- Cervus alces Linnaeus, 1758
Moose, Elk, Eurasian Elk, Eurasian Moose, European Elk, Siberian Elk
Value of species
Game (hunting) species
Grubb (in Wilson and Reeder 2005) recognized Eurasian Elk (Alces alces) and Moose (Alces americanus) as distinct species, citing sources that documented differences in karyotype, body dimensions and proportions, form of premaxilla, coloration, and structure and dimensions of antlers (Geist 1998, Boyeskorov 1999). There is still some debate surrounding whether Moose comprise one or two species. Groves and Grubb (1987) called them "semi-species". Boeskorov (1997) proposed that the chromosomal races of Alces alces were different species, however, Bowyer et al. (2000) cautioned that chromosome numbers might be a poor designator of species among large mammals. Based on cited sources that documented differences between Eurasian Elk and Moose, Geist (1998) recommended separation at the subspecies level (i.e., Alces alces alces Linneaus, 1758 and Alces alces americanus Clinton, 1822). Geist (1998) noted a broad zone of hybridization between the two forms in central and eastern Siberia but Boeskorov (2003) identifies the Yenisei River as the boundary between the ranges of the two putative species. There are no differences in antler morphology on either side of the river, however, suggesting substantial gene flow (Kolesnikov and Kozlovskii 2014). Moreover, genetic analyses have generally supported distinguishing the two at the subspecific level (Hundertmark et al. 2002 a,b; Udina et al. 2002; Hundertmark and Bowyer 2004); further research is needed before a consensus would support species-level classification.
Hundertmark et al. (2002b) report that analysis of mtDNA revealed three haplogroups, one entirely Asian, one primarily European and one primarily North American. Eight extant subspecies are recognised here: A. a. alces - Scandinavia, Finland, Baltic states and Poland E to Yenisei River, A. a. americana - E Canada (C Ontario to Newfoundland), A. a. andersoni - British Colombia to Minnesota and Ontario, A. a. buturlini - NE Siberia and Kamchatka, A. a. cameloides - N Mongolia, Ussuriland, N Manchuria, A. a. gigas - Alaska and Yukon, A. a. pfizenmayeri - C Siberia and Stanovoy and Cherskiy Mountains, A. a. shirasi - S Alberta to Wyoming and Utah.
In North America it occurs in Alaska and Canada south through the Rocky Mountains, northern Great Lakes, and New England. The species is estimated to have arrived in North America from Asia about 11,000-14,000 years ago, shortly before flooding of the Bering land bridge (Hundertmark et al. 2002b). The species' range has decreased over the past 100 years in the southern boreal forest regions in the eastern provinces of Canada (e.g., Beazley et al. 2006), but has expanded in other areas. In recent decades, it has expanded its range westward into the coastal temperate rainforests of British Columbia and some coastal islands (Darimont et al. 2005). These changes have been due to habitat changes caused by humans in boreal and rainforest ecosystems.
The species has a range in north Eurasia from Scandinavia, Poland, N Austria, and S Czech Republic (vagrant in Croatia, Hungary, and Romania), east to the Yenisei River (Siberia) and south to Ukraine, N Kazakhstan, N China (N Sinkiang), and possibly northern Mongolia (Wilson and Reeder 2005). It has been extinct in the Caucasus region since the 19th century (Wilson and Reeder 2005) and was introduced but now extirpated in New Zealand (Nugent et al. 2001, Boyeskorov 1999, Grubb in Wilson and Reeder 2005).
In Europe, it has a continuous distribution extending through Norway, Sweden, Finland, Russia, the Baltic states, Belarus, Poland and northern Ukraine. There was a small population in northern Austria, which is now extinct (although there may still be occasional migrants). There are three isolated subpopulations in southern Czech Republic, and the species is occasionally recorded in Germany, Croatia, Hungary and Romania. It has been extending its range southwards along the rivers into the northern Caucasus lowlands. It ranges from sea level up to at least 1,500 m in Europe (H. Henttonen pers. comm. 2006), and up to 2,500 m in the Altai mountains of Central Asia (Nygrén 1986). The species occurs east to the Yenisei River (Siberia) and south to Ukraine, N Kazakhstan, N China (N Sinkiang and the Altai region of Xinjiang, and possibly adjacent parts of Mongolia.
It is a widespread and abundant species in most range states with the exception of China where it is extremely rare and of limited distribution (Piao et al. 1995, Sheng and Ohtaishi 1993, Yu et al. 1993). It is common in Siberia and North America. In 2003, the species was declared endangered in Nova Scotia, Canada (Beazley et al. 2006). Populations have expanded and increased in western British Colombia since the mid-1900s (Darimont et al. 2005).
