Poaching is a key phenomenon in the overexploitation of natural resources, driving many species towards extinction (Milner-Gulland et al., Reference Milner-Gulland and Bennett2003). There are various incentives for poaching, which occurs at a variety of scales, from hunting by local communities for subsistence to selling wild meat in urban markets and international trafficking of wildlife or wildlife parts for income generation (Sutherland, Reference Sutherland2000). The illegal and cryptic nature of poaching, and a lack of systematic wildlife monitoring schemes, make it difficult to detect population declines from poaching (Singh & Milner-Gulland, Reference Singh and Milner-Gulland2011; Nuno et al., Reference Nuno, Bunnefeld, Naiman and Milner-Gulland2013), and a lack of information on wildlife trends may hamper timely conservation responses (Milner-Gulland et al., Reference Milner-Gulland and Bennett2003).
Enforcement is probably the most widely practised measure against poaching, and intensification has a positive influence on wildlife (Hilborn et al., Reference Hilborn, Arcese, Borner, Hando, Hopcraft and Loibooki2006; Ghoddousi et al., Reference Ghoddousi, Hamidi, Soofi, Khorozyan, Kiabi and Waltert2016a). However, enforcement requires social acceptability and proper sanctions to be effective (Milner-Gulland & Rowcliffe, Reference Milner-Gulland and Rowcliffe2007). Integrated conservation and development projects may therefore be influential in reducing poaching by targeting economic or non-economic incentives, although they may fail if the wrong incentives are targeted (Duffy et al., Reference Duffy, St John, Büscher and Brockington2016). Knowledge about incentives for poaching at the local level is required to guide the choice of appropriate conservation measures for reducing poaching pressure (Nuno et al., Reference Nuno, Bunnefeld, Naiman and Milner-Gulland2013; Challender & MacMillan, Reference Challender and MacMillan2014).
Hunting has a long history in livelihoods and culture in Iran (Firouz, Reference Firouz2005), but with the onset of the 20th century the availability of off-road vehicles and sophisticated firearms led to wildlife massacres across the country (Firouz, Reference Firouz2005). The Caspian tiger Panthera tigris virgata and Asiatic lion Panthera leo persica have gone extinct in Iran in the past century and the Persian fallow deer Dama mesopotamica had been considered to be extinct prior to its rediscovery in 1957 (Firouz, Reference Firouz2005). Wildlife populations declined significantly in most areas until the first modern hunting control was introduced in the 1950s (Moore, Reference Moore1976; Firouz, Reference Firouz2005).
Regulation of hunting in Iran began in 1956 with the establishment of the Game Council of Iran (renamed the Department of Environment in 1974) and the creation of the first network of protected areas. These efforts initiated recovery of wildlife in some areas (Moore, Reference Moore1976; Firouz, Reference Firouz2005). Since then the Department of Environment has continued to increase the number of protected areas, and the area under its protection now comprises > 10% of the country's land (Kolahi et al., Reference Kolahi, Sakai, Moriya and Makhdoum2012). However, political turbulence since 1979, lack of acceptance of conservation laws by local communities and the insufficient resources of the Department of Environment have resulted in widespread poaching in most protected areas (Tatin et al., Reference Tatin, Darreh-Shoori, Tourenq, Tatin and Azmayesh2003; Kiabi et al., Reference Kiabi, Ghaemi, Jahanshahi and Sassani2004; Ghoddousi et al., Reference Ghoddousi, Hamidi, Soofi, Khorozyan, Kiabi and Waltert2016a).
Ungulates are a major target of hunting in Iran (Firouz, Reference Firouz2005), with a diverse range of species, including the bezoar or wild goat Capra aegagrus, chinkara or jebeer gazelle Gazella bennettii, goitered gazelle Gazella subgutturosa, mouflon or wild sheep Ovis orientalis, urial Ovis vignei, onager Equus hemionus onager, Persian fallow deer, red deer Cervus elaphus, roe deer Capreolus capreolus and wild boar Sus scrofa. Six of these species are threatened globally (IUCN, 2016; note synonymy of urial and mouflon on the Red List and the new classification of Rezaei et al., Reference Rezaei, Naderi, Chintauan-Marquier, Taberlet, Virk and Naghash2010). Apart from the Persian fallow deer, which lives in semi-captive conditions, and the wild boar, the consumption of which is prohibited on religious grounds, all other ungulate species are threatened by poaching and have declined since the 1970s (Karami et al., Reference Karami, Hemami and Groves2002; Tatin et al., Reference Tatin, Darreh-Shoori, Tourenq, Tatin and Azmayesh2003; Kiabi et al., Reference Kiabi, Ghaemi, Jahanshahi and Sassani2004; Shams Esfandabad et al., Reference Shams Esfandabad, Karami, Hemami, Riazi and Sadough2010; Ghoddousi et al., Reference Ghoddousi, Hamidi, Soofi, Khorozyan, Kiabi and Waltert2016a; Soofi et al., Reference Soofi, Ghoddousi, Hamidi, Ghasemi, Egli and Voinopol-Sassu2017).
