Arctic Ecozone+ Status and Trends Assessment
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J. Eamer,Footnote 1 G. Henry,Footnote 2 A. Gunn,Footnote 3 and L. HardingFootnote 4
Ecosystem Status and Trends 2010
Technical Ecozone+ Report
Library and Archives Canada Cataloguing in Publication
Arctic Ecozone+ status and trends assessment.
Issued also in French under title:
Évaluation de l’état et des tendences de l’écozone+ de l’Arctique.
Electronic monograph in PDF format.
Cat. no.: En14-43/01-1-2014E-PDF
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This report should be cited as:
Eamer, J., Henry, G., Gunn, A and Harding L. 2014. Arctic Ecozone+ Ecozone+ Status and Trends Assessment. Canadian Biodiversity: Ecosystem Status and Trends 2010, Technical Ecozone+ Report. Canadian Councils of Resource Ministers, Ottawa, ON. xii + 260 pp. Technical Reports
© Her Majesty the Queen in Right of Canada, 2014
Aussi disponible en français
The Canadian Councils of Resource Ministers developed a Biodiversity Outcomes Framework (Environment Canada, 2006) in 2006 to focus conservation and restoration actions under the Canadian Biodiversity Strategy (Federal-Provincial-Territorial Biodiversity Working Group, 1995). Canadian Biodiversity: Ecosystem Status and Trends 2010 (Federal, Provincial and Territorial Governments of Canada, 2010) was the first report under this framework. It presents 22 key findings that emerged from synthesis and analysis of background technical reports prepared on the status and trends for many cross-cutting national themes (the Technical Thematic Report Series) and for individual terrestrial and marine ecozones+ of Canada (the Technical Ecozone+ Reports). More than 500 experts participated in data analysis, writing, and review of these foundation documents. Summary reports were also prepared for each terrestrial ecozone+ to present the ecozone+-specific evidence related to each of the 22 national key findings (the Evidence for Key Findings Summary Report Series). Together, the full complement of these products constitutes the 2010 Ecosystem Status and Trends Report (ESTR) (Figure 1).
This report is the technical report for the Arctic Ecozone+ (Figure 2 and Figure 3) which includes all or portions of Yukon, Northwest Territories, Nunavut, Quebec, and Newfoundland and Labrador. A range of authors and reviewers contributed to the report from government, academia, non-governmental, and consulting sectors. No claim is made that the information presented is exhaustive. As in all ESTR products, the time frames over which trends are assessed vary – both because time frames that are meaningful for these diverse aspects of ecosystems vary and because the assessment is based on the best available information, which is over a range of time periods.
Information about the broader ESTR project and its other reporting products is available at: Canadian Biodiversity: Ecosystem Status and Trends 2010.
Sources of information for this report include research results published in books, scientific journals, and government reports, and records from monitoring and management-related studies. Information from experts is also included (as personal communications) for some subjects when, for example, study results have not yet been published. An important source of information throughout this report is the series of thematic reports prepared for the ESTR. In most cases, these reports are national in scope, with status and trends synthesized both nationally and by ecozone+. The reports include analyses of specific datasets, set in the context of the broader scientific literature. In this report, the thematic reports are summarized and discussed in terms of the Arctic Ecozone+. Key trend analyses are reproduced. Note, however, that analyses in the thematic reports cover a range of dates and have not been updated. This report also draws on summaries and features prepared for the main ESTR report Canadian biodiversity: Ecosystem status and trends 2010 (Federal, Provincial and Territorial Governments of Canada, 2010) and on Arctic Council’s Arctic Biodiversity Assessment (CAFF, 2013), a source of information on Canadian Arctic biodiversity and on the circumpolar context of status and trends.
Ecological classification system – ecozones+
A slightly modified version of the Terrestrial Ecozones of Canada, described in the National Ecological Framework for Canada (Ecological Stratification Working Group, 1995), provided the ecosystem-based units for all reports related to this project (Figure 2). Modifications from the original framework include: adjustments to terrestrial boundaries to reflect improvements from ground-truthing exercises; the combination of three Arctic ecozones into one; the use of two ecoprovinces – Western Interior Basin and Newfoundland Boreal; the addition of nine marine ecosystem-based units; and, the addition of the Great Lakes as a unit. This modified classification system is referred to as “ecozones+” throughout these reports to avoid confusion with the more familiar “ecozones” of the original framework (Rankin et al., 2011).
The southern boundary of the Arctic Ecozone+ corresponds with the northern edge of the boreal forest. Adjustments to the Taiga Shield/Arctic boundary, based on ground-truthing, were made in the ecozone+ classification. As noted above, three Arctic terrestrial ecozones were combined into the Arctic Ecozone+. This is a vast and diverse ecozone+ and the original three Arctic ecozones represent distinct ecological units. Where appropriate they are discussed separately and referred to as the Southern Arctic, Northern Arctic, and Arctic Cordillera regions of the Arctic Ecozone+.
- Lead authors
- J. Eamer, G. Henry, A. Gunn and L. Harding
- Contributing authors
- R. Brown, S. Carrière, B. Collins, S. Elmendorf, G. Gauthier, J. Goulet, C. Haas, E. Jacobsen, J. Paquet, T. Prowse, A. Shabbar, R. Smith and L. Toretti
- Authors of ESTR Thematic Technical Reports from which material is drawn
- Ecological classification system for the ecosystem status and trends report: R. Rankin, M. Austin, and J. Rice
Guidance for the preparation of ESTR products – classifying threats to biodiversity: C. Wong
Guidance for the preparation of ESTR products – land classification scheme: J. Frisk
Large-scale climate oscillations influencing Canada, 1900–2008: B. Bonsal and A. Shabbar
Canadian climate trends, 1950–2007: X. Zhang, R. Brown, L. Vincent, W. Skinner, Y. Feng, and E. Mekis
Wildlife pathogens and diseases in Canada: F.A. Leighton
Trends in permafrost conditions and ecology in northern Canada: S. Smith
Northern caribou population trends in Canada: A. Gunn, D. Russell, and J. Eamer
Landbird trends in Canada, 1968–2006: C. Downes, P. Blancher, and B. Collins
Trends in Canadian shorebirds: C. Gratto-Trevor, R.I.G. Morrison, B. Collin, J. Rausch, M. Drever, and V. Johnston
Monitoring biodiversity remotely: a selection of trends measured from satellite observations of Canada: F. Ahern, J. Frisk, R. Latifovic, and D. Pouliot
Biodiversity in Canadian lakes and rivers: W.A. Monk and D.J. Baird
Review conducted by scientists and renewable resource and wildlife managers from relevant territorial and federal government agencies through a review process recommended by the ESTR Steering Committee. Additional reviews of specific sections conducted by university researchers in their field of expertise at the request of the authors. Peer review of the final draft coordinated by the Canadian Society for Ecology and Evolution. The report was greatly improved by the reviews of D. Berteaux, P. Hale, and one anonymous reviewer.
Direction provided by the ESTR Steering Committee composed of representatives of federal, provincial and territorial agencies.
Editing, synthesis, technical contributions, maps and graphics, and report production by the ESTR Secretariat, J. Eamer, M. Osmond-Jones, P. Lilley, K. Badger, E. Jacobsen, and M. Connolly.
Aboriginal Traditional Knowledge compiled from publicly available sources by D. Hurlburt.
- Footnote 1
Environment Canada (to 2011); current: SLR Consulting (Canada) Ltd., 200-1620 West 8th Avenue, Vancouver, BC
- Footnote 2
Department of Geography, University of British Columbia, Vancouver, BC
- Footnote 3
368 Roland Road, Salt Spring Island, BC
- Footnote 4
SciWrite, 2339 Sumpter Drive, Coquitlam, BC
List of Figures
- Figure 1. Ecosystem status and trends (ESTR) report series.
- Figure 2. Canada’s ecozones+.
- Figure 3. Overview map of the Arctic Ecozone+.
- Figure 4. Reconstructed Canadian Arctic vegetation: 18,000, 8,000, and 5,000 years ago.
- Figure 5. Regions of the Arctic Ecozone+: the Arctic Cordillera, Northern Arctic, and Southern Arctic.
- Figure 6. Major drainage areas of the Arctic Ecozone+.
- Figure 7. Surficial materials in the Arctic Ecozone+: percentage of total land area. See Appendix 1 for descriptions of classes of surficial materials.
- Figure 8. Map of surficial materials of the Arctic Ecozone+.
- Figure 9. Human population trends for the Arctic Ecozones+, 1971–2006.
- Figure 10. Regions covered by comprehensive land claim settlements in the Arctic Ecozone+.
- Figure 11. Wall map from a DEW Line station, showing the line and location of main stations.
- Figure 12. Global increases in annual temperatures for 2001–2005 relative to 1951–1980, showing the greatest warming over land and at high latitudes in the Northern Hemisphere.
- Figure 13. Mean annual global and circumpolar Arctic temperatures (64 to 90°N latitude), 1880 to 2012.
- Figure 14. Change in mean temperature, 1950–2007, for: a) spring (March–May), b) summer (June–August), c) fall (September–November), and d) winter (December–February).
- Figure 15. Growing season length in control plots of a long-term warming experiment in a Cassiope heath tundra at Alexandra Fiord, Ellesmere Island.
- Figure 16. Change in the amount of precipitation, 1950–2007, for: a) spring (March–May), b) summer (June–August), c) fall (September–November), and d) winter (December–February).
- Figure 17. Time series of the annually averaged Arctic Oscillation Index (AO) for the period 1950–2005.
- Figure 18. Northern Hemisphere surface temperature anomalies averaged over three periods with different types of dominating pattern of natural variability.
- Figure 19. Maximum summer thaw depth for a site at Baker Lake.
- Figure 20. Mean annual ground temperatures at two boreholes and two depths at Alert, Nunavut, 1978 to 2011.
- Figure 21. Monthly ground temperature at a depth of 5 m at Environment Canada’s borehole at Iqaluit, 1988 to 2002.
- Figure 22. Change in snow cover duration ( the number of days with greater than or equal to 2 cm of snow on the ground), 1950–2007.
- Figure 23. Change in the maximum annual snow depth, 1950–2007.
- Figure 24. Absolute change in the ratio of snow to total precipitation in Canada, 1950–2007.
- Figure 25. Trends in terrestrial spring snow cover for the Eurasian and the North American Arctic, 1967–2012.
- Figure 26. Extent of Arctic sea ice over the annual cycle, recent years compared to 1979–2000 average.
- Figure 27. Trends in mean September sea ice cover, 1979–2013, showing the recent rapid decline.
- Figure 28. Distribution of winter and summer sea ice, 2012, and median ice edge for both seasons.
- Figure 29. Annual net surface mass balance of four glaciers in the Queen Elizabeth Islands, 1960 to 2011.
- Figure 30. Location map and ice cap retreat on Baffin Island, 1949 to 2005.
- Figure 31. Ice break-up and freeze-up trends for six Arctic lakes, 1985–2004, and average trends compared with average trends for Canadian lakes south of the Arctic Ecozone+.
- Figure 32. Map showing trends in 1- day minimum river flow in natural rivers across Canada, using data for hydrological years 1970–2005.
- Figure 33. Overall trends in river discharge for six major Eurasian rivers (1936 to 2009) and four major North American rivers (1970 to 2010) rivers flowing to the Arctic Ocean.
- Figure 34. Area burned by large fires in the Arctic Ecozone+, 1960–2007.
- Figure 35. Active layer detachments on sloped tundra, Melville Island, Nunavut.
- Figure 36. Trends in small mammals (lemmings, voles, and shrews) at sites grouped according to apparent synchrony and ecological regions, 1990–2012.
- Figure 37. Geographic ranges for protostrongylid parasites in northern ungulates.
- Figure 38. Meta-analysis of the effect of experimental warming on leaf bud break at all International Tundra Experiment (ITEX) sites, circumpolar Arctic.
- Figure 39. Effect of experimental warming on net ecosystem exchange of CO2 at four ITEX sites in North America, including Alexandra Fiord, Ellesmere Island.
- Figure 40. Relative net primary productivity of major world ecosystems.
- Figure 41. Primary production in Lost Pack Lake, as measured by variations in chlorophyll a concentration in lake sediment.
- Figure 42. Trends in Normalized-Difference Vegetation Index (NDVI) for the Arctic Ecozone+, 1985–2006.
- Figure 43. Trends in primary productivity, based on remote sensing, 1982–2008.
- Figure 44. Vegetation map of dominant plant functional types of the Canadian Arctic.
- Figure 45. Above-ground plant biomass in the Canadian Arctic.
- Figure 46. Net primary productivity in the Canadian Arctic.
- Figure 47. Results of meta-analysis of plant community changes at circumpolar ITEX sites after 1 to 6 years of experimental warming.
- Figure 48. Change in above-ground and below-ground wet sedge meadow biomass over 25 years at Alexandra Fiord, Ellesmere Island.
- Figure 49. Changes in point cover (biomass index) of major plant functional groups in a snow bed heath community at Alexandra Fiord, Ellesmere Island, 1995 to 2007.
- Figure 50. Mean annual air temperature at four sites in the Canadian High Arctic, 1970–2005.
- Figure 51. Bylot Island and research station location.
- Figure 52. Temporal trends in lemming abundance and Arctic fox and snowy owl reproductive activity.
- Figure 53. Cumulative thawing degree-days (above 0°C) in the Qarlikturvik Valley, Bylot Island, 1989 to 2011.
- Figure 54. Temporal trend in above-ground live dry biomass of graminoid plants in wetland of Bylot Island, 1990–2010.
- Figure 55. Annual mean goose laying date and relationship between goose laying date and date of 50% snow cover, Qalikturvik Valley, Bylot Island, 1989 to 2012.
- Figure 56. Atmospheric carbon dioxide trends A) measured at Alert, Nunavut (1975–2010) and B) thousand-year trends based on Antarctic ice cores.
- Figure 57. General land cover classes (biomes) in the Arctic Ecozone+. See Appendix 2: Detailed land cover classes for more information.
- Figure 58. Detailed land cover in the Arctic Ecozone+.
- Figure 59. Changes in area of the tundra climatic zone around the circumpolar Arctic since 1982, based on NDVI and Köppen Climate Classification System.
- Figure 60. Spatial distribution of difference in areas of tundra climate using the Köppen climate classifications between the two years 1978 and 1998.
- Figure 61. Changes in NDVI across the annual range of the Porcupine Caribou Herd in northern Yukon and adjacent Alaska, 1986 to 2005.
- Figure 62. Changes in shrub cover in an area near Kangiqsualujjuaq, Nunavik, Quebec, 1964 to 2003, based on air photo analysis.
- Figure 63. Diatom stratigraphy from a sediment core from a pond on Ellef Ringnes Island, showing the marked changes in species assemblages since 1850, especially in very recent years.
- Figure 64. Biome-wide changes in vegetation height and abundance, circumpolar sites.
- Figure 65. Relationship between vascular plant species richness and mean July temperature in the Canadian Arctic.
- Figure 66. Terrestrial vertebrate species richness in Canada.
- Figure 67. Peary Caribou population estimates.
- Figure 68. Global distribution of Peary caribou.
- Figure 69. Home ranges of satellite-collared Peary caribou relative to favourable and unfavourable winters, Bathurst Island Complex, NU, 1993/94 and 1994 to 1997.
- Figure 70. Dolphin and Union Caribou population estimates.
- Figure 71. Circumpolar distribution of the polar bear, showing status and trends of subpopulations
- Figure 72. Declining fall weights of female polar bears, western Hudson Bay, 1980 to 2007.
- Figure 73. Approximate boundaries of current and historic (19th century) distribution of the grizzly bear in North America, showing current densities.
- Figure 74. Northern caribou subspecies and groupings.
- Figure 75. Current distribution and status designations of Rangifer in North America.
- Figure 76. Ranges and recent trends of northern caribou populations in Canada.
- Figure 77. The exponential rate of increase and decline of major tundra dwelling caribou herds in Canada.
- Figure 78. The relative size of tundra dwelling wild Rangifer herds (Canada).
- Figure 79. Muskox world distribution.
- Figure 80. The rise and fall of the muskox fur trade in Canada: total numbers of muskox hides purchased annually by the Hudson’s Bay Company, 1862 to 1916.
- Figure 81. Estimated muskox abundance in Canada, 1967–2012.
- Figure 82. Muskox trends, Bathurst Island, 1961 to 2001.
- Figure 83. Muskox trends, Banks Island, 1982 to 2010.
- Figure 84. Annual Inuvialuit total muskox harvest, 2006 to 2011.
- Figure 85. Reduction in ranges for large carnivores in North America.
- Figure 86. Wolves sighted during aerial surveys of Peary caribou and muskoxen, Banks Island, NWT.
- Figure 87. Trends in the status of breeding seabird populations in Canada, 1980s to 2000s.
- Figure 88. King eiders migration counts at Point Barrow, Alaska, 1953 to 2003.
- Figure 89. King eider nesting populations, western Victoria Island, 1990s and 2000s.
- Figure 90. Pacific common eider migration counts at Point Barrow, Alaska, 1976 to 2003.
- Figure 91. Pacific common eider counts on nesting grounds in Bathurst Inlet area, 1995, 2007, and 2008.
- Figure 92. Tundra swan (eastern population) abundance, based on January surveys on wintering grounds, 1982–2012.
- Figure 93. (A) Nesting lesser snow geese estimated through photo-inventories of major breeding colonies 1970s to 2000s, and (B) midwinter abundance indices, 1970 to 2012.
- Figure 94. Summary of population trends for Arctic-breeding shorebirds, 2003.
- Figure 95. Estimated trends in Arctic-breeding shorebird fall migration counts, 1974–1998.
- Figure 96. Annual indices of population change for the rough-legged hawk, snowy owl, Harris’s sparrow, and snow bunting, 1966 to 2005, based on Christmas Bird Count results for North America.
- Figure 97. Participation in subsistence activities by region.
- Figure 98. Participation in subsistence activities, comparing residents of major population centres with other residents of other areas: (a) Western Arctic and (b) Nunavut.
- Figure 99. Lack of a trend in wolf sightings and pack size during peak abundance and the decline of the Bathurst Caribou Herd, 1987 to 2008.
- Figure 100. Adjusting management based on herd size and population trend: Harvest management plan for the Porcupine Caribou Herd in Canada
- Figure 101. Canada’s annual greenhouse gas emissions, 1990–2011, projected emissions with current mitigation measures, and Canada’s international commitment for reduction by 2020.
- Figure 102. Protected areas in the Arctic Ecozone+, 2009.
- Figure 103. Growth of protected areas in the Arctic Ecozone+, 1927–2009.
- Figure 104. West Kitikmeot/Slave Study area.
List of Tables
- Table 1. The three major regions of the Arctic Ecozone+.
- Table 2. Overview of climate trends since 1950 across the Arctic Ecozone+.
- Table 3. Surface area of glaciers and ice caps, Canadian Arctic Archipelago, mid-1970s.
- Table 4. Summary of recent research articles exploring statistical trends in streamflow and runoff in Canadian rivers draining to northern latitude oceans.
- Table 5. Frequency of active-layer detachment failure at monitoring sites on the Fosheim Peninsula, Ellesmere Island.
- Table 6. Cruise ship visits to Arctic National Parks and nearby destinations
- Table 7. Species with a Canadian distribution restricted to the Arctic Ecozone+ and/or Arctic marine ecozones+ that are listed under Canada’s Species at Risk Act (SARA) as being of conservation concern or are assessed as at risk by the Committee on the Status of Endangered Wildlife in Canada (COSEWIC).
- Table 8. Polar bear abundance estimates for Canadian subpopulations
- Table 9. Relative imminent risk to polar bear subpopulations from climate change effects on ice habitat and known levels of toxic contaminants
- Table 10. Trends in body condition for the Baffin Bay polar bear subpopulation and relationship with sea ice conditions, 1977 to 2010.
- Table 11. Muskox population estimates by region
- Table 12. Population trend assessments for Arctic-breeding shorebirds.
- Table 13. Conservation status of tundra-nesting shorebirds as classified in the U.S. Shorebird Conservation Plan.
- Table 14. Trends in annual abundance of selected Arctic Ecozone+ landbirds, 1966 to 2006, based on Christmas Bird Count results for North America
- Table 15. Examples of documented observations by Inuit regarding range shifts of species native to Canada.
- Table 16. Trends in mercury and persistent organic pollutants from Northern Contaminants Program datasets.
- Table 17. Examples of industrial projects and caribou issues, 2012.
- Table 18. Observed and potential impacts of climate change on migratory tundra caribou populations, based primarily on research conducted on the Porcupine Caribou Herd.
- Table 19. Comparison of detailed land cover classes in the Arctic Ecozone+ with generalized land cover classifications.
Overview of the Arctic Ecozone+
- Long-term perspective: the last ice age and its retreat
- Ecozone+ characteristics
- Humans in the Arctic Ecozone+
Long-term perspective: the last ice age and its retreat
At the height of the last glaciation of the Pleistocene, about 20,000 years before present, the Laurentide ice sheet covered the mainland and some of the southern Arctic islands, including Victoria and Baffin islands, while smaller ice caps covered Melville and Bathurst islands and the islands to the northeast (Pielou, 1991). Sea level was about 150 m lower than now. The area that became Banks Island (excluding the southwestern part of the island) and parts of the western Queen Elizabeth Islands were, as they are now, polar desert (Adams and Faure, 1997).
By 13,000 years before present, Banks Island had been separated from the mainland by rising seas. Melville Island, which was connected by land to Prince Patrick Island, as well as some smaller islands to the northeast, had become ice-free (Dyke, 2004). Rising water levels isolated caribou and muskox populations from those in Beringia and, to some extent, from each other, although the animals could still walk across winter ice.
