| Citation |
Garshelis, D. & Steinmetz, R. 2016. Ursus thibetanus (errata version published in 2017). The IUCN Red List of Threatened Species 2016: e.T22824A114252336. http://dx.doi.org/10.2305/IUCN.UK.2016-3.RLTS.T22824A45034242.en. Downloaded on 08 August 2019. |
Description |
RANGE DESCRIPTION
Fossil remains of the Asiatic Black Bear have been found in various sites in Europe, as far north as the Ural Mountains and Germany and west to France, dating from the early Pliocene to late Pleistocene (Erdbrink 1953, Kosintsev 2007, Baryshnikov and Zakharov 2013, Fourvel et al. 2014); however, in historic times the species has been limited to Asia. The western range limit is in southeastern Iran, inhabited by the so-called Baluchistan bear (U. t. gedrosianus) (Ahmadzadeh et al. 2008, Ghadirian et al. 2012). This small population is likely connected to the Baluchistan bear population in southern Pakistan. Disjunct populations of Asiatic Black Bears also occur in the more mountainous regions of northern Pakistan (Khan et al. 2012) and Afghanistan (Ostrowski et al. 2009). Eastward they continue within a narrow band along the foothills and south side of the Himalayas (up to treeline) across India, Nepal, and Bhutan, and then more widely distributed at lower elevations (generally >70 m but occasionally to 20 m) in the hill states of northeastern India (Sathyakumar and Choudhury 2007). They occur across mainland Southeast Asia, stretching south in Myanmar and Thailand to ~200 km north of the Malaysian border (Kanchanasakha et al. 2010); there are no records of Asiatic Black Bears ever existing in Malaysia. Over half the total range area of this species exists in China, especially in the south-central and southwestern parts of the country. This distribution includes portions of Tibet, from which the specific name, thibetanus, is derived. Smaller, remnant populations occur in eastern China. Another population cluster exists in northeastern China, the southern Russian Far East, and North Korea. A small isolated population exists in southern South Korea. They also live on the southern islands of Japan (Honshu and Shikoku) and on Taiwan and Hainan. Although they have been extirpated from large portions of their range, they remain in all 18 historic range countries.
The distribution of the Asiatic Black Bear roughly coincides with forest distribution in southern and eastern Asia (FAO 2010), except that in central and southern India this species is replaced by the Sloth Bear (Melursus ursinus), in Malaysia it is replaced by the Sun Bear and north and west of the Russian Far East it is replaced by the Brown Bear (Ursus arctos). However, the Asiatic Black Bear overlaps the ranges of each of these species, especially the Sun Bear in a large portion of Southeast Asia and small portions of northeast India. It also greatly overlaps the range of Giant Pandas (Ailuropoda melanoleuca) in south-central China. In Afghanistan, Pakistan, and India, the Asiatic Black Bear range overlaps the Brown Bear (at elevations >3,000 m) in the Himalayas (but curiously, does not appear to overlap Brown Bears in Nepal (A. Aryal, Massey University, New Zealand, pers. comm, 2016; Bista and Aryal 2013). It also overlaps Brown Bears in south-central and northeastern China, North Korea, and the Russian Far East. In India, Asiatic Black Bear range overlaps the Sloth Bear at low elevations (<1,000 m) in some protected areas including Corbett Tiger Reserve and Rajaji National Parks, Uttarakhand (Bargali 2012). However, there is no evidence of overlap with Sloth Bears in neighbouring Nepal (Garshelis et al. 1999). In North Karbi Anglong wildlife sanctuary in Assam, northeast India, Asiatic Black Bear range overlaps both Sloth Bears and Sun Bears—one of the few places in the world where all three of these species coexist, although all are reported to be rare (Choudhury 2011, Choudhury and Chand 2012).
DESCRIPTION
The only rigorous population estimates for this species have been produced in Thailand, based on mark–recapture with camera traps, using distinctive chest patterns (Ngoprasert et al 2010; Higashide et al. 2012, 2013) to identify individual bears. This work yielded density estimates in three study sites in Khao Yai National Park, ranging from 8-29 bears per 100 km² (Ngoprasert et al. 2012, 2013). Bear sign density was found to be correlated with estimates of bear density, so sign surveys conducted across the country may, in the future, yield a reasonably reliable country-wide population index and trend information.
In the Lower Dachigam area of Dachigam National Park and surrounding agricultural landscapes in Kashmir, India, genetic analysis of hair samples revealed the presence of 109 unique individuals in an area of 650 km², or17 bears per 100 km² (Mukesh et al. 2015). Notably, this is a particularly food-rich area.