Despite widespread poaching, the scale of these declines in Iranian protected areas is unknown, as robust ungulate monitoring techniques are largely lacking. Furthermore, knowledge of the incentives for poaching remains limited. Subsistence, monetary profit, cultural values and conflict with the Department of Environment are the main incentives for poaching of ungulates in Bamu National Park (Ashayeri & Newing, Reference Ashayeri and Newing2012), but whether this is also the case in other protected areas in Iran is unclear.
We measured ungulate population trends in Golestan National Park, for which ungulate abundance data are available from the 1970s (Decker & Kowalski, Reference Decker and Kowalski1972; Kiabi, Reference Kiabi1978; Kiabi et al., Reference Kiabi, Ghaemi, Jahanshahi and Sassani2004). We assessed the populations of four species (bezoar goat, red deer, urial and wild boar) and compared them with their earlier status. There is no information on large-scale migrations, diseases or other environmental conditions, which might have affected the populations of these four species during this period (Ghoddousi et al., Reference Ghoddousi, Soofi, Hamidi, Lumetsberger, Egli and Khorozyan2016b). There is no competition with livestock in this area, as grazing is banned inside the Park and illegal grazing occurs only at a limited scale along the periphery (Ghoddousi et al., Reference Ghoddousi, Soofi, Hamidi, Lumetsberger, Egli and Khorozyan2016b). Moreover, no major habitat destruction or development projects have reduced wildlife habitats in the Park in recent decades. However, poaching has been widely reported as being one of the main threats to ungulate species in the Park (Kiabi et al., Reference Kiabi, Ghaemi, Jahanshahi and Sassani2004; Hamidi et al., Reference Hamidi, Ghoddousi, Soufi, Ghadirian, Jowkar and Ashayeri2014; Ghoddousi et al., Reference Ghoddousi, Hamidi, Soofi, Khorozyan, Kiabi and Waltert2016a; Soofi et al., Reference Soofi, Ghoddousi, Hamidi, Ghasemi, Egli and Voinopol-Sassu2017).
The severity of penalties, likelihood of capture by rangers, and incentives are the most important factors in poaching decisions (Milner-Gulland & Leader-Williams, Reference Milner-Gulland and Leader-Williams1992). Hunting is illegal in Golestan National Park and incurs fines or imprisonment. The density of rangers in the Park (c. 1 per 29.1 km2; authors, unpubl. data) is deemed sufficient to control illegal activities according to international recommendations (one ranger per 23.8 km2; Jachmann & Billiouw, Reference Jachmann and Billiouw1997). As understanding incentives may help managers find solutions to curb poaching (Milner-Gulland & Leader-Williams, Reference Milner-Gulland and Leader-Williams1992), we also evaluated the incentives for local poachers of ungulates in the context of existing disincentives.
Golestan National Park was established in 1957 in north-eastern Iran (Fig. 1). It encompasses Hyrcanian montane forests, steppes and arid plains in an area of 874 km2 (Akhani, Reference Akhani2005). From west to east, elevations span 450–2,411 m, with mean annual precipitation of 866–142 mm, thus creating a variety of habitats (Akhani, Reference Akhani2005). The Park holds six species of ungulates, which is one of the highest numbers of ungulate species in Iranian protected areas (Ghoddousi et al., Reference Ghoddousi, Soofi, Hamidi, Lumetsberger, Egli and Khorozyan2016b). Urial occur in steppes in the east and north of the Park, and roe deer and red deer inhabit forests in western and central parts. Wild boar are present throughout the Park (with the exception of a 25 km2 arid plain), with higher densities in forests. A population of goitered gazelles occupies narrow plains in the east and north. Bezoar goats occur on cliffs across the Park. There are no villages within the Park but there are 15 villages, with c. 8,660 inhabitants, within 2 km of the Park borders. These communities rely mainly on crop and livestock farming.