Proxy records from the Arctic Ocean floor and from surrounding land indicate that sea ice has covered at least part of the Arctic Ocean for the past 13 to 14 million years and has been widespread over the past 2 to 3 million years (Polyak et al., 2010). During periods when Earth was warmer, ice cover was reduced in extent. The most recent of these warm periods (prior to the warming of recent years) started near the beginning of the Holocene, about 10,000 years before present (Polyak et al., 2010). Based on records from the Agassiz Ice Cap (Ellesmere Island), average temperatures reached about 3°C above those of the mid-20th century, resulting in loss of multiyear landfast ice (Polyak et al., 2010). Expanses of open water during much of the year isolated islands from the mainland, creating conditions for increasing biological diversity in wildlife and in other terrestrial ecosystem components, either by genetic drift or through adaptations to local environments. Caribou and muskoxen may have thrived during this period because the vegetation changed from polar desert to dry tundra (Adams and Faure, 1997). Thick, multiyear sea ice returned to the Canadian Arctic by about 5,500 years ago and has been present since, based on evidence from Ellesmere Island (England et al., 2008). Figure 4 shows the sequence of vegetation change from the last glacial maximum through the early Holocene.
Overview by region
The Arctic Ecozone+ is characterized by low air and soil temperatures, continuous permafrost, a short growing season and limited vegetation productivity (Bolen, 1998). The area of land is about 3,148,000 km2 and the combined area of lakes, ponds and rivers is about 80,000 km2. The combined length of all the coastlines in the ecozone+ is 179,950 km, almost three-quarters of Canada’s total coastline (Natural Resources Canada, 2010). This extensive and complex interface between terrestrial and marine environments has a dominant influence on Arctic ecosystems and the wildlife and people who live there.
The Arctic Ecozone+ includes three major regions (shown in Figure 5 and described in Table 1), each an ecozone in the National Ecological Classification System (see Preface : Ecological classification system – ecozones+).
|Arctic Cordillera||Northern Arctic||Southern Arctic|
|Northeastern fringe of Nunavut and Labrador, including northeastern Baffin Island, eastern Devon Island,Ellesmere, Axel Heiberg, and Bylot islands, and the Torngat Mountains.||Non-mountainous areas of the Arctic islands plus parts of mainland Nunavut and northern Quebec.||Stretches across mainland Canada, from the Yukon coastal plain to Ungava Bay in northern Quebec.|
|A vast mountain chain (the only major Canadian mountain ranges outside of the Western Cordillera); polar ice fields, and alpine glaciers.||Low, rolling plains covered in frost-patterned soils and broken rock debris left by glaciers; broad plains in coastal zones.||Dominant features are hills and plains, ponds and lakes.|
|Dominated by lichens and some mosses at high elevations and tundra at lower elevations.||Lichens and herbs dominate, with dry tundra interspersed with wetland species.||Low shrubs mixed with herbs, lichens, and sedges such as cotton-grass. Major river valleys support scattered clumps of stunted “krummholz” trees.|
|Sparse; main communities are Pond Inlet, Clyde River, and Qikiqtarjuaq.||Iqaluit, on Baffin Island, is the largest community. Other communities with populations over one thousand include Baker Lake, Cambridge Bay, Igloolik, and Pangnirtung.||Communities are found throughout the ecozone, primarily along the coast. Those with populations over one thousand include Rankin Inlet, Arviat, Puvimituq, and Salluit. Inuvik, in the Mackenzie Delta, is just south of the ecozone+.|
Sources: Environment Canada (Environment Canada, 2005); Ecological Stratification Working Group (1995)
Rivers draining the ecozone+ flow either to Hudson Bay, Ungava Bay, or directly to the Arctic Ocean (Figure 6). The eastern slope of northern Labrador drains to the Atlantic Ocean. Because of its position along the northern rim of the Canadian mainland, the ecozone+ includes the lower reaches and marine estuaries of rivers with drainage basins extending south of the ecozone+. The vast Mackenzie River delta is partly in the Arctic Ecozone+ and partially in the Taiga Plains Ecozone+.
Topography and soils
Descriptions in this section are based on the report of the Ecological Stratification Working Group (1995).
The Arctic Cordillera has rugged mountains over 2,000 m high with massive ice fields and valley glaciers. Nunataks (mountain peaks bare of snow in summer and surrounded by valley glaciers) are common. U-shaped valleys and deep fjords extend far inland. Almost 75% of the landscape is ice or exposed bedrock. Soils are dominated by colluvial and morainal debris.
The western Northern Arctic is mainly lowland plains covered with glacial moraine, marine deposits and outcrops of Paleozoic and Mesozoic sedimentary bedrock. East of Prince of Wales and Somerset islands, granitoid Precambrian bedrock prevails. The permafrost may extend to depths of several hundred metres.
The Southern Arctic is mostly underlain by Precambrian granitic bedrock that occurs as outcrops, except in the west from Great Bear Lake to the Firth River, Yukon, where Cretaceous shales predominate. Soils are dominated by discontinuous glacial deposits except near the coasts where fine-textured marine sediments are more common. The undulating landscape has innumerable lakes, ponds and wetlands and sinuous eskers (raised ridges of gravel deposited by rivers running under the melting Laurentide ice sheet) up to 100 km long. Seasonal thaw layers are of variable thickness, and differences in moisture give rise to a variety of landforms and, consequently, plant and animal habitats.
Because of the low relief in the Southern and Northern Arctic, the high proportion of ice caps and bedrock in the Arctic Cordillera, the generally frozen soils throughout the ecozone+ , and the long seasonal period during which streams are frozen and flow is low or non-existent, stream erosion has not been significant. The melting of ice lenses exposed by fluvial action along rivers and wave action along lake and sea shorelines, however, can cause large-scale slumping and sediment discharge.
Percentages (Figure 7) and distribution (Figure 8) of surficial materials in the Arctic Ecozone+ are presented below and classes of surficial materials are described in Appendix 1. See also the related analysis of land cover in the section on Ecosystem structure (page 88).
Source: Geological Survey of Canada (Geological Survey of Canada, 1994)
Long description for Figure 7
This bar graph shows the following information:
|Surficial materials||Percent of Arctic Ecozone+|
|Fine grained (Glacio) Lacustrine||0|
|Coarse grained (Glacio) Lacustrine||0|
|Coarse grained (Glacio) Marine||2|
|Lag (Glacio) Marine||2|
|Fine grained (Glacio) Marine||5|
The Arctic landscape is shaped by the complex relationships between climate, permafrost, and vegetation (Walker et al., 2003). Permafrost is defined as soil, rock, or sediment that remains at or below a temperature of 0°C for at least two consecutive years. It can contain ice as pore ice, ice wedges, or massive ice bodies. In the Arctic Ecozone+ permafrost is continuous and may be several hundred metres thick and have temperatures colder than –5°C (Heginbottom et al., 1995; Smith et al., 2001a).
Above the permafrost, the soil profile has an active layer that is seasonally frozen. The thickness of the active layer is highly variable depending on aspect, vegetative cover, and the ambient temperature regime. Moisture and gas fluxes are generally confined to this seasonally thawed active layer. Thus permafrost’s presence and influence is through physical and chemical processes in the active layer which operate as thermal and hydrological gradients. Plant roots and nutrients accessible to plants are largely confined to the active layer, as are the burrowing activities of invertebrates and vertebrates.
Humans in the Arctic Ecozone+
(This section is based on the following sources, with additional references as noted, McGhee, 1978; Bone, 1992; Black, 2002; Bonesteel, 2006; Nunatsiavut Government, 2009; Freeman, 2012; Natcher et al., 2012)
The Arctic Ecozone+ is sparsely populated, though its human population approximately tripled between 1971 and 2006 (Figure 9). The majority of the people are Inuit and the ecozone+ covers most of the four Inuit regions established through Canadian comprehensive land claim agreements: 1) the Inuvialuit Settlement Region (parts of the Yukon and the Northwest Territories); 2) Nunavut; 3) Nunavik (part of Quebec); and 4) Nunatsiavut (part of Newfoundland and Labrador). The land and marine areas covered by comprehensive land claim settlements in the ecozone+ are shown in Figure 10. While the terms of these agreements differ, they all provide for wildlife and habitat management by boards or councils with input from Aboriginal regional and local governance bodies, as well as federal and territorial governments. These co-management regimes are a distinguishing element of the Canadian Arctic and central to all aspects of ecological management, monitoring, and research in the Arctic Ecozone+.
SSource: based on population data for the three Arctic ecozones compiled from Statistics Canada (Statistics Canada, 2000; Statistics Canada, 2008b)
Long description for Figure 9
This bar graph shows the following information:
Settlement and economic history
Humans have occupied the western side of Beringia for at least 30,000 years and are known to have crossed into North America by 14,000 years ago, either following the retreat of the Cordilleran and Laurentide ice sheets or routes along the coast of Alaska (Kitchen et al., 2008). About 4,500 years ago the first wave of human inhabitants is thought to have moved across the eastern Arctic from Alaska (Friesen, 2004). Based on distinctive material culture, archeologists have assigned these people several separate cultural phases, the last of which is the Dorset culture. The Pre-Dorset culture was defined by a highly mobile lifestyle, limited tool kits, and limited reliance on food storage (Milne et al., 2012). Between 1000 and 1200 AD the second wave of settlers, the Thule, entered the Arctic from the Bering Strait region. The Thule were the ancestors of the contemporary Inuit and are traditionally thought to have displaced the Dorset culture (Friesen and Arnold, 2008). There is some evidence of intermixing (Helgson et al., 2014). By 1200 to 1300 AD, the Thule had spread across the Arctic to Greenland (Kallreuth et al., 2012). With the Thule culture came a shift to a more sedentary way of life, the introduction of boats, and a high diversity of specialized tools (Anderson, 2004).
Europeans (the Norse) sailed to Baffin Island around 1000 AD, but did not settle. European exploration recommenced in the 16th century, with Martin Frobisher claiming Baffin Island for England in 1576. The 19th century commercial whaling boom began the widespread transition to a mixed economy for the Inuit, with guns and dry goods being received in exchange for food and guiding services. Whaling had a profound influence on the Inuit, introducing alien diseases including typhus that spread rapidly and killed many people, and through impacts such as depletion of caribou for food for whaling crews, and an increasing reliance on European goods. Whaling was replaced in the early 20th century by the fur trade boom centred around the Hudson’s Bay Company trading posts and accompanied by an expansion of the mixed economy to include trapping and trading for a wider selection of goods. Government posts, policing, and missions were also established. The Inuit altered their hunting practices and annual cycles to adapt to these changes, and suffered economic and social hardship when the price of furs fell drastically in the 1930s and 1940s.
The Arctic attained a profile in North American defense during and after the Second World War (Figure 11 ). Mineral and energy resources began to be discovered and developed, creating a perception of the Arctic as a vast storehouse of resources. Starting in the 1950s the Canadian government resettled many Inuit into permanent communities, provided social services and required education in English, with the aim of improving the well-being of the Inuit through assimilation into the mainstream culture and economy. Residential schools were established under this policy of assimilation, leading to many instances of physical and emotional abuse, with lifetime consequences for the children of this era and for their communities. In the 1960s Inuit political movements arose in response to the family, social, cultural, and economic disruption that had resulted from decades of upheaval and hardship. Goals included restoring and reviving native languages and traditional lifestyles and regaining control over the management of lands and resources, the basis of Inuit culture and prosperity.
The Distant Early Warning (DEW) Line
Construction of the Distant Early Warning (DEW) line of 42 radar stations along the Canadian Arctic coast in the late 1950s provided some wage employment and introduced new technology (such as snowmobiles) to more communities. The rapid influx of shifting populations of people with very different cultural perspectives and the rapid introduction of new technologies, infrastructure, and goods also resulted in major social upheaval to the Inuit (Neufeld, 2002). The DEW Line was a joint project of the U.S. and Canada and a product of the Cold War: its purpose was to provide early warning of attacks on North America from the Soviet Union (Neufeld, 2002; Bonesteel, 2006). In the early 1960s, the smaller sites were decommissioned. The remaining 21 sites were operated by Canada to various dates up to 1993. The new North Warning System, which includes eight of the old radar stations, replaced the DEW line in the 1990s. Clean-up of the sites has been plagued by problems and escalating costs related to contamination, including from PCBs, DDT, hydrocarbons, and lead. Clean-up, nearing completion in 2013, has a goal of keeping contaminants out of the Arctic food chain and will be followed by a 25-year monitoring plan (DND, 2001; Aglukkaq, 2012); there have been local instances of contamination of biota and continuing concerns expressed by Arctic residents about impacts (e.g., Gamberg et al., 2005b).
The Inuvialuit Settlement Region
In the Western Arctic, commercial bowhead whaling stations in the latter part of the 19th century introduced infectious diseases that led to epidemics and the death of most indigenous Inuit (the “Mackenzie Eskimos”). Iñupiat from Alaska migrated eastward and integrated with the remaining residents in the Mackenzie Delta region. The Inuvialuit, as an Inuit group, are a newer group, formed by the Inuit from the Delta region uniting with those from Banks and Victoria islands in the 1970s to work together as a political and economic force, partly in response to the pressures imposed by Beaufort Sea oil and gas development. The Inuvialuit were successful in negotiating the Western Arctic agreement (Inuvialuit Final Agreement) in 1984.
The Inuvialuit Settlement Region (ISR) covers 435,000 km2 of land and sea (Figure 10). Most of the land is part of the Northwest Territories, but it also includes the Yukon North Slope, designated under the agreement to fall under a special conservation regime. Known mineral deposits in the ISR include petroleum, natural gas, zinc, and nickel. The Mackenzie River is a commercial artery and there is an all-season road to Inuvik and seasonal ice roads from Inuvik to Aklavik and Tuktoyaktuk. Air access is important for all communities. National parks form 29% of the ISR.
Inuvik, the largest community of the ISR and its main administrative centre, and the town of Aklavik, are just south of the Arctic Ecozone+. Within the ecozone+ are the communities of Tuktoyaktuk, Sachs Harbour, Ulukhaktok (Holman), and Paulatuk. Tuktoyaktuk was an important centre for offshore oil drilling in the 1970s and 1980s. There are no year-round communities in the Yukon portion of the ISR. Wage employment is the main source of income throughout the ISR, but unemployment is high. The major wage employment is in the public sector. The Inuvialuit Development Corporation includes several firms involved in oil and gas activity and the Inuvialuit are involved in the development of a range of new economic initiatives.
Oil and gas activities were important in the region in the 1970s and 1980s. Renewed interest and activity in onshore and offshore exploration and in development of onshore fields in the Mackenzie Delta region began in the early 2000s. In addition to the conventional petroleum resources, there is good potential for unconventional petroleum resources in this region. Exploration has been undertaken on gas hydrate accumulations beneath the Mackenzie Delta, and, more recently, industry is beginning to evaluate the potential of extensive shale formations in the Mackenzie Valley.
Hunting remains important economically, socially, and culturally and provides the main source of meat. Caribou (Rangifer tarandus), muskoxen (Ovibos moschatus), polar bears (Ursus maritimus), beluga (Delphinapterus leucas), and seals are hunted, with variation in main species harvested from community to community. The introduction of gill nets in the 1830s led to the development of river fisheries, which remain important in Inuvialuit subsistence economies. Other economic activities include trapping, tourism, guided sport hunting (for polar bears and muskoxen), and the production of arts and crafts.
The Nunavut project, first proposed in 1976, eventually led to a final Nunavut Land Claims Agreement in 1993 and the creation of the Territory of Nunavut in 1999. Nunavut, at 2.12 million km2, is the largest political unit in Canada, covering one-fifth of the country’s landmass. The population is about 84% Inuit (Statistics Canada, 2008a). There are 29 communities, of which Baker Lake is the only one that is located away from the coast. Iqaluit is the only city (population 6,669 in 2011 (Statistics Canada, 2012)) and the territorial capital. Household income is from the wage economy, government transfers, and the land-based economy. Government is the largest employer, with about 50% of jobs being in the public sector. Tourism is the fastest growing economic sector but remains relatively small. There is no permanent road infrastructure linking Nunavut with the rest of Canada. The longest road, 21 km in length, connects Arctic Bay and Nanisivik.
Mineral resource exploration and extraction have historically dominated the private sector economy. Nanisivik, a lead-zinc mine on north Baffin Island, opened in 1976 and closed in 2002. The North Rankin Nickel Mine operated from 1957 to 1962 in Rankin Inlet. Polaris, a lead-zinc mine on Little Cornwallis Island, operated from 1982 to 2002. Lupin Gold Mine opened in 1982 and closed 2005. Nunavut’s only diamond mine, Jericho Mine, operated from 2006 to early 2008 and is currently (2013) in care and maintenance mode. Lupin and Jericho mines are close to the border with the NWT and were served by a winter road from Yellowknife when they were operational. Meadowbank Gold Mine, near Baker Lake, started production in 2009 while Doris North Gold Mine, near Bathurst Inlet, obtained mine approval in 2006 but has yet to operate and has since gone into care and maintenance. The Mary River iron ore deposit on Baffin Island was approved for mine development in 2012 but is currently (2013) seeking an amendment to allow for changes to the project and has yet to go into production. Potential mines in the review process are the Meliadine gold project near Rankin Inlet and the Kiggavik uranium project near Baker Lake. A gold deposit at Back River, east of Lupin, is currently in the advanced exploration stage and other major exploration projects include base metal deposits at Izok Lake, High Lake, and Hackett River (west, north and east of Lupin, respectively).
Nunavik covers 443,684 km2, about one-third of the province of Quebec. The 2006 population was 11,627, 90% Inuit (Statistics Canada, 2008a). There are 15 communities along the coasts of Ungava Bay and the eastern shore of Hudson Bay. The major settlement is Kuujjuaq.
Pre-Dorset and Dorset artifacts are found along the coasts of both Hudson and Ungava bays, dating to around 4,000 years ago. These cultures were succeeded by the Thule culture. Commercial whaling started in the 18th century in Hudson and Ungava bays, and fur trade rose in the region. The hunter-gatherer economy was little changed for many years, until consumer goods became more widely traded. The distribution of villages today is influenced by locations of Hudson’s Bay Company trading posts.
Major sources of energy and minerals attracted attention in the 1960s and 1970s. Hydro development proposals for James Bay and La Grande River (in the Taiga Shield Ecozone+) led to widespread opposition by the Inuit and the Cree and the signing of the James Bay and Northern Quebec Agreement in 1975. The agreement addressed outstanding issues of indigenous rights and formed the Kativik Regional Government and Makivik Corporation, the political and fiscal arms of the Inuit of Nunavik. In 2006 the Nunavik Inuit Land Claims Agreement was signed. Applying to eligible beneficiaries of the James Bay and Northern Quebec Agreement, it provides protection for land and resource rights and mechanisms for involvement in management on areas traditionally used by the Inuit of Nunavik, including marine areas and lands in Nunavut and northern Labrador (Figure 10).
The lineage of the Inuit of Nunatsiavut--the Sikumiut, ‘people of the sea’--traces back to Thule Eskimos who migrated from Baffin Island. The long European settlement history (since the 17th century) in coastal Labrador influenced the Sikumiut through intermarriage and cultural change. Traditionally the economy was based on sea mammal harvest and caribou hunting, supplemented with fishing and hunting waterfowl. When whales were depleted after 1800, seals became increasingly important. In the 1930s government began to supplant the influence of the fur trade posts and missions. Confederation in 1949 brought profound social change but no recognition of the Sikumiut as Aboriginal peoples under the Canadian constitution. The Sikumiut were the last Inuit in Canada to achieve the major step of a land claim settlement.
The Labrador Inuit land Claims Agreement received Royal Assent in 2005. The agreement covers about 72,500 km2 of land and 49,000 km2 of sea (Figure 10). Intended to create a stable environment for land use and investment and contribute to development of the Labrador Inuit, it also led to formation of Torngat Mountains National Park and the establishment of mechanisms for involvement in environmental assessment and for co-management. A regional Inuit government, the Nunutsiavut government, was created by the agreement and provisions were established for sharing of revenue from mineral resources.
The population of Nunatsiavut was 2,427 in 2006, 89% Inuit (Statistics Canada, 2008a), mainly in five communities: Nain, Hopedale, Makkovik, Postville, and Rigolet, all located south of the Arctic Ecozone+.
Description of the Condition of the Arctic Ecozone+
- Abiotic drivers
- Ecosystem functions/processes
- The Cryosphere
- Rivers and lakes
- Natural disturbance
- Community and population dynamics
- Nutrient cycling
- Carbon storage and release
- Primary productivity
- Human stressors on ecosystem functions and processes
Introduction to climate-related drivers
In the Arctic, as elsewhere, ecosystems are determined to a large extent by abiotic drivers such as mean and extreme seasonal temperatures and precipitation, number of degree-days and frost-free days, insolation (amount of sunlight reaching the surface), wind and albedo (amount of sunlight reflected from the surface, rather than being absorbed). These factors, plus soil moisture, determine the extent of permafrost (perennially frozen ground), which limits root penetration and has other profound effects on plants and animals. Terrestrial habitats are further influenced by microclimates determined by such physical features as slope, aspect, soil type, and proximity to large water bodies.
The combined influence of these abiotic drivers creates an environment unfavourable to tree growth and generally defines Arctic ecosystems as “tundra,” a mosaic of wet and dry habitats dominated by varying proportions of mostly perennial grasses, forbs and low-growing shrubs. Because of the variety and dominance of shrubs in the heather family (Ericaceae), shrublands are often called “shrub-heath” or “dwarf shrub-heath” habitats, although other families, such as willows (Salicaceae) and dwarf birches (Betulaceae), are also prominent. Spring growth, flowering, and seed-setting periods are brief, and animal life histories are closely linked to vegetation phenology. Tundra closely follows the limits of continuous permafrost, while discontinuous permafrost characterizes the transition between tundra and taiga.
