Closely related to L. limosa, and elsewhere previously considered conspecific. Monotypic.

Various efforts to monitor the populations of this large long-distance migrant shorebird indicate that a significant decline is taking place, most severely noted in numbers recorded at migratory sites in North America but evident in count data assembled for key wintering areas. Rates of reduction could well exceed 30% over three generations and accordingly the species is assessed as Vulnerable. At present the drivers of such declines are uncertain, although the loss of and disturbance to key non-breeding and stopover sites is suspected to impact populations and climate change driven habitat shifts may be causing prey mismatches that affect reproductive output, as well as altering fine-scale habitat suitability.

The species breeds in the subarctic and boreal regions in northern Canada and Alaska (United States of America) (Walker et al. 2024). It undertakes a long-distance migration to non-breeding areas primarily in Argentina and Chile, with small numbers along the Atlantic coast in Uruguay and Brazil and the Pacific coast to northern Peru (Fink et al. 2023, Walker et al. 2024). On northbound passage most pass through the centre of the USA from Louisiana and Texas to North Dakota and Saskatchwan (Fink et al. 2023, Walker et al. 2024).

The population size is estimated to be 41,000 to 70,000 mature individuals.

Simultaneous surveys of the Atlantic portion of the non-breeding range in January 2019 recorded a mean total of 56,276 (9,326-113,221) individuals, using two step ‘hurdle’ generalised linear models to predict abundance in different habitat classes (Faria et al. in press). The wide confidence intervals are a cause for caution in using this value and the total number of individuals observed in the production of this estimate is only 1,317 (Faria et al. in press), in contrast to the 45,500 individuals counted during aerial surveys in the 1980s (Morrison and Ross 1989). These caveats aside, it is a considerable simultaneous effort to sample the non-breeding range that hosts more than 70% of the population. To account for the large uncertainty in this data this assessment uses the second quartile range of values (25-50%) as a precautionary position. This is considered necessary especially in light of data indicating extremely rapid declines in monitored migration sites in North America (Smith et al. 2023). The lower bound of the second quartile is 28,959 individuals, hence a cautious range is given for this estimate of 29,000-56,300 individuals. Faria et al. (in press) did not estimate numbers for Chilean sites, hence it appears the birds occurring at Bahia Lomas in Chile (Morrison unpublished data, in COSEWIC 2019, Morrison and Ross 1989) are not included in these counts. Hence the c. 18,000 individuals counted at this site in 2018 (Morrison unpublished data, in COSEWIC 2019) need adding to this estimate, increasing it to 47,000-74,300 individuals. There have been subsequent aerial counts of the Bahia Lomas non-breeding population: 20.423 ± 639 individuals in January 2022 (Norambuena et al. 2022b) and 24,038 ± 148 individuals in January 2023 (Norambuena et al. 2023): but the proximity of the site to those covered by Faria et al. (in press) suggests that the precautionary position is to use the values closest in time in case of interannual dispersal between non-breeding sites.

Almost all the remaining non-breeding birds occur around Chiloe Island on the Pacific coast of Chile (Morrison and Ross 1989, Espinosa et al. 2005, Andres et al. 2009, García Walther 2016, Fink et al. 2023, Walker et al. 2024). This has most recently been estimated at 18,640 individuals (excluding potentially inflationary extrapolation to non-surveyed sites) from surveys in 2014 (García Walther 2016). This latter total is comparable in method to the 20,961 individuals reported by Andres et al. (2009) for 2007-8, and the 21,282 individuals in 1997 (Espinosa et al. 2005).Summing the most recent values for the two regions gives a range of 65,599 – 92,916 individuals, rounded to 65,600-93,000 for the region that hosts nearly all non-breeding birds. While small numbers are present further north as far as Piura, Peru (Senner 2008, Fink et al. 2023), these are not thought to add significantly to this estimate.The median value is very similar to the 77,000 individuals given by Andres et al. (2012), which was based on an earlier estimate of 56,000 individuals for the Tierra del Fuego (and Samborombon Bay) non-breeding population (Morrison et al. 2006) plus 21,000 individuals for the Chiloe population (Andres et al. 2009).These estimates, all based on non-breeding counts, include immature birds. Assuming these represent between one quarter (COSEWIC 2019) and one third of the counted individuals then the population size falls between 43,000 and 70,000 mature individuals.