Others have used a variety of ad hoc procedures to generate rough country-wide estimates. Largest populations occur in China (government estimate: ~28,000; Gong and Harris 2006), Japan (government estimate: 12,000-19,000, with much wider confidence intervals; Ministry of the Environment 2011), India (5,000-7,000; Sathyakumar and Choudhury 2007), and Russia (5,000-7,000; Aramilev 2006). Although these country-wide estimates are large, small, isolated populations also exist. For example, Japan has one large metapopulation as well as a few, disjunct small populations (Horino and Miura 2000, Yamamoto et al. 2012). Northern Pakistan may contain as many as nine separate populations (Abbas et al. 2015). Thailand has a number of small populations inside protected areas, separated by expanses of agriculture (Kanchanasakha et al. 2010).
Countries with the smallest total numbers of Asiatic Black Bears are Iran (~100-200) and South Korea. Iran’s population has a relatively wide but disconnected distribution and occurs at very low density, due to the scarcity of suitable habitat. This population’s low genetic diversity and distinct genetic clade indicate that it has been small and disjunct for a long time (Yusefi 2013). South Korea’s population was nearly extirpated, but saved by a reintroduction program in Jirisan National Park, started in 2004 (Jeong et al. 2011). All of the reintroduced animals were radio-collared, so their fate is known. It is believed that there were about 40 bears in the Republic of Korea population in 2015.
Population trend has been deduced from sign, interviews, changes in habitat, and evidence of poaching. Republic of Korea and Japan report stable to increasing populations, Bhutan and Thailand report stable populations, and India reports stable to declining populations. All other countries report probable declining numbers (no report from Democratic Peoples Republic of Korea). The most severe declines, estimated at >60% in the past 30 years, were reported in Viet Nam and Bangladesh. However, Asiatic Black Bears have fared better than Bangladesh’s other native bear species, which have been extirpated (Sloth Bears) or nearly extirpated (Sun Bears) due to severe habitat deterioration (Islam et al. 2013). Likewise, Sun Bears have disappeared from the only portion of their range in China (Yunnan province), whereas Asiatic Black Bears still remain there (Wen and Wang 2013).
HABITAT AND ECOLOGY
Asiatic Black Bears occupy a variety of forested habitats, both broad-leaved and coniferous, from near sea level to an elevation of 4,300 m (in northeastern India and Sikkim; Sathyakumar and Choudhury 2007, Sathyakumar et al. 2011). They also infrequently use open alpine meadows. A photo-capture was made in the alpine region of Nanda Devi Biosphere Reserve, Uttarakhand, India at 4,500 m (>1,000 m above the mean tree line; S. Sathyakumar, Wildlife Institute of India, pers. comm., 2014). In some areas of Nepal, local people have reported Asiatic Black Bears at higher than normal elevations, possibly a result of climate change (Aryal et al. 2012).
Individual bears move to different habitats and elevations seasonally (Izumiyama and Shiraishi 2004, Hwang et al. 2010), tracking changes in food abundance. In seasonal climates, foods include succulent vegetation (shoots, forbs and leaves) in spring, turning to insects and a variety of tree and shrub-borne fruits in summer, and hard mast (nuts) in autumn (Bromlei 1965, Reid et al. 1991, Hashimoto 2002, Hwang et al. 2002, Huygens et al. 2003, Koike 2010). In the tropics and subtropics, fruits are the mainstay year-round (Steinmetz et al. 2013). The diet may vary year to year with differences in food availability (Koike 2010), and this opportunistic species appears to be able to adapt its diet to gradually changing habitat conditions (Koike et al. 2013). In some places the diet includes a sizeable portion of meat from wild mammalian ungulates (which they either kill or scavenge, including tiger kills; Hwang et al. 2002, Seryodkin et al. 2005, Narita et al. 2006), livestock (Abbas et al. 2015), ants (which they may forage upon for 7-8 hours per day; Yamazaki et al. 2012), or bees (Hashimoto 2002).
Asiatic Black Bears also use regenerating forests, which may have a high production of berries or young bamboo shoots (Takahata et al. 2014). They also feed in plantations, where they may damage trees by stripping the bark and eating cambium (Yamazaki 2003, Yamada and Fujioka 2010), and in cultivated areas, especially corn and oat fields and fruit orchards (Carr et al. 2002, Mizukami et al. 2005, Gong and Harris 2006, Abbas et al. 2015, Mukesh et al. 2015).
In southeastern Iran and Pakistan, this species occupies a very dry, sparsely-forested landscape (often called steppe forest or steppe woodland). This is the driest landscape inhabited by this species. Bears here use riparian areas, abandoned groves of date palm, and wild olive and pistachio forests (Ahmadzadeh et al. 2008, Ghadirian et al. 2012a). In some parts of this region, access to anthropogenic foods, particularly orchards (e.g., date palm, apricots, figs, walnuts), enable them to persist in a habitat with scarce natural food (Ghadirian 2012b). They use rock caves as shelters during daytime, and feed at night, possibly to avoid the sun or encounters with people in this very exposed habitat (Fahimi et al. 2011). By contrast, in more forested parts of the species’ range, they tend to be diurnally active (Hwang and Garshelis 2007) or crepuscular (Sharma et al. 2010).