Large-scale changes in abiotic drivers, principally temperature and precipitation, can have profound effects on distribution of tundra, and on the distribution and composition of habitats within it. For example, as noted in the introduction, a warming period beginning about 8,000 years ago changed the Northern Arctic from polar desert to dry tundra.
The Earth has warmed in recent decades, largely due to greenhouse gas emissions. The degree of warming has been unprecedented since earlier interglacial periods in the Pleistocene Epoch (IPCC, 2007). The Arctic Climate Impact Assessment (ACIA, 2005) and the IPCC Assessment Report 4 (IPCC, 2007) both reported that surface air temperatures in the circumpolar Arctic have increased over the past few decades at almost twice the rate of the global mean, with warming being particularly pronounced since the early 1990s. Figure 12 maps temperature anomalies for 2001–2005 relative to 1951–1980. This illustrates the greater warming in the circumpolar Arctic and also the regional trend of greater warming in the western part of the North American Arctic. Figure 13 shows the trends in mean annual global and Arctic temperatures from 1880 to 2012.
Source: Overland et al. (2007), based on Hansen et al. (2006)
Long description for Figure 12
This map shows mean surface temperature anomalies for 2001–2005 relative to 1951–1980 in the Northern Hemisphere. The map illustrates there was greater warming in the circumpolar Arctic (generally 1.2–2.1°C) and also the regional trend of greater warming in the western part of the North American Arctic.
Land-surface air temperature anomalies. Based on meteorological station measurements, the annual means include the Canadian stations discussed below. Data are anomalies, calculated on the base period 1951 to 1980.
Source: ecozone+ data provided by authors of Zhang et al. (2011)
Long description for Figure 13
This graphic represents a line graph showing the following information:
A major consideration for the Arctic Ecozone+ is the role of feedback mechanisms in amplification of climate change in the Arctic--particularly as these mechanisms involve changes in ecosystems. For example, reductions of surface albedo from the retreat of sea ice has led to enhanced warming of the lands peripheral to the Arctic Ocean (Walsh, 2008); shorter snow cover duration has resulted in greater surface warming and increased shrub growth, further reducing surface albedo and leading to additional regional warming (Sturm et al., 2001; Chapin III et al., 2005; Walsh, 2008). Increased shrub growth also increases the ability to retain a deeper snow cover which impacts the ground thermal and hydrological regimes, microbial activity, and nutrient and carbon cycling (Myers-Smith, 2007).
Observations by Canadian Inuit about climate change are documented in many sources, including the Arctic Climate Impact Assessment (ACIA, 2005) and studies based on interviews in Inuit communities (Kassi, 1993; McDonald et al., 1997; e.g., Fox, 1998; Fox, 2004; Huntington et al., 2005). A common theme is the increased variability of the climate. For example, people in Igloolik report that weather is more unpredictable and sudden weather changes have forced hunters to spend extra unplanned nights on the land (Ford et al., 2006). People in Nunavut report experiencing warmer temperatures year-round accompanied by changes in the length and timing of the seasons (Government of Nunavut, Department of Environment, 2005). The communities of the Inuvialuit Settlement Region have been observing changes associated with warming in their region for a longer period than those living in the Eastern Arctic communities and these changes appear more pronounced (Furgal et al., 2002; Nickels et al., 2002; Furgal and Seguin, 2006). Milder winters, warmer summers, a shorter fall and a slower and later freeze-up were among the many weather-related changes observed by Banks Island Inuvialuit. Fluctuations in the seasons were noted, particularly the earlier arrival of spring (Ashford and Castleden, 2001).
Climate variability and change is the unifying theme of this report and the material outlined above is expanded upon in the relevant sections throughout the report.
Climate trends since 1950
This section is based on records from Canada’s network of climate stations, analysed for the country as a whole, and on an ecozone+ basis for this report. Results of the analysis for Canada are presented in the ESTR thematic report Canadian climate trends, 1950–2007 (Zhang et al., 2011). A base or reference period (1961–1990) was used to construct both a regionally averaged and station-by-station anomaly series of precipitation and temperature trends. The results of these trend analyses for the Arctic Ecozone+ are summarized in Table 2 and presented and discussed below in relation to the Canadian trends from the same analysis and with reference to Canadian and circumpolar Arctic climate trends discussed by other authors. Zhang et al. (2011) also analysed trends related to snow, reported in the section on Ecosystem functions/processes (page 25).
Temperature and precipitation stations are widely distributed across the Arctic Ecozone+, but station locations are biased to coastal areas and the station density is low. The ecozone+ average values for temperature and precipitation cover at least three distinct climatic regions [Canadian Arctic Archipelago (also known as the Arctic Archipelago), Western Arctic, and Eastern Arctic] and long-term variations of climate in these regions may be different. This means that ecozone+ averages are not necessarily representative of status and trends at different locations within the ecozone+. Maps showing trends at individual stations are included to provide a better indication of regional trends.
See Zhang et al. (2011) for more information on methodology of this analysis.
|Climate variable||Trends from 1950 to 2007|
|Temperature||Significant increases in ecozone+ seasonal average temperatures in summer and fall. Significant increases at many climate stations in all seasons (Figure 14). No significant cooling trends at any station in any season.|
|Precipitation||Increased significantly in all seasons across the ecozone+, with the greatest increase being in winter. Significant increases at many climate stations in fall, winter and spring (Figure 16). Significant increase at only 1 station in summer.|
Annual mean temperature increased about 1.4°C between 1950 and 2007 across Canada, with warming trends evident in most parts of the country. Although for the country as a whole the greatest temperature increases occurred during winter and spring, warming in the Arctic Ecozone+ was most consistent in summer and fall, with significant temperature increases of 0.9°C in the summer and 1.7°C in the fall. These means mask the greater changes that have occurred at some locations, as can be seen on the maps showing magnitude and significance of temperature trends for each Arctic climate station included in the data set (Figure 14).
Growing season in this national analysis is defined as the period during which the mean daily temperature exceeds 5°C (based on a running average of 5 days). Significant increases in growing season length were observed in many places in Canada, largely due to earlier starts to the growing season. The selection of 5°C as the minimum temperature is arbitrary and the characteristics of the actual growing season vary among species, with cold-adapted tundra plants starting growth just above the freezing point. As this 5°C threshold for the start of the growing season was not met at many Arctic locations, trends in growing season in the Arctic were not detected through this analysis. The records, however, show significant increases in temperatures at many Arctic stations, across all seasons (Figure 14). Temperature measurements taken continuously since 1995 as part of the International Tundra Experiment (ITEX) site at Alexandra Fiord, Ellesmere Island, show an increase in the growing season length (measured as days above 0°C) of about 1.5 days per year over the 12 years of measurement (Figure 15). Plant productivity has increased across the ecozone+, indicating both an increase in peak productivity and in growing season length (Zhou et al., 2001; Goetz et al., 2007) – see the section on Primary productivity (page 71).
Growing season is defined here as the number of days with mean daily air temperature greater than 0°C. Air temperature was measured at 10 cm above the ground surface. Data are means (n=2) with range bars. The trend line is a linear regression showing an increase of ca. 1.5 days per year over the 12 years.
Source: G. Henry, unpublished data
Long description for Figure 15
This line graph shows the following information:
|Year||Average growing season length (days)|
Precipitation increased at the majority of stations across Canada between 1950 and 2007, with stations in Arctic Canada experiencing the greatest relative increases. Precipitation increased significantly in all seasons in the Arctic Ecozone+, showing increases from 16 to 73% of the base period mean. Fall, winter, and spring precipitation showed signs of a regime shift in the mid- to late 1970s while summer appeared to be more stable--though still significantly increasing, but by only 16% of the base period mean. The abrupt change in precipitation during the fall, winter, and spring in the late 1970s may be related to a change in the atmospheric circulation, specifically a shift in the Pacific North America teleconnection pattern that occurred in 1976 (Brown and Braaten, 1998). Figure 16 shows the magnitude and significance of changes in precipitation for each season at each station.
While it is not yet clear what is responsible for the precipitation changes in Canada, a study by Min et al. (2008) suggests that precipitation increases over Northern Hemispheric high latitudes (north of 55°N) may have been a result of anthropogenic influences on climate. The observed trend towards more precipitation in Canada, especially in the North, is consistent with climate model projections of future changes in precipitation, and is thus likely to continue in the future (Zhang et al., 2011). As projected increases in precipitation are strongly correlated with projected warming, precipitation increases in the Arctic are likely to be much larger than global mean increases (Kattsov et al., 2007).
The Arctic Climate Impact Assessment (2005) reported an increase in precipitation of 1.4% per decade from 1900 to 2005 for the Arctic (60 to 90°N)--considerably greater magnitude of change than has been observed in this analysis for the Arctic Ecozone+. The larger increases in winter precipitation observed in this analysis are consistent with both model projections and with observations for the Arctic as a whole (ACIA, 2005; Lemke et al., 2007).
The section is based on the ESTR thematic report Large‐scale climate oscillations influencing Canada, 1900–2008 (Bonsal and Shabbar, 2011) with the addition of information specific to the Arctic, as noted.
Observed trends and variability in Canadian climate are influenced by large-scale atmospheric and oceanic oscillations known as teleconnections. There are several identified teleconnection patterns used to describe various circulation features across the globe (Bonsal and Shabbar, 2011). Relationships between these teleconnections and Canadian climate are strongest during the cold season (late autumn through spring) although some connections with summer conditions have been identified. In addition, they have strongest and more consistent impacts on temperature variables and, to a lesser extent, on precipitation-related factors. The Arctic Ecozone+ is particularly influenced by the Arctic Oscillation (AO) and by North Pacific patterns of oscillation, for example the Pacific Decadal Oscillation (PDO), though other climate oscillations such as the El Niño/Southern Oscillation (ENSO) also affect Arctic temperatures.
The impact on cold-season temperatures over various regions of Canada has led to significant relationships between teleconnection patterns and several spring climate-related variables. These include the duration of lake and river ice in western Canada (Bonsal et al., 2006), the timing of snowmelt and spring peak streamflow across western North America (Stewart et al., 2005), and the timing of spring runoff in the Mackenzie River (related to the PDO and other indices, but not the AO or ENSO) (Burn, 2008).
Indices that represent these large-scale oscillations are correlated with fluctuations and cycles in Arctic ecosystems. The linkages are not always straightforward, with ecosystem variability often being linked with more than one teleconnection, and with the relationship varying from region to region. Some examples of associations with Arctic biotic communities:
- High values of the AO are associated with periods of reduced plant growth and reduced reindeer populations on Svalbard (Norway) (Aanes et al., 2002).
- Large-scale climate patterns play a role in caribou population fluctuations, but the associations vary from region to region: in Alaska, the PDO is correlated with changes in western Alaska herds, while herds in northeastern Alaska are more affected by the AO (Joly et al., 2011).
- Mortality of greater snow goose young during summer on Bylot Island was shown to be lower at both extremes of the AO: in years when the index is very positive (low summer temperatures) and in years when it is very negative (high summer temperatures) (Dickey et al., 2008).
Arctic Oscillation Index
The Arctic Oscillation (AO) represents atmospheric circulation variability over the Northern Hemisphere north of the tropics where sea-level pressure over the polar regions varies in opposition with that over middle latitudes (about 45°N) (Thompson and Wallace, 1998). It is strongly correlated with the North Atlantic Oscillation (NAO) that influences climate in eastern North America and Europe. When the AO is negative, winter temperatures are lower, ocean waters are fresher, sea ice coverage is greater, and sea ice is thicker than when the AO is positive. During 1950–1980, the AO shifted frequently between positive and negative, and remained positive from 1989 to 1995 (Richter-Menge et al., 2006) (Figure 17). Features of the positive AO phase include warmer-than-average conditions over large regions in northern Europe and Asian Russia, with opposite anomalies in the eastern half of North America (Buermann et al., 2003). The AO has been linked with a variety of physical and ecological features in Arctic and subarctic latitudes. For example, positive AO levels are associated with less winter and spring season snow extent, especially in Eurasia (Bamzai, 2003), earlier melt onset, especially in the European Arctic (Tedesco et al., 2009), higher river discharge to Bering Strait, but lower river discharge to Hudson Bay (Déry et al., 2005), and declines in some Alaskan caribou herds, including the Porcupine Herd whose range extends into Canada (Stenseth et al., 2003).
The different phases in large-scale teleconnections often lead to larger-amplitude regional responses in observed climate. As a result, a number of the observed 20th-century climate changes can be attributed, at least in part, to changes in these various teleconnection patterns (Solomon et al., 2007). For example, Hurrell (1996) found that the North Atlantic Oscillation, ENSO, and variations in North Pacific circulation collectively explained a significant portion of Northern Hemisphere winter temperature variability during the 20th century. Specifically for North America, the mid-1970s climate shift to positive Pacific Decadal Oscillation and more frequent El Niño events appear to have led to contrasting changes across the continent, with the west warming more than the east for a period thereafter (Trenberth et al., 2007). This shift to a positive PDO has therefore also been associated with the trend toward warmer winter and spring temperatures over western Canada, including in the Arctic.
It is not currently known how increases in greenhouse gas concentrations have affected the occurrence of these large-scale climate oscillations. Furthermore, the effects of projected future climate change on the major teleconnection patterns remain uncertain since there is a lack of agreement among the various climate models concerning the future frequency and structure of large-scale atmospheric and oceanic modes. Climate oscillations do not account for recent warming patterns in the Canadian and circumpolar Arctic (Figure 18).
In 1973, Dunbar (1973) published a landmark paper, “Stability and fragility in Arctic ecosystems”. He examined the apparent dichotomy that Arctic ecosystems are susceptible to damage, even though species that live there all year are extraordinarily tough. Others migrate huge distances and habitats undergo extreme annual fluctuations in living conditions. Arctic terrestrial animals, such as lemmings (), can crash to extremely low population levels or even be locally extirpated, but can recover as long as there is a vast area available in which to harbour recolonizing stocks. Species are protected from invasion by ecological equivalents, as well as their predators and pathogens, and by climatic limitations that maintain tundra habitats. He pointed out that “stability” has to be defined in terms of coherence and continuity of annual and multi-annual cycles, and that “fragility” relates not so much to low resistance to small-scale physical damage as to restricted growth and turnover rates that limit resilience. He gave this example of vulnerability due to slow turnover:
“If an Arctic lake full of char or lake trout is fished out in a season or two, which has happened many times, the damage is either permanent or will take many years to repair... Lakes, especially small lakes, in the Arctic, are probably the most vulnerable part of the landscape... Many of them harbour only one species of herbivore (a copepod crustacean), and one species of primary carnivore, usually a mysid or an amphipod. These are analogous to the lemming of the tundra, but without the protection of large geographic scale. They represent extreme cases of ecosystem simplicity, in which the removal of one species means ecological disaster.”
Since then, our understanding of resilience has improved, so that, for example, we know that, when populations drop below a minimum viable population level (a threshold), their recovery becomes dependent on conditions that enhance resilience for that species: for example, low mortality from predation or harvest, or sufficient range size and sufficient connectivity to allow other populations to act as sources for recolonization. Stochastic events such as unusual winter storms, industrial disturbance, or introduction of a new predator or pathogen may pose high probabilities of extinction or extirpation. Peary caribou (Rangifer tarandus pearyi) are a case in point, as discussed in the Ecosystem composition - Peary caribou section (page 107). After a decade of intensive research, scientists are beginning to understand how a changing climate may destabilize Arctic ecosystems--and, indeed, have already detected signs of destabilization--but we have much to learn. The science on cumulative impacts, including understanding of interactions between climate change and industrial disturbance, is in its infancy. The following sections focus on processes that may indicate or foretell ecosystem perturbations. Two important characteristics of the Arctic Ecozone+ strongly influence the resilience of its ecosystems to perturbation: 1) the huge scales (temporal and geographical) at which most Arctic ecosystems operate (at least for vertebrates) in the processes of annual movements, population cycles, and recolonization events; and 2) the dominant evolutionary role of winter in driving adaptive traits and habitat structure. Understanding the influence of scale the specific adaptations and habitat structures characteristic of Arctic wildlife, in the light of current change, is essential for conservation and enhancement of Arctic ecosystem resilience.
Changes to the cryosphere (permafrost, snow, sea ice, glaciers, and lake and river ice) could also be classified as abiotic drivers (secondary to the primary climatic drivers of temperature and precipitation). We include these trends in the Ecosystem processes section to highlight and discuss their linkages with ecosystem disturbance and other ecosystem processes such as hydrology and primary productivity. Rapid shrinking of all aspects of the cryosphere has been a dominant trend in Arctic ecosystems that began over 30 years ago and is particularly evident over the past decade. The implications of this loss of frozen water and frozen ground, and the interconnections with other ecosystem processes, are profound and, in many cases, poorly understood. A major source of the increasingly high rate of change cryosphere shrinkage is the positive feedback resulting from reduction in surface reflectivity (albedo) due to loss of ice and snow. This reduction in albedo increases the capacity to absorb and store heat in Arctic ecosystems, enabling more melting. In the face of projections for continued global warming, further losses and changes in timing and other characteristics of ice and snow in their various forms are expected (AMAP, 2011; Jeffries et al., 2012).
The sections below on permafrost trends are based on Trends in permafrost conditions and ecology in northern Canada (Smith, 2011), an ESTR national thematic report. Some information has been updated, as noted.
Information on recent trends in permafrost temperatures in the Arctic Ecozone+ comes from a number of monitoring sites from the Western Arctic to the Eastern Arctic and the High Arctic. In general, changes in shallow permafrost temperatures over the last decade are greater in the Arctic Ecozone+ compared to areas below the treeline (taiga and boreal ecozones+). Tundra vegetation, with less of a vertical structure than forest, allows more direct light penetration to the soil and provides less insulation against re-radiation of absorbed heat. The presence of colder permafrost also means that phase change and the presence of unfrozen water do not obscure the climate signal. There is, therefore, a more direct link between changes in air temperature and changes in permafrost temperature.
Permafrost temperature data collected since the 1980s from the northern Mackenzie Basin (Smith et al., 2005b) indicate that warming of permafrost has occurred since the early 1990s. On the Tuktoyaktuk Peninsula for example, at a depth of 28 m, permafrost temperatures increased between 1990 and 2002 at a rate between 0.02 and 0.06°C per year. Analysis by Burn and Kokelj (2009) indicates that near-surface ground temperatures in the tundra uplands of the Mackenzie Delta region increased by 1 to 2°C from the early 1970s to 2007. Modelling analysis conducted for a permafrost monitoring site at Herschel Island (in the Beaufort Sea off the Yukon coast) indicates that permafrost temperature at a depth of 20 m has increased by 1.9°C over the past 100 years (Burn and Zhang, 2009). Recent field observations at this monitoring site also indicate an increase in active layer thickness since 1985.
Central Southern Arctic
Permafrost temperatures to 3 m depth have been collected since 1997 at Baker Lake, Nunavut. Between 1997 and 2007 a general increase in thaw depth (Figure 19) has been observed, although there is some interannual variability within the short record (Smith et al., 2005b; Throop et al., 2008). The largest increase in thaw depth occurred between 1997 and 1998 and this was related to the longer thaw season in 1998 (Smith et al., 2001b).
Permafrost temperature data collected since 1978 at CFS Alert Nunavut can be used to characterize trends in permafrost in the High Arctic. Although air temperatures have increased since the 1980s, distinct warming in shallow permafrost temperatures has only been observed since the mid-1990s. Between 1994 and 2001, an increase in permafrost temperatures of about 0.15°C per year occurred at a depth of 15 m (Smith et al., 2005b). Although some cooling of permafrost was observed between 2000 and 2002, more recent data indicates that warming of permafrost is continuing.
Figure 20 shows trends in annual mean temperatures at two depths and two boreholes up to 2011. Increases in winter temperatures appear to be the main factor in the recent ground temperature increases, especially at High Arctic sites with relatively low snow cover (and thus little insulation of the ground during winter) (Smith et al., 2012).
Shallow permafrost temperatures at Iqaluit also began to warm in 1993, with warming continuing through the 1990s (Figure 21). Temperatures of shallow permafrost (measured at a depth of 5 m) increased at a rate of 0.4°C per year between 1993 and 2000. A similar trend has been observed in northern Quebec, where cooling of about 0.10°C per year was observed between the mid-1980s and mid-1990s at a depth of 10 m (Allard et al., 1995). An increase in air temperatures commencing in 1993 in northern Quebec has been associated with warming of permafrost since 1996 to depths of 20 m (Allard et al., 2002; Ouranos, 2004; Chouinard et al., 2007) and an increase in active layer thickness (Brown et al., 2000).
Recent increases in shallow ground temperatures have also been observed in other Arctic regions such as Scandinavia and Svalbard (Isaksen et al., 2007a; Isaksen et al., 2007b; Harris and Isaksen, 2008). Although snow cover is generally thin at these High Arctic sites, its variability can be an important factor affecting the response of permafrost temperatures to changes in air temperature (Smith et al., 2003). Changes in snow cover may counteract changes in air temperature occurring over the same period such that permafrost temperatures may increase in the High Arctic during periods of higher snow cover but lower air temperature (Taylor et al., 2006).
