1The trade in wild animals for meat, or ‘bushmeat’, is perceived as one of the most important threats to wildlife in the tropics. Unsustainable bushmeat extraction also threatens the loss of livelihoods. However, the long-term persistence of the bushmeat trade, documented in Africa over several centuries, suggests that the trade can be sustainable. In this study, we investigate sustainability in a mature bushmeat market in West Africa (Takoradi, Ghana).
2Our study, conducted over January and February 2000, combined biological and socio-economic approaches. Offtake data, including information on species identity, capture location and sales price, were collected in a market survey. Species biological data, and the historical price of bushmeat and its substitutes (fish and domestic meat), were taken from the literature. The theoretical sustainable yield for each species was estimated using standard algorithms.
3We tested the hypothesis that the current trade is unsustainable with four predictions: that (1) the number of animals extracted exceeds a theoretical sustainable yield, (2) larger taxa are depleted more heavily close to the city, (3) the price of bushmeat has outstripped inflation and (4) the price of alternatives, such as domestic meat and fish, has fallen relative to the price of bushmeat. None of these predictions were supported. There was therefore no evidence of unsustainability.
4Analysis of market profiles and hunter reports suggest that the present pattern of sustainability is the result of a series of non-random extinctions from historical hunting. Vulnerable taxa (slow reproducers) have been depleted heavily in the past, so that only robust taxa (fast reproducers), such as rodents and small antelope, are now traded. These robust taxa are supplied from a predominantly agricultural landscape around the city.
5Synthesis and applications. The bushmeat trade can have a severe impact on species that are vulnerable to overexploitation. However, once these species have disappeared, the remaining species may be harvested sustainably. Bushmeat management policy might therefore be improved by adopting a two-pronged approach in which vulnerable species are protected from hunting, but robust species are allowed to supply a sustainable trade. The productivity of agricultural landscapes for many bushmeat species indicates that these areas may play an important role in supporting a sustainable bushmeat trade.
Over recent years, a growing number of studies have described bushmeat harvesting as unsustainable (e.g. Robinson & Bennett 2000), particularly in Africa where the volume of extraction is exceptionally high (Fa, Peres & Meeuwig 2002). The term bushmeat has therefore become synonymous with overexploitation. However, the link between bushmeat extraction and unsustainable use is likely to be more complicated. African societies have harvested and traded bushmeat for centuries (e.g. Lewicki 1974; Mendelson, Cowlishaw & Rowcliffe 2003) and this trade must have possessed some element of sustainability, at least until recent times, otherwise bushmeat and all its associated animal species would have disappeared a long time ago. In reality, the extent to which the trade is unsustainable is likely to be variable, contingent on a variety of supply-and-demand factors such as the available habitat for bushmeat species and the local human population size. An improved understanding of this variability in the bushmeat trade is urgently needed to enhance our understanding of its impacts and to improve the efficacy of conservation action.
In light of this need, the purpose of the present study was to investigate sustainability in a mature (long established) urban bushmeat market in West Africa. Analysis of such a market should be particularly insightful because urban markets tend to be associated with unsustainable exploitation, but a mature market implies some degree of sustainability.
Our first step was to establish whether the local bushmeat trade is unsustainable. Sustainability is notoriously hard to assess in such systems, owing to the difficulties in reliably monitoring offtake across all prey species at the necessary spatial and temporal scales. Consequently, we used a suite of indirect measures. These do not allow us to prove sustainability outright, but do provide evidence for or against a lack of sustainability. These measures encompass both biological and socio-economic data to produce a number of indicators. Individually these measures are not fully informative, but cumulatively they indicate whether there is evidence for unsustainable use.
According to the overexploitation hypothesis, the current harvesting of species is unsustainable and is therefore associated with a decline in species populations and bushmeat availability. Under these conditions, our indicators should show the following patterns.
1. The observed levels of offtake for each taxa will exceed their theoretical sustainable yield (prediction 1·1).
2. Across taxa there will be a positive correlation between body mass and average capture distance from the city (prediction 1·2). This is a common pattern associated with unsustainable hunting around population centres (Alvard et al. 1997).
3. The market price of bushmeat will outstrip inflation as a result of bushmeat becoming increasingly scarce (prediction 1·3).
4. The growing scarcity of bushmeat will make it increasingly more expensive than its alternatives such as beef, mutton and fish (prediction 1·4). These last two predictions ensue from both the short-term effects of reduced supply relative to demand and the long-term effects of a rise in harvesting costs with increasing scarcity.