In the last 50 years European populations have increased dramatically, there are at least 440,000 individuals of which about 50% are harvested annually (Wilson and Mittermeier 2011). Numbers have increased markedly in Scandinavia in recent decades, and the range is expanding in the Caucasus. European populations show fluctuations over a multi-year cycle (Bauer and Nygrén 1999). Population estimates for European countries include the following: Czech Republic - maximum of 50 animals, Estonia - 10,000 individuals, Finland - at least 110,000 individuals (60-80,000 shot annually), Poland - 2,800 individuals, Sweden - 340,000 individuals (Pielowski and Jaworski 2005, Ruusila and Kojola 2010).
In North America the estimated population is approximately one million, with an annual harvest in the late 1990s of about 85,000 individuals (Wilson and Mittermeier 2011).
Alces alces is found in a range of woodland habitats, both coniferous and broadleaved, from the tundra and taiga southwards through boreal to temperate zones. This species prefers a mosaic of second-growth boreal forest, openings, swamps, lakes and wetlands. It thrives in secondary growth, and its population expansion in Scandinavia has been linked to the replacement of natural taiga forest by secondary woodland after logging (Bauer and Nygrén 1999). It is also found in open country in the lowlands and mountains, including farmland, if there is forest nearby. The species avoids hot summer conditions by utilizing dense shade or bodies of water. Mineral licks may be an important sodium source, used in early summer in Canada and Alaska. It feeds on vegetative parts of various broadleaf trees, preferring birch, ashes and willow in the spring and summer and the twigs of these species as well as of fir, alpine, and juniper in the autumn and winter. It also eats shrubs, such as blueberry and heather, dwarf shrubs, herbs, and aquatic plants and can be a pest of agriculture and forestry in at least parts of its range (Ma Yiqing pers. comm; (Ruusila and Kojola 2010). In the search for food, some populations migrate during the year - covering up to 180 km (110 miles) in North America and 300 km (180 miles) in Asia. Whether sedentary or migratory, A. alces utilize specific home ranges, varying in size from 3.6 to >259 km² (Hundertmark 1997).
The species may become sexually mature after one year (Schwartz 1997), and the maximum life span is about 20 years (Peterson 1977). It is active throughout the day and night, although there are peaks at dawn and especially at dusk.
Populations can be limited or regulated by complex interactions of ecological factors that vary from population to population, or ecosystem to ecosystem. Winter weather (snow accumulation) may strongly affect populations, even more so than wolf density (Mech et al. 1987); however, Messier (1991) found that competition for food, but not wolf predation and snow, had a regulatory impact on the species. Van Ballenberghie and Ballard (1994) found that in some naturally regulated ecosystems, predation by bears and wolves often is limiting and may be regulating under certain conditions. Messier (1994) developed population models of A. alces-wolf interactions. Under favourable conditions, populations are capable of large annual increases (20-25%) in population size; large populations may degrade habitat. In the presence of relatively few predators, Albright and Keith (1987) documented high calf-survival despite poor physical conditions during winter in a study of population dynamics of introduced populations in Newfoundland.
See Nudds (1990) for discussion of relation between the species, white-tailed deer, and meningeal (brain) worms. Brain worm (Parelaphostrongylus tenuis) may limit populations in areas where white-tailed deer are common (Nudds 1990). Deer are not negatively impacted by the brain worm, the larval stage of which is passed in deer faeces. Snails, often inadvertently ingested by A. alces feeding on vegetation, are the intermediate host for the worm. Deer, through worm-mediated impacts, commonly are believed to exclude A. alces and caribou from areas where deer occur; however, an analysis by Schmitz and Nudds (1994) concluded that the species may be able to coexist with deer, albeit at lower densities, even in the absence of habitat refuges from the disease. Whitlaw and Lankester (1994) found that the evidence that brain worm has caused A. alces declines is weak. The species is also severely impacted by another parasite of white-tailed deer, the winter tick (Samuel et al. 2000).
Where there is no human, wolf or bear predation, A. alces may alter the structure and dynamics of boreal forest ecosystems. At Isle Royale, Michigan, browsing by the species prevented saplings of preferred species from growing into the tree canopy, resulting in a forest with fewer canopy trees and a well-developed understory of shrubs and herbs; also, browsing may have caused an increase in spruce and a decrease in balsam fir (McInnes et al. 1992).
Ferguson et al. (2000) reviewed the dynamics of 15 Canadian A. alces populations, and reconciled some major factors that influence various populations differently. Populations that live in greater forest cover (i.e., greater primary productivity) and with greater natural predation had more predictable population trends from year to year. Populations living in areas with low primary productivity and with low natural predation experienced more density-independent population change with lower predictability in population size.
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