Ecological consequences of changes in permafrost
The observed trends of warming permafrost and increased thickness of the active layer are projected to continue and accelerate as the Arctic climate warms (ACIA, 2005; Intergovernmental Panel on Climate Change, 2007). The implications for ecosystem processes and ecosystem structure are huge. Melting of permafrost is also implicated in feedback loops that: 1) lead to increases in air temperatures in the region through changes in vegetation communities and albedo (Chapin III et al., 2005); and 2) lead to increases in warming at the global scale through changes in the carbon balance of the Arctic landscape. Research provides clear links between permafrost conditions and many ecosystem characteristics. Though it can be hard to demonstrate and measure trends in these interlinked impacts at the level of change to populations and ecosystem services, the observed trends in permafrost are likely now influencing Arctic ecosystem processes, structure, and composition. Warming and melting of permafrost projected for the decades to come will be accompanied by major ecosystem change. The discussion below outlines some of the ways in which permafrost drives and influences ecosystem processes. See also the section on Permafrost disturbance (page 54).
Patterned ground caused by permafrost provides diversity in microclimates for Arctic plants and arthropods (Strathdee and Bale, 1998; Vonlanthen et al., 2008; Walker et al., 2008a). The annual freeze-thaw cycles of the active layer lead to ground patterns such as frost boils and ice-wedge polygons which are a defining feature of the tundra. Frost boils are patches of barren or sparsely vegetated soil where differential frost heave forms distinct patterns of stones and sediments (Walker et al., 2004). Frost boils cycle organic matter and nutrients vertically within the boil (Walker et al., 2004) and are favourable sites for seed germination (Sutton et al., 2006), contributing to the diversity and amount of vegetation available to herbivores (Walker et al., 2001).
One of the most productive tundra ecosystems is non-acid tussock tundra which is a key to migratory tundra caribou herds and their associated predators (Walker et al., 2001). The productivity is driven by the freeze-thaw cycles of the active layer which induces the migration of cations essential to plants. The near-surface ice-rich zone of permafrost is a sink for cations such as calcium and magnesium which reach the active layer and become available for plant growth. When permafrost melts, more of these nutrients are contributed to the active layer (Kokelj and Burn, 2005).
The active layer is kept moist to water-logged by underlying permafrost. Correspondingly, microbial activity is reduced and the microbial release of nitrogen from decomposing organic matter is slow, limiting plant growth (Weintraub and Schimel, 2005). Net primary productivity is related to nutrient availability and is highest on deeply thawed soil, dominated by nitrogen fixers and disturbed by animals that release nutrients.
The depth of the active layer varies from tens of centimetres to 1 to 2 metres and constrains the activities of burrowing animals in summer. Frost boils and frost cracks are preferred sites for lemming burrows (Potter, 1972), possibly because they are warmer than burrows deeper in the active layer. Lemmings spend most of their time in their burrows in summer to reduce the risk from avian predators (Boonstra et al., 1996). The burrows and associated latrines add to the micro sites of disturbed ground and enhanced nutrient availability for plant growth. Other burrowing mammals, including Arctic ground squirrels (Spermophilus parryii), select sites without permafrost, such as river banks (Batzli and Sobaski, 1980).
Permafrost is the main factor influencing where Arctic foxes (Vulpes lagopus) den. The foxes select south-facing, well-drained sites with early snowmelt, conditions indicative of a deeper active layer (Szor et al., 2008). Pregnant polar bears denning on the Hudson Bay coast select thawed peat banks to dig dens. The re-use of the dens over hundreds of years as measured by annual tree ring chronologies suggest the importance of the thawed peat banks for the bears (Scott and Stirling, 2002).
Recent techniques have revealed bacteria and fungal spores within permafrost (Steven et al., 2008). Permafrost’s frozen sediments and ice have carbon dioxide, methane, oxygen, and nitrogen in pore spaces, as well as a thin film of water, and evidence points to active microbial ecosystems within permafrost (Steven et al., 2006). This has implications for global warming. Steven et al. (2006) drew attention to “the considerable microbial biomass believed to inhabit the significant areas of terrestrial permafrost, [and] this may have important implications on global nutrient cycling and biogeochemical processes, such as the C, N, and S cycles”.
This section is based on analysis of ground-based snow cover measurements for Canada as a whole and for the Arctic Ecozone+ as analysed for the ESTR report Canadian climate trends, 1950–2007 (Zhang et al., 2011), with reference to additional studies, as noted.
Snow cover duration has significantly decreased almost everywhere in Canada, with the greatest changes occurring in spring over western and northern Canada. This is a consequence of warmer springs observed over these regions, and is part of a Northern Hemispheric trend toward spring warming and earlier melting of snow and ice (Lemke et al., 2007; Derksen and Brown, 2012).
Across the Arctic Ecozone+, snow cover duration decreased by an average of 9 days in the fall and 8.5 days in the spring from the 1950/51 to the 2006/07 snow seasons. Figure 22 shows the fall and spring trends in snow cover duration at each station. Annual maximum snow depth also decreased: by an average of 13.2 cm over the Arctic Ecozone+ from 1950 to 2007 (Figure 23). This is part of a general tendency to lower maximum snow depths over Canada, but is less significant and less consistent from place to place than the decreases in snow cover duration. In southern Canada, the decrease in annual maximum snow depth is linked to a trend toward a smaller fraction of annual precipitation falling as snow. However, there were no significant trends in this ratio when averaged across the Arctic Ecozone+--in fact, increases in the ratio were recorded at many stations (Figure 24), in line with the greater increase in precipitation during the snow season than during summer (Figure 16). The shorter snow season (later onset of snow cover and earlier melt) is likely offsetting the effects of this increase in cold-season precipitation. This is consistent with analysis by Brown and Mote (2009), which shows that, under a scenario of warming and increasing precipitation, maximum snow accumulation still decreases.
The Arctic snow depth trend results include greater uncertainty than snow cover duration because: 1) snow depths exhibit much stronger local-scale variability than snow cover duration (Brown et al., 2007b); and 2) the snow depth observing network in the Canadian Arctic is sparse and biased to coastal locations. In addition, snow depth measurements are made at open sites near airports that may not be representative of surrounding terrain. With increasing shrubbiness in the Arctic, one would expect to see more snow being retained on the ground in areas of shrub expansion. Snow cover duration, however, is less affected by station conditions.
Circumpolar snow trends
The greatest and most rapid decreases in snow water equivalent and snow cover duration have occurred in high-precipitation maritime regions of the Arctic (Callaghan et al., 2011a). There are also differences between Eurasia and North America: declines in snow depth and snow cover tend to be more recent in Eurasia, mainly after about 1980, while declines are evident since the 1950s in North America. Snow depth, however, is increasing in parts of Eurasia (Callaghan et al., 2011a). Arctic spring snow cover in both Eurasia and North America has decreased greatly, especially over the past decade (Derksen and Brown, 2012) (Figure 25).
Ecosystem consequences of changes in snow
A reduction in duration of snow cover has been shown to have significant positive feedbacks to the Arctic climate system through reduced albedo (Chapin III et al., 2005; Screen and Simmonds, 2012). This mechanism is stronger over tundra than over taiga because of the shallow snowpack, high amounts of incoming solar radiation in the spring melt period, and the characteristics of the vegetation. Tundra vegetation has little vertical structure and therefore little capacity to shade the ground. Chapin et al. (2005) concluded that reductions in the length of the snow-cover period in Alaska over the past few decades have contributed substantially to local atmospheric heating in the summer--increasing heating by about 3 Watts per m2 per decade. Across the entire Arctic region, feedback from changes in snow cover over the period 1970–2000 was simulated to have increased atmospheric heating by 0.9 Watts per m2 per decade (Euskirchen et al., 2007). Natural and anthropogenic impurities in snow are also contributing to reduced albedo. Flanner et al. (2008) show that the surface deposition of black carbon (soot particles) reduces snow surface albedo (a process termed “darkening”), which contributes to melting of snow and ice and increases atmospheric heating.
Snow characteristics, including extent, depth, duration, and timing--but also the characteristics of the snow pack determined by events such as freeze-thaw cycles--strongly influence Arctic terrestrial and freshwater ecosystem processes including permafrost dynamics, hydrology, and primary production. Snow characteristics also directly affect biotic communities and many Arctic species. Observed and predicted effects from changes in snow are assessed in recent Arctic Council projects (AMAP, 2011; Callaghan et al., 2011b; CAFF, 2013) and discussed in several sections of this report.
Status and trends of sea ice are discussed more fully in the Arctic Marine report (Niemi et al., 2010). A summary is provided here because of the importance of sea ice to some of the animals discussed in this report (notably polar bears) and because of the implications to Arctic terrestrial ecosystems of the major changes now occurring in extent and quality of sea ice.
Sea ice extent throughout the year has decreased significantly over the period of record, as measured by remote sensing (Figure 26). The summer minimum ice cover (occurring in September) has declined particularly rapidly in the past few years (Figure 27). Extending the record further back, using ice observation networks and proxy data, indicates that the recent and current rate and extent of decline of summer ice extent is unprecedented at least over the past 1,450 years (Kinnard et al., 2011). Increased summer ice melting has led to a loss of multi-year ice. The percentage of the winter (March) ice cover composed of first-year ice increased from 58% in 1988 to 75% in 2012, while the percentage of ice aged four or more years decreased from 26% in 1988 to 7% in 2012 (Perovich et al., 2012).
Climate modellers project ice-free summers in the Arctic within 30 years (Wang and Overland, 2012). During an ice-free summer, there would still be remnants of land-fast ice and likely ice in sheltered waters between islands of the Canadian Arctic Archipelago (Wang and Overland, 2012). The spatial distribution of sea ice in winter and summer 2012 is shown in Figure 28.
Source: updated from Fetterer et al. (2010) with data from National Snow and Ice Data Center (2013)
Long description for Figure 26
This line graph shows the following information:
The September 2012 minimum ice extent was 48% below the 1979–2000 average. The cooler summer of 2013 resulted in a slightly greater ice minimum extent, similar to 2009 and 24% below the 1979–2000 average.
Source: updated from Fetterer et al. (2010) with data from National Snow and Ice Data Center (2013)
Long description for Figure 27
This line graph shows the following information:
|Year||Extent (million km2)|
Ecological consequences of rapidly changing sea ice
Major consequences of this regime shift in Arctic sea ice for Canada’s Arctic marine ecosystems are covered in the ESTR report Ecosystem status and trends report: Arctic marine ecozones (Niemi et al., 2010). Consequences at the circumpolar and global scale were assessed through recent Arctic Council projects (AMAP, 2011; CAFF, 2013; Eamer et al., 2013).
The extent and quality of sea ice cover has major implications for terrestrial ecosystems and wildlife, as well as for people. Open water has a much lower albedo (reflectivity) than ice, and consequently absorbs more sunlight. Reduction of sea ice has led to warming in adjacent coastal areas (Serreze and Barry, 2011) with consequences for tundra vegetation (Bhatt et al., 2010), as well as increased coastal erosion, affecting coastal habitats and coastal communities, especially along the Beaufort Sea coast (Lantuit and Pollard, 2008; AMAP, 2011). Sea ice is a major platform for travel and hunting for Inuit, and changes over the past decade have affected communities (Niemi et al., 2010; Eamer et al., 2013).
While the polar bear is the species (of those covered in this report) most clearly at risk from the decline of sea ice, terrestrial animals are also vulnerable. This discussion below on sea ice and tundra ecosystems is based partly on the Conservation of Arctic Flora and Fauna (CAFF) report Life linked to ice: A guide to sea-ice-associated biodiversity in this time of rapid change (Eamer et al., 2013).
Winter sea ice provides foxes and wolves with a means to access and colonize remote islands. A study of the genetics of Arctic foxes throughout their circumpolar range showed that the occurrence of sea ice is the main factor in determining how similar fox populations are to one another: the less sea ice, the more genetically distinct are the populations (Geffen et al., 2007). If island populations become isolated due to loss of winter ice, some populations will be at risk of decline or extinction due to loss of genetic diversity and inbreeding (Geffen et al., 2007; Noren et al., 2011). Wolf (Canis lupis) populations on Banks, Ellesmere and Devon islands have declined and been reestablished in the past through colonization over the sea ice from other islands (Carmichael et al., 2008). As with foxes, wolf populations are at risk of reduction if winter ice conditions alter sufficiently to compromise movement over ice. As winter sea ice is expected to persist in the region of the Canadian Arctic Archipelago, this potential impact is more likely to occur in other circumpolar regions, including parts of the European and Russian Arctic.
Changes in sea ice also affect the movements of some caribou herds. Peary caribou have been observed moving from island to island over the sea ice (Jenkins et al., 2011). The Dolphin and Union Herd migrates annually across the Dolphin and Union Strait between Victoria Island and the mainland, one of the areas where the expansion of open water has been most extensive (COSEWIC, 2004). Late freeze-up of the strait will delay migration to the winter ranges.
Both Arctic foxes and wolves feed on marine resources during winter, travelling long distances to hunt and scavenge (Tarroux et al., 2010). Arctic foxes scavenge seals killed by polar bears (Roth, 2003; Geffen et al., 2007) and may also prey on ringed seal pups, as observed on sea ice off the coast of northern Labrador (Andriashek et al., 1985). They may rely heavily on marine foods during years of low lemming abundance (Tarroux et al., 2012). This suggests that the loss of sea ice could affect coastal populations of Arctic foxes during years of low terrestrial food abundance (Pamperin et al., 2008). Wolves are also known to forage on sea ice in winter, likely scavenging seal carcasses, based on an Alaskan study (Watts et al., 2010). The main changes in sea ice and their potential impacts on these primarily terrestrial carnivores are: 1) delayed onset of ice cover in autumn, leading to the carnivores having to spend longer on land in what could be a relatively food-poor situation (for example, when rodents are at low densities and birds have left); and 2) reduced summer ice cover leading to reduced polar bear populations and possibly reduced seal populations, and therefore reduced density of scavenging or hunting opportunities for the terrestrial carnivores when they are on the ice.
Recent studies using satellite tracking show that both gyrfalcons (Burnham and Newton, 2011) and snowy owls (Therrien et al., 2011) spend substantial periods of time far offshore in winter, presumably preying on marine birds and roosting on ice floes. Snowy owls have been observed hunting seabirds that congregate in patches of open water in Hudson Bay in winter (Gilchrist and Robertson, 2000). Satellite tracking over two winters (2007–2009) of nine adult female snowy owls that had been fitted with transmitters while on their summer range on Bylot Island showed that most spent several weeks on the sea ice between December and April. Analysis of high-resolution satellite images showed that they spent most of their time on the ice around open-water patches frequented by seabirds (Therrien et al., 2011). Changes in winter sea ice extent and polynya formation that affect seabirds could also affect the food intake of snowy owls. As owls are small-mammal predators during breeding season, changes in snowy owl populations would affect the tundra food web (Therrien et al., 2011).
Glaciers and ice caps
During the past 10,000 years, glaciers and ice caps--the remnants of the great ice sheets of the last ice age--have been melting, with intermittent periods of re-advance. The last major advance was during the Little Ice Age from the 14th to 19th centuries. Since then, glaciers around the globe have been shrinking, with increasing rates of ice loss since the mid-1980s. On a time-scale of decades, glaciers in some regions (including the Arctic) have shown intermittent advances, but overwhelmingly the trend is to increasingly rapid melt. The outlook, based on climate change scenarios, is for deglaciation of many mountain regions within the 21st century (Gerland et al., 2007).
About 300,000 km2 of glaciers and ice caps (excluding Greenland) drain to the Arctic Ocean, most directly (Dyurgerov and Carter, 2004). Over half of this areal extent of land ice is in Canada’s Arctic Archipelago (Table 3). These glaciers and ice caps fluctuate in mass and extent with temperature and snowfall (UNEP and WGMS, 2008; Colgan and Sharp, 2008), with the strongest linkage being between mass balance and air temperature (Koerner, 2005). A second important control on extent and volume changes of land ice has been iceberg calving that occurs as flux across the grounding line and retreat/advance of tidewater margins. Iceberg calving accounts for a large portion (up to 40%) of mass loss from ice caps in the Canadian Arctic (Burgess et al., 2005; Mair et al., 2009; Mair, 2012).
Glaciers in the Canadian Arctic Archipelago reached their last maximum extent relatively late: in the latter part of the 19th century. Time series of glacier fluctuations are sparse, with the only long-term mass balance series being for the White and Baby glaciers and the Devon, Meighen, and Melville ice caps and Agassiz Ice Field, all in the Queen Elizabeth Islands. Based on these time series and other evidence, there has been a general melting trend for Canadian Arctic Island glaciers since the late 19th century, with the trend slowing down for a period in the mid-20th century (UNEP and WGMS, 2008), and with accelerated loss in the past 25 years. Analysis of ice cores from the Devon Ice Cap (Devon Island), Penny Ice Cap (Baffin Island), and the Agassiz and Prince of Wales ice fields (Ellesmere Island) indicates that the melt rates since the mid-1990s are the highest in at least the last two millennia (Fisher et al., 2012). The longer record for the Agassiz Ice Field shows that this recent melt rate is the highest in 4,200 years and resembles the melt rate during the Holocene thermal maximum over 9,000 years ago (Fisher et al., 2012).
|Ellesmere Ice Shelf||500||Meighen||85|
On Baffin Island, changes in the extent of a few small plateau ice caps (currently less than 50 km2 in size) have been measured since 1975. This record has been extended back in time through the use of chemical and vegetation studies (Miller et al., 1975; Anderson et al., 2008). Some of these ice caps have disappeared already and researchers predict that all small ice caps will have disappeared completely by 2070. Figure 30 shows the measured and inferred rate of decrease in extent since 1940.
Mass loss of glaciers and ice caps due to surface melt and iceberg calving contributes significantly to sea level rise. As the Canadian Arctic Archipelago is the largest glaciated area in the world outside of Greenland and Antarctica, the changes in Arctic Ecozone+ glaciers and ice caps are of global significance.
A study published in 2009 estimates that glaciers and ice caps around the world will contribute 373 ± 21 mm of sea-level rise over the next 100 years, nearly half of the projected rise of 800 mm for the period (Bahr et al., 2009). An analysis of glacier mass balance based on both ground studies and remote sensing (Gardner et al., 2013) estimated that, between 2003 and 2009, glaciers and ice caps in northern Arctic Canada experienced a mass balance loss of a 33 ± 4 Gt/year and southern Arctic Canada glaciers lost 27 ± 4 Gt/year. The total for Arctic Canada is about 28% of the global glacier mass loss (excluding Antarctica and Greenland).
About half of sea-level rise is due to thermal expansion of the ocean. The next biggest source is meltwater from the Greenland and Antarctic ice sheets, followed by meltwater from the world’s smaller glaciers and ice caps (IPCC, 2007). Although the contribution from the smaller glaciers and ice caps will, in the long run, be far less than from the Greenland and Antarctic ice sheets, they react faster to changes in climate and have more immediate effects (Bahr et al., 2009). From 2003 to 2009, global glacier mass loss, excluding Antarctica and Greenland, would have contributed 0.71 ± 0.08 mm/year to sea-level rise, which is about 29% of the observed sea level rise over this period (Gardner et al., 2013).
One of the major ecosystem changes will be the increase in land available for vegetation and the development of tundra ecosystems. This is a process that has been occurring since the late 19th century. The actual area of newly exposed land is not currently large, but could become increasingly important as ice masses disappear from the High Arctic islands (see Table 3 for areas covered). Areas exposed since the glacial maximum at the end of the Little Ice Age (about 1870) are visible as lichen-free zones near the glaciers. Satellite imagery of the Queen Elizabeth Islands (which includes the glaciers shown in Figure 29) shows that the area of terrestrial ice in the islands has decreased by 37% (62,387 km2) since the last glacial maximum (Wolken et al., 2008). A study of ecosystem development in front of a glacier at Alexandra Fiord that has retreated over 300 m since 1980 described patterns of fairly rapid vegetation succession, with diverse plant communities being established, but with sparse plant cover after over 40 years (Jones and Henry, 2003). Areas achieving over 80% plant cover had been ice-free for over 300 years.
Lake and river ice
The duration of Arctic lake and river ice is determined by the dates of autumn freeze-up and spring break-up. In general, rivers experience much more dynamic events than those on lakes, as the formation, growth, and ablation of lake ice primarily occurs in situ, while that on rivers is controlled by the hydraulic effects of flowing water. Overall, however, the timing of such events on both systems is strongly controlled by climate (Prowse et al., 2011).
Analysis of changes in the remains of algae and invertebrates buried in lake sediments has been used by many researchers to identify warming trends and the historical presence/absence of ice cover on northern lakes (Douglas et al., 1994; Korhola et al., 2002; Sorvari et al., 2002; Michelutti et al., 2003; Ruhland et al., 2003; Smol et al., 2005), based on the relationships between ice-cover duration, growing season length, and annual primary production (Rouse et al., 1997; Douglas and Smol, 1999; Keatley, 2007). In general, around the circumpolar Arctic, this evidence points towards warming temperatures and shorter ice durations since the end of the Little Ice Age (around 1850), with greater changes observed in northernmost areas when compared to more temperate locations (Smol et al., 2005). (See section on Changes in algal and invertebrate species assemblages in lakes and ponds on page 96.)
In-situ monitoring datasets are not long enough to evaluate trends in lake and river ice in the Arctic Ecozone+ and remote sensing provides an alternative. Latifovic and Pouliot (Latifovic and Pouliot, 2007) used ground-based monitoring datasets and remote sensing records to test comparability of the methodologies and examine trends in lake ice for 36 lakes across Canada (Latifovic and Pouliot, 2007). In doing so they developed records based just on remote sensing for six Arctic lakes. Break-up trends are shown in Figure 31 A and freeze-up trends in Figure 31 B. They concluded that, based on this limited set of lakes examined, the ice-free period increased significantly for lakes across all of Canada, but that the changes were of greater magnitude in the Arctic, at least in the past 20 years. Rates of change are shown in Figure 31 C.
