between study areas and buffers. Markov-Chain Monte Carlo iterations were set between 60,000 and 100,000 with a burn-in period between 10,000 and 65,000 iterations and a thinning rate of 1. The data augmentation numbers were c. 20–40 times the number of animals identified in each regional survey and varied between 170 and 1,500. We assessed chain convergence with the Geweke diagnostic test produced within the SPACECAP output in the form of z-score values. Z-scores between −1.6 and +1.6 implied adequate convergence and confirmed that the Markov-Chain Monte Carlo analysis was run with a sufficiently long burn-in period. The SPACECAP output also produces a Bayesian P-value to provide additional assessment of the model fit where P-values close to 0 or 1 imply that the model is inadequate. The adequacy of the data augmentation number can be checked in the ‘density plot for psi’ and ‘density plot for N’ files included within the output files. All densities obtained with SPACECAP produced z-scores that achieved convergence, as well as sufficient model fit and data augmentation number. The output also included pixel-specific density estimates (animals per km2), which we used to create a pixel density map of the study areas and compare leopard densities across various land covers and uses to illustrate relationship with anthropogenic factors.
Camera-trap data and sampling effort
The total sampling effort for the Overberg (November 2011–March 2012) was 2,880 camera-trap nights across a total of 32 camera-trap locations (64 cameras). This yielded 118 leopard photographs (captures), with eight individuals identified in 40 of them. These individuals were detected at nine of the 32 camera-trap sites (Table 1).
The total sampling effort for the two phases of the Langeberg survey (April 2012–December 2012) consisted of 9,480 camera-trap nights across a total of 79 camera-trap locations (158 cameras). This yielded 454 leopard captures, from which 42 individual leopards were identified in 250 photographs, and these individuals were detected at 58 of the 79 camera-trap sites (Table 1).
The total sampling effort for the three phases of the Garden Route survey area (May 2013–December 2014) was 3,600 camera-trap nights across a total of 40 camera-trap locations (80 cameras). This yielded 142 leopard captures, of which 21 individual leopards were identified in 86 photographs, and these individuals were detected at 33 of the 40 camera-trap sites (Table 1). The total sampling effort was 13,590 camera-trap nights across 151 camera-trap locations (302 cameras), with 714 leopard captures and 71 identified individuals (Table 1).
The spatially explicit capture–recapture density estimates varied by region and between regional phases, and estimates from secr were lower than those from SPACECAP. The effect of using various buffers within the two programmes, particularly SPACECAP, demonstrated the sensitivity to buffer width and data augmentation size (Kalle et al., Reference Kalle, Ramesh, Qureshi and Sankar2011). We calculated mean density estimates from stabilized buffer values to ensure the study area was large enough to avoid capturing any individuals residing outside the buffered region during the survey.
In the Overberg, secr density estimates stabilized at 0.17 leopards/100 km2 (CI 0.06–0.48). Both phases in the Langeberg had density estimates that were stable across all five buffers with a phase one density estimate of 0.29 leopards/100 km2 (CI 0.2–0.41) and a phase two estimate of 0.7 leopards/100 km2 (CI 0.57–1.76). The Garden Route's first phase had an estimated density of 0.5 leopards/100 km2 (CI 0.19–1.32) with densities stable across all buffers, whereas phase two had an estimated density of 0.34 leopards/100 km2 (CI 0.18–0.63) and phase three produced a density estimate of 0.29 leopards/100 km2 (CI 0.13–0.67; Table 2). We calculated the mean density for multi-phase study areas resulting in 0.17 leopards/100 km2 in the Overberg, compared to 0.5 leopards/100 km2 (CI 0.39–1.09) in the Langeberg and 0.38 leopards/100 km2 (CI 0.17–0.87) along the Garden Route. The highest estimated density was Langeberg's phase two with 0.7 leopards/100 km2, and the lowest was 0.17 leopards/100 km2 in the Overberg (Table 2).
1The area total is without 35 km buffer land overlap between survey areas.
SPACECAP produced an Overberg density estimate of 0.69 leopards/100 km2 (CI 0.39–1.28), Langeberg phase one and two density estimates of 1.67 (CI 1.21–2.18) and 2.11 (CI 1.46–2.82) leopards/100 km2, respectively, and Garden Route phase one, two and three of 1.44 leopards/100 km2 (CI 0.55–2.54), 0.92 leopards/ 100 km2 (CI 0.16–1.22) and 0.51 leopards/100 km2 (CI 0.36–0.7), respectively. Again, we calculated the mean density for multi-phase areas; the Overberg was 0.69 leopards/100 km2, compared to 1.89 leopards/100 km2 (CI 0.89–2.5) in the Langeberg and 0.96 leopards/100 km2 (CI 0.52–1.49) along the Garden Route (Table 2).
Analysis with secr and SPACECAP resulted in an overall estimated leopard density of 0.35 leopards/100 km2 (CI 0.21–0.81) and 1.18 leopards/100 km2 (CI 0.60–1.76) respectively. The output from SPACECAP includes a pixel density file, which we converted into a fine-scale map using QGIS, showing the variation of estimated animal densities across each of the potential home range centres (Fig. 3). Using these pixel density values, we also created study area graphs depicting the sum of pixel densities across land cover categories (CapeNature, 2014) and protected areas (Department of Environmental Affairs, 2017; Fig. 4).