Given that we found no evidence of unsustainable use, we then investigated those factors that might lead to sustainability. We hypothesized that the market has experienced a historical period of overexploitation, leading to the depletion of the most vulnerable species and their concomitant disappearance from the marketplace, such that only robust species that can sustain high levels of offtake remain. Sustainability therefore arises as a consequence of an ‘extinction filter’ (Balmford 1996). We tested this historical depletion hypothesis using two predictions: that those species most vulnerable to overexploitation, i.e. those with the slowest reproductive rates, would be absent from the market (prediction 2·1), and that local hunters would report historical declines but current stability in the abundance of prey species (prediction 2·2).
Our study focused on the mature urban bushmeat market of Sekondi-Takoradi (hereafter Takoradi), Ghana's third largest city, located in the Upper Guinea Forest global biodiversity hotspot (Myers et al. 2000). Takoradi has several centuries of recorded settlement and has grown rapidly over the last 100 years through the development of gold mining, railways and harbour facilities. This economic growth has also transformed the surrounding hinterland into an agricultural farmbush matrix, consisting of a mosaic of plantations, mixed bush fallow (predominantly cocoa, coconut and oil palm) and remnant tropical forest (including secondary forest). Bushmeat has been documented as part of the local diet for centuries (Grubb et al. 1998). Data were collected in January–February 2000, a period representative of annual bushmeat trading: it avoided both the peak hunting season (May–July inclusive: Holbech 1998) and the closed season (August–November inclusive), and was described consistently as a typical month by the traders interviewed.
Data were collected using a combination of direct observation and semistructured interviews (following Magrath 1992), from a representative sample of all actors in the market (farmer hunters, commercial hunters, wholesalers, market traders and chopbars). These data describe a total of 2430 bushmeat transactions reported by 70 different actors encompassing 16 different taxa. Data collected for each transaction included: taxon identity, weight and condition (fresh/smoked), identity of purchaser (e.g. wholesaler, the public) and sale price (in Ghanaian cedis, ¢). For a subset of transactions, additional information included the supplier of the meat (n = 1745 transactions) and the capture location (n = 438 transactions between hunters and market traders). Distances between capture location and the city market were calculated on the basis of road distance (the distance travelled by the hunter). Transaction weight was either measured directly (n = 1094) or reported by the seller (n = 1362): both methods produced the same mean carcass weights and were combined for this analysis. Supplementary data were collected on actor perceptions of changes in bushmeat availability (n = 12 hunters) and the weight, price and type of fish and domestic meat sold in the market (n = 1138). All our informants were open and relaxed, and their responses showed a high degree of consistency when subjected to direct observation and when cross-referenced to other actors. None of the recorded trade was illegal, although illegal trade usually takes place openly where it does occur in Ghana (Ntiamoa-Baidu 1998). Further information on the Takoradi bushmeat trade and details of data collection are given in Cowlishaw et al. (2005) and Mendelson et al. (2003).
Our analyses focused on the 10 terrestrial mammals in the trade comprising 84% of the total biomass sold. All analyses involving sales price per kilogram were based on the market value of smoked meat (85% of all retail sales). Historical data describing 1963 Takoradi market prices are taken from Asibey (1966). To investigate changes in the real value of bushmeat (inflation-corrected), prices were adjusted according to the Consumer Price Index (CPI) taken from IMF (1980, 2001). The CPI measures the cost of a standardized basket of market goods and is the most widely used measure of inflation. The prices of bushmeat and other types of meat in 2000 were determined by averaging the mean sales price across traders. Our expectations about price changes (predictions 1·3 and 1·4) assume that bushmeat is a luxury, or superior, good. Otherwise, it would not become more expensive when scarce because consumers would switch to less expensive alternatives. Although we were unable to explore the precise elasticity of bushmeat consumption in Takoradi (data describing local household income and expenditure on bushmeat are unavailable), our research indicates that bushmeat is a luxury item. Bushmeat was more expensive per kg than fish, beef or mutton in 1963 (Asibey 1966) and continued to be in 2000 (see below), and during our study consumers in Takoradi preferred to eat bushmeat over these alternatives when they could afford it (Mendelson et al. 2003).
