Justification of Red List Category
This species has been listed as Vulnerable because of an apparent severe decline detected in the wintering population in the Baltic Sea between the early 1990s and late 2000s. This rate of decline implies that the global population will undergo at least a ≥30 decline over three generations (1993-2020), even when factoring-in uncertainty regarding the sizes and trends of other populations. Improved knowledge regarding populations outside the Baltic Sea might lead to the species being uplisted to Endangered if the overall rate of decline can be confidently shown to be very rapid.
The global population is estimated to number 3,200,000 to 3,750,000 individuals (Wetlands International 2017).
Surveys of the wintering population in the Baltic sea indicate that the species has undergone a precipitous decline there, from c.4,272,000 individuals in 1992-1993 to c.1,486,000 individuals in 2007-2009 (Skov et al. 2011). There is considerable uncertainty over the trends of smaller populations in Europe outside the Baltic sea, in Greenland and Iceland and East Siberia and North America, rendering the estimation of its global trend very difficult. The European wintering population is estimated to be declining by 30-49% (BirdLife International 2015). However, the overall rate of decline is likely to approach 50% over three generations (27 years), from 1993 until 2020. The population size in winter is estimated to be decreasing by 30-49% in 27 years (three generations).
This species has a circumpolar range, breeding on the Arctic coasts of North America (Canada, Alaska, U.S.A. and Greenland), Europe (Iceland and Norway), and Asia (Russia). It winters at sea further south, as far as the United Kingdom, South Carolina and Washington in the United States, Korea on the Asian Pacific coast, and other areas including the Black Sea and Caspian Sea (del Hoyo et al. 1992). According to monitoring data from the Baltic Sea, where the western Siberian and northern European populations winter, the population there has declined significantly since the early 1990s at least (Hario et al. 2009, Ellermaa et al. 2010, Nilsson and Månsson 2010, Skov et al. 2011). An estimated total of c.4,272,000 individuals was counted there in 1992-1993, falling to c.1,486,000 individuals in 2007-2009, which suggests that a decline of c.65% has occurred over a period of 16 years (Skov et al. 2011). Research on the tundra of eastern European Russia since 1973 suggests that a steep decline may have occurred there (Y. and O. Mineev in litt. 2012). There is no evidence to suggest that the geographical distribution of this population has shifted sufficiently to account for this decline (Skov et al. 2011, L. Nilsson in litt. 2011, J. Bellebaum in litt. 2012, M. Ellermaa in litt. 2012), though small increases in wintering numbers have been recorded in Finland. The estimation of a real and severe decline is supported by data from migration bottlenecks in the Gulf of Finland where the numbers observed on migration have fallen dramatically since the early 1990s at least (A. Lehikoinen et al. in litt. 2012). North America holds the second largest population (c.1 million birds [Wetlands International 2017]). The Waterfowl Breeding Population and Habitat Survey (WBPHS) indicates large declines from the early 1950s but a relatively stable population since 1989 (Flint 2013); on a shorter time-scale, the population appears to have declined from 1980 to 2002 and then increased from 2007 to 2012 (Bowman et al. 2015). Population trends vary among regional surveys; breeding numbers on the Arctic Coastal Plain were stable during 1986-2012 but increasing 2003-2012; on the Yukon-Kuskokwim Delta breeding numbers were stable during 1988-2012 but decreased during 2003-2012; on western Victoria Island, Northwest Territories, breeding numbers probably decreased from 1992-94 to 2004-05; and on Kodiak Island, Alaska, wintering numbers were stable during 1991-2005 (Raven and Dickson 2006; Zwiefelhofer et al. 2008; Stehn et al. 2013; Bowman et al. 2015). However, this species is especially poorly monitored because its breeding distribution is largely outside the area covered by breeding waterfowl surveys, and because of its offshore distribution during winter and the lack of comprehensive winter surveys (T. Bowman in litt. 2012). No trend data are available for the third largest population (c.1,600,000 birds [Wetlands International 2017]), which breeds in western Siberia and winters off eastern Asia. The fourth and smallest population (c.36,000–99,000 birds [Wetlands International 2017]), which breeds in Greenland and Iceland and winters in the north Atlantic, is poorly monitored and trends are uncertain, although it may have been stable until the 1990s (Wetlands International 2017).