Estimation of ungulate populations
We used a variety of methods to estimate population size, depending on the detection probabilities and habitat characteristics of the ungulate species (Ghoddousi et al., Reference Ghoddousi, Soofi, Hamidi, Lumetsberger, Egli and Khorozyan2016b). We excluded the goitered gazelle and roe deer from our study because the former has a limited distribution in the Park and there are insufficient data on the latter to facilitate estimation. Our estimates of abundance for 2011–2014 were compared to data for 1970–1978, which were based on comparable systematic monitoring methods (Decker & Kowalski, Reference Decker and Kowalski1972; Kiabi, Reference Kiabi1978; Kiabi et al., Reference Kiabi, Ghaemi, Jahanshahi and Sassani2004). We are not aware of any other studies on populations of these ungulates in the Park during this time frame. When more than one estimate for a given species was available, or if a population range was given, we calculated the arithmetic mean.
We used a double-observer point count during November–December 2014 to estimate the abundance of bezoar goats (Suryawanshi et al., Reference Suryawanshi, Bhatnagar and Mishra2012). We identified 53.6 km2 of rugged landscape as bezoar goat habitat and surveyed 16 sampling points at random within this area, with a minimum distance of 3 km between two points (Fig. 1). Two observers counted groups of goats, from vantage points 200–500 m away from the sampling points. The data were analysed using DOBSERV (Nichols et al., Reference Nichols, Hines, Sauer, Fallon, Fallon and Heglund2000). The sampled area was calculated as the overlap of observable areas from vantage points and the identified goat habitat, using the viewshed function in ArcGIS 10.1 (ESRI, Redlands, USA). A detailed description of our sampling and modelling approaches is provided by Ghoddousi et al. (Reference Ghoddousi, Soofi, Hamidi, Lumetsberger, Egli and Khorozyan2016b). The earlier estimate of the bezoar goat population in the Park (4,000–4,500) was based on full-day observations in sample areas during 1976–1978 (Kiabi, Reference Kiabi1978).
To estimate the red deer population size we used dung counts (faecal standing crop approach) and camera-trap data (randomized encounter model) in 422 km2 of forests and grasslands (Buckland et al., Reference Buckland, Anderson, Burnham, Laake, Borchers and Thomas2001; Rowcliffe et al., Reference Rowcliffe, Field, Turvey and Carbone2008). For the dung counts we estimated red deer defecation rates by observing 20 individuals for 8 days in a 0.02 km2 enclosure with habitat comparable to that in the Park. Prior to the survey we estimated the dung decay rate by monitoring 80 fresh dung samples across red deer habitats in the Park and using binary logistic regression to assess the influence of time and habitat types on the survival of dung samples. We then surveyed 18 strip transects of 2 km length and 2 m width across red deer habitats during January–February 2013 (Fig. 1). For the randomized encounter modelling we used data from an earlier study, gathered from 37 camera traps during May–December 2011 (Fig. 1) (Hamidi et al., Reference Hamidi, Ghoddousi, Soufi, Ghadirian, Jowkar and Ashayeri2014). We calculated mean group size from observations of 57 red deer groups by park rangers during the period of camera trapping. Given the lack of red deer movement data from Golestan National Park or elsewhere in Iran, we used a mean of daily range estimates from other studies of 3.36 ± SE 0.23 km per day (Soofi et al., Reference Soofi, Ghoddousi, Hamidi, Ghasemi, Egli and Voinopol-Sassu2017). The camera-related parameters required by the randomized encounter model were obtained from a previous study that used a similar brand of camera trap (Deercam DC300; Non Typical Inc., De Pere, USA; Rowcliffe et al., Reference Rowcliffe, Field, Turvey and Carbone2008). Details of the sampling and the application of both methods are provided by Soofi et al. (Reference Soofi, Ghoddousi, Hamidi, Ghasemi, Egli and Voinopol-Sassu2017). Red deer abundance during 1976–1978 was estimated by Hahn's census method via line transect surveys, and dung counts, as 1,897 and 2,096 individuals, respectively (Kiabi et al., Reference Kiabi, Ghaemi, Jahanshahi and Sassani2004).