Determination of the current annual extraction of the 10 taxa across the Takoradi catchment took two steps. In the first step, we calculated species extraction for the city itself, in three discrete stages. (i) To determine the number of urban retailers, the number of market stalls, nm, was recorded through a complete census of market places while the number of chopbars, nc, was estimated by extrapolation from nm and the average per capita bushmeat biomass purchased by each chopbar from market traders (Bc,m) and sold by market traders to each chopbar (Bm,c) (calculated from 187 and 375 independently sampled transactions, respectively). This extrapolation was based on the assumption that nc · Bc,m = nm · Bm,c · (ii) To determine the annual urban sales by these traders, we multiplied nm and nc by the average number of species carcasses we observed each selling (per capita) in a typical 1-month period, and multiplied this by 12. (iii) To estimate annual bushmeat extraction for all urban sales, we accounted for the additional informal trade between hunters and consumers, on the basis that such sales may comprise up to 18% of urban sales (Ntiamoa-Baidu 1998). In the second step, we calculated total extraction from the entire catchment in two discrete stages. (i) We first accounted for rural sales of bushmeat, on the basis that only 17% of sales by farmer hunters are urban sales (Falconer 1992). (ii) We then accounted for bushmeat that hunters eat themselves or give away, on the basis that farmer hunters sell either a consistent fraction equal to 69% of all captures (Method 1) (Falconer 1992) or an inconsistent fraction of captures ranging from 17% to 100% (median 66%) depending on the species in question (Method 2) (Ntiamoa-Baidu 1998). These calculations thus produced two estimates of the annual extraction for each species. These estimates indicate that the formal urban bushmeat sales (160 000 kg fresh mass) are only 14% of total extraction (1130 tonnes: mean, Methods 1 and 2) for these taxa in the Takoradi catchment.
The two estimates of annual extraction for each species were compared to the estimated sustainable production for each species according to two sustainability indices (Milner-Gulland & Akçakaya 2001). These indices calculate a species sustainable production, P, according to its population density (either at its current population size, N, or at carrying capacity, K), its intrinsic rate of population increase, rmax, and its mortality or recovery factor, F. According to the Robinson and Redford algorithm:
PRR = 0·6K(rmax − 1)FRR
In contrast, the US National Marine Fisheries Service algorithm states that:
PNMFS = 0·5N(rmax − 1)FNMFS
To determine K and N, species population densities, d, were first taken from Fa & Purvis (1997) or estimated through standard allometric relationships (Rowcliffe, Cowlishaw & Long 2003). The catchment area for each species was then defined as the area around Takoradi between its minimum and maximum recorded capture distances. The landscape around Takoradi is characterized by a farmbush matrix: a representative habitat type for all species in the market (habitat preferences from Grubb et al. 1998). Nevertheless, species estimates for K and N within the catchment were set to 0·75d and 0·50d, to account for habitat heterogeneity and historical hunting in the matrix, respectively. To err on the side of caution, we also repeated the analysis with lower values (K = 0·50d, N = 0·25d), but similar results were obtained. Species values of rmax were also estimated by allometry (Rowcliffe et al. 2003). FRR was set at 0·6, 0·4 and 0·2 for very short-lived species (rodents, 0·5–4 kg), short-lived species (small ungulates, 4–14 kg) and long-lived species (bushbuck, 43 kg), respectively (Robinson 2000); FNMFS was set at 0·5 throughout (Milner-Gulland & Akçakaya 2001).
These algorithms are commonly used to determine the maximum sustainable production and thus whether an observed yield is unsustainable. However, they are less reliable at pinpointing a sustainable yield because production is not always maximal (e.g. Robinson 2000). We therefore made conservative estimates of production and sustainability by using modest values of K & N. Stephens et al. (2002) reported recently that PRR may overestimate sustainable yields in social species. However, their conclusions were based on models of marmots, Marmota marmota, which live in larger and more complex social groups than any of the species entering the Takoradi market. In addition, Milner-Gulland & Akçakaya (2001) found that PNMFS outperforms PRR in its ability to correctly predict sustainability. Here we use PRR because it provides an alternative estimate of sustainable production and because it is also the most widely used index for estimating sustainability in the field (thus providing comparability with previous studies). Nevertheless, it should be noted that the PNMFS figures are likely to be the more accurate of the two measures.