Behaviour This species is fully migratory although its movements are poorly understood (del Hoyo et al. 1992, Scott and Rose 1996). It breeds from late-May onward in single pairs or loose groups (Madge and Burn 1988, del Hoyo et al. 1992, Kear 2005), the males leaving the females soon after the start of incubation (between late-June and early-September) to gather in small flocks for a flightless moulting period (Madge and Burn 1988, Scott and Rose 1996). Some populations undergo extensive moult migrations of up to 1,000 km, while others moult on waters near the breeding grounds (Madge and Burn 1988). Females moult between early-August and early-October on the breeding grounds (Scott and Rose 1996), often abandoning their young at the start of the moult which then gather into large parentless groups (Johnsgard 1978). The southward autumn migration occurs from September to October after the post-breeding moult (Scott and Rose 1996) and non-breeders may oversummer in the wintering areas (Madge and Burn 1988). Outside of the breeding season the species is highly gregarious, in winter gathering into large aggregations of perhaps several tens of thousands of individuals to roost or to feed in inshore and offshore waters (Johnsgard 1978, del Hoyo et al. 1992). The species is diurnal and regularly dives to depths of 3-10 m when foraging, with a maximum dive depth of 50-60 m (Kear 2005).
Habitat Breeding The species breeds on marshy grass tundra in the high Arctic, especially where habitat mosaics are formed by hummocks and ridges together with moist depressions, freshwater lakes, bogs, slow rivers or pools of standing water, it is common among willows or dwarf birch in the arctic-alpine zone (Scandinavia) (Flint et al. 1984, Madge and Burn 1988, del Hoyo et al. 1992, Snow and Perrins 1998). It generally avoids wooded tundra (Johnsgard 1978, Snow and Perrins 1998, Kear 2005). The species also breeds on small rocky islands off mainland Arctic coasts and on larger offshore islands, using promontories, deltas, coastal inlets and islets in fjords (Greenland) (Madge and Burn 1988, del Hoyo et al. 1992, Snow and Perrins 1998). Non-breeding The species winters at sea, generally far offshore in waters 10-35 m deep, as well as in saline, brackish or fresh estuarine waters, brackish lagoons , and inland (very rarely) on large, deep freshwater lakes (Johnsgard 1978, del Hoyo et al. 1992, Scott and Rose 1996).
Diet The species shows a preference for marine foods during both the breeding and non-breeding seasons (Johnsgard 1978), its diet consisting predominantly of animal matter such as crustaceans (e.g. amphipods and cladocerans), molluscs, other marine invertebrates (e.g. echinoderms, worms) and fish (Johnsgard 1978, del Hoyo et al. 1992, Snow and Perrins 1998, Kear 2005). The species also takes freshwater insects and insect larvae ands well as plant material such as algae, grasses, and the seeds and fruits of tundra plants (del Hoyo et al. 1992).
Breeding site The nest is a natural depression on dry ground positioned in the open, amongst vegetation, partially hidden by overhanging boulders or under low shrubs (e.g. willows or dwarf birch) usually close to water (Johnsgard 1978, Flint et al. 1984, Madge and Burn 1988, del Hoyo et al. 1992). Although it is not a colonial species some pairs may nest in loose groups, and the species may also nest in association with Arctic Terns (Kear 2005).
A major decline observed in the wintering population of the Long-tailed Duck in the Baltic Sea coincides with heavy bycatch mortality in gillnet fisheries, with an estimate of at least 90,000 birds (of different species) killed annually (Žydelis et al. 2009). This species is highly susceptible to gillnet mortality and is the most frequently recorded victim in the eastern and south eastern Baltic Sea (Skov et al. 2011), with annual mortality estimates ranging from 1 to 5% for the total Baltic Sea population (Bellebaum et al. 2013). Age ratios of gillnet victims in the southern Baltic suggest a decline in breeding success of c.75% over the 20th century (J. Bellebaum in litt. 2012).
The effects of chronic oil pollution from small-scale oil discharges in non-breeding areas is a significant threat as Long-tailed Ducks form large aggregations that overlap with major shipping and oil transportation routes. Further expansion of such activity in the northern part of the range is likely to increase the scope of this threat (Hearn et al. 2015). Large, catastrophic oil spills still occur, although the frequency of incidence has reduced (Hearn et al. 2015). Industrial activity within the breeding range poses a threat through direct mortality from oil pollution (Gorski et al. 1977, Kirby et al. 1993, Y. and O. Mineev in litt. 2012, Carboneras and Kirwan 2017) as well as the degradation and loss of wetland habitat due to petroleum pollution, drainage and peat-extraction (Grishanov 2006). Sand and gravel extraction destroys habitats of benthic organisms, which are key prey items. In the Baltic, these activities are currently occurring at insufficient levels to be responsible for declines in waterbirds (Skov et al. 2011). A mortality event in 2006-2007 ascribed to pollution resulted in ‘several thousand’ dead Long-tailed Ducks being washed up along the Hatpudirskaya bay coast (Y. and O. Mineev in litt. 2012). Development of large offshore wind farms could have negative effects on the species, through restricting access to foraging areas and by altering the suitability of benthic habitat.