We used line transects to estimate the urial population size in 340 km2 of steppes in the east and north of the Park (Buckland et al., Reference Buckland, Anderson, Burnham, Laake, Borchers and Thomas2001). We surveyed 17 3-km transects during January–February and August–September 2013, and February 2014 (Fig. 1), and analysed the data using Distance 6.0 (Thomas et al., Reference Thomas, Buckland, Rexstad, Laake, Strindberg and Hedley2010). A detailed description of the methodology used to estimate the urial population is provided by Ghoddousi et al. (Reference Ghoddousi, Hamidi, Soofi, Khorozyan, Kiabi and Waltert2016a). The urial population in 1970 was estimated by total counts in 12 sampling units and the extrapolation of recorded densities over the steppes of the Park (Decker & Kowalski, Reference Decker and Kowalski1972). The estimated abundance was c. 15,000 individuals (Decker & Kowalski, Reference Decker and Kowalski1972). In a separate survey that used direct counts on line transects, there were estimated to be 10,000–11,000 urial in the Park during 1976–1978 (Kiabi, Reference Kiabi1978).
We estimated the abundance of wild boar using randomized encounter modelling based on camera-trap surveys conducted during January–December 2011 (Rowcliffe et al., Reference Rowcliffe, Field, Turvey and Carbone2008; Hamidi et al., Reference Hamidi, Ghoddousi, Soufi, Ghadirian, Jowkar and Ashayeri2014). We used data from 67 camera traps installed throughout the Park, excluding a 25 km2 arid plain (Fig. 1) (Hamidi et al., Reference Hamidi, Ghoddousi, Soufi, Ghadirian, Jowkar and Ashayeri2014). The mean group size was estimated from observations on line transects during 2013–2014 (see above). Given the lack of information on the daily range of this species from the study site, we used an estimate of 6.8 ± SE 0.57 km per day from a study with similar habitat conditions (Podgórski et al., Reference Podgórski, Baś, Jędrzejewska, Sőnnichsen, Śnieżko, Jędrzejewski and Okarma2013). Details of our study design and analysis are provided by Hamidi et al. (Reference Hamidi, Ghoddousi, Soufi, Ghadirian, Jowkar and Ashayeri2014) and Ghoddousi et al. (Reference Ghoddousi, Soofi, Hamidi, Lumetsberger, Egli and Khorozyan2016b). We used an earlier estimate of 2,500–3,000 wild boar from line transect surveys during 1976–1978 (Kiabi, Reference Kiabi1978).
To assess the effects of uncertainty in population estimates we conducted sensitivity analysis using various combinations of 95% confidence intervals (2011–2014) and population ranges (1970–1978) for each species.
Interview survey and arrest records
Taking into account that poaching is a sensitive subject, we collected data on incentives for poaching by conducting semi-structured interviews with local people (Newing, Reference Newing2011). We identified poachers in villages in the vicinity of the Park through a process of chain referral (Newing, Reference Newing2011). Local poachers agreed to participate in the study after we collaborated with them in joint wildlife monitoring programmes (Hamidi et al., Reference Hamidi, Ghoddousi, Soufi, Ghadirian, Jowkar and Ashayeri2014; Ghoddousi et al., Reference Ghoddousi, Hamidi, Soofi, Khorozyan, Kiabi and Waltert2016a), built mutual trust, and explained the purpose of the study to them. Not being affiliated to any governmental organizations facilitated the process of data gathering and communicating with poachers. We asked each individual to give the main reasons for poaching ungulates in the Park (Ashayeri & Newing, Reference Ashayeri and Newing2012). We encoded and categorized incentives based on similar elements in responses (Ashayeri & Newing, Reference Ashayeri and Newing2012). We assured interviewees that their data would remain anonymous and interviewees gave their verbal consent to participation in the survey. We held a focus group meeting with five poachers in December 2012 to elicit information about the incentives for poaching. We also used the findings of a social study on poachers in the Park (Ashayeri, Reference Ashayeri2014), conducting informal qualitative interviews with 15 poachers during June 2013–February 2014. Interviews were continued until data reached the level of saturation, meaning that no further information could be extracted from new interviews (Newing, Reference Newing2011). Details of the interview procedure and analytical approaches used are in Ashayeri (Reference Ashayeri2014). The interviewees in both studies were all men, 29–66 years old, from 10 villages around the Park (Fig. 1). We also used seizure records for 2007–2014 to assess the frequency of poaching of various species in the Park. We calculated the poaching rate for each species as the number of hunted individuals as a proportion of their population size (2011–2014). As rangers use a sit-and-wait approach or tip-offs from local informants to detect poachers, we believe that seizure data are not biased towards a certain species, and represent the distribution of hunted species in the Park (authors, unpubl. data).