Parametric statistical tests were always employed where the data under investigation did not differ from a normal distribution (Kolmogorov–Smirnov tests: P > 0·05). Variables were loge-transformed where this improved the fit to a normal distribution (e.g. body mass). Where the data were skewed and could not be transformed satisfactorily, non-parametric tests were used. Given the inherent patterns of statistical non-independence in these data (e.g. one hunter will be responsible for several transactions, which in turn can involve multiple market traders), we used average species values across all transactions, rather than each individual transaction, as the unit of analysis. All statistical tests were two-tailed.
testing the overexploitation hypothesis
First, we compared the mean estimates of observed production (average of Methods 1 and 2) and sustainable production (average of PRR and PNMFS) for the terrestrial mammal taxa in the Takoradi catchment (Fig. 1). Across species, actual production is consistently lower than sustainable production (paired-sample t-test: t9 = −5·0, P = 0·001). On average, observed extraction is only 20% of what might be sustainable (median, n = 10 species). In no case does the actual yield exceed the sustainable yield, although in two cases the actual yield falls within the range of estimates for the sustainable yield: the extraction of these species, Atherurus africanus and Cricetomys emini, may therefore be at the limit of what is sustainable. Nevertheless, these results suggest that the current extraction of bushmeat in the Takoradi catchment does not exceed a sustainable harvest, contrary to prediction 1·1.
Secondly, we explored the relationship between taxon body mass and harvesting distance from Takoradi (Fig. 2). Contrary to prediction 1·2, there was no significant positive relationship between these two variables, either across all taxa (Pearson correlation: rp = 0·29, n = 10, P = 0·42) or separately across rodents (rp = −0·12, n = 5, P = 0·84) and ungulates (rp = −0·82, n = 5, P = 0·09). There was therefore no evidence of differential depletion of larger prey in closer proximity to the city. Our result was not confounded by smaller species being more valuable as the carcasses of larger taxa always obtained higher prices (rp = 1·00, n = 10, P < 0·001).
Thirdly, the real price of bushmeat (inflation-controlled) has not increased but rather decreased over the past 37 years. Contrary to prediction 1·3, the Takoradi market price of bushmeat has increased by 6052% between 1963 and 2000, while inflation has increased by 11 627% in the same period. The real price of bushmeat has thus declined by nearly one-half (48%) in the 37-year period preceding this study.
Fourthly, contrary to prediction 1·4, a plot of market prices in 2000 against market prices in 1963 indicates that the price of bushmeat has also failed to increase in comparison to the price of other types of meat sold in Takoradi over this period (Fig. 3). Rather, the price of bushmeat has declined relative to the price of both mutton and beef, while remaining approximately constant relative to the price of fish.
testing the historical depletion hypothesis
These results indicate that the current bushmeat trade may be sustainable. To determine whether this is the result of historical overexploitation, we identified a set of potential bushmeat taxa (rodents, primates and ungulates ≥ 700 g body mass) that should occur in the Takoradi market: each has a previously recorded local distribution, exists in farmbush matrix habitat and could be hunted legally during the study. These taxa are scored for presence/absence in the market and ranked by their intrinsic rate of population increase rmax in Table 1. This analysis indicates that all low-rmax species are absent while all high-rmax species are present. This finding supports prediction 2·1.
Table 1. Local farmbush matrix speciesa listed on the Second Scheduleb of the Wildlife Laws of Ghana (Government of Ghana 1998): primates, ungulates and rodents
Species distributions and habitat preferences in Ghana are taken from Grubb et al. (1998). Species listed in order of declining rmax.
Species that cannot be hunted during the closed season, but can be hunted as adults at all other times. Three further species occur in the market but are absent from the Second Schedule because they are hunted without restriction: Atherurus africanus (brush-tailed porcupine), Cricetomys emini (giant rat) and Thryonomys swinderianus (cane rat, or grasscutter).
An alternative explanation would be that slow reproducers are naturally rare and their absence reflects the rarity with which they enter the market. Because the number of carcasses for each taxon found in the market can be predicted by its body mass (linear regression: r2 = 0·31, F1,8 = 5·0, P = 0·056), we can estimate the number of carcasses that should be present for each absent species on Table 1. In each case, the predicted number is relatively high (median = 15). Even in the case of African buffalo, the rarest species, seven carcasses should have been recorded during the study period. Across these missing species, the number expected in the market was significantly greater than the number observed (i.e. 0) (one-sample t-test: t6 = 3·46, P = 0·013). Thus natural rarity does not appear to explain the absence of these species.