The species has previously suffered heavy losses from avian cholera outbreaks (Friend 2006) and is susceptible to avian influenza (Melville and Shortridge 2006). Future outbreaks of disease may be expected and may compound observed declines in parts of the range. The introduced Round Goby Neogobius melanostomus has become established in the Baltic where it feeds on benthic invertebrates and could affect food supplies for Long-tailed Ducks, with unknown severity of impact (Almquist et al. 2010, Dagys 2017). The Round Goby appear to considerably reduce mussel density in shallow water (less than 20 m) in the Baltic Sea, and the numbers of Long-tailed Duck in these areas have decreased correspondingly (Hearn et al. 2015).
The species is still hunted across the majority of its range (Hearn et al. 2015, Carboneras and Kirwan 2017). The mortality from hunting has fallen considerably in line with the recorded decline in the species, partly due to reduced numbers of birds available but also as a result of tighter regulation and fewer active hunters (Hearn et al. 2015). However, bag size has recently increased in Finland (19,400 birds in 2013), which may reflect greater numbers present during winter (Hearn et al. 2015). The sustainability of present-day hunting levels is unclear, but these populations are already subject to significant mortality from other threats so require close monitoring. Previous hunting levels cannot be assumed to have been sustainable, and better quantification of additive mortality from all threats is needed to be confident in the size of any harvestable surplus (Hearn et al. 2015).The results of autumn migration monitoring at various Baltic sites show that juveniles now represent a very low proportion of the population (Hario et al. 2009, Ellermaa et al. 2010), indicating that insufficient young are being raised to compensate for adult mortality (low recruitment). This low breeding success has been linked to the collapse of the formerly distinctive 3-4 year cycle of Arctic rodent abundance since the mid-1990s as a consequence of climate change, which may lead to greater rates of predation on alternative prey (Hario 2009, Iles et al. 2013). However, breeding success also appears to have declined in locations where rodents are absent, with poor female condition leading to bypassed breeding (A. Kondratyev in litt. 2012). In addition, ocean acidification may impact mollusc density and quality, further reducing prey availability (Steinacher et al. 2009). More information about the effects of climate change on lower trophic levels is required, but prey density and size is likely to be impacted by changes in sea temperature, and key foraging ecologies may be altered, but there is uncertainty over whether resources are lost or have shifted their distribution.
Conservation Actions Underway
CMS Appendix II. EU Birds Directive Annex II. Some of the species's habitat is protected. Efforts are on-going to monitor populations of this species in many parts of its range. AEWA Action Plan adopted in 2015. Working group to oversee implementation in process of being established. New coordinated survey of Baltic conducted in January 2016 (results expected in 2017), plus development of other surveys, demographic monitoring and migration studies. Some new restrictions on hunting have been introduced recently. Actions to reduce bycatch ongoing in several countries. Various protected areas implemented recently, especially marine SPAs for wintering birds.
Conservation Actions Proposed
Conduct surveys to monitor the least-known populations in East Asia, improve monitoring in North America, and continue to assess trends in Europe. Carry out research to identify the causes of the decline in the Baltic Sea. React to improved knowledge of threats to the species by implementing actions to mitigate their impacts. Study the causes of reduced breeding productivity.
Text account compilers
Moreno, R., Palmer-Newton, A., Stuart, A., Taylor, J., Ekstrom, J., Butchart, S., Fjagesund, T., Ashpole, J, Hermes, C., Malpas, L., Calvert, R., Martin, R.
Ellermaa, M., Kondratyev, A., Bowman, T., Lehtiniemi, T., Below, A., Kharitonov, S., Bellebaum, J., Pessa, J., Bianki, V., Kharitonova, I., Solovyeva, D., Hearn, R., Kontiokorpi, J., Rajasarkka, A., Tiainen, J., Hario, M., Fefelov, I., Mineev, O., Nilsson, L., Mikkola-Roos, M., Mineev, Y., Valkama, J., Pihl, S., Lehikoinen, A., Lehikoinen, E., Grishanov, G.
BirdLife International (2020) Species factsheet: Clangula hyemalis. Downloaded from http://www.birdlife.org on 11/08/2020. Recommended citation for factsheets for more than one species: BirdLife International (2020) IUCN Red List for birds. Downloaded from http://www.birdlife.org on 11/08/2020.