During 64 scans of 15 minutes each we observed 39 bezoar goats in seven groups. The model with equal detection probability between the observers estimated an abundance of 519 individuals (confidence interval CV = 31.3%; 95% CI = 201–807; Table 1). A comparison of recent estimates with the mean population size from 1976–1978 indicates an 88% decline of the bezoar goat population (Fig. 2). The results of sensitivity analysis show a decline of 79–96% during this period (Table 1).
1 Arithmetic mean of two population estimates/ranges
2 Kiabi (Reference Kiabi1978)
3 Ghoddousi et al. (Reference Ghoddousi, Soofi, Hamidi, Lumetsberger, Egli and Khorozyan2016b)
4 Kiabi et al. (Reference Kiabi, Ghaemi, Jahanshahi and Sassani2004)
5 Soofi et al. (Reference Soofi, Ghoddousi, Hamidi, Ghasemi, Egli and Voinopol-Sassu2017), from random encounter models
6 Decker & Kowalski (Reference Decker and Kowalski1972)
7 Ghoddousi et al. (Reference Ghoddousi, Hamidi, Soofi, Khorozyan, Kiabi and Waltert2016a)
From 1,676 dung samples we estimated a defecation rate of 10.5 dung piles per individual per day. Given the lack of knowledge regarding the variation in defecation rates among red deer individuals it was not possible to calculate the standard error. The age-based model estimated a red deer dung decay rate of 142 ± SE 15 days. Fifty red deer dung samples were detected from a survey effort of 36 km. The faecal standing crop method estimated an abundance of 194 red deer (CV = 28.4%; 95% CI = 103–285; Table 1). We captured 10 photographs of red deer from 1,585 camera-nights of effort in forests and grasslands of the Park (Fig. 1). Using the randomized encounter modelling approach we estimated a red deer population of 257 individuals (CV = 35.3%; 95% CI = 91–423; Table 1). Comparison of the mean red deer populations in 2011–2013 and 1976–1978 indicates an 89% decline (Fig. 2). We used the wider 95% CI from the randomized encounter model for red deer sensitivity analysis, and the results indicated a decline rate of 78–96% compared to 1976–1978 (Table 1).
From a total survey effort of 186 km we observed 1,981 urial in 70 groups. The half-normal key detection function of Distance 6.0 estimated a population of 4,275 individuals (CV = 35.5%; 95% CI = 2,117–8,632; Table 1). Comparison between 2013–2014 abundance estimates and the mean of the two estimates for 1970 and 1976–1978 indicates a 66% decline in urial abundance (Fig. 2). The results of sensitivity analysis indicated a 14–86% decline in the urial population (Table 1).
From our observations of 38 groups of wild boar on line transects we estimate a mean group size of 3.1 ± SE 0.9 individuals. We captured 386 photographs of wild boar during 2,777 trap-nights across the Park, and using the randomized encounter modelling approach we estimated an abundance of 6,478 individuals (CV = 27.0%; 95% CI = 3,050–9,906; Table 1). Compared to 1976–1978, the wild boar population increased by 58% (Fig. 2). The results of the sensitivity analysis indicated a population increase of 2–75% (Table 1).
Poaching incentives and frequency of hunted species
The results of our interview surveys and an earlier study (Ashayeri, Reference Ashayeri2014) revealed five main categories of incentives for poaching, although we concluded that a poacher's decision to go on a hunting trip was influenced by a combination of incentives. Reported incentives were poverty/livelihoods, hunting for meat market/trade, pleasure/love of hunting, tradition/habits, and hunting for revenge or as a result of conflict with conservation regulations and organizations. Our approach did not facilitate ordinal ranking of incentives based on their importance but showed a mixture of social, economic and policy-related motivations for poaching in the Park (Fig. 3). Our data on law enforcement records for 2007–2014 indicate 113 arrests of poachers, 38 before hunting had taken place. In the remaining 75 cases 113 individuals of eight species had been hunted. Urial accounted for the highest proportion of hunted species (58%), followed by red deer (12%) and bezoar goats (9%). The rate of poaching relative to population size was highest for red deer (5.75%), followed by bezoar goats (1.92%), urial (1.54%) and wild boar (0.10%). Circa 27.4% of the arrested poachers were not from the villages in the vicinity of the Park.