Finally, of the 12 Takoradi hunters interviewed about bushmeat species trends, only those who had been hunting for at least 8 years reported a decline in prey abundance. In support of prediction 2·2, all hunters who had been operating for less than 8 years (3–7 years, median 5 years, n = 4) perceived no change in abundance, whereas all hunters who had been active for a longer period (8–24 years, median 16 years, n = 8) reported a decline.
These analyses suggest that the bushmeat trade in Takoradi is currently in a sustainable phase. This sustainability appears to be the result of historical overexploitation that has eliminated the vulnerable species from the market. Such ‘extinction filters’ have been widely described in island ecosystems, where the current fauna comprise only those species able to survive the anthropogenic impacts associated with island discovery and colonization (Balmford 1996). Although extinction filters have not been linked previously to the bushmeat trade, our findings are corroborated by a substantial literature on hunter gatherers from both archaeological and anthropological sources. These studies report a consistent transition towards smaller-bodied (more robust) prey species in hunter-gatherer harvests over time (e.g. Cannon 2000; Jerozolimski & Peres 2003). Our results also highlight the emerging biodiversity crisis in West Africa, where overhunting has led to the possible extinction of Miss Waldron's red colobus monkey Procolobus badius waldroni (Oates et al. 2000; McGraw & Oates 2002) and widespread local extinctions among large mammals in Ghana's national parks (Brashares, Arcese & Sam 2001).
The pattern of post-depletion sustainability that has emerged from this study is based on a set of indirect measures that cannot alone provide conclusive proof of sustainability. Further research in this area would therefore be valuable, both in Takoradi and elsewhere. Additional information that would help to substantiate our findings would include more detailed data on the abundance and productivity of prey populations in the Takoradi catchment and the scale of rural extraction. In addition, further data on temporal changes to the size of the catchment area and the overall number of hunters in the system (both of which should remain constant under stable conditions), and on the costs and revenues for hunter trips (which should also remain equal when the system is stable), would be informative and relatively easy to collect.
It is also important to consider the generality of these results. Takoradi appears to be a typical Ghanaian urban market (Cowlishaw et al. 2005), but are there any unusual aspects of this system that might limit the application of these results elsewhere? The coastal location of Takoradi may be important in this respect, because an abundance of fish in the market might reduce the consumption of bushmeat and thus be responsible for making the bushmeat trade sustainable (cf. Brashares et al. 2004). However, bushmeat and fish are unlikely to be directly interchangeable in Takoradi, because the former is a luxury good whereas the latter is not. Moreover, the finding that bushmeat consumption in Takoradi is 0·01 kg per capita per day (Cowlishaw et al. 2005), a typical figure for urban dwellers across Central Africa (Chardonnet et al. 1995), suggests that consumer demand for bushmeat in Takoradi is characteristic of other cities in tropical Africa.
In addition, similarities between the bushmeat trade in Takoradi and other localities suggest that the historical depletion of vulnerable species, followed by an ongoing trade in more robust species, may be a common pattern across Ghana. The same five bushmeat species that comprise 67% of the Takoradi market biomass (cane rat Thryonomys swinderianus, brush-tailed porcupine Atherurus africanus, Maxwell's duiker Cephalophus maxwelli, bushbuck Tragelaphus scriptus and black duiker Cephalophus niger) also comprise 70% of market biomass across 15 localities in five different Regions (Ntiamoa-Baidu 1998). Similarly, there is no evidence elsewhere of a scarcity-driven increase in the price of bushmeat, either in absolute or relative terms: the real price of bushmeat has declined in the capital Accra (1975–1993: Tutu, Ntiamoa-Baidu & Asuming-Brempong 1996), and the price of beef and mutton have increased more rapidly than the price of bushmeat in both Accra (1990–1993: Tutu et al. 1996) and Kumasi, Ghana's second largest city (1980–1986: Manu 1987). In addition, the pattern of species depletion in Takoradi matches hunter perceptions of species vulnerability elsewhere in the country: brush-tailed porcupine A. africanus and Maxwell's duiker C. maxwelli are widely thought not to have been affected by hunting, in contrast to yellow-backed duiker Cephalophus sylvicultor, giant hog Hylochoerus meinertzhagene and red-capped mangabey Cercocebus torquatus (Western Region: Holbech 1998), while the largest of the slow reproducers, giant hog H. meinertzhagene and African buffalo Syncerus caffer, are also reported to be locally extinct elsewhere (Ashanti region: Ntiamoa-Baidu 1998). Finally, evidence that local species depletion is not a contemporary phenomenon, but rather occurred some years previously, is also found elsewhere in Ghana. Falconer (1992) documented that, around Kumasi, village elders reported a decline in bushmeat species in their lifetime, but contemporary hunters stated that prey populations were stable and local traders reported that bushmeat supply had not declined over the previous 10 years.