Much of the above data was used in the COSIEWIC assessment of Hudsonian Godwit in 2019 (COSEWIC 2019), which gave a revised population estimate of 41,000 mature individuals. This differed in using a lower estimate for the Tierra del Fuego wintering population of 32,400 ± 5,900 based on the average count over 2011-2018 (Morrison unpublished data, in COSEWIC 2019) plus a separate estimate for Samborombon Bay in Argentina of 1,070 ± 290 individuals (Martínez-Curci and Isacch 2017). Averaged count values can be underestimates for a target species that can switch sites between years, often leading to apparent large fluctuations in numbers at small geographic scales: this was noted for counts in the Chiloe region by Espinosa et al. (2005). But this lower value (based on count data rather than modelled abundance) is comparable with the lower quartile value used as a precautionary bound above. The counter suggestion is that a modelling method can perform poorly for species that aggregate in large flocks, hence occur only patchily across apparently suitable habitat. This is certainly the case for this species. Given that the estimate falls so close to the lower bound of that derived from the above approach, this estimate is accepted as the lower bound of the population size.

A separate population estimate of 70,710 ± 30,385 individuals was derived from 2002-2003 spring migration counts in the U.S. Prairie Pothole region (Skagen et al. 2008). It is assumed that almost all godwits heading to the breeding grounds (a proportion remain in non-breeding areas [Navedo and Ruiz 2020]) are mature individuals. It is notable that this value is comparable to that derived from the non-breeding data. The value of 7 days for stopover duration used was a generic value based on a small number of observations relating to other species of shorebird (Skagen et al. 2008). Morrison et al. (2006) considered a conservative approach assuming the true value lies within 1 SE below the mean, which gives an estimate from this data of 40,300-70,700 mature individuals.

This is a wholly migratory species for which there do not appear to be any barriers to dispersal, and there is no known variation in morphology. Despite this, significant genetic divergence was reported between the Hudson Bay and Mackenzie Delta breeding individuals, though based on a sample of only 20 individuals (Haig et al. 1997). It appears there is also some evidence of migratory connectivity of Alaskan breeding birds and wintering areas on the Pacific coast of Chile, and separately the remaining breeding population wintering on the Atlantic coast and Tierra del Fuego (Senner 2010). However, tracking data has shown that individuals may move between these areas prior to northbound migration (COSEWIC 2019, Walker et al. 2024) and it is considered most likely that there is regular individual dispersal between different breeding areas. If the Alaska/Pacific population is a separate subpopulation, the current estimate of the population size is 12,300 mature individuals (García Walther 2016), which would be 30% of the minimum population size.

Hudsonian Godwit breeds on marshy tundra close to the treeline in Alaska and Canada. During their migration and over winter in South America, the species occupies mudflats, shallow marshes, tidal pools, shallow freshwater lakes and flooded fields (Van Gils et al. 2019).
Variation in survival rates over the annual cycle indicate lowest survival during breeding and southbound migration, but that breeding condition is affected by site quality in the non-breeding season (Swift et al. 2020).