We estimate there has been a 66–89% population decline of urial, red deer and bezoar goats in Golestan National Park since the 1970s. These were the preferred species of poachers (authors, unpubl. data) and accounted for the majority of hunted species. Such a trend in the absence of any reported migrations or mass mortalities may represent the effects of poaching in the Park. The fact that the population of wild boar, consumption of which is prohibited by Islam, has increased by 58% during the same time frame further supports our claim. Poachers avoid hunting this species in the Park (authors, unpubl. data).
The greatest declines were in bezoar goat (88%) and red deer (89%) populations. Bezoar goat habitat is restricted to patches of cliffs within the Park, where the goats are exposed and vulnerable to poaching. The Park is one of the last population strongholds of red deer in the Caspian forests (Kiabi et al., Reference Kiabi, Ghaemi, Jahanshahi and Sassani2004) but despite dense vegetation and the elusive behaviour of red deer, this species is vulnerable to poaching, especially during the rutting season (authors, unpubl. data). In September and October each year, poachers imitate stag calls to attract deer to within shooting range. Without immediate conservation action both the bezoar goat and the red deer may go extinct in the Park in the near future. The urial population has also declined (66%), and the species is almost extirpated from some of its former range in the Park (Decker & Kowalski, Reference Decker and Kowalski1972; Ghoddousi et al., Reference Ghoddousi, Hamidi, Soofi, Khorozyan, Kiabi and Waltert2016a). However, urial still occur in higher densities in the vicinity of the ranger stations (Ghoddousi et al., Reference Ghoddousi, Hamidi, Soofi, Khorozyan, Kiabi and Waltert2016a). It appears that the lack of regular systematic monitoring coupled with low detection probability of some species may have created an illusion of plenty among park managers, who may underestimate the decline in populations of hunted ungulates.
In Golestan National Park and other Iranian protected areas species living in open landscapes have been routinely monitored by annual total counts. Total counts do not follow a systematic sampling approach and the assumption of observation of all individuals in large areas can rarely be met (Buckland et al., Reference Buckland, Anderson, Burnham, Laake, Borchers and Thomas2001). Moreover, this method does not provide a measure of variance, which is necessary for assessing population trends over time (Suryawanshi et al., Reference Suryawanshi, Bhatnagar and Mishra2012). Therefore, it is necessary to adopt monitoring methods that are suitable for rugged landscapes and are sufficiently robust to detect trends in exploited populations at low densities (Singh & Milner-Gulland, Reference Singh and Milner-Gulland2011). We used a variety of monitoring methods successfully, and we recommend that the Department of Environment should initiate capacity-building programmes for rangers and invest in the equipment required to conduct similar surveys on a regular basis.
The differences in monitoring methods used in our study and those used to gather historical data are a potential source of bias. However, we are not aware of any other systematic surveys of these species since the establishment of the Park (Decker & Kowalski, Reference Decker and Kowalski1972; Kiabi, Reference Kiabi1978; Kiabi et al., Reference Kiabi, Ghaemi, Jahanshahi and Sassani2004). Although our camera trapping was designed to target leopards Panthera pardus (Hamidi et al., Reference Hamidi, Ghoddousi, Soufi, Ghadirian, Jowkar and Ashayeri2014), we assume it did not produce a major bias, as movement patterns of herbivores are independent of those of carnivores (Cusack et al., Reference Cusack, Dickman, Rowcliffe, Carbone, Macdonald and Coulson2015). Moreover, randomized encounter modelling of data for red deer and wild boar produced comparable results to other tested methods (Ghoddousi et al., Reference Ghoddousi, Soofi, Hamidi, Lumetsberger, Egli and Khorozyan2016b; Soofi et al., Reference Soofi, Ghoddousi, Hamidi, Ghasemi, Egli and Voinopol-Sassu2017).