Break-up and freeze-up dates are based on remote sensing as part of a larger study for all of Canada that found good agreement between ground-based records and the remote sensing methodology used. There were no ground-based records for the analysis for the Arctic lakes. Lakes are (west to east): Sitidgi, Washburn, Stanwell-Fletcher, Hazen, Angajurjualuk, and Amadjuak.
A. Break-up: only the trend for Lake Hazen was statistically significant (p<0.1)
B. Freeze-up: trends for 3 of the 6 lakes were statistically significant (p<0.1)
C. Comparison with lakes south of the ecozone+: rates for the 30 lakes south of the Arctic were based on ground-based monitoring and remote sensing. Note that, because of the different time periods, the difference in rates of change reflects both the influence of greater warming in the past 2 decades and the effect of higher latitude.
Source: adapted from Latifovic and Pouliot (2007)
Long description for Figure 31
This figure shows two maps of the Arctic Ecozone+ with icons representing the timing of ice break-up and ice freeze-up for six Arctic lakes. The figure also includes a bar graph that shows the following information:
|Ice break-up||All Canada|
|Longer ice-free period||0.39||1.75|
Rivers and lakes
Trends in river discharge
Changing river discharge is a consequence of changing precipitation patterns, as well as changing temperature patterns that affect water storage features such as soil moisture, permafrost depth, and glacier melt. Arctic residents have noticed changes in river flows and water levels that may reflect climate change. For example, residents in Chesterfield Inlet reported a decline in water levels (rivers, streams, ponds, and lakes) by as much as a metre (Nunavut Research Institute, 2004). They observed that spawning charr had more scratches and bruises caused by moving through shallow waters.
An assessment of Canada’s lake and river systems undertaken for this report, reported in the ESTR thematic technical report Biodiversity in Canadian lakes and rivers (Monk and Baird, 2011) and in Monk et al. (2011), examines status and trends across the country and by ecozone+ where there are sufficient long-term datasets. Only two hydrometric stations in the Northern Arctic and three in the Southern Arctic were of sufficient consistency and length to be included in the analysis of river flows. No stations were included from the Arctic Cordillera. Figure 32 shows the distribution of stations across all ecozones+ and trends in one-day minimum runoff across Canada. Trends are generally grouped regionally and increases in minimum runoff are evident in two out of the five stations within the Arctic Ecozone+. The same analysis was conducted for one-day maximum runoff, resulting in less clear regional trends across the country and no consistency in trends in the few Arctic stations (figure not shown).
With few long-term data sets for the smaller rivers that are wholly or mainly within the Arctic Ecozone+, little is known about trends in discharge, especially in the Arctic Archipelago. An analysis conducted as part of International Polar Year research extended the sparse hydrometric records for the Arctic Archipelago through extrapolation of observed streamflow attributes to provide estimates of current discharge. The uncertainty associated with the estimates is high because of the sparse data coverage. The authors concluded that the current freshwater flux for the region, and if or how it is changing, are unknown (Spence and Burke, 2008). They highlight the importance of this information for domestic water management decisions and for meeting international obligations to quantify freshwater inputs to the oceans.
An evaluation of hydrometric monitoring for 76 river systems across the Canadian Arctic (1950 to 2008) was undertaken by Mlynowski et al. (2011). The peak in Arctic land area that was monitored through streamflow gauges was 64% in 1990, decreasing to 56% in 2008. Larger river systems typically had the most data available. The peak in record length was in 1998, when 22 river systems had more than 30 years of continuous records. The authors concluded that major improvements were needed in Arctic hydrometric monitoring, especially for the relatively small rivers.
About three-quarters of the landmass of Canada is drained by rivers discharging into the Arctic Ocean, Bering Strait, Hudson/James Bay, and the Labrador Sea (Déry and Wood, 2005). This flow involves almost half (47.9%) of the total discharge of Canadian rivers (Monk and Baird, 2011). Thus hydrological processes and aquatic ecosystem health of rivers that cross briefly through the Arctic Ecozone+ at the far northern end of their courses are influenced most strongly by climatic conditions, terrain, and stressors in ecozones+ to the south. A prime example is Canada’s largest river basin, the Mackenzie River, which drains a total area of 1,787,000 km2, or 20% of the nation’s area (Culp et al., 2005). The river collects drainage from a number of other important rivers, including the Athabasca, Peace, Liard, Slave, Arctic Red and Peel, before draining through the Mackenzie Delta in the Southern Arctic (The Atlas of Canada, 2008a; The Atlas of Canada, 2008b).
Results of trend analyses are very dependent on the subset of stations and the time period used. Long-term trends are further masked by decadal-scale variations related to climate oscillations. Déry and Wood (2005), for example, found significant links between the Arctic Oscillation, El Niño/Southern Oscillation, and Pacific Decadal Oscillation with total annual freshwater discharge of rivers in northern Canada. Results from recent analyses of trends in discharge of Canadian rivers draining to various combinations of high-latitude oceans are shown in Table 4. Three studies that examined trends to the early 2000s (Déry and Wood, 2005; Déry et al., 2005; McClelland et al., 2006a) showed significant decreases in streamflow. Examination of these trends by regional river basins, however, shows that the decreases were more pronounced in the Hudson Bay and Labrador Sea drainages and that there was no significant trend from the 1960s to the early 2000s for rivers draining directly to the Arctic Ocean (Déry and Wood, 2005). Analyses that include data into the 2000s show a reversal of earlier declines in discharge, including in Hudson Bay, and a significant increase in annual average flows since 1989 (Déry et al., 2009; Déry et al., 2011).
|Study catchment||Analysis period||Results / direction of trends||Reference|
|64 hydrometric sites in Canada draining to high-latitude oceans||1964–2003||Déry and Wood (2005)|
|45 rivers in northern Canada||1964–2007||Déry et al. (2009)|
|56 rivers across North America (14 flow into Arctic Ocean and 42 flow into Hudson, Ungava and James bays)||1964–2000||McClelland et al. (2006a)|
|42 rivers draining into Hudson, James and Ungava bays||1964–2000||Déry et al. (2005)|
|23 rivers draining to Hudson and James bays||1964–2008||Déry et al.|
Freshwater discharge to these northern seas is significant not just in terms of the terrestrial/freshwater ecosystems in Canada’s North but also to ocean processes, including sea ice dynamics and thermohaline circulation, which are strong influences on regional and global climates and of increasing interest as key components in understanding climate change. Piecing together the trends around the Arctic Ocean and subarctic seas puts both the Canadian trends and the importance of Canadian monitoring into perspective. An analysis of freshwater sources to all high latitude oceans (Peterson et al., 2006) concluded that there was a 5.3% increase in river discharge to the Arctic Ocean and an 8.0% decrease in river discharge to Hudson Bay in the 1990s relative to a 1936–1955 baseline. There was a shift in the late 1960s to early 1970s, marking the start of increasing pan-Arctic river discharge to the Arctic Ocean and decreasing river discharge to Hudson Bay. This increase in discharge to the Arctic Ocean is strongly influenced by annual discharge from the six largest Eurasian rivers, which increased by 7% from 1936 to 1999 (Peterson et al., 2002) and reached record highs in 2002 (Peterson et al., 2006) and 2008 (Figure 33).
These analyses include rivers affected by hydroelectric projects, including rivers discharging to James and Hudson Bay. Hydroelectric dams and diversions have profound impacts on annual discharge patterns and water quality parameters like temperature and salinity, but do not have major effects on trends in total annual discharge. Changes in annual discharge related to periods when reservoirs were filling were accounted for in the analyses (Déry and Wood, 2005; McClelland et al., 2006b).
Trends in lake area
While remote sensing allows a broad view of change in lake areas, data for such analyses are limited to recent years. Carroll et al. (2011) examined the change in lake area across Canada from 60° to 70°N latitude from 2000 to 2009, based on analysis of MODIS satellite data. They concluded that, overall, lake area declined over the decade. They interpret this net loss of lake area to be related to longer ice-free periods and increased evapotranspiration, an effect that is particularly noticeable in the small, shallow lakes in Nunavut. Working counter to this trend, climate warming also can increase lake area as a result of melting permafrost causing collapsing and flooding, as has been occurring, for example, in the Ungava Peninsula and James Bay area, south of the treeline (Taiga Shield Ecozone+) (Payette et al., 2004b; Vallée and Payette, 2007; Thibault and Payette, 2009).
Ecological processes in lakes and rivers and linkages with climate change
The Arctic climate directly affects physical, chemical, and biological processes in aquatic systems. It also indirectly affects ecological processes through the control of terrestrial hydrologic systems and processes, particularly those associated with cryospheric components such as permafrost, freshwater ice, and snow accumulation and ablation (Prowse et al., 2006). In moist areas in the spring, thaw lakes and ponds have increased in amount and extent, evidently as a result of permafrost melting and increased precipitation (Stow et al., 2004). However, in drier regions of the Northern Arctic and Arctic Cordillera ecozones, ponds have been reduced in extent and some have disappeared.
Ponds on Ellesmere Island that had been permanent water bodies for millennia (Douglas et al., 1994) dried up completely in the warm summers of 2005 and 2006. Surrounding moss and grass wetlands also dried, with the loss of seasonal standing water, and the vegetation became drier and more subject to fire. This drying of the landscape is attributed to increased evaporation related to high summer temperatures and decreased ice cover, based on analysis of specific conductance measurements taken in the ponds since 1983 (Smol and Douglas, 2007a). More permanent ponds and seasonal ponds and wetlands can be expected to be lost through desiccation as the climate warms further. Certainly many aspects of the physical, chemical, and biological characteristics of these sites will be affected.
Changes in hydrology and water temperature greatly affect the life cycles of aquatic organisms either directly, through temperature, day length, or photoperiod changes, or indirectly, through changes in water quality and available nutrients. Many studies have shown recent changes in aquatic ecosystems that are linked with, or portend, climate change. Effects include regime shifts, or widespread species changes and ecological reorganizations (Smol et al., 2005) (See section on Changes in algal and invertebrate species assemblages in lakes and ponds on page 96). Shifts projected for fish populations will range from positive to negative in overall effect, differ among species and also among populations within species depending upon their biology and tolerances, and will be integrated by the fish within their local habitats (Reist et al., 2006b).
Shorter periods of ice cover on lakes and rivers bring with them many ecosystem changes. For example, less ice leads to warmer water, changing lake mixing regimes, and the distribution of nutrients and oxygen. Changing river ice conditions alter hydrological events like the ice-jam driven spring floods that are a critical part of maintaining wetland and riparian habitat (Peters et al., 2006). Changes in ice duration have impacts at all trophic levels, as outlined in several assessments and reviews (Walsh et al., 2005; Wrona et al., 2005; Vincent et al., 2008). These changes are complex--for example, the increased abundance of food available for fish in river systems (from greater productivity), and the increased habitat availability with less ice may cause otherwise anadromous species to remain in rivers year round. Feeding at sea has been linked to larger sizes in fish and larger populations --thus the increased productivity from less lake ice may ultimately lead to decreased fish yields in lakes and rivers (Reist et al., 2006a).
Extreme weather events
In the Arctic Ecozone+ severe weather events that influence the timing, amount, or quality of snow can have major ecological impacts on vegetation, small mammals, and ungulates. Deep snow and frozen snow layers block access to vegetation and prey or make food sources difficult to reach, leading to poor body condition and poor reproduction, as well as direct mortality (Miller and Barry, 2009; Hansen et al., 2013). While these conditions are a normal part of the Arctic winter, extreme weather events can mean that the conditions are more intense or last for longer than normal periods. In this section we have not attempted to define when an event is considered “extreme”, but rather we consider the continuum from normal variability to events that, due to their severity, pose risks of major mortality or reproductive failure to wildlife populations. For tundra ecosystem organisms, the major extreme events tend to occur in the shoulder seasons, autumn and spring, when the combination of temperatures and the timing of snow arrival and melt can result in major disturbances to their populations.
In the autumn, if the snow arrives too late or is too thin to insulate the ground, temperatures at the surface can be extreme for small mammals and invertebrates that depend on this insulation. On the other hand, rapid increases in snow depth in the autumn can make it difficult for large mammals to access forage plants or prey animals. It is for protection and insulation that small mammals seek deeper snow areas in tundra as overwintering habitat (Reid et al., 2012). Heavy snowfall events in August of 1997 and again in August 2000 in the Eureka region of Ellesmere Island covered vegetation, resulting in snow-free seasons of about half the normal length (Mech, 2004). Coincident population crashes of muskoxen, hares, and wolves were documented in the area and are attributed to increased energetic costs to the herbivores (Mech, 2004). In spring, a rapid snowmelt followed by a cold period can have strong effects on small mammals unable to get under the snow for insulation.
Extreme events falling within critical seasons are of particular ecological significance. In a classic paper on winter ecology, Pruitt (1957) described the temperature environment of a central boreal forest site in Alaska and related the changes to small mammal populations. He identified a fall critical period from the date that mean daily air temperatures went below the surface temperature to the date when a depth of about 20 cm of snow dampened daily surface temperature fluctuations. Similarly, the period from when snow depths become less than 20 cm to when mean daily air temperatures rise above surface temperatures is known as the spring critical period (Pruitt, 1957).
Another type of extreme weather event important in tundra systems is icing caused by rain on snow during the freeze period. A short period of temperatures above 0 °C with rain will result in ice layers either on the tundra surface or in the snow. These icing events can be devastating to wildlife populations if they seal the tundra surface making it very difficult for the animals to access forage. Miller and Barry (2009) noted that Peary caribou populations south-central Queen Elizabeth Islands were devastated (over 60% reduction) by exceptionally heavy snow and icing events in the four winters of 1973/74, 1994/95, 1995/96, and 1996/97. These events are projected to become more frequent and more widely distributed with climate change (Putkonen and Roe, 2003). The current distribution and frequency of icing events in the Canadian Arctic are not well-known and are difficult to reconstruct from data from the sparse network of climate stations (Grenfell and Putkonen, 2008). A rain-on-snow event in October 2003 on Banks Island contributed to a reduction of the muskoxen population by 20,000 over a three-year period. This event was analysed by satellite passive microwave imagery, developing a technique to map areas that have internal water and ice layers in the snowpack (Grenfell and Putkonen, 2008).
For tundra vegetation, icing events and extreme weather during the critical periods identified for small mammals are not as important--prolonged changes in temperature and snow regimes are of more significance. The distribution of vegetation along exposure gradients is based on the responses of plant species to the long-term site conditions, including snow depth and summer moisture availability (e.g., Walker, 2000). Sustained changes in soil moisture regimes as a result of flooding or permafrost disturbance (e.g., thermokarst; see section below) can greatly impact tundra vegetation (Vincent et al., 2011).
Additional types of extreme weather events with potential impacts on terrestrial ecosystems are likely increasing in frequency and/or intensity. Extreme heat events, for example, could result in increased fire frequency, loss of vegetation cover, and heat stress effects on tundra fauna, including outbreaks of disease. Intense storms in late summer and fall in the Arctic Ocean are known to alter ocean mixing and to play a role in reduction of sea ice (Jeffries et al., 2012), resulting in impacts on climate of the adjacent land, as well as increased coastal erosion.
Fire does not currently have a significant impact on Arctic ecosystems, as fires are rare and those that occur tend to be small, due to lack of fire-prone fuels and discontinuity of fuels (as patches of tundra are broken by barren lands, lakes, ponds, and rivers) (Krezek-Hanes et al., 2011). Records and research related to tundra fire ecology in Canada are sparse. Monitoring and analysis of satellite imagery indicates that there were only five large fires in the Arctic Ecozone+ from 1960 to 2007 (Figure 34), though this low frequency may be partly a function of poor monitoring capacity in remote regions (Krezek, 2009, pers. comm.). A 1990 study of fire records, satellite imagery, and observations collected through a questionnaire concluded that fire was more frequent in the western Southern Arctic, rarer to the north and east, and rare to absent in the Arctic Archipelago, related to trends in biomass, human settlement, and climatologic conditions (Wein, 1990).
Most reported fires were small, with larger ones occurring nearer the forest-tundra zone, such as the 1968 fire north of Inuvik (the large fire shown in Figure 34), which spread 15 km from the forest into the tundra (Wein, 1990). The site of this fire was monitored for changes in active layer depth and vegetation until 1993, showing quite variable responses among monitoring locations (Mackay, 1995). At all locations where the active layer deepened (due to the burn or to increased summer air temperatures), the underlying ice-rich permafrost thawed to produce thaw settlement (see next section on Permafrost disturbance). In general, tundra fires remove the vegetation cover and result in a short-term deepening of the active layer (Mackay, 1995).
Based on Canadian Large Fire Database from 1960 to 1996 and on analysis of remote sensing imagery from 1997 to 2007.
Source: data provided by Canadian Forest Service; methodology presented in Stocks et al. (2003) and Krezek-Hanes et al. (2011).
Long description for Figure 34
This bar graph shows the following information:
In northern Alaska a large tundra fire occurred in the exceptionally dry summer of 2007, burning 103,900 ha over a period of two months (Mack et al., 2011). This was the largest tundra fire recorded anywhere. Prior to this fire, researchers at the nearby Toolik Lake Long Term Ecological Research Network site had observed only two very small fires in the region over a 33-year period (LTER Network, 2008). The tundra fire was significant in terms of carbon balance: it was estimated to have released 2.1 teragrams of carbon to the atmosphere, an amount similar to the average annual carbon sink for the entire Alaska tundra biome (Mack et al., 2011).
Although tundra fires occur infrequently in modern times, analysis of charcoal and pollen in lake sediments in Arctic Alaska shows that the shrub tundra between 14,000 and 10,000 years ago burned with a comparable frequency to modern boreal forests (Higuera et al., 2008). These records, along with climate and vegetation reconstructions and analysis of tundra fires in Alaska since 1950 indicate that low effective moisture in summer (from a combination of higher temperatures and/or lower precipitation) and shrub-dominated vegetation cover are key factors in the fire regime. Several analyses now suggest that increases in tundra fires will likely accompany climate warming and the temperature-related increase of shrub vegetation cover that is now being observed in parts of the circumpolar Arctic, including Canada (Steven et al., 2006; Higuera et al., 2008; Hu et al., 2010; Myers-Smith et al., 2011a; Rocha et al., 2012; Elmendorf et al., 2012a).
Higher temperatures are thawing permafrost resulting in increased frequency and magnitude of slope failures and areas with thermokarst ponding (Lantz and Kokelj, 2008; Schuur et al., 2008; Lantz et al., 2009; Lamoureux and Lafreniere, 2009). These disturbances are restricted to areas with ice-rich permafrost in fine sediment. In the Mackenzie Delta region, retrogressive thaw slumps (Burn and Lewkowicz, 1990) have increased in numbers and size since the early 1970s (Lantz and Kokelj, 2008). Increases in these disturbances have also been observed in the High Arctic on the Fosheim Peninsula, Ellesmere Island (G. Henry, pers. observation, and Table 5) and on Melville Island (Lamoureux and Lafreniere, 2009) (Figure 35). The slumps result in areas with bare sediment, which can provide areas for different combinations of species to become established leading to greater landscape diversity. Lantz et al. (2009) found that green alder (Alnus viridis) had greater growth and reproductive effort on slumps than in undisturbed areas.
|Period||Duration (years)||Number of detachment failures||Annual rate at which detachments are initiated|
|Pre-1975||37–87Note * of Table 5||230||2.6 to 6.2|
Source: Lewkowicz and Harris (2005)
Notes of Table 5
- Note * of Table 5
assumes slope failures had taken place in the past 50 to 100 years, based on the length of time they are estimated to remain visible on the landscape
In the Low Arctic, these disturbances can expose previously frozen carbon to oxidation and change the tundra sites from carbon sinks to sources (Schuur et al., 2009). Many of the slumps in the Mackenzie Delta region occur on slopes above small lakes and the increased sediment load can significantly change the biogeochemistry of the aquatic environment and alter the biotic communities in the lakes (Kokelj et al., 2009; Thienpont et al., 2013). Lakes affected by slumping arising from permafrost thawing experience shifts in nutrient, light, and phytoplankton relationships, and slump-affected lakes generally have lower nutrients and lower productivity than lakes unaffected by slumps (Thompson et al., 2012).
In addition to the increases in slope failures, thawing of ice-rich permafrost can lead to subsidence and development of thermokarst ponds (Laurion et al., 2010). The conversion of terrestrial systems to ponds is a profound landscape change, with important implications for the carbon balance of these regions. Thermokarst ponds have been found to be important sources of methane and carbon dioxide (Laurion et al., 2010; Abnizova et al., 2012).
Deeper active layers also contribute to the drying of tundra soils, as water percolates further into the soil and may become unavailable for plants. Warming temperatures also increase evaporation from soils and water bodies, leading to further drying if the loss in moisture is not compensated for by increased precipitation and/or increased moisture from melting permafrost. One of the manifestations of this effect is seen in the drying of Arctic lakes and ponds (Smith et al., 2005a). Smol and Douglas (2007a) describe the drainage of ponds on eastern Ellesmere Island in 2006 which had been permanent water bodies through the Holocene, the ultimate threshold change for these systems. Loss of lakes can result both from increased evaporation and from drainage through the weakened melting permafrost (see also the section on Ecological processes in lakes and rivers and linkages with climate change on page 50). Drying tundra will have important effects on the system response to climate change and will depend on the initial conditions (Shaver et al., 2000).