Asiatic Black Bears feed on a wide array of fleshy fruits and nuts. They were observed to feed on fruits from over 30 species of woody plants within a small, mountainous study site in Japan (Koike 2009). An equal dietary diversity was observed in a mountainous site in northern India (Sharma et al. 2014). In tropical Thailand, dietary diversity was even greater: they were found to feed on fruits from least 30 families of trees (Steinmetz et al. 2013).
Fruiting tree density is a good predictor of the occurrence and relative density of Asiatic Black Bears (Ngoprasert et al. 2011, Steinmetz et al. 2011). In Thailand, these bears were noted to seek out rare fruits in the forest, possibly to diversify their diet (Steinmetz et al. 2013). Asiatic Black Bears are likely an important disperser of the seeds of some fleshy fruits (Sathyakumar and Viswanath 2003; Koike et al. 2008, 2010). They climb trees to eat fruits, and also eat fruits that drop to the ground. Sympatric Sun Bears also eat these same fruits. In Thailand, both species commonly feed on fruits in the cinnamon (Lauraceae) and pea (Fabaceae or Leguminosae) families. Both species live together in lowland habitats (<1,200 m), but Asiatic black bears predominate at higher elevations (Steinmetz et al. 2011).
In temperate forests, Asiatic Black Bears rely heavily on hard mast in autumn, in part to attain sufficient fat reserves for winter denning (hibernation). Therefore, these bears tend to focus their activities in habitats with high abundance of oak acorns, beechnuts, walnuts, chestnuts, hazelnuts, or stone pine seeds (Schaller et al. 1989, Reid et al. 1991, Hwang et al. 2002, Hashimoto et al. 2003, Huygens et al. 2003). When Asiatic Black Bears feed in hard mast trees they often break branches and pile them up in the canopy, forming what appears to be a platform or “nest”. Males may socially exclude females, juvenile bears, and Sun Bars from rich stands of hard mast (Huygens and Hayashi 2001, Hwang et al. 2010, Steinmetz et al. 2011, Koike et al. 2012). When hard mast is poor, these bears expand their home ranges to find alternate fall foods (Hwang et al. 2010, Kozakai et al. 2011, Koike et al. 2012). In Japan, hard mast failures have led to massive intrusions of bears into residential areas, where they seek anthropogenic foods as a substitute (Oka et al. 2004, Oi and Furusawa 2008).
In northern latitudes, where food becomes unavailable in winter, or in high altitudes covered by snow, both sexes hibernate. Therefore, Asiatic Black Bears hibernate throughout their range in Russia, Korea, Japan and northeastern China, and in high elevation areas of more southerly range states. Bears enter dens as early as October and as late as late-December; when hard mast is poor, den entry is earlier (Kozakai et al. 2013). They exit dens as early as mid-March to as late as the end of May (Seryodkin et al. 2003, Koike and Hazumi 2008). They den in rock crevices, hollow trees or stumps, under upturned trees, in dug-out earthen dens, natural cavities under roots, or in ground nests. In a mature natural forest in Japan, they were reported denning inside hollow trees, with entrances above ground level (Hazumi et al. 2001). In Russia, Asiatic Black Bears selected flat river bottoms for denning (Seryodkin et al. 2003), whereas in central China they moved to high elevation rocky outcrops on steep slopes (Reid et al. 1991). Likewise in Japan, dens tended to be in remote, difficult to access mountainous areas (Huygens et al. 2001, Koike and Hazumi 2008). Hunters often have knowledge of the sorts of places and types of dens that the bears tend to use, and have been reported finding and killing bears in dens. Denning and active Asiatic Black Bears are also subject to predation by other Asiatic Black Bears, Brown Bears, and Tigers (Seryodkin et al. 2005).
Asiatic Black Bears do not hibernate from the Himalayan foothills southward where food is available all year and not covered by snow. However, pregnant females den throughout the range (even where food is available and other bears remain active) because they give birth to altricial cubs during the winter (Hwang and Garshelis 2007).
Asiatic Black Bears generally breed during June-July and give birth during November-March; however, timing of reproduction is not known for all portions of the range. Age of first reproduction is typically 4-5 years old, and they normally produce litters of 1 or 2 cubs every other year (at most) (Yamanaka et al. 2011). Maximum lifespan is over 30 years, but average lifespan is less in the wild. No data are available on survival rates or causes of mortality.
THREATS
Habitat loss due to logging, expansion of agriculture and plantations, roadway networks and dams, combined with hunting for skins, paws and especially gall bladders are the main threats to this species.
Habitat loss due to logging and conversion to agriculture is a major threat to bears in 9 of 18 range countries, including Afghanistan, Bangladesh, Cambodia, India, Lao PDR, Myanmar, Nepal, Pakistan, and Russia. Of these countries, Myanmar stands out as it encompasses a relatively large portion (about 12%) of the global range, so habitat loss here will have a large impact on the global status of the species.