Our results demonstrate that conservation laws and enforcement measures have failed to stop poaching since the 1970s and therefore require reconsideration. We identified a combination of economic and non-economic incentives for poaching of ungulates in the Park, which may guide the selection of appropriate anti-poaching schemes. The variety of incentives to poach suggests that single policies are unlikely to succeed in deterring poaching, and that a combination of approaches is therefore required (Duffy et al., Reference Duffy, St John, Büscher and Brockington2016). Poverty and the existence of a market for meat can be considered economic incentives, and creating alternative livelihoods for local communities is a common approach to tackle poaching stemming from such economic incentives (Duffy et al., Reference Duffy, St John, Büscher and Brockington2016). However, subsistence poachers normally lack the skills, education and cultural capacities required for employment in many sectors (Nuno et al., Reference Nuno, Bunnefeld, Naiman and Milner-Gulland2013). Thus, integrated conservation and development projects could potentially explore livelihood opportunities in developing ecotourism or facilitating the establishment of community-based reserves, benefiting from poachers' local ecological knowledge. Awareness-raising campaigns against consumption of wild meat in urban areas could be considered, to target the demand side (Challender & MacMillan, Reference Challender and MacMillan2014). In addition, the distribution and efficiency of law enforcement efforts in Golestan National Park should be improved (Ghoddousi et al., Reference Ghoddousi, Hamidi, Soofi, Khorozyan, Kiabi and Waltert2016a).
The existence of hunting incentives related to tradition and pleasure suggests that alternative livelihood programmes alone may fail to address the poaching problem (Waylen et al., Reference Waylen, McGowan and Milner-Gulland2009), but community outreach programmes aimed at building trust, awareness, motivation and opportunities have proven to be influential in controlling poaching in South-east Asia (Steinmetz et al., Reference Steinmetz, Srirattanaporn, Mor-Tip and Seuaturien2014). As the limited number of hunting permits issued annually by the Department of Environment is insufficient to satisfy demand, establishing community-based reserves could provide legal hunting opportunities for local communities. Integrated conservation and development programmes could investigate opportunities for creating such reserves.
As in a previous study in Iran (Ashayeri & Newing, Reference Ashayeri and Newing2012), conflict with conservation bodies and regulations was stated to be an incentive for poaching. The non-participatory and top-down approach to protected area management (Zendehdel et al., Reference Zendehdel, Rademaker, De Baets and Van Huylenbroeck2010), coupled with hostile encounters between rangers and local communities, causes conflict between the two parties. Additionally, hiring non-local rangers may overlook local ecological knowledge, leave local communities out of decision-making and cause conflict between local communities and conservation authorities. Nevertheless, we presume that conflict may exacerbate poaching but is not a root cause of it.
The future of hunted species in Golestan National Park and protected areas elsewhere in Iran is unclear. We recommend that the Department of Environment should adopt participatory conservation strategies, improve law enforcement practices and cooperate with international experts in resolving the poaching problem nationwide.
We acknowledge the support of the Iranian Department of Environment, the Golestan provincial office of the Department of Environment, and the management of Golestan National Park. We appreciate the cooperation of the Persian Wildlife Heritage Foundation in granting access to their camera-trap and social survey data and for providing logistical support. This study would not have been possible without the collaboration of rangers, local guides and volunteers. AG was funded by the German Academic Exchange Service (DAAD) and Panthera's Kaplan Graduate Award, and MS was funded by Erasmus Mundus/SALAM. We acknowledge Turkmen Ecolodge for hosting the focus group meeting with local poachers, and we thank K. Seyed Emami, M. Festa-Bianchet and an anonymous reviewer for their feedback.
AG, MS, AKH, SA, LE and MW conceived and designed the research. AG, MS, AKH, SA, LE and SG conducted the surveys. AG, MS, SA and LE analysed data. AG, MS, AKH, SA, LE, JS, IK, BHK and MW wrote the article.
Arash Ghoddousi works on poaching mitigation, law enforcement practices, wildlife monitoring techniques and human–wildlife conflict. Mahmood Soofi studies human–wildlife interactions at the landscape level. Amirhossein Kh. Hamidi is involved in participatory conservation practices. Sheyda Ashayeri conducts research on human dimensions in conservation. Lukas Egli investigates interactions between land-use change and biodiversity. Siavash Ghoddousi studies ecotourism and regional development. Julian Speicher focuses on ecology and management of wild animals in natural and modified habitats. Igor Khorozyan works on human–wildlife conflict and carnivore conservation. Bahram H. Kiabi studies relationships between wildlife populations and environmental factors. Matthias Waltert is interested in wildlife conservation research and teaching, and has experience in the Old World tropics and the Middle East.