Permafrost thaw has caused significant disturbances to Arctic landscapes and the combination of climate change and increasing industrial development has the potential to greatly increase rates and magnitudes of permafrost disturbance. These changes have important implications for the structure and function of Arctic terrestrial ecosystems, especially for carbon balance and feedbacks to the atmosphere.
Community and population dynamics
Arctic foods webs are characterized by short food chains, a few dominant animal species that often have cyclic population fluctuations, the central role of small mammals in driving populations at higher trophic levels, and a high degree of responsiveness to regional climatic oscillations and climatic variables such as precipitation, snow depth, and temperature (Elton, 1924; Norrdahl, 1995; Hudson and Bjornstad, 2003; Krebs et al., 2003; Gunn, 2003a; Hörnfeldt et al., 2005; Van Bogaert et al., 2007). They are influenced by a multitude of large-scale disturbances and are particularly vulnerable to global changes that may affect the sustained production of plants and animals on which Arctic cultures depend (Elmqvist et al., 2004). Arctic people are accustomed to large annual fluctuations in the wildlife on which they depend. However, a long-term decline in harvested species, such as caribou, could significantly affect their ability to maintain traditional lifestyles (see the section on Ecosystem goods and services on page 167).
Understanding trends in community and population dynamics requires examining ecosystem processes and integration of trends at all trophic levels, along with trends in drivers and stressors. There are few long-term monitoring programs in Arctic Canada that allow for this integration, though this was a focus of several Canadian-led International Polar Year research programs. For example, the International Tundra Experiment (ITEX) and the CircumArctic Rangifer Monitoring and Assessment (CARMA) networks work to consolidate existing research and monitoring and enhance ongoing observations for the Canadian and the circumpolar Arctic. At the broad-picture level, the Circumpolar Biodiversity Monitoring Program, a project of Conservation of Arctic Flora and Fauna (CAFF) (a working group of the Arctic Council), is working to improve coordination and integration of ecosystem monitoring.
Integration of monitoring and research through intensive study at research sites improves understanding of status and trends in community and population dynamics. These programs track and investigate trends and linkages to understand ecosystems. Examples for the circumpolar Arctic are the research and monitoring programs at Toolik Lake in Alaska, Zackenberg in Greenland, and Abisko in northern Sweden. Canada has not invested significantly in this approach to ecological monitoring and research in the Arctic and consequently understanding of Canadian trends in community and population dynamics is poor. Results from one Canadian site with long-term integrated research are summarized in the “Case study on ecosystem functions and processes: Bylot Island” (page 81). The general discussion on trophic dynamics below is developed further through specific examples in the case study. The bottom trophic level, primary producers, is addressed in the section on Primary productivity (page 71).
Dynamics of predators in Arctic tundra ecosystems are based mainly on lemmings and other small rodents (Callaghan et al., 2005b). Some predators, including snowy owls and weasels, specialize in rodents and their reproductive success is closely linked to small mammal cycles. Other predators, including foxes, wolverines (Gulo gulo), wolves, and grizzly bears (Ursus arctos), prey on a wider range of species and are able to switch to alternative prey in years of low abundance of small mammals. This means that alternative prey species, such as geese and shorebirds, become indirectly linked to small mammal cycles. Research in the eastern and central Canadian Arctic shows that fluctuations in lemming populations affect their main predators, snowy owls and Arctic foxes, which in turn indirectly affect snow geese, and possibly shorebirds, through shared-predator interactions (Kerbes et al., 2006).
Compared with boreal ecosystems, large predators are not abundant in the Arctic tundra and predation impacts on ungulates are usually low unless the ungulates are at low densities (Callaghan et al., 2005b; Legagneux et al., 2012). In a study across 12 sites in Arctic Canada, 79% of production of small to mid-sized herbivores, including voles, lemmings, and Arctic hares, was consumed by predators, while predators consumed only 9% of combined caribou and muskox production (Krebs et al., 2003).
Wolves and tundra grizzly bears depend on caribou, although the regulatory role of predation for caribou dynamics is uncertain. In the mid-1990s in the Southern Arctic, the Bathurst Herd of 350,000 caribou was estimated to support some 1,500 wolves (Cluff, 2004, pers. comm.) that likely annually killed 40,000 caribou. In addition the herd was estimated to support 500 to 1,000 grizzly bears, with caribou making up 80% of their diet. A grizzly bear may eat between 9 and 18 adult caribou in a year (Gau et al., 2002). For the Northern Arctic Ecozone, Krebs et al. (2003) argue that the dominant ecosystem type is more one driven by variance in weather (bottom-up driven) rather than one controlled by predation (top-down driven). Lagagneux et al. (2012) contrast the predator-driven Bylot Island ecosystem, which lacks large ungulates, with the stronger controls exerted by bottom-up effects on systems in which caribou play a large functional role.
Herbivores and decomposers
Small and mid-sized herbivores are critical in Arctic food webs. The three- to five-year population cycles of small herbivorous mammals are well-known in the ecological literature. The cycles of the various species of lemmings [brown lemmings (Lemmus trimucronatus)andcollared lemmings (Dicrostonys spp.)] (Krebs, 1996; Predavec et al., 2001; Gilg et al., 2003) and voles [tundra voles (Microtus oeconomus)and northern red-backed voles (Myodes rutilus)] (Krebs et al., 2002; Krebs, 2011) have been shown to have a high degree of synchrony across the Canadian Arctic (Krebs et al., 2002). Larger herbivores such as caribou may also cycle with a periodicity of 40 to 70 years (Gunn, 2003a) in the Arctic, but this is not well-documented because of the length of the cycles. Traditional knowledge does, however, provide ample evidence for the cycles in several regions. For example, one account of traditional knowledge passed along through generations of Chipewayan hunters indicates that the Bathurst Herd has been fluctuating in size with a cycle length about 30 years over the past 120 years (Nesbitt and Adamczewski, 2009).
Many small mammal populations throughout the circumpolar Arctic exhibit population cycles of more or less regularity. Some populations do not cycle but instead fluctuate without a pattern or remain at low population levels, with regular outbreaks of high densities being inhibited by conditions like patchy habitat, high predation rates, or particularly harsh winter conditions (for example, lemmings and collared voles showed no signs of cyclic abundance in a study on the Yukon North Slope (Krebs et al., 1995)). Scandinavian small mammal cycles have been well-studied. However, as most Norwegian lemmings (Lemus lemus) live in alpine tundra, the mechanisms driving their population cycles may not be broadly applicable to the Canadian Arctic. Data from northern Scandinavia indicate that lemming cycles may be dampened or lost for decades, while voles maintain regular cycles. Lemming populations in northern Scandinavia remained at low densities in recent decades, followed by recent population outbreaks, while vole cycles have weakened--observations attributed by researchers to the greater sensitivity of lemmings to variations in climate and predation (Ims et al., 2011). Of more relevance to the Canadian Arctic, 20-year studies of small mammals and predator-prey interactions in eastern Greenland indicate that climate change has the effect of increasing the length of the lemming cycle and decreasing maximum population densities, an impact that is detrimental to populations of predators reliant on lemmings (Gilg et al., 2009).
Using an indirect method of studying collared lemmings at 17 sites in the Canadian High Arctic and one site in Alaska, Predavec et al. (2001) were unable to detect a regular periodicity in fluctuations of lemming populations, although variability from year to year was high. Trends in small mammal densities from all monitored sites under the Northwest Territories/Nunavut Small Mammal Survey show, however, that population fluctuations can be synchronized over a large region (Figure 52), at least during the short period over which they have been recorded. Long-term monitoring covering approximately 10 cycles is necessary for a rigorous analysis of cyclic populations (Predavec et al., 2001).
There are no clear trends of long-term changes in the cycles in Canada. At Bylot Island, brown lemmings show cyclic, large-amplitude fluctuations in density with a periodicity of three to four years, though not collared lemmings. There is little evidence of a temporal trend in lemming density, but trend detection is made difficult by the large inter-annual variability. To smooth out those variations, Gauthier et al. (2013) looked at the three-year running mean of density. This analysis suggested a possible decreasing trend in summer lemming density. Lemming populations were especially low during the period 2002 to 2009, but the most recent peak (2011) was relatively high (Gauthier et al., 2013). Continuous, site-specific long-term monitoring of lemmings is still rare in the Canadian Arctic (Krebs et al., 2002), so it remains uncertain to what degree findings from Bylot Island are particular to the site. Only three peaks with similar abundance levels were detected between 1994 and 2012 at the only other site where long-term data on lemmings and voles exist, the Daring Lake Research Station in the central barrenlands, a site that is representative of the tundra-taiga ecotone (Figure 36).
Food webs in the boreal and Arctic ecozones+ have relatively few links. The links, however, can be complex in terms of the dynamics of their inter-relationships. These northern ecosystems are nutrient-limited as so much carbon is inaccessible because only a shallow active layer of the soil thaws each year. Caribou, through their forage intake and output (faecal pellets), have complex and cascading effects, strongly patterned over time and space (Kielland et al., 2006). As well, caribou support a diverse group of other species, including external parasites such as blood-feeding mosquitoes. Mosquitoes, in turn, through the filter-feeding of their larvae, are a key element in nutrient cycling in aquatic systems. Further up the food webs, caribou support large-bodied and medium-sized predators and scavengers. Earlier debates about top-down (predator) or bottom-up (forage) regulation for caribou are now replaced by an appreciation of how nutrition and predation interact (Brown et al., 2007a).
Relationships between plants and caribou include the plants’ responses to caribou’s selective foraging. Caribou strongly select for individual plant species and forage for buds and unfolding leaves to maximize nutritional value (White and Trudell, 1980; Russell et al., 1993). The gregarious and migratory behaviours of migratory tundra caribou force their role in ecosystem structure and functioning to be strongly scale dependent (Griffith et al., 2002). Caribou convert plant tissue into body mass and faecal pellets. Through their local foraging movements and their seasonal migrations, they re-distribute nutrients within and across ecozones+. In the taiga ecozones+, the effects of caribou herbivory lag a season, as caribou are foraging during winter when most plant growth and nutrient cycling are quiescent due to sub-zero temperatures. Over the time scale of decades, caribou winter ranges expand and contract and the herds cycle from high to low abundance. Abundance can vary three-fold, with cascading effects on plants and nutrient cycling as the plant communities shift from one state to another. Succession of plant communities as a response to density of foraging may include, for example, lichen-dominated tundra shifting to greater dominance of moss, and then to increased dominance of grass (Van der Wal, 2006).
Nitrogen is a limiting factor for plant growth. Caribou summer browsing can increase the rate of soil nitrogen cycling through modifying the amount of plant litter, which changes the soil microclimate for decomposition and mineralization processes, and through adding soluble nitrogen from faecal pellets and urine (Olofsson et al., 2004). The changes vary with season and time, and with intensity of browsing (Kielland et al., 2006).
Decomposition rates are low in tundra ecosystems (see section on Permafrost on page 27 and section on Nutrient cycling on page 67). Most primary production enters the detrital food chain (Smith and Smith, 2001). A study on Devon Island showed that lemmings consumed about 3–4% of the standing crop most years, with most of the rest of the vegetation eventually being channelled to a variety of microbes and soil organisms, dominated by bacteria and fungi (Smith and Smith, 2001).
Wildlife diseases and parasites
This section, which draws material from the ESTR technical thematic report Wildlife pathogens and diseases in Canada (Leighton, 2011) and the ESTR technical thematic report Northern caribou population trends in Canada (Gunn et al., 2011c), covers disease and parasites related to Arctic ungulates and implications of climate change. Arctic ungulates, caribou in particular, are key components of Arctic terrestrial ecosystems and are vulnerable to changes in disease and parasite infestation. These changes could be related to disease type and frequency and also changes in the ranges of ungulates and of other parasite hosts. Warm weather can promote outbreaks of disease in cold-adapted animals and this may occur more frequently as the climate changes. For example, an introduced muskox population in Norway was struck by an outbreak of pasteurellosis in 2006, killing about 20% of the population. The outbreak was attributed to unusually hot and humid conditions (Ytrehus et al., 2008). An unusual outbreak of yersiniosis, a bacterial disease, in muskoxen on Banks Island in 1986 may also have been related to a period of unusually hot summer weather (Blake et al., 1991).
Brucellosis is the name given to all diseases caused by infection with any of the several different species of the bacterial genus Brucella. The clinical manifestations of brucellosis are many, but the most common are infection and inflammation of the female and male reproductive tracts with resulting abortion and male infertility, and infection of joints and tendon sheaths resulting in progressive lameness. Infection persists, often for the lifetime of the animal. People are similarly susceptible to infection with Brucella sp., and brucellosis in animals with which people have contact is a public health risk (Chan et al., 1989; Forbes, 1991; Thorne, 2001).
In the Arctic Ecozone+ infection with Brucella sp. is widespread and of potential ecological and public health significance in barren ground caribou populations (and one introduced herd of reindeer near Tuktoyaktuk, Northwest Territories). Brucella suis biotype 4 infects caribou across the Arctic, Taiga Cordillera, Taiga Plains, Taiga Shield, and the northern edges of the Boreal Plains, Boreal Shield, and Hudson Plains ecozones+ (Forbes, 1991). Brucellosis is widespread in Arctic caribou, with 20–50% of animals in some herds infected (Leighton, unpublished data; Koller-Jones, 2006, pers. comm.). However, its ecological impact, if any, on infected populations is not known. Infection of northern people with this bacterium occurs and is associated with consumption of caribou (Chan et al., 1989; Forbes, 1991). Whether or not B. suis biotype 4 is a naturally-occurring pathogen in North America or a pathogen introduced from Europe in imported reindeer also is not known. There are no records of this infection in woodland caribou (Rangifer tarandus caribou), including in the George River herd of northern Quebec.
A serological survey of a large herd of reindeer in the western edge of the Arctic Ecozone+ and of a barren-ground caribou (Rangifer tarandus groenlandicus) herd (Qamanurjuaq) in the Taiga Shield and adjacent Arctic regions of Manitoba and Nunavut in the 1960s found only 9% of reindeer and 4% of caribou infected (Broughton et al., 1970). The more recent infection rates of 20 to 50% may represent a trend of increasing prevalence and/or increasing surveillance. Any environmental changes that increase the overlap of barren-ground caribou with boreal caribou carry the risk that Brucella suis biotype 4 may become established in boreal caribou populations.
Brucellosis is considered to be responsible at least in part for the recent decline of the Southampton Island Caribou Herd. Since 1968, susceptibility to disease and parasites due to low genetic differences has been a likely catalyst for the widespread infection with brucellosis first detected in the Southampton Herd in 2000. Prevalence of the disease rose to 58.8% in 2011 and may be responsible for the drop in pregnancy rates in Southampton caribou since 2000, with the greatest decline in pregnancy rates being since 2008 (Department of Environment, 2013a).
Host–parasite systems are particularly sensitive to climate change because many macroparasites have life cycles with free-living stages whose development and survival are strongly dependent on temperature and moisture conditions. Small changes in climate may have a large impact (Kutz et al., 2004) by influencing the development and survival rates of these free-living life stages as well as the abundance and activity of their arthropod vectors and mollusc intermediate hosts. Projected impacts and some empirical observations predict a longer season for development and transmission of parasites, increased infection levels in host populations, and increased outbreaks of disease (Kutz et al., 2005; Kutz et al., 2008).
Umingmakstrongylus pallikuukensis is a parasitic nematode worm prevalent in muskoxen in the western mainland parts of the Arctic Ecozone+. This lungworm requires a gastropod intermediate host to develop to the infective stage, which historically took two years. An empirically based model showed that temperatures have been warm enough since the early 1990s for the parasite to develop in one season. This could lead to potentially increased infection pressure on muskoxen, with possible impacts on fecundity and survival rates of muskoxen (Kutz et al., 2005). The extent to which this has affected muskoxen populations, however, is not known.
In Canada, Besnoitia tarandi infects caribou and reindeer, and probably muskoxen, across their ranges. It has been documented in caribou since 1922 in both Arctic and subarctic Canada (Ducrocq et al., 2013). Infection is common in barren-ground caribou and has been described in woodland caribou. Infection rates in muskoxen are not known (Choquette et al., 1967; Wobeser, 1976; Gunn et al., 1991a; Ayroud et al., 1995). Although occasional severe manifestations of infection on the skin have been seen, most infections appear to have little or no health consequences for these species. Hunter observations, confirmed by veterinary investigations, suggest that the protozoan Besnoitia has recently emerged as a disease-causing agent in caribou herds in Quebec and Labrador (Kutz et al., 2009; Ducrocq et al., 2013). In 2005, focus group discussions in Inuvialuit, Gwich’in and Dene communities in the Southern Arctic and taiga ecozones, some hunters reported increasing lesions (Kutz, 2007), possibly associated with Besnoitia cysts or even warble larvae (Kutz et al., 2009).
Climate change can also lead to changes in ranges of pathogens, both through increasing the area over which temperatures are warm enough for the pathogens to thrive, and through range changes in the host animals. Current knowledge regarding range extents of four ungulate protostrongylid nematodes is shown in Figure 37 . Model projections indicate that, as temperatures rise, U. pallikuuensis will likely expand its range to the north and east (Kutz et al., 2005). Recent observations suggest that this parasite may have expanded its geographic range significantly in the last decade (Kutz et al., 2009) although interpretation of the data is complicated by increased surveillance for parasites. Protostrongylids have not been detected in ungulates from the Arctic islands and Greenland and may be excluded from high latitudes under current climate conditions but could invade these regions under warmed conditions if their host populations of moose, muskoxen, or caribou were to expand northward (Hoberg et al., 2008). Parelaphostrongylus odocoilei, a parasite with a range south of the Arctic Circle, may expand further north. In 1999, following a year with the warmest recorded annual temperature for the Mackenzie District, fatal pneumonia associated with P. odocoilei was first detected in Dall’s sheep (Ovis dalli) in the Mackenzie Mountains (just south of the ecozone+) (Hoberg et al., 2008).
Emergence of disease may follow climate change, but for macroparasites in particular, there are likely to be lag times determined by the life history of the parasites, including the period of development in the hosts (Hoberg et al., 2008). Cascading and cumulative long-term effects of climate change, including shifts in host–pathogen relationships, may be among the factors contributing to large-scale changes in abundance and distribution observed in keystone wildlife (e.g., woodland and barren-ground caribou) in northern North America.
Rabies persists in Arctic foxes and red foxes (Vulpes vulpes) in the circumpolar Arctic (Mork and Prestrud, 2004). It is difficult to assess trends in northern Canada, partly because samples are not routinely sent for testing once it has been established that rabies is present in the area during a particular winter (J. Kush, Rabies Lab, CFIA-Lethbridge, pers. comm.). Samples are submitted for rabies testing if a potentially infected animal is suspected of having been in contact with a person or a domestic dog. Although rare, rabies has also been documented in polar bears (P. Hale, Government of Nunavut, pers. comm.).
Little information exists on parasite communities of terrestrial Arctic carnivores. Without empirical baseline data on commonly occurring parasites in these animals, few studies will be able to assess impacts and changes of hosts and parasites as a result of climate change (Brooks and Hoberg, 2007).
Using mathematical models to predict change, it has been postulated that the prevalence of Toxoplasma gondii infection in humans will increase due to environmental as well as anthropogenic factors related to climate change (Meerburg and Kijlstra, 2009). With increasing temperatures favouring survival of T. gondii in the environment, as well as changes in ecological distributions of vertebrate hosts, (more animals moving into Arctic areas), climate change will likely increase the prevalence of T. gondii in the Arctic (Hueffer et al., 2013).
Infection with Toxoplasma gondii is common in wildlife, domestic animals, and humans worldwide, including those in the Canadian Arctic (Dubey and Beattie, 1988; Mcdonald et al., 1990; Philippa et al., 2004). While some information exists on the occurrence of T. gondii in wild animals in the Arctic, the prevalence of the parasite in areas with a cold climate is low compared to humid, temperate, and tropical areas (Tenter et al., 2000). Infection of, or antibodies specific to, T. gondii have been found in some wild animals with ranges that include or are close to Arctic areas of North America, including moose (Kocan et al., 1986; Zarnke et al., 2001), wolves (Zarnke et al., 2001), wolverines (Reichard et al., 2008; Dubey et al., 2010) , Canadian lynx (Lynx canadensis) (Labelle et al., 2001; Zarnke et al., 2001; Philippa et al., 2004) , muskoxen (Kutz et al., 2000) , and caribou (Kutz et al., 2001).
In the Arctic, Canadian lynx are the most likely wild definitive hosts for T. gondii and are responsible for contaminating the environment with oocysts. Studies conducted on Canadian lynx reported 1 of 5 (20%) from Nova Scotia, 47 of 106 (44%) from Quebec (Labelle et al., 2001), and 39 of 255 (15%) from Alaska (Zarnke et al., 2001) had antibodies to T. gondii. Canadian lynx generally prefer to stay below the treeline, but are occasionally observed and harvested north of treeline.
Close proximity to the treeline was shown to be an important variable for exposure to T. gondii. Surveys conducted by Kutz et al. (2000) to determine the seroprevalence of antibodies to T. gondii in muskox showed a higher proportion of T. gondii exposure in a mainland herd (closer to the treeline) compared to island populations (farther from the tree line) as well as a higher prevalence in adults and juveniles compared to calves. A significant difference was not found in the seroprevalence of T. gondii antibodies between male and female muskoxen (Kutz et al., 2000). In caribou, Kutz et al. (2001) showed that seroprevalence of T. gondii antibodies was significantly higher in mainland populations compared to island populations. However, there were no statistical differences in the seroprevalence of T. gondii antibodies among age classes or between sexes of caribou (Kutz et al., 2001). Reichard et al. (2008) showed that exposure to T. gondii was common in wolverines harvested in the Kitikmeot Region of Nunavut. Prevalence of T. gondii antibodies did not differ significantly between sex and age of wolverines, nor was exposure associated with proximate to the treeline at time of harvest (Reichard et al., 2008).