Habitat loss and degradation is most severe in the southern portion of the range. In India, <10% of the species’ range is within protected areas (PAs), and areas outside PAs are subject to development projects and extraction of wood for fuel and livestock fodder (Sathyakumar 2006, Sathyakumar et al. 2012). In Bangladesh, where forest cover is now <7% of the land area, Asiatic Black Bears survive only in small remnant patches in the east, generally near the Myanmar and Indian borders. Cambodia and Myanmar, although still well forested (57% and 48%, respectively), are third and fourth in the world in the annual rate of loss of forested area (among countries occupied by black bears; North Korea and Pakistan have higher rates: FAO 2010). Thailand has lower forest cover (<30%), but most of its remaining forests are within PAs, and three-fourths of these are occupied by black bears (Kanchanasakha et al. 2010). Forest area has recently been increasing in Viet Nam, but much of the present remaining natural forest is highly degraded from both legal and illegal lumbering (Nguyen Xuan Dang 2006, FAO 2010).
Forest area is increasing rapidly in China, which is now first in the world in terms of area gained per year. This increasing forest area stems from mandated government programs aimed mainly toward reducing flooding and erosion; the replanted trees may or may not be particularly suitable for bears. However, good forest habitat does persist in northeastern China, Taiwan, Korea, Russia, and Japan. In Japan, Asiatic Black Bear range has expanded with increasing forest area and diminishing rural human populations (Oi and Yamazaki 2006, Takahata et al. 2014).
Direct killing represents another potential threat. Local people in Nepal, for example, use meat, organs, and bile for food, for healing, to alleviate pain, or even to deter ghosts (Yadav et al. 2012). The sustainability of this practice has never been measured.
Commercial poaching is believed to be a major threat to bears in at least half the range countries, including China, Taiwan, Russia, India, and the five range countries in Southeast Asia (Cambodia, Lao PDR, Myanmar, Thailand, and Viet Nam). Commercial poaching is likely to have especially severe synergistic effects where it coincides with logging and habitat encroachment (Cambodia, Myanmar, Lao, Russia), which increase access for poachers and concentrate bears in ever smaller and fragmented pockets.
Poachers primarily seek gall bladders and paws. The market for bear paws appears to be increasing commensurate with an increasing number of wealthy people who find it within their means to indulge in this very expensive delicacy (Burgess et al. 2014). In 2013, two seizures in two months recovered parts of 81 individual Asiatic Black Bears being smuggled (from Russia) into or within China (Servheen 2013). Such seizures probably represent a small portion of the actual trade in bear paws.
The demand for these bear products has fuelled a growing network of international trade throughout Southeast Asia, fed through poaching of wild bears (Shepherd and Nijman 2008, Foley et al. 2011). Though difficult to quantify because of its covert nature, bear poaching for the wildlife trade is clearly widespread. Hunting of bears for the wildlife trade was reported in over 70% (10 of 14) of the villages interviewed in a northeastern Lao National Protected Area, facilitated by wildlife dealers from Viet Nam who comb rural areas to place orders, supply weapons and ammunition, and transport the off-take (Scotson 2010). In parts of Nepal and Myanmar near the Chinese border, Asiatic Black Bears are among the top preferred species that are hunted commercially for sale of their body parts to China (Rao et al. 2005, Yadav 2012).
It is difficult to evaluate the actual impacts of this trade on wild populations because reliable population estimates and assessments of trends of wild bear populations are unavailable (Garshelis 2002, 2006). Nevertheless, there is evidence that this commercially-driven trade in bear parts is unsustainable and therefore causing populations to decline. In Sichuan Province, which likely has the highest numbers of Asiatic Black Bears in China, bear abundance and distribution have been in steady decline according to opinions of local people, driven predominantly by illegal poaching for bear parts (Liu et al. 2009, 2011). Demand for wild bears drives poaching throughout Southeast Asia. Local inhabitants of a large Thai reserve perceived a 50% decline in Asiatic Black Bear abundance over the previous 20 years, due almost solely to commercial poaching (Steinmetz et al. 2006). Across four countries in Southeast Asia (Myanmar, Thailand, Cambodia, Lao PDR), relative abundance of Asiatic Black Bears determined from camera trapping was half that of Sun Bears (with which they are widely sympatric) on average, indicating much lower densities (Steinmetz 2011). Notably, areas with the most depleted black bear populations are also closer to the Chinese border (Steinmetz 2011). The Asiatic Black Bear is the most highly valued species for consumers of bear parts, and seizures of bear parts in Southeast Asia are almost exclusively from Asiatic Black Bears (Shepherd and Nijman 2008).