Estimated population numbers
Combining this study's total area with each region's 35 km buffer width (Garden Route 6,680.34 km2; Overberg 7,910.04 km2; Langeberg 19,063.42 km2) with the regional secr mean density estimates results in an estimate of 25 leopards (CI 11.4–58.1) in the Garden Route, 13 leopards (CI 4.7–38.0) in the Overberg and 93 leopards (CI 74.3–205.9) in the Langeberg (Table 2). The total estimate for all areas combined was 132 leopards (CI 90.4–302.0; Table 2). In comparison, combining the total non-overlapping area of all three study area’ 35 km buffers (29,258.43 km2) with the overall secr mean density produces a population estimate of 102 leopards (CI 61.4–237.0) in the total study area (Table 3).
With the regional SPACECAP mean density estimates there would be c. 64 leopards (CI 34.7–99.5) in the Garden Route, 55 leopards (CI 30.8–101.2) in the Overberg and 360 leopards (CI 169.7–476.6) in the Langeberg (Table 2), yielding a total estimate of 479 leopards (CI 235.3–677.4; Table 2). Comparatively, using the total area of all three regions’ 35 km buffers (without overlap) and the overall SPACECAP mean density provides a population estimate of 345 leopards (CI 175.6–515.0) for the entire study area (Table 3).
Our study area of 29,258.43 km2 comprises 23% of the Western Cape's total area (129,460 km2). To provide a population outlook at the provincial level we extrapolated an ecologically useful Western Cape abundance estimate using our secr and SPACECAP mean density estimates and Swanepoel et al.’s (Reference Swanepoel, Lindsey, Somers, Hoven and Dalerum2013) estimated 49,850 km2 of remaining suitable leopard habitat in Western Cape (38% of the province). This suggests that the entire Western Cape could harbour as few as 175 (CI 104.7–403.8) and as many as 588 (CI 299.1–877.4) leopards (Table 3).
This study covered c. 30,000 km2 and is one of the most extensive leopard camera-trap density estimate surveys conducted in the Western Cape province of South Africa. We compared the use of two spatially explicit capture–recapture methods for a regionally imperilled disjunct leopard population. Our comprehensive density analysis investigated leopard persistence in a predominately human-modified and privately-owned landscape with varying degrees of conflict and persecution.
It is important to reiterate that our study area incorporates variation across vegetation types, landscape topography and degree of landscape modification and fragmentation. Because of its highly cultivated agricultural landscape (mainly grain production, as well as livestock farming), the Overberg is commonly considered as the breadbasket of the Cape. The Langeberg's Breede River Valley is encircled by the Cape Fold Mountain ranges and is dominated by fynbos, vineyards and orchards, whereas the Garden Route is a long stretch of the south-western coast situated between mountains to the north and the Indian Ocean to the south, which harbours a unique mixture of fynbos, indigenous temperate forest and forestry plantations. The agricultural land in all three areas is utilized by leopards, as evidenced by capture and recapture locations and suitable habitat extending beyond natural vegetation (Fig. 4) and protected area boundaries (Fig. 5). None of our identified territorial adult leopards remained strictly within protected areas, and most did not utilize any protected land within their home range. Although protected areas play a crucial role in the conservation of natural landscapes, they do not encompass the entirety of remaining natural vegetation or existing leopard habitat. In the Western Cape 79% of conservation areas contain suitable leopard habitat but only 30% of leopard habitat occurs within conservation areas (Swanepoel et al., Reference Swanepoel, Lindsey, Somers, Hoven and Dalerum2013).
These findings, together with estimated leopard densities within highly modified, cultivated and non-protected land (Figs 4 & 5), suggest that non-protected, mostly privately owned land plays an important role in sustaining the Western Cape leopard population. Both spatially explicit capture–recapture methods suggest that the Langeberg winelands region supports the highest leopard density and the Overberg's agricultural landscape the lowest. The sum of each study area's pixel densities is consistent with these findings (Fig. 4). Although densities estimates vary between study areas, the composition of land-cover and protected land utilized within each area is comparatively consistent (Supplementary Tables 1 & 2). The Langeberg's leopard pixel densities on agricultural land are higher than in the other two study areas, and equal to the Garden Route's natural vegetation (Fig. 4). This can be attributed to the predominantly non-conflict crop cultivation (vineyard and orchard) in a valley surrounded by natural mountain vegetation. Additionally, the majority (89%) of the estimated leopard pixel density on agricultural land across all study areas occurs in cultivated commercial fields and 11% occurs in forestry plantations, which can generally be considered non-conflict agriculture.
All three areas demonstrate a discrepancy between suitable habitat and land designated as protected for the conservation and sustainability of biodiversity (Figs 4 & 5). The vast majority of each area's estimated leopard pixel density occurs on non-protected land with the highest percentage occurring in the Overberg (89%) (Fig. 5; Supplementary Table 2).