A number of threats have been identified that may be impacting the population but no clear driver of declines is evident. Climate change impacts may affect the availability or productivity of non-breeding or stopover habitat throughout the range (COSEWIC 2019), but no direct link to the species is evident. However, the species breeding habitat in the sub-Arctic and boreal regions is being affected by climate change and the amount of precisely suitable habitat is predicted to decline (Swift et al. 2017), as already noted for Whimbrel N. phaeopus through the advancement of scrub and trees (Ballantyne and Nol 2015) and projected for other arctic breeding waders (Wauchope et al. 2016). In addition there is evidence of climate variability causing a significant phenological mismatch between breeding and peak prey availability in some parts of the breeding range (Senner et al. 2017, Wilde et al. 2022). Chick survival was lower when invertebrate prey (especially larger items needed by larger chicks) was scarcer, but notably years with especially severe mismatches may pass tipping points in the species' ability to adapt: productivity in the study area was almost zero in 2014 and 2015 where the mismatch was particularly large (Wilde et al. 2022). 
Habitat is also being lost to direct conversion and degradation from a variety of sources. Natural wetlands in the Great Plains, key migratory stopover areas for the species, have been extensively converted to agriculture, although recent and ongoing changes are thought minimal (COSEWIC 2019), but there remain plausible ongoing impacts from nutrient enrichment of wetlands, especially from cattle (A.P. Nunes in litt. 2024). Overgrazing by greatly increased populations of Canada Goose Branta canadensis and Snow Goose Anser caerulescens have created large bare areas in breeding habitat that are avoided by Hudsonian Godwits, but there is no evidence this has affected breeding density (Swift et al. 2017). There is concern that a large expansion of offshore wind turbines in Bahia Lomas and Tierra del Fuego may cause displacement or collision mortality (Norambuena et al. 2022a). Infrastructure associated with oil extraction in the vicinity of Bahia Lomas pose a risk of oil spills (Espoz et al. 2022), and the proximity of major shipping routes there and in Tierra del Fuego poses the risk that a major oil spill could affect a large number of individuals (Senner 2010), but population impacts of oil contamination have not been demonstrated in the species. On Isla Chiloé (Chile) salmon farms, peat mining and seaweed collection threaten the foraging grounds (Senner 2010), Anon. 2015) but their most severe impact may be via physical disturbance of foraging and roosting flocks (Senner 2008). Human disturbance may also affect stopover sites that are popular tourist beaches (Senner 2010). Disturbance by dogs is also thought notable in this area and around Rio Grande, Argentina (COSEWIC 2019). Close to Churchill, Manitoba, Common Raven have become established through reliance on human resources but predate nests in the region to a level that may suppress reproductive output here (COSEWIC 2019). 
Historically hunting was considered the main driver of significant population declines during the 19th Century (Walker et al. 2024). The species is now protected in Canada and the USA. In investigations of shorebird hunting in the Caribbean and northern South America this species is rarely mentioned and it is not one of the main target species (e.g. AFSI Harvest Working Group 2017, Andres et al. 2022), but a potential biological removal value was generated by Watts et al. (2015) of 1,945 ± 573. On Martinique bag limits for the species of 15 birds for the season have applied (AFSI 2020), but any killing of the species was outlawed on 4th October 2021 (Anon. 2021). An unknown but likely small number are likely to still be taken in the Caribbean each autumn. There is still a level of traditional subsistence harvest by the James Bay Cree during migration, but this is thought unlikely to be at significant levels (COSEWIC 2019); it is not a target of subsistence harvest in Alaska (Naves et al. 2019) but may be rarely taken if mistaken for Bar-tailed Godwit L. lapponica (Senner 2010).

Historically hunting was considered the main driver of significant population declines during the 19th Century (Walker et al. 2024). The species is now protected in Canada and the USA. In investigations of shorebird hunting in the Caribbean and northern South America this species is rarely mentioned and it is not one of the main target species (e.g. AFSI Harvest Working Group 2017, Andres et al. 2022), but a potential biological removal value was generated by Watts et al. (2015) of 1,945 ± 573. On Martinique bag limits for the species of 15 birds for the season have applied (AFSI 2020), but any killing of the species was outlawed on 4th October 2021 (Anon. 2021). An unknown but likely small number are likely to still be taken in the Caribbean each autumn. There is still a level of traditional subsistence harvest by the James Bay Cree during migration, but this is thought unlikely to be at significant levels (COSEWIC 2019); it is not a target of subsistence harvest in Alaska (Naves et al. 2019) but may be rarely taken if mistaken for Bar-tailed Godwit L. lapponica (Senner 2010).

Conservation Actions in Place
Protected in Canada under the Migratory Birds Convention Act, 1994 (Government of Canada 2017) and in the United States under the Migratory Bird Treaty Act (USFWS 2017), through prohibitions on harm to birds, nests or eggs. It breeds within a number of protected areas and many important sites for migration stopovers and non-breeding areas are covered by the Western Hemisphere Shorebird Reserve Network (WHSRN undated).
Tracking studies are underway on Isla Chiloé to investigate migratory routes (Walker et al. 2024). A conservation action plan for Isla Chiloé was completed in 2010 with this as one of the focal species (Delgado et al. 2010). 

Conservation Actions Proposed
Continue research into the migration routes and staging sites of Hudsonian Godwit. Regular surveys at key staging sites: Yukon-Kuskokwim Delta, Upper Cook Inlet, Hudson and James Bays, Quill Lakes would give insight into the status of the different breeding areas (Walker et al. 2024). Monitor the population size and trends at breeding and non-breeding sites. Monitor habitat changes at important sites and how these affect Hudsonian Godwits. Protect important sites from habitat degradation. Protect islands and tidal areas that are important for shorebirds from development. Raise awareness about the effects of agriculture and aquaculture practices on shorebirds. Introduce laws and regulations to manage human activities around important shorebird sites. Regulate farming and land management practices to conserve important habitat types for Hudsonian Godwits. Enforce regulations within National Parks to protect important wetland habitats. Raise awareness of the importance of the conservation of Hudsonian Godwit populations. Promote the benefits of Godwit populations for bird watching and tourism opportunities. Convene a working group to cooperate and organise research and conservation efforts for the Hudsonian Godwit (Senner 2010).