The start and the duration of the growing season in northern terrestrial systems are largely determined by snow cover. Changes in the depth and duration of snow cover over the past 50 years are discussed above in the section on Snow (page 32). The length of the snow-free season has increased significantly in most regions of the circumpolar Arctic, including Canada. In the Canadian Arctic, snowmelt dates have shifted earlier by an average of about 8.5 days from 1950 to 2007 (Zhang et al., 2011).
Tundra plants can be divided into two groups based on their flowering phenology: early or late flowering (Molau, 1993). The flowering phenology is related to their reproductive strategies as either risking pollination (early-flowering species) or seed production (late-flowering species) (Molau, 1993). Early-flowering species dependent on insects for pollination may flower earlier or later than the peak of the insect populations, and any disconnect in timing may be exacerbated by climate change (Molau, 1997). Earlier snowmelt has resulted in earlier flowering by many species (Callaghan et al., 2011b; Oberbauer et al., 2013), and experimental warming has been shown to advance flowering (Arft et al., 1999; Aerts et al., 2006). Warmer temperatures in eastern Greenland have resulted in earlier flowering and a shorter flowering period, with a concomitant decrease in the number of potential insect pollinators (Høye et al., 2013). However, this is the only long-term monitoring study of the effects of warming on both flowering and insect phenology, and there is a need for more systematic studies on effects of earlier snow melt on pollinators in tundra systems. Late-flowering species risk lower seed production because of the unpredictability of late-season weather, especially in the High Arctic. In general, the end of the growing season and the onset of snow cover have not changed as much as snowmelt (Callaghan et al., 2011a), although experimental warming has been shown to delay fall senescence in some tundra species (Marchand et al., 2004). In many coastal locations, at least in the western Canadian Arctic, however, the predominant change in growing season length is in the autumn, with later onset of snow (due to the warming effect of the ocean) (Dye, 2002).
Leaf phenology also varies by species in deciduous shrubs and forbs, although leaf bud break occurs a few days after snowmelt in most species (Shaver and Kummerow, 1992). Experimental warming has shown that leaf bud break is occurring earlier in most species (Figure 38) (Henry and Molau, 1997; Arft et al., 1999; Oberbauer et al., 2013), and remote sensing has shown that the Arctic tundra regions are snow free and green-up earlier than 20 years ago (Zhou et al., 2001; Parmesan, 2007; Bhatt et al., 2010). A recent synthesis of phenological responses at the plot level over the past 30 years has shown that tundra plants are flowering and leafing-out earlier and that this corresponds with the general warming over the same time period (Oberbauer et al., 2013).
Tundra ecosystems are nutrient limited, as shown by relatively strong responses to low and moderate levels of fertilization (Henry et al., 1986; Chapin et al., 1995; Shaver and Chapin, 1995). This ultimately stems from the low temperature and low resource environment of the Arctic, with low rates of net primary production, decomposition, and mineralization (Hobbie, 1996; Nadelhoffer et al., 1997; Cornelissen et al., 2007). The uptake by plants of organic forms of nitrogen, for example amino acids, is important in tundra soils and is undoubtedly a response to the low rates of decomposition and mineralization (Kielland and Chapin III, 1992; Schimel and Chapin III, 1996). Research on the effects of climate change on nutrient cycling in tundra systems has shown that availability of inorganic and organic nitrogen can increase in experimentally warmed plots (Schmidt et al., 2002; Rolph, 2003; Aerts et al., 2006).
The soil microbial communities and their functioning are likely to be altered by climate change, both directly by changes in temperature and moisture, and indirectly through changes in vegetation, such as the increase in shrub abundance in the Arctic (Sturm et al., 2001; Bigelow et al., 2003; Stow et al., 2004; Deslippe et al., 2005; Tape et al., 2006; Myers-Smith et al., 2011a). The soil temperature regime has major consequences for Arctic ecosystems. In Alaska, winter biological processes are contributing to the conversion from tundra to shrub communities through a positive feedback that involves the snow-holding capacity of shrubs, the insulating properties of snow, a soil layer that has high water content because it overlies nearly impermeable permafrost, and hardy microbes that can maintain metabolic activity at low temperatures (Sturm et al., 2005; Myers-Smith et al., 2011a). Increasing shrub abundance leads to deeper snow, which promotes higher winter soil temperatures, greater microbial activity, and more plant-available nitrogen (Schimel et al., 2004). High levels of soil nitrogen favour shrub growth the following summer.
Experimental warming has been shown to alter microbial community structure (Walker et al., 2008b; Deslippe et al., 2012) and snow depth changes have been shown to alter microbial processes in some cases in the Low Arctic (e.g., Schimel et al., 2004). While experimental warming caused differences within sites in microbial composition, based on frequency and abundance of genotypes involved in nitrogen transformations, there were greater differences between sites. The greatest effects of warming were found to occur in wet sedge tundra communities (Walker et al., 2008b). However, resistance to change in microbial communities has been found in studies of warming and combined warming and fertilization of tundra communities in the High Arctic (Deslippe et al., 2005; Lamb et al., 2011).
In wet sedge meadow soils, microbial biomass and nutrient availability peaked early in the spring freeze-thaw phase, but then crashed after soil temperatures rose above 0°C, implying that competition for nutrients from roots results in the collapse of the microbial pool (Edwards et al., 2006). Earlier spring snowmelt and warmer temperatures would obviously alter the timing of these changes, but the vascular plant growth would still depend on available nutrients and soil moisture (Nadelhoffer et al., 1997).
Nutrient uptake in many northern plant species depends on mycorrhizal associations. Deslippe et al. (2011) have shown that long-term experimental warming changes the diversity of ectomycorrhizal fungal families on root tips of dwarf birch (Betula nana), a deciduous shrub. Similarly, Fujimura et al. (2008) found a change in the density of genotypes of the fungal community associated with the roots of arctic willow (Salix arctica) in warmed plots of three High Arctic tundra communities, although there was no shift in relative abundance. The effects of these changes in mycorrhizal communities are unknown, but they will likely affect the ability of shrubs and other plant species to absorb nutrients in the warmer soils.
Arctic animals play a major role in nutrient cycling. Caribou, the main large herbivore in the Arctic, have the net effect of forage removal, production of greenhouse gas, and return of nutrients through faecal pellets. Based on energetics modelling (Russell et al., 2005; Gunn et al., 2011c), a caribou, annually, removes 900 kg of food (2.5 kg/day), produces 20 kg of methane (55 g/day) and returns to the ecosystem, nutrients in the form of faecal pellets, 270 kg (30 g x 25 times a day). At the herd scale, annually, 170,000–350,000 caribou remove 140–320 million kg of forage, produce 3–7 million kg of methane, and return 38–77 million kg of faecal pellets spread over the annual range (150–300 kg/km2). As caribou travel and rest on frozen waterways, the nutrient return from faecal pellets is to aquatic as well as terrestrial ecosystems (Gunn et al., 2011c).
Geese, as well as cycling nutrients through their role as tundra grazers, import energy and nutrients into Arctic ecosystems when they return to their breeding grounds after a winter of feeding in fields and wetlands far to the south. These imported resources move through the food chain when the geese are preyed on, for example, by Arctic foxes (Giroux et al., 2012a). Migratory birds may also export nutrients and energy, for example through removing nitrogen from tundra ecosystems when young produced locally migrate south and die (Gauthier et al., 2011). While locally significant, the importance overall of these resource exchanges to tundra ecology is not known (Gauthier et al., 2011; Giroux et al., 2012a).
Carbon storage and release
Permafrost soils are estimated to contain 1,400 to 1,859 Pg of carbon in frozen and seasonally unfrozen surface layers, which amounts to 50% of the soil carbon in the world (McGuire et al., 2009; Tarnocai et al., 2009). In Canada, Arctic tundra systems were estimated in 2008 to contain about 76 Gt of soil organic carbon in the upper metre of the soil profile (Ping et al., 2008), but, given more recent research (Tarnocai et al., 2009; Grosse et al., 2011; Kuhry et al., 2013), this figure may be low. Soil carbon (C) might react to near-term climate change (Clein et al., 2000). Tundra ecosystems have been sinks for carbon over tens of thousands of years (Ping et al., 2008). This “service” was possible because of the cold soils, which restricted decomposition rates and the accretion of permafrost, which helped to freeze the carbon stored in the soil. There is concern that the warming climate will result in a switch in tundra ecosystems from sink to source of carbon for the atmosphere (Oechel and Vourlitis, 1997; Ping et al., 2008; Grosse et al., 2011; Lund et al., 2012).
Large amounts of C and nitrogen (N) could be released in inorganic forms as Arctic soils warm, the active layer deepens, decomposition rates increase, and the growing season lengthens (Nadelhoffer et al., 1997; Grosse et al., 2011). A large release of carbon dioxide (CO2) from these soils would increase atmospheric CO2, enhancing the rate and magnitude of climate change (a positive feedback). The release of other greenhouse gases, especially methane and nitrous oxide, could be increased from the permafrost soils as they warm, which would contribute to the positive feedback (Vincent et al., 2011).
A coordinated study of the net ecosystem production of Arctic tundra sites in the Canadian Arctic found that all sites were sinks for CO2 during the summer, and the differences in rates followed expected patterns due to latitude and proximity to cold coastal conditions (Lafleur et al., 2012). Annual variability in net ecosystem production was considerable at both a wet and a mesic (moderately moist) tundra site in the Low Arctic, and was strongly related to differences in climate during the growing season (Humphreys and Lafleur, 2011).
Warming experiments to date have shown that wet tundra systems remain as sinks when warmed, as both photosynthesis and respiration are increased; however, warming greatly increased the loss of carbon from dry tundra ecosystems (Oberbauer et al., 2007) (Figure 39). In addition, there was a gradient in net carbon exchanges, with High Arctic systems remaining as sinks but with increasing loss of carbon from Low Arctic systems. There are very few studies of carbon flux in response to experimental warming and other environmental perturbations in Canadian tundra ecosystems.
Phenological changes and changes in animals that act as pollinators could affect vegetation. This may not be as great a concern for the Arctic Ecozone+ as for other ecozones+, as many Arctic flowering plants can self-pollinate or are pollinated by the wind (Callaghan et al., 2005b). Nonetheless, most tundra flowering plants are dependent on insect pollination to set seeds (Kevan, 1972), and outcrossing through insect pollination is important to maintain genetic diversity. Insect pollination is an important part of the reproductive success of flowering Arctic plants because it provides more seed genetic variability than self-pollination or asexual reproduction. Insect pollination is more successful at pollen transfer than wind pollination or transfer by animals (Kevan, 1972; Crawford, 2008).
The main pollinators at the few Arctic locations where pollinators have been studied are flies (Diptera), with bumblebees and butterflies also being important (Ellesmere Island: Kevan, 1972; Zackenberg Greenland: Klein et al., 2008; Høye et al., 2013). Studies in Greenland indicate that most insect pollinators are generalists, though some preferences for certain plants are exhibited (Klein et al., 2008). It is not known to what extent changes in plant and insect species distributions, arrival of new species to the Arctic, as well as changes in phenology, might affect pollination (Klein et al., 2008). However, Høye et al. (2013) have found that the flowering season has become shorter and insect flower visitors have declined as temperatures increased over the past 15 years at Zackenberg, Greenland.
There has been little study on pollination relationships in the Canadian Arctic and little is known about the status and current trends related to the changing climate. The synchronization between insects and flowering Arctic plants is complex and it is hard to predict how these relationships will be affected by climate change, but it is expected that impacts will vary with species (Danks, 1992; Danks, 2004). Insects are highly dependent on the microclimates where they live and in the Arctic they are already at the limits of their ability to adapt in relation to cold hardiness, solar dependence, and sensitivity to desiccation (Danks, 2004). The northward movement of shrubs into new areas creates new microhabitats for insects, in particular pollinators that may be adapted to invade Arctic environments (Klein et al., 2008). Insects have the ability to adapt quite readily to environmental change by moving to areas with more favourable conditions and they have previously shown rapid movement into new habitats created by the retreat of glaciers at the end of the last ice age (Danks, 2004; Klein et al., 2008).
Primary productivity is low in the Arctic compared with other ecosystems (Figure 40). Primary production ranged from a low of 3 kg dry mass/ha/yr to a high of 334 kg dry mass/ha/yr at 12 sites across the Canadian Arctic (Krebs et al., 2003). Within the Arctic, the standing crop of vascular plants was largest in the Western Arctic, while the standing crop of mosses was largest at High Arctic sites such as Melville Island and Ellef Ringnes Island (Krebs et al., 2003).
Source: adapted from Schultz (2005)
Long description for Figure 40
This bar graph showing the following information:
|Major world ecosystems||Relative productivity|
Evidence is accumulating that the Arctic is getting greener and that the net productivity of Arctic ecosystems is increasing.
In lakes and ponds
In Arctic lakes and ponds, primary production has been shown in several studies to have increased, accompanied by changes in algal species assemblages (Smol et al., 2005; Antoniades et al., 2007). In a study designed to evaluate long-term trends in Arctic lake primary production (Michelutti et al., 2005), researchers used reflective spectroscopy to infer chlorophyll a concentrations in sediment cores. Results from the survey of six Baffin Island lakes indicate that there have been pronounced 20th century increases in primary production (see text box below and accompanying Figure 41 showing long-term changes in an Arctic lake). The changes appear to be synchronized with the record of recent climate change. See also discussion in section on Changes in algal and invertebrate species assemblages in lakes and ponds (page 96).
Primary production in Arctic lakes
Lost Pack Lake, Nunavut
The figure shows chlorophyll a reconstructions from Lost Pack Lake, one of six Baffin Island lakes examined for long-term trends. All lakes show dramatic increases of inferred primary production within the most recently deposited sediment, following prolonged periods of comparatively low values (Michelutti et al., 2005). Dating of the sediment cores indicates that these rapid increases started in the late 19th century and continue to the present. The increases are a departure, in most lakes, from relatively stable levels of primary production that persisted for millennia. A widespread increase in freshwater production over much of northern Canada is also inferred from major shifts in species composition of algae in ponds and small lakes in many areas (also detected from studies of sediment cores) (Smol et al., 2005; Antoniades et al., 2007).
The best explanation for this change in algae is climatic warming leading to longer ice-free growing seasons and associated changes in lake ecosystems (Antoniades et al., 2005; Smol and Douglas, 2007b). The changes are most pronounced in the High Arctic, but similar shifts in algal species are found in many locations in the Northern Hemisphere--with changes being more recent in temperate latitudes (Rühland et al., 2008).
Source: Federal, Provincial and Territorial Governments of Canada (2010)
On land, changing primary production and changing physical structure of biomass, represented by increases in distribution and height or vigour of shrubs (Elmendorf et al., 2012b), will affect herbivores and, consequently, predators. Increasing productivity and advancing tree and shrub cover have the potential to alter predator-prey relationships and to facilitate range extensions for some biota, such as certain perching birds that cannot otherwise occupy the tundra, as well as their predators, diseases, and parasites. Ecological changes such as these may be involved in some of the declining trends noted in the section on Ecosystem composition (page 102).
Several measures of primary productivity show marked and widespread increases, and ground observations and experimental studies provide confirmation and understanding of the nature of this change. This section examines trends in productivity from several studies, and the section on Ecosystem structure (page 88) discusses related changes in the tundra biome.
A study undertaken for this report (Ahern et al., 2011) used analysis of the Normalized-Difference Vegetation Index (NDVI) to measure trends in productivity for all of Canada’s ecozones+. NDVI is a measure of the photosynthetic capacity of plant cover from space-based observation. Dense plant canopies have positive values of NDVI while snow and ice have negative NDVI values. The resulting map (Figure 42) shows significant trends in NDVI for the Arctic from 1985 to 2006. Only a small percentage of Canada showed a negative NDVI trend, while 22% of the nation’s land area showed a positive trend, with the largest positive trends being in regions of Arctic tundra, alpine tundra, the Pacific coast, and the eastern prairies.
Trends for the Arctic Ecozone+
Arctic Cordillera: increasing trend in 9.2% of area, decreasing in 0%
The only area that has significant amounts of vegetation is the Labrador Peninsula. The entire peninsula exhibits a strong positive trend in NDVI, particularly the lower elevations bordering Ungava Bay. The vegetation in this area is low vegetation characteristic of tundra regions.
Northern Arctic: increasing trend in 6.8% of area, decreasing in 0.1%
Areas that are particularly notable for increases in NDVI are the northern portion of Banks Island, the Dundas and Sabine peninsulas of Melville Island, the south shore of Bowman Bay on Baffin Island, and the area along the northwestern shore of Hudson Bay. All of these areas are dominated by tundra vegetation.
Southern Arctic: increasing trend in 23.8% of area, decreasing in 0.3%
The Southern Arctic shows extensive areas of increasing NDVI, most notably along the northwestern shore of Hudson Bay, northeast of Great Bear Lake, and the southern portion of the Ungava Peninsula, where the trend is particularly pronounced. All of these increases are in areas of tundra vegetation.
These changes can be seen in a circumpolar perspective in Figure 43. This figure shows the regional trends in increases in tundra and taiga ecosystem NDVI--with major increases being in the Canadian Western Arctic, Alaska, and Siberia.
Primary production (biomass) in tundra ecosystems
Primary production is typically low in tundra ecosystems and depends on latitude and local topographic position (Bliss and Matveyeva, 1992). Gould et al. (2003) mapped the vegetation of the Canadian Arctic (Figure 44) and provided estimates of the biomass (Figure 45) and net production of tundra ecosystems (Figure 46). They estimated that net primary production ranges from less than 20 g/m2/yr in graminoid/forb barrens in the polar desert of the High Arctic, to as much as 1,000 g/m2/yr in the lowland riparian areas of the Low Arctic (Gould et al., 2003).
Experimental warming research at Alexandra Fiord, Ellesmere Island, has shown that a slight increase in summer temperature (about 1°C) results in significant increases in plant growth (Jones et al., 1997; Jones et al., 1999). Significant increases in growth were also noted in similar studies throughout the tundra biome as part of the International Tundra Experiment (ITEX) (Henry and Molau, 1997; Arft et al., 1999; Hollister et al., 2005; Jónsdóttir et al., 2005; Wahren et al., 2005; Walker et al., 2006; Elmendorf et al., 2012a). A meta-analysis of plant community change across the ITEX network showed that up to eight years of experimental warming caused significant increases in growth of shrubs and graminoids (Walker et al., 2006). After nearly 20 years of experimental warming at some sites, the same general patterns were found by Elmendorf et al. (2012a) (Figure 47).
Recent results from long-term studies at Alexandra Fiord, Ellesmere Island, show that there has been a significant increase in biomass (net production) in Canadian High Arctic tundra over the past 20-plus years in response to climate change (Hill, 2006; Hudson and Henry, 2009; Hill and Henry, 2011):
- Wet sedge tundra biomass increased both above and below ground, with increases from the early 1980s to 2005 of 158% in above-ground vegetation biomass, 67% in root biomass, and 139% in rhizome biomass (Figure 48) (Hill and Henry, 2011).
- Snow-bed heath community biomass (at the same site) increased in biomass by 160%, from 33 g/m2 in 1981 to 87 g/m2 in 2008 (Hudson and Henry, 2009). Figure 49 shows the changes in major plant groups in this heath community since 1995.
The increase in biomass in the heath community was in contrast to the effects of long-term experimental warming at the same site, where no increases in overall biomass were found in response to summer season warming of about 1°C (Hudson and Henry, 2010). The increase in average air temperature in the vicinity of the study area over the same time period has been nearly 2°C, which may indicate that certain tundra plant communities will be resistant to relatively small increases in temperature (Figure 50, and see also Figure 14 in the section on Climate trends since 1950 on page 18).
Case study on ecosystem functions and processes: Bylot Island
This case study was prepared by G. Gauthier and D. Berteaux for the Arctic biodiversity assessment (CAFF, 2013) and has been revised based on Gauthier et al. (2013).
Bylot Island is located in the Canadian Arctic Archipelago, at 73°N, 80°W. The 11,067-km2 island is covered by mountains culminating at 1,905 m, an icecap, and several glaciers. The southern part of Bylot is a 1,600-km2 plain gently sloping from about 400 m elevation near the mountains to sea level at the coast, and is bisected by several glacial rivers and creeks (Figure 51). Wet polygons abound in low-lying areas. This plain is covered by tundra vegetation that is rich for its latitude, owing to its southern exposure and protection from northerly winds by high mountains. It is dominated by prostrate dwarf-shrubs, graminoids, forb tundra, and, in polygonal areas, sedge/grass moss wetlands.
Small to intermediate body size species dominate the wildlife assemblage as caribou (Rangifer tarandus) and muskox (Ovibos moschatus) are absent. The bird fauna is especially rich (71 recorded species, 45 as breeders) and comprises a snow goose colony (Chen caerulescens) of about 20,000 pairs. Geese, brown lemmings (Lemmus trimucronatus), and collared lemmings (Dicrostonyx groenlandicus) are the most abundant herbivores. Several species of avian predators, Arctic fox (Vulpes lagopus), and ermine (Mustela erminea) are present.
Average annual temperature is –14.5°C: 4.5°C in summer and -32.8°C in winter (Gauthier et al., 2011). From 1976 to 2010, the area experienced a strong warming trend in the autumn (September to November), increasing 4.3°C over a 35-year period, and in spring and summer, increasing 2.8°C over the same period, but not in winter (December to February) (Gauthier et al., 2011). The Intergovernmental Panel on Climate Change (IPCC) (2007) projects a 3 to 6°C increase in annual average surface air temperatures from 1980–1999 to 2080–2099 for the area.