Our secr and SPACECAP estimates are some of the lowest national density estimates for leopards and are comparable to three other studies in the Western Cape (Martins, Reference Martins2010; Mann, Reference Mann2014; Devens et al., Reference Devens, Tshabalala, McManus and Smuts2018; Table 4). The difference between the estimates from the two programmes was greater than expected, which could be attributed partially to the extra Maxent habitat mask covariate modelling incorporated into the SPACECAP analyses. The likelihood method requires substantially less computation time than the Bayesian approach (seconds or minutes vs hours or days), is less sensitive to buffer width and data augmentation size and has greater convenience and versatility for customization with covariates and models. SPACECAP entails checks to verify that convergence is achieved, model fit is sufficient, and data augmentation is adequate. Hence, it is more appropriate and robust for small sample sizes. Our mean SPACECAP density estimate of 1.18 leopards/100 km2 is directly comparable to results from Devens et al. (Reference Devens, Tshabalala, McManus and Smuts2018), who reported mean density estimates of 0.95 and 0.9–1.11 leopards/100 km2 for SPACECAP and two GPS methods, respectively (Table 4). The GPS methods used data from 21 collared leopards and incorporated home range size (home range density estimate) and home range overlap of same-sex neighbouring leopards (socially considerate density estimate). Therefore, we suggest our SPACECAP results are more accurate and reliable for determining the spatial requirements of species. We recommend spatially explicit capture–recapture methods for future research, with the specific incorporation of GPS collar data, because these methods are the most robust at capturing a species’ spatial ecology within a population.
1HRDE, home range density estimate; SCDE, socially considerate density estimate.
We extrapolated an ecological density (number of individuals per useable area) using our density estimates and a prediction of 49,850 km2 suitable leopard habitat remaining in the Western Cape (Swanepoel et al., Reference Swanepoel, Lindsey, Somers, Hoven and Dalerum2013). Our findings suggest that there are 175–588 leopards remaining (Table 3). However, this may be an optimistic estimation. Habitat modelling is not infallible and identified suitable habitat is not necessarily useable for wide-ranging species such as the leopard. The available habitat estimate included fragmented habitat, small isolated pockets of habitat and areas of habitat subject to edge effects and anthropogenic pressures. These factors can lead to habitat areas being unable to accommodate viable populations (Woodroffe & Ginsberg, Reference Woodroffe and Ginsberg1998). In addition, our maximum Western Cape population estimate of 588 individuals is well below the minimum viable population size necessary to maintain genetic diversity (Traill et al., Reference Traill, Bradshaw and Brook2007). The genetic population structure of leopards in the Western Cape and Eastern Cape provinces indicates very low to moderate gene flow between the three subpopulations (McManus et al., Reference McManus, Dalton, Kotzé, Smuts and Dickman2015).
Although highly adaptable, the leopard is a widely persecuted species that experiences varying levels of anthropogenic threats and habitat loss. Our results suggest that protected areas are inadequate to secure the long-term conservation of leopards in the Western Cape. By establishing density and population size estimates in an increasingly fragmented and modified landscape, we increase the understanding of how leopards can persist in human-dominated areas, influencing conservation planning for the species. For adaptive and wide-ranging species such as large carnivores, non-protected and human-dominated areas are becoming increasingly important for genetic dispersal and land- scape connectivity (Boron et al., Reference Boron, Tzanopoulos, Gallo, Barragan, Jaimes-Rodriguez, Schaller and Payán2016). The threats and conservation conflicts affecting the leopard in South Africa's Cape region are affecting all apex carnivores globally. Accurate density estimates are critically important as numerous anthropogenic interests continue to threaten leopards, their resources and habitat.
We thank the ABAX Foundation, Development Bank South Africa, Green Fund, United Nations Environmental Program, Global Environmental Facility, Henry and Iris Englund Foundation, National Lotteries Distribution Trust Fund, Mones Michaels Trust, Arne Hanson and the Deutsche Bank South Africa Foundation for assisting in funding this research. MJS was supported by the National Research Foundation. We acknowledge Landmark Foundation for providing the resources to enable this research, and thank private landowners and CapeNature for their cooperation and assistance.
Study conception: CHD, BS, JM; methodology design, data collection and writing: CHD; data analysis: CHD, TT; revisions: MS, MH, AD, TT, BS.
Conflicts of interest
All authors have abided by the Oryx guidelines on ethical standards, and fieldwork was conducted with the necessary approvals and permits from appropriate institutions and statutory authorities. Cape Nature Research Permit to collect fauna specimens for scientific research in the Western Cape was granted to CHD in August 2014 (permit no.: AAA007-00130-0056) and prior camera-trap data was covered under a Cape Nature Permit for JM. These permits were granted for the purposes of collaring leopards. CHD was also granted an Animal Ethics Committee Approval Certificate from the University of Pretoria in February 2014–December 2016 (project number: EC005-14). We did not collect specimens for this study, and all fieldwork was conducted ethically with the cooperation of local conservation authorities and landowners.