Melanitta deglandi

Melanitta fusca was formerly (AOU 1983, 1998) considered conspecific with M. deglandi and M. stejnegeri, but separated on the basis of color and pattern differences, including bill structure; tracheal differences (Miller 1926); a lack of known hybridization in areas of parapatry and co-occurrence; and a lack of rationale for the original merger by Hartert (1920). M. fusca and M. deglandi had been previously considered distinct (AOU 1895 through AOU 1957) (AOS 2019).

This species breeds in North America from northern Alaska, northern Yukon, northwestern and southern Mackenzie, southern Keewatin, and northern Manitoba south to central Alaska, southern Yukon, interior British Columbia, southeastern Alberta, southern Saskatchewan, northern North Dakota (formerly), southern Manitoba, northern Ontario, and western Quebec, occurring in summer to northeastern Mackenzie and from Hudson Bay east to Labrador and Newfoundland. It winters in North America on the Pacific coast from the Aleutians and Alaska Peninsula south to central California, less commonly south to northern Baja California, on the Great Lakes, and on the Atlantic coast from the Gulf of St. Lawrence and Newfoundland south to New Jersey, less commonly south to North Carolina and rarely south to Florida. It migrates regularly through Utah, North Dakota, the Great Lakes region, and the Mississippi and Ohio valleys. Casual on Melville Island, through the interior of North America south to Baja California, Arizona, Sonora, New Mexico, southern Texas, and the Gulf coast (east to Florida), in Greenland, and in northwestern Europe (most records from Iceland, the Faeroes, and Denmark) (AOS 2019).

The following is from BirdLife International (2018):

Hunting poses a significant threat to the species, with an estimated harvest of 6,000 – 14,000 individuals in the 10-year period from 2004-2014 in US and Canada, but with a downward trend (Canadian Wildlife Service Waterfowl Committee 2015). A recent estimate suggests that the harvest of White-winged Scoters may be 29-37% over the maximum sustainable rate (Koneff et al. 2017).

The species feeds at depths of 30–40 m and is therefore highly susceptible to bycatch in fishing nets, as documented by Good et al. (2009) in the Salient Sea, Washington. Moulting and wintering populations are also vulnerable to oil spills as they often occur in high densities in close proximity to oil transportation routes. One catastrophic spill could impact a large proportion of the population. Oil pollution can also cause the long-term disruption of food supplies which may have more serious effects on populations in the long-term than direct mortality from oiling. Lance (2001) documented evidence for a long-term negative effect as a result of the 1989 Exxon Valdez accident. The disturbance and habitat degradation caused by increasing levels of oil and gas exploration in the Arctic will likely have only minor or negligible impacts on this species (Poland et al. 2003, Henderson and Loe 2014). In some locations in the Pacific Northwest, sea ducks could be exposed to toxicologically significant levels of cadmium associated with mussels foraged from aquaculture structures, which raises the possibility that such exposure could be contributing to observed population declines (Bendell 2011). Climate change is an additional threat, with remote sensing showing the shrinkage of ponds in subarctic Alaska (Riordan et al. 2006), with these wetlands providing important breeding grounds for the species at present. In general, climate change is predicted to cause dramatic habitat changes in the Arctic region (Fox et al. 2015). Decreasing spring snow cover duration in boreal regions has been linked to population declines of Scoters, likely due to trophic mismatch, projected to cause a 31.0% population decline between 1980 and 2080 (Drever et al. 2011). Ocean acidification may lead to declines in molluscs which form a large part of White-winged Scoter diet (Steinacher et al. 2009, del Hoyo et al. 2017).