Monitoring and research
Two 100 km2 study areas centered on glacial valleys have been the focus of continuous monitoring and intensive observational and experimental studies in the fields of climatology, animal and plant ecology, geomorphology, and limnology since 1989. Snow geese, lemmings, Arctic foxes, snowy owls (Bubo scandiacus), long-tailed jaegers (Stercorarius longicaudus), Lapland longspurs (Calcarius laponicus) and, more recently, several species of shorebirds, are the most intensively studied species. Annual plant production in wetlands, plant phenology, and seasonal insect abundance are also monitored. Traditional knowledge on foxes and geese has been collected from members of the community of Pond Inlet located south of Bylot (Gagnon and Berteaux, 2006; Gagnon and Berteaux, 2009).
The trophic dynamic on Bylot Island is dominated by regular, three- to four-year cycles in lemming abundance (Figure 52). Brown lemming populations show strong fluctuations (greater than forty-fold) but collared lemmings show relatively weak fluctuations (about four-fold) (Gruyer et al., 2008). Predators like foxes, ermines, owls, and jaegers quickly track these fluctuations and their number, reproductive activity, and lemming consumption rate increase dramatically in peak lemming years (Therrien, 2012; Tarroux et al., 2012) . This, in turn, has indirect effects on other species like snow geese due, in part, to prey switching by shared predators (Bety et al., 2002). Goose grazing has a significant impact on the wetland vegetation during the summer, but lemming grazing appears to have little impact on plant production, even in years of high abundance (Gauthier et al., 2004). A trophic balance model showed that less than 10 % of the total annual primary production is consumed by herbivores, but 20 to 100% of the herbivore production is consumed by predators (Legagneux et al., 2012). This suggests that predation plays a key role in the functioning of this ecosystem.
Allochthonous subsidies (energy and resources brought in from outside of the ecosystem) may be important to maintain high predator populations. For instance, high goose populations, which are in part fuelled by an agricultural food subsidy obtained in winter, may help to sustain fox populations, especially in low lemming years (Giroux et al., 2012a). In winter, predators like snowy owls and Arctic foxes use the sea ice for extensive periods, though this may be variable among years for foxes (Therrien et al., 2011; Tarroux et al., 2012), and thus they may depend upon the marine environment for their survival.
The Bylot Island environment has experienced significant changes in temperature and snow characteristics over the study period (Gauthier et al., 2013). Cumulative annual thawing degree-days increased by 37% (Figure 53) from 1989 to 2011 and average temperatures increased in spring and summer, including an increase in June of 3°C per decade, over the same time period. Snowmelt date advanced by 4 to 7 days from 1989 to 2012 and snow depth increased by 48% from 1994 to 2010. Mean annual ground temperature at 10 cm depth showed no trend.
The strongest temporal trend detected in the Bylot Island ecosystem is an almost doubling (87% increase) of annual above-ground graminoid production (mostly Dupontia fisheri and Eriophorum scheuchzeri) in wetlands over a 20-year period (Figure 54). This is largely due to an increase in summer temperature--the sum of thawing degree-days explains a significant proportion of the annual variation in plant growth (Gauthier et al., 2011). The proportion of the primary production consumed by herbivores also showed a decreasing trend over time. Annual climatic variation is the most important driver of the annual production of young in several migratory birds, including snow geese, as warm spring temperatures increase their breeding effort and advance their phenology (Dickey et al., 2008; Morrissette et al., 2010).
Despite these strong links with temperature and the observed warming trend, we have yet to see climate-induced trends in most wildlife populations (Gauthier et al., 2013). We have no evidence that lemming cycles have dampened or disappeared in recent years, as has occurred at other sites. Recent analyses suggest, however, that snow depth and quality (density) can affect winter abundance of lemmings and the amplitude of cyclic fluctuations (Duchesne et al., 2011; Bilodeau et al., 2013a; Bilodeau et al., 2013b). The snow goose population has increased significantly over the past 25 years, but this is due to events unrelated to what is happening in the Arctic, such as change in agricultural practices on their wintering ground (Gauthier et al., 2005).
Nonetheless, the more rapid response of lower trophic levels (plants) than higher levels (herbivores and predators) to climate warming may lead to a trophic mismatch. We have evidence that this is already occurring in geese. In years with an early spring, gosling growth is reduced because plants mature too rapidly and the young hatch too late, after the peak in plant nutritive quality (Dickey et al., 2008; Gauthier et al., 2013) (M. Doiron, unpublished data). The median date that geese lay their eggs is related to snowmelt: the earlier the snow melts, the earlier the geese lay their eggs (Figure 55) (Gauthier et al., 2013). The geese undercompensate, however--for an advance in snowmelt date of 10 days, they advanced their laying date by only 3.8 days. Plant growth tracked changes in snowmelt much more closely, indicating a potential for mismatch in timing between hatching and the supply of high-nutrient food in years of early snowmelt.
Other aspects of ecosystem change include changes in species distribution and in wetlands. The red fox (Vulpes vulpes) invaded the Bylot area in the late 1940s (Gagnon and Berteaux, 2009) and is now reproducing regularly. Drainage and loss of some productive wetlands due to the rapid thermal erosion of ice wedges forming polygons during periods of high spring run-off is a source of concern in this area because such events are likely to increase with climate warming (Fortier et al., 2007). Most of Bylot has been part of Sirmilik National Park since 2001, and this study was a main source of data for evaluating potential changes to the ecological integrity of the park (McLennan et al., 2012). However, the Nunavut Field Unit of Parks Canada has now switched to a much less detailed protocol of ecological integrity monitoring that does not include wildlife monitoring on Bylot Island (D. Berteaux, pers. comm.).
Human stressors on ecosystem functions and processes
Climate change is a major stressor for tundra ecosystems in general (ACIA, 2005). At local scales, effects of roads and other potential human disturbances are also important.
While climate change is expected to affect the entire globe, general circulation models consistently indicate that the most severe climate warming will occur in polar latitudes (ACIA, 2005; IPCC, 2007). The Arctic is currently warming at about double the rate of the global average, and some of the most marked changes have occurred in parts of the Canadian Arctic (IPCC, 2007; AMAP, 2011). Paleoecological data confirm that Arctic climate warming during the last century has been well outside the normal range of the previous 400 years and, though there is a natural component to climate change, including at regional and temporal scales related to oscillations, it is clear that the rise in greenhouse gases in the atmosphere due to human activities is a significant human stressor on Arctic ecosystem functions and processes. The impacts of this stressor are woven throughout the discussion in this and other sections of this report. See also the section on Climate trends since 1950 (page 18).
Climate change is a result of anthropogenic increases in greenhouse gases (IPCC, 2007). Trends in the main greenhouse gas, carbon dioxide, are shown in Figure 56 . Air samples from Alert, Nunavut, distant from interference by local sources, are representative of global atmospheric carbon dioxide levels. Annual fluctuations are due to seasonal release of carbon dioxide by plant growth in the Northern Hemisphere. Ice core gas samples from Antarctica show the rapid increase of previously stable carbon dioxide concentrations starting in the late 19th century.
Other global-scale stressors
There are other human stressors related to large-scale change that have known or potential impacts on biodiversity in the Arctic, often by interacting with climate change. A full discussion of these is beyond the scope of this report. Changes to the atmosphere that potentially affect biodiversity include ozone depletion and consequent rise of UV-B levels (ACIA, 2005), and Arctic haze (see box below). Of recent concern is the impact of black carbon (soot) from natural and industrial sources. The net effect of black carbon (both in the atmosphere and settling on snow) on the Arctic climate is complex and not well understood, with most studies being based on modeling of interacting factors. Model-based analyses show that the effect of atmospheric black carbon varies considerably depending on the altitude at which it is present in the atmosphere (Flanner, 2013).
By the 1990s, an unusual haze was noticed in the Arctic that was subsequently linked to emissions of sulphurous and nitrous aerosols. The principal source was air emissions from iron smelting, particularly in Russia. With improved international emission standards, sulphur sources have declined. Non-ferrous metal production remains the dominant source of emissions of acidifying gases to the atmosphere within the Arctic (AMAP, 2006). Other significant anthropogenic sources of sulphur emissions within or close to the Arctic include energy production plants and mining industries. Sources of nitrogen emissions within the Arctic include transportation, in particular shipping and oil and gas activities. Monitoring at Alert showed that, although sulphur aerosols have declined, levels of nitrate aerosol are increasing during the haze season. Forest fires produce soot, which also contributes to haze.
The causes and the effects of acidifying air pollutants and Arctic haze are closely linked to other environmental problems, though these relationships are not well understood. It is not clear, for example, how climate change will influence future acidification and Arctic haze pollution. The effects of haze-producing aerosols on the Arctic climate are complicated by feedbacks between aerosols, clouds, radiation, snow and ice cover, and vertical and horizontal transport processes. Whether the pollutant aerosols cause an overall warming or an overall cooling, or whether the haze itself may affect Arctic ecosystems, is not yet known (AMAP, 2006).
Description of the Condition of the Arctic Ecozone+
- Ecosystem structure
- Changes in the extent and quality of important biomes
- Species assemblages and plant communities
- Major human stressors on ecosystem structure
- Ecosystem composition
- Species diversity
- Trends in species of conservation concern
- Trends in species of special interest
- Insects and pathogens
- Major range shifts of species native to Canada
- Human stressors on ecosystem composition
Changes in the extent and quality of important biomes
Status: land cover classification and biomes
There are several classification systems that define biomes slightly differently. This report uses a generalized land cover classification system derived from the Millennium Ecosystem Assessment (Millennium Ecosystem Assessment, 2005; Frisk, 2011). In the Arctic Ecozone+, polar barrens, polar tundra, and wetlands cover the largest areas (Figure 57). Lakes and rivers, which comprise a small proportion of the total area of the ecozone+, were excluded from the analysis. The five broadly defined biomes in the Arctic ecozone+ are defined as:
- Polar tundra – treeless regions with greater than 50% vegetation cover
- Polar barren – treeless region with less than 50% vegetation cover
- Wetlands – areas saturated for sufficient time to promote wetland or aquatic processes
- Mountains – steep and high lands with specific criteria related to elevation and slope
- Snow/ice/glaciers – masses of perennial snow and ice with definite lateral limits, typically flowing in a particular direction
Climatically, “Arctic” is defined as the region north of the 10°C summer isotherm; ecologically, it is the area north of the treeline.
A more detailed analysis of land cover in the Arctic Ecozone+ (Olthof et al., 2005) has been developed by the Canadian Centre for Remote Sensing and was analysed for this report (Ahern, 2010) (Figure 58). This analysis uses 14 land cover classes, 12 of which are found in the Arctic Ecozone+ (see Appendix 2: Detailed land cover classes for more information) (Ahern, 2010; Frisk, 2011; Ahern et al., 2011). This analysis classifies the Arctic Cordillera as mostly snow and ice, mountains, and sparsely vegetated bedrock or till colluvium. The Northern Arctic is dominated by barren ground or sparsely vegetated bedrock, till colluviums, or barren ground. The Southern Arctic is dominated by low and dwarf shrubs.
Analyses conducted for the production of the circumpolar Arctic vegetation map (Walker et al., 2005) produced the following summary results:
Within the Arctic (total area = 7.11 x 106 km2), about 5.05 x 106 km2 is vegetated. The remainder is ice-covered.
About 26% of the vegetated area is erect shrublands, 13% peaty graminoid tundra, 13% mountain complexes, 12% barrens, 11% mineral graminoid tundra, 11% prostrate-shrub tundra, and 7% wetlands.
Canada, among Arctic nations, has the largest portion of Arctic terrain (36%) and has by far the most in the High Arctic ecosystem categories (63% of the global total), mostly associated with abundant barren types and prostrate dwarf-shrub tundra, whereas Russia has the largest area in the Low Arctic, predominantly low-shrub tundra.
Habitat diversity is also an important component of ecosystem structure. Some habitat types assume a particular importance because they are rare and are important, even critical, to the life cycle of particular species. The following examples illustrate the patchy distribution and importance of small habitat elements in a large landscape:
- Eskers occupy a small proportion of the Arctic but may be critically important as denning habitat for wolves, grizzly bears, and ground squirrels (Mueller, 1995; McLoughlin et al., 2004).
- Grizzly bears in the Mackenzie Delta can only den in river cut-banks in consolidated, upland soils of Pleistocene origin (or, rarely, pingos), not in more recent fluvial deposits of Holocene origin (Harding, 1976).
- Nesting shorebirds and geese tend to associate with wetland and riparian habitats that are unevenly distributed on the coastal plain (Brown et al., 2007c).
Trends in polar biomes
Reduction in tundra
There is evidence of approximately a 20% decrease over the past 25+ years in the extent of Tundra Climate, a measure of the cold temperature and low precipitation conditions that support the polar tundra, barrens and ice and snow biomes (Figure 59) (Wang and Overland, 2004). The steady decline in the area of Tundra Climate is matched by reductions in the areas with a tundra signature in primary productivity as measured by NDVI and the strongest changes in Tundra Climate have occurred in Northwest Canada in the past 25+ years (Wang and Overland, 2004; Bhatt et al., 2010, and clearly visible in Figure 63).
The warming climate may have been the cause of increased biomass (inferred from increased NDVI) since 1986 on the Porcupine caribou range in northern Yukon and adjacent Alaska (Henry et al., 2012) (Figure 61). Structural changes south of the Arctic Ecozone+ also influence ecosystems within the ecozone+. For example, increased fire frequency is believed to be the reason for a decrease in the cover of forest tundra woodland in the winter range of the Bathurst caribou herd (Chen et al., 2013). As discussed in the section on Primary productivity (page 71), increases in biomass have been measured since 1980 in permanent plots at a High Arctic site (Figure 48 and Figure 49). These increases in biomass are likely responses to the warming climate (Figure 50).
Comparisons of historical and contemporary aerial photographs also provide evidence that Arctic vegetation has undergone significant shifts in recent decades. Increased shrub cover has been confirmed in two repeat photography studies in northern Alaska (Sturm et al., 2001; Tape et al., 2006) and one on Herschel Island in Canada (Myers-Smith et al., 2011b). A study in the Mackenzie Delta region of Canada’s Southern Arctic showed that there has been an increase in deciduous shrubs in the region over the past 40 years (Lantz and Kokelj, 2008). A similar dramatic increase in shrub cover was documented by Tremblay et al. (2012) for the region around the community of Kangisuujuaq in Nunavik east of Ungava Bay (Figure 62).
Plot-level data offer a more detailed view of vegetation changes than either remote sensing or aerial photography. A synthesis of resampled plots in tundra ecosystems around the world showed changes in vegetation over the past 10–30 years that are consistent with responses to warming, but with considerable variation among the sites (Elmendorf et al., 2012b). The responses were strongly controlled by the initial conditions of the site and the changes in soil moisture (Callaghan et al., 2011c; Elmendorf et al., 2012b). Resampling of plots established in the 1970s in Alaska yielded results consistent with a warming and drying trend, in which moist and wet community types tended to be replaced by dry community types through time (Hinzman et al., 2005). After resampling of control plots from a 13-year experiment at Toolik Lake, Alaska, Shaver et al. (2001) found that graminoids, mosses, and lichens were decreasing and evergreen shrubs were increasing in abundance. In a similar resampling study at a High Arctic site on Ellesmere Island, Hill and Henry (2011) found a significant increase in biomass of wet sedge tundra between 1980 and 2005 (see the section on Primary production (biomass) in tundra ecosystems on page 75).
Increased growth and cover by shrubs and infilling by tree species also threatens tundra at its southern margins in the forest-tundra ecotone. Correlations between temperature and treelines throughout the paleoecological record strongly support the notion that climate warming will advance treelines. For example, during the Little Ice Age, shrub tundra increased while treelines retreated (Tinner et al., 2008). As the northern boundary of tree distributions is often temperature limited, northern migration of trees might seem almost inevitable. However, caution in predicting the rate and locations of treeline expansion is warranted, as rates of treeline expansion may lag in some areas due to seed availability, disturbance frequency, permafrost changes, and moisture constraints (Hofgaard and Harper, 2011; Walker et al., 2012; Henry et al., 2012).
White spruce trees along the northern Quebec–Labrador treelines show different responses according to their position relative to the sea. Along the coast, invading spruce are taking hold above the current treeline, while in the interior recent warming has not been strong enough to change the regressive tree line trajectory (Payette, 2007). Treelines in the forest-tundra areas of Quebec have risen slightly either through establishment of white spruce from seed or through growth of stunted spruce already established on tundra hilltops (Gamache and Payette, 2005; Caccianiga and Payette, 2006).
Treeline is advancing in other regions of the circumpolar Arctic that are increasing in productivity (Ims et al., 2013). Expansion of white spruce forests into areas previously occupied by tundra has been documented in numerous locations in Alaska (Suarez et al., 1999; Lloyd and Bunn, 2007), while recent studies have also shown that moisture is an important limiting factor in treeline advance (Ohse et al., 2012). Treeline advance into the tundra has also been documented in mountainous areas of Siberia (Kharuk et al., 2010; Kirdyanov et al., 2012) and northwestern Siberia, where forests have expanded into tundra at the rate of 3 to 10 m per year (Kharuk et al., 2006).
Vegetation is a key component in surface energy balance, regulating climate at regional and global scales. The observed and projected increases in the cover and density of shrub species and the increasing density of trees in forest tundra areas (Sturm et al., 2001) will alter the surface energy balance of northern terrestrial systems by decreasing the albedo and increasing absorption of solar radiation (Chapin III et al., 2005; Sturm et al., 2005). The increased absorption of solar radiation will increase the heating of the atmosphere. Chapin et al. (2005) estimated that a change from tundra to boreal forest will increase the heating of the atmosphere by approximately 4 W/m2, which is similar to the effect of doubling the CO2 concentration in the atmosphere. This will be an increasingly important positive feedback.
Species assemblages and plant communities
Changes in algal and invertebrate species assemblages in lakes and ponds
Striking and often unprecedented changes in the biota of Arctic lakes since the1850s have been linked to ecological shifts consistent with climate warming (Smol et al., 2005). Many algal and invertebrate communities of Arctic lakes and ponds have crossed ecological thresholds and a plausible explanation is recent warming (Smol and Douglas, 2007b). Other potential causes including ultraviolet radiation, nutrient enrichment, and atmospheric transport of pollutants cannot account for the scale, the nature, and the timing of the observed changes in algal communities. The best explanation is climatic warming leading to longer ice-free growing seasons and associated limnological changes including thermal stratification, habitat availability, and lakewater chemistry (Antoniades et al., 2005; Smol and Douglas, 2007b).
Analysis of sediments to identify and measure types of fossil diatoms, as well as other indicators, has detected major changes in species assemblages starting in the mid-19th century. Sediments deposited before this time show that diatom communities were relatively stable over centuries, and in some cases millennia, compared to the recent changes. Analysis of cores from 55 circumpolar Arctic lakes (Smol et al., 2005) showed that lakes have become more productive with more diverse algal and invertebrate communities. Lakes in the Canadian High Arctic experienced abrupt shifts in favour of diatoms that thrive in littoral habitats and mossy substrates, consistent with longer growing seasons and habitat changes under warming conditions. In deeper lakes on Baffin and Ellesmere islands and in lakes throughout the subarctic, the earlier onset and longer duration of thermal stratification becomes increasingly important, with pronounced increases in planktonic diatoms that fare better under conditions of less ice cover and greater thermal stratification, and concurrent decreases in benthic diatom taxa associated with cold conditions and extensive ice cover. In contrast, no major shifts in diatom assemblages were observed over the last two centuries in lakes in Labrador and northern Quebec, areas that have experienced little warming until very recently.
Local topographic and geologic factors also influence the response to warming (Smol et al., 2005). Sediments from lakes on Ellef Ringnes Island showed a transition to different, more diverse diatom communities starting about 1850 (Figure 63), typical of results reported from previous research (Smol et al., 2005). The same shift occurred in lake sediments near Alert, Ellesmere Island, but later, a difference that is attributed to varying sensitivity of the lakes to environmental change (Antoniades et al., 2005). For example, as reviewed by Smol and Douglas (2007b), lakes of different size would be expected to change at different time scales to climatic changes, with larger lakes typically responding more slowly than shallow ponds.
In Canada’s most northerly lake, Ward Hunt Lake, located at 83°N latitude and with perennial ice cover, diatoms appear only in the top 2.5 cm of sediment, corresponding to about the past 200 years of the 8,500 years of sediment records examined (Antoniades et al., 2007). This was accompanied by an increase in photosynthetic pigments (a measure of the standing crop of algae) of two orders of magnitude.
This pronounced shift in diatom assemblage composition is not a trend exclusive to the Canadian Arctic, or even just in northern latitudes. A synthesis of diatom-based paleolimnological studies of over 200 lake sediment records from around the Northern Hemisphere (Rühland et al., 2008) showed a general pattern of species-specific shifts in diatom assemblage composition since the mid-19th century that often followed millennia of relative stability in the diatom communities. Their analysis found that these changes occurred significantly earlier (by about 100 years) and often more dramatically in Arctic lakes than in lakes from temperate latitudes. They attributed this earlier and more rapid shift in Arctic lakes to the greater degree of warming that has occurred in the Arctic. The diatom species shifts observed at all latitudes were remarkably similar and occurred in conjunction with observed changes in freshwater habitat and structure that were linked to substantially warmer temperatures and changes in ice cover.
Remains of invertebrates in sediments show