These results emphasize the dynamic nature of bushmeat harvesting systems and the severity of the bushmeat crisis. They also have two important implications for conservation policy. First, if mature urban bushmeat markets are in a sustainable phase, having already lost their vulnerable species, then those markets should not be priority targets for conservation action. Rather, the priority targets should be new markets or those that are supplied from new catchments (e.g. logging camps: Auzel & Wilkie 2000). There are perhaps only two circumstances in which such mature urban markets might justify significant attention: (1) where a monitoring programme is desirable, e.g. because a shift is expected in local socio-economic, demographic or ecological conditions that might precipitate an elevated demand for, or reduced supply in, bushmeat; or (2) where conservation action is required to assist those species that are too heavily depleted to appear in the market but still persist in low-density remnant populations susceptible to extinction (in this study it has not been possible to establish whether the depleted species that are absent from the market are locally extinct or not).
Secondly, and most importantly, if bushmeat markets can be supplied sustainably solely from robust species existing in the farmbush matrix, it is possible in principle to protect vulnerable species and habitats without threatening the livelihoods of those people who depend on the bushmeat trade, many of whom already live in poverty (e.g. de Merode et al. 2004). Thus far, the importance of the farmbush matrix as a source of bushmeat has been largely neglected, with conservation attention focusing on hunting in low-productivity primary forests (e.g. Robinson & Bennett 2000). The farmbush matrix, in contrast, might be much more productive. This is due partly to the abundance of crops (Falconer 1992) and partly to the patches of high-productivity secondary forest (Cowlishaw & Dunbar 2000) in these agricultural mosaics. This productivity, and also the need to control crop raiders, helps to explain how bushmeat extraction in secondary forest can match or exceed that in primary forest (Wilkie 1989) and why Ghanaian hunters prefer hunting in farmbush matrix over other habitats (Falconer 1992; Holbech 1998). Moreover, this pattern does not appear to be unique to tropical Africa. The potential value of agricultural landscapes for wildlife populations, and thus for the production of wild meat, has also been highlighted recently in studies in Costa Rica and Peru (Daily et al. 2003; Naughton-Treves et al. 2003).
The possibility that city bushmeat markets can be supplied solely and sustainably by robust species from an agricultural landscape is an encouraging result. It suggests that the conservation goals of vulnerable species and habitat protection are not always in conflict with human needs. However, our ability to implement management steps that allow for the coexistence of vulnerable species with an active bushmeat trade will be more difficult in practice, due to the limited institutional capacity of those countries where bushmeat is traded (e.g. Smith et al. 2003). In Ghana, wildlife laws already exist to regulate the bushmeat trade, many of which are consistent with the policy recommendations made here, such as the protection of vulnerable species and the permitted hunting of robust species. Nevertheless, both public awareness and state enforcement of these laws is extremely limited (Mendelson et al. 2003). Similarly, Rowcliffe et al. (2004) have demonstrated that the existence of wildlife laws in Democratic Republic of Congo is in itself insufficient to regulate bushmeat hunting without effective enforcement. The development of effective conservation management on the basis of these research findings is therefore likely to involve wider issues of capacity building and good governance (Davies 2002). This remains a significant challenge for the future.
We thank Lars Holbech, Katherine Homewood, Francis Hurst, Catherine MacKenzie, Candy Mends and Paul Symonds for their assistance in this research; and Lars Holbech, E. J. Milner-Gulland, John Robinson, David Wilkie and two anonymous referees for their helpful comments on this paper. We are also grateful to the many actors in the Takoradi bushmeat commodity chain who generously contributed to the study. The project was funded by NERC and ESRC. G. C. is currently in receipt of a NERC Advanced Fellowship. The fieldwork was carried out in affiliation with the Protected Area Development Programme of the Wildlife Department, Ministry of Lands and Forestry, Republic of Ghana. This paper is a contribution to the ZSL Institute of Zoology Bushmeat Research Programme.