Use of cable snares to capture wildlife is widespread in Southeast Asia, and is particularly intensive in Viet Nam, Lao PDR, Myanmar, and Taiwan. In Lao PDR and Vietnam, 1,000s to 10,000s of snares have been removed from single protected areas within a few months (Scotson and Brocklehurst 2013; B. Robichaud, pers. comm). Intensive snaring poses a serious threat to bears because they are often caught inadvertently even if snares are targeted at other species. For example, 8 of 15 bears observed by researchers in Taiwan had missing toes or paws from snares set for wild pigs and other wildlife (Hwang et al. 2010). The recent discovery in Lao PDR of commercially-motivated snare lines designed specifically for bears (Scotson and Hunt 2012) points to increasing pressure on wild populations driven by international demand from the trade in bear parts.
The capture of live bears presents yet another threat to this species. In several Southeast Asian countries Asiatic Black Bears are routinely confiscated from people attempting to raise them as pets, possibly originating from the poaching of their mother for the gall bladder and paws. In Pakistan, several thousand bears were taken from the wild for exhibitions (referred to as bear baiting) in which individual bears (with canines and claws removed) fight with dogs. This practice was made illegal in 2001, but continues to some extent surreptitiously. More commonly now, bears are used for “dancing”.
Conflict between humans and black bears is a widespread problem that occurs across the range of the species: Iran (Ghadirian et al. 2012), Pakistan (Dar et al. 2009, Khan et al. 2012), India (Charoo et al. 2011, Choudhury 2013), Nepal (Stubblefield and Shrestha 2007), Bhutan (Sangay and Vernes 2008; Sonam Wangchuk, Wildlife Conservation Division, Bhutan, pers. comm., 2014), Lao PDR (Scotson et al. 2014), Thailand (Ngoprasert et al. 2011), China (Liu et al. 2011), and Japan (Kishimoto 2009). Human-bear conflicts have negative effects on both people and bears?they include losses to livestock, crops, and apiaries; human injuries and deaths; and killing of bears in defence or retribution. Crop raiding is often extensive especially where farm fields and orchards are close to forest, as in India, Bhutan, and Lao PDR, and large numbers of farmers have reported crop damage from black bears (Charoo et al. 2011, Rinzin et al. 2009, Scotson et al. 2014). Killing of Asiatic Black Bears in crop fields may be motivated more by the value of their parts than the protection of the crops (Liu et al. 2011, Yadav et al. 2012, Scotson et al. 2014).
Human injuries and deaths sometimes occur at tragically high rates, especially in densely settled areas with abundant bears: in Sikkim state, India, at least 25 people were killed or injured by black bears during 2008-2013 (Choudhury 2013). In Honshu, Japan, about 80 people are injured annually, including a few deaths (Japan Bear Network 2011). Such conflicts commonly give rise to negative attitudes towards black bears (Wang et al. 2006, Charoo et al. 2011, Liu et al. 2011, Sakurai and Jacobson 2011). Human-bear conflict can be a source of significant bear mortality when local people or government authorities kill bears to reduce damage; in an extreme case, over 4,000 bears were killed as pests in Japan in 2006 (Kishimoto 2009), representing a significant portion of the country’s population of black bears.
Protected areas are often too small to contain resident black bears. Black bears have large home ranges (>100 km² in some areas; Hwang et al. 2010, Chongsomchai 2013), and their wide-ranging movements may bring them near or beyond the borders of protected areas, where they are more vulnerable to poaching and more likely to come into conflict with farmers (Hwang et al. 2010, Ngoprasert et al. 2011, Choudhury 2013). The availability of natural foods also has a strong effect on whether bears roam outside protected areas. Periods of mast failure or low fruit availability can induce bears to seek human foods outside the forest (Kishimoto 2009, Ngoprasert et al. 2011). Black bear raiding of corn fields adjacent to a park in Thailand coincided with annual periods of low fruit availability, and resulted in some being killed by farmers (Ngoprasert et al. 2011).
USE AND TRADE
Bear bile has been an important component of traditional medicine in Eastern Asia for millennia. The first written account of such use was recorded in the first pharmacopeia of China in 659 A.D. (Feng et al. 2009). Bear bile is used in Traditional Chinese Medicine (TCM) for reducing fever and inflammation, detoxifying the liver, arresting convulsions, improving eyesight, and dissolving gall stones.
The medicinally active ingredient of bear bile is ursodeoxycholic acid (UDCA). In controlled, clinical trials, this compound (and its associated conjugates) has been shown to have many of the medicinal properties claimed in TCM. In western societies it has been approved as a drug to treat certain liver diseases. More recently it has been shown to prevent retinal degeneration (Boatright et al. 2006), protect against Type I diabetes (Engin et al. 2013), and to have therapeutic effects for a number of neurodegenerative diseases, including amyotrophic lateral sclerosis, Alzheimer’s disease, Parkinson’s disease, and Huntington’s disease (Vang et al. 2014).
Despite the availability of synthesized UDCA, and the widespread use of this product in China, Japan, and South Korea, as well as a large number of herbal alternatives to bear bile in the Chinese pharmacopeia, many TCM practitioners (in Asia and elsewhere) prefer using bear bile because it is thought to be more effective and more natural than synthetic products. Some have claimed that other compounds in bear bile aid in the medicinal functions of UDCA, although no clinical trials have compared the efficacy of bear bile versus synthesized UDCA. Evidence suggests that UDCA in bear bile, which is a potent inhibitor of cell death (apoptosis), may serve to protect bears during hibernation (Solá et al. 2006). Asiatic Black Bears produce high levels of UDCA. Historically, the Asiatic Black Bear has been sought after for its bile more than any other bear species. It is sold in various forms including whole gallbladders, raw bile, pills, powders, flakes, and ointments (Foley et al. 2011).
Increased demand, stemming from burgeoning human populations, coupled with more effective hunting of wild bears and increased ability to sell and transport products has led to the over-exploitation of many Asiatic Black Bear populations, especially in China and Southeast Asia. This trade was further fuelled by an increasing use of bear paws in a high-priced soup (having nothing to do with TCM; Burgess et al. 2014).
In the late 1970s, a technique was developed to extract bile from captive bears without killing them. This practice, called bear bile farming (or bear farming), started in Korea, then spread to China in 1984, and later to Vietnam, and most recently to Lao PDR and Myanmar. In Republic of Korea, Viet Nam, Lao PDR, and Myanmar, bear bile farming is technically illegal. In China, farmed bile is legally sold for use in TCM; it has been argued that this helps to satisfy public demand for bile while reducing the illegal hunting and sale of gall bladders of wild bears.
Bear farming, though, has raised a number of concerns, both about animal welfare and conservation. In Viet Nam, many small-scale bile farms were stocked by thousands of cubs removed from the wild. The condition in which these bears were kept precluded successful breeding and cub rearing. Moreover, although this practice has been illegal since 1992, with regulations strengthened in 2002, the number of wild-caught farmed bears in Viet Nam continually increased, peaking at over 4,000, but declined by over 70% since the mid-2000s with stricter governmental efforts. In part as a result of the decline in farms in Viet Nam, farming operations sprang up in Lao PDR in 2000, and have been increasing rapidly (~120 bears by 2012; Livingstone and Shepherd 2016). Notably, prices of wild bear bile in Lao increased 180x from pre-farming years (1990s) to 2012, suggesting that the availability of farmed bile was actually enticing more users into the market, some of whom were willing to pay considerably more for the wild product (Livingstone and Shepherd 2016). As there is no evidence of breeding of farmed bears anywhere in Southeast Asia, the farms have a negative impact on wild populations both by taking cubs from the wild and by promoting more demand for products of wild bears.
In China, it appears that many bear farms now have active breeding programs. It is unknown what degree of restocking occurs with cubs from the wild, but cases are still reported of people being apprehended with over 20 bear cubs apparently destined for a Chinese bear farm. The biggest unknown issue in China is whether the legal supply of bile has indeed reduced the demand for (illegal) wild bile, and hence reduced poaching of bears, or, as in Lao PDR, has stimulated more demand.
For farming to reduce pressure on wild populations consumers must regard wild and farmed products as similar in quality or value—they must be substitutable (Phelps et al. 2013). In the case of bear bile, interview surveys in both China and Viet Nam show that many consumers regard wild bile as superior to farmed bile, are willing to pay higher prices for it, and are disinclined to use substitutes (Drury et al. 2008, Dutton et al. 2011). Further, the widespread availability of farmed bile may draw more people into the market via more prescriptions by TCM doctors and more promotions by the farming industry; some of these new users may be prompted to use wild bile for some illnesses, considering it to be more potent and pure. Farmed bile has also been promoted (via advertising) for non-medicinal uses, such as general well-being (administered in lotions, shampoos, cosmetics, tonics, etc.) or hangovers, although such uses are technically illegal. Because farmed bile is cheap, some people also use it recreationally (Vu 2010).
Currently there are 67 registered bear farms in China, with an undisclosed number of bears (recently estimated at >17,000). With China’s consent, the World Conservation Congress in 2012 approved an IUCN Recommendation (WCC-2012-Rec-139-EN Bear farming in Asia, with particular reference to the conservation of wild populations) that (1) halts any increase in the number of farms and bears on farms, and (2) calls for a scientifically independent and peer-reviewed “situation analysis” to assess whether bear farming has a positive or negative effect (or no effect) on wild bear populations (Garshelis and Scotson 2012). Meanwhile, the recommendation adopted by the IUCN calls for the other four bear farming countries to phase out bear farming.
All international trade in bile from Asiatic Black Bears, including farmed bile, is illegal under CITES. Despite this, a wide and complex network of international bear bile trading exists, which includes China, Republic of Korea, Japan, Taiwan, and all of the Asiatic black bear range countries in Southeast Asia (Foley et al. 2011).
CONSERVATION ACTIONS
Protection of forested habitats would be an important conservation measure for this species. China, Thailand, and Viet Nam have imposed various sorts of logging bans, but with varying effects (Durst et al. 2001). In some cases this has resulted in trees being obtained (often illegally) from neighbouring countries, or in creating plantations which do not provide food resources for bears. However, in 2010 Russia banned the felling of Korean Pine, a key bear food source (I. Seryodkin, Russian Academy of Sciences, pers. comm. 2014).
In most countries, closer government collaboration with local people and communities would help conserve natural forests. Community-managed forests have become increasingly important tools for conservation of large mammals in Asia: they can provide breeding habitat, create corridors that link isolated state managed protected areas, facilitate dispersal, and increase habitat that supports an overall larger number of animals. Community-managed forests in Nepal, for example, have created habitat that now holds breeding populations of one-horned rhinos and tigers (Gurung et al. 2008). These community-managed forests may provide similarly valuable habitat for bears, although this remains unexplored. The value of community-managed forests for bears should be investigated further, and expansion of community forest networks promoted where appropriate. The use of existing community-managed forests by bears should also be monitored, such as through sign surveys or camera trapping.
The most beneficial conservation measure for Asiatic Black Bears would be to substantially lessen the demand for bear products, and thus reduce hunting and trade. Large scale, multifaceted, long-term campaigns are urgently required to change social norms regarding consumption of bear parts. Such campaigns are especially important in the main consumer nations of China and Viet Nam, and have indeed been initiated in both those countries by NGOs. As a result, public awareness of bear conservation issues has expanded greatly in the past 10 years in China and Vietnam and attitudes and behaviour are starting to change. In 2014 business leaders in China pledged to cease the practice of corporate gifting of wildlife products and committed to setting new trends for others to follow (TRAFFIC 2014). As of 2011 over 100,000 Vietnamese citizens had pledged to not consume bear bile, and a number of bear farms had closed due mainly to intensified social pressure (ENV 2012).
The governments of Republic of Korea and Viet Nam are attempting to phase out bear bile farming. Conversely, bear bile farming is on the rise in Lao PDR and Myanmar, stocked by cubs from the wild. Authorities must recognize that these farms have become established due to loopholes in the laws and misreporting of the activities (Livingstone and Shepherd 2016). Meanwhile, China contends that farming must continue or poaching would increase; whereas no solid data are available to evaluate that view, it remains clear that poaching is still the primary threat, and that this will lessen only if the public demand for bile is reduced. Promotion of synthesized UDCA (or a product that includes UDCA as well as other bile compounds) is likely to be the best course for long-term reduction in demand for bear bile.
The Asiatic Black Bear is protected under both international and national laws, but often these laws are not enforced. The Asiatic Black Bear has been included on CITES Appendix I since 1979. In most range countries Asiatic Black Bears are listed as a protected species. For example, they are protected under Class 2 of Chinas Wildlife Protection Law (a limited number of permits are issued to kill nuisance animals), and under Schedule I of the Indian Wild Life (Protection) Act. In South Korea they are designated as a national monument (No. 329) within the Cultural Properties Protection Law and also as an Endangered Wild Animal. In Japan, this species is listed under the Law for Conservation of Endangered Species of Wild Fauna and Flora, which for trade requires certification of legal take; however, gall bladders and paws are exempted. Throughout Southeast Asia this species is totally protected in every range country, with the exception of Myanmar, where it is classified as “normally protected”, meaning that it may be killed with a special license, although such licenses are rarely issued (Saw Htun, Wildlife Conservation Society, Myanmar, pers. comm., 2014). The Baluchistan subspecies was listed as Critically Endangered in the 1996 IUCN Red List, and is nationally listed as critically endangered in Pakistan and Iran.
Despite these measures of protection, unabated poaching puts this species at risk across most of its range. Much greater efforts are needed to reduce poaching and snaring pressure, especially where populations are small or fragmented. This can be partly accomplished through anti-poaching patrols, which are typically conducted by trained reserve staff but may also include village volunteers. Such patrols are important for finding and physically removing snares, particularly in Lao PDR and Viet Nam where snaring pressure is intense (Scotson and Brocklehurst 2013). Another indispensable function of patrols is to assess and monitor hunting pressure, in order to gauge the effectiveness of conservation interventions. However, the ability of patrols to actually deter poaching and snaring from happening in the first place, while widely assumed, is little supported by empirical data (Steinmetz et al. 2014). Patrolling is particularly difficult in the remote, mountainous environments that serve as key, remnant habitats for this species. Additional approaches are required.
Poaching could also be reduced by building conservation and co-management partnerships with local communities around parks and reserves, and by supporting community-based conservation in the absence of designated protected areas. Such approaches seek to build trust, raise awareness, provide motivation, support local resource management institutions, use local (as well as outside) knowledge, offer opportunities for action, boost confidence to act, and generate social pressure against poaching—behavioural change is promoted when these conditions converge. Such approaches have been demonstrated to reduce overall poaching pressure (Steinmetz et al. 2014) and increase tolerance for crop-raiding bears (Khan et al. 2012), improving the status of local bear populations (Khan et al. 2012; Steinmetz, unpublished data). Much more effort should be devoted to enhancing the involvement of local communities as conservation partners in protected areas where bears occur.
Sport hunting of Asiatic Black Bears is legal only in Japan and Russia. Russia reports a legal harvest of 75-100 bears/year and an estimated illegal take of about 500 bears/year. Sport harvests of black bears in Japan average about 500/year and have been slowly declining since the late 1980s due to diminishing interest in hunting (Oi and Yamazaki 2006). However, a higher number (generally 1,000-2,000) of nuisance black bears are killed annually (using mainly traps) in towns or agricultural areas of Japan.
The Republic of Korea has been partially successful in restoring their wild bear population through restocking, initially with captive-born bears, and later with orphaned wild bears from Russia (which are genetically similar to native Korean bears; Kim et al. 2011). These bears are now reproducing. Some Southeast Asian countries, like Cambodia and Thailand are also considering reintroducing bears from captivity, in part to augment wild populations, and in part to relieve pressures on captive facilities which cannot sustain the steady influx of confiscated animals.
Throughout much of the southern portion of the range of this species, efforts to reduce habitat degradation outside PAs and to increase the number and/or area of PAs would be highly beneficial. An increasing number of PAs are being established in China, Russia, India, and a few other countries within the range of Asiatic black bears (Chape et al. 2008, Sathyakumar et al. 2012), mainly to protect other species, but serving as well to increase protection for black bears. Additionally, the recently amended (2003) Indian Wild Life (Protection) Act provides options for new categories of PAs that could be established to form travel corridors between existing PAs.
More attention is needed on the issue of human-bear conflicts. A few recent studies have examined mitigation methods to reduce conflicts (Charoo et al. 2011, Scotson et al. 2014) and to create risk models to predict where conflicts are most likely to occur (Honda et al. 2009, Takahata et al. 2014). However, solutions to this issue remain largely elusive, because conflicts are in part due to varying food conditions in the wild, and in part to nearby availability of human food sources. Participation of local people in designing solutions and monitoring outcomes will be critical to success. Projects will probably need to be implemented for a number of years to produce useful outcomes; long-term funding will thus be needed to advance the science of human–bear conflict resolution. Where sufficient human resources and funding exist, conflict rapid-response teams (staffed by NGOs, government personnel, and local people) can be effective. Safety guidelines can be devised and disseminated, based on knowledge of bear behaviour and ecology in an area, to help people avoid bears or minimize chances of being attacked. Better methods to deter bears are also needed. Additionally, educational efforts to increase awareness and promote greater tolerance of some level of crop-raiding and livestock loss are likely to be key to success (Khan et al. 2012).
Aside from the global Status Survey and Conservation Action Plan for this species, published by the IUCN in 1999 (Servheen et al. 1999), which is incomplete and now quite outdated, two range countries, Taiwan and India, have developed National Conservation Action Plans for bears (Hwang et al. 2012, Sathyakumar et al. 2012). These are the first such conservation action plans for bears in Asia (although other countries have developed less extensive plans that aid conservation, as for example to deal with human-bear conflicts). Both of these country action plans were developed through a wide range of consultations with stakeholders. The Taiwanese plan concerns only Asiatic Black Bears, since this is the only bear species in the country (and also an endemic subspecies, U. t. formosanus). The Indian plan covers all four species present in that country: Asiatic Black Bears, Sloth Bears, Brown Bears, and Sun Bears. Both plans emphasize the same key themes: reduction in illegal hunting, mitigation of human–bear conflicts, increased habitat management, enhanced research and information gathering, capacity building, and improved communication and education. Both plans stress actions aimed against illegal hunting, including intelligence gathering, incentivized enforcement, community awareness, restriction of guns, and creation of a database to monitor trade. However, the root of the hunting issue is different in the two countries: in India hunters are motivated to sell bear parts, some of which are traded internationally, whereas in Taiwan bears are taken mainly opportunistically or as by-catch from illegal snaring for ungulates, and the parts (including meat) are used or sold locally (Hwang 2003). Other priority actions in the action plans include: strengthening methods of crop and livestock protection; reducing bear-caused human injuries (India), creating rapid response teams to investigate conflicts with bears (and possible claims for monetary compensation); enhancing human tolerance toward bears; identifying critical habitats and corridors used by bears, especially those outside PAs; increasing habitat protection and restoring degraded habitats outside PAs; reducing dependency of local communities on resources needed by bears; discouraging shifting agriculture; developing methods to track population size and trends; involving communities in bear monitoring programs; equipping forest and wildlife staff with adequate knowledge and modern equipment to manage all types of human-bear interactions; and developing an advocacy program for bear conservation through active communication to the public and pressure on corporations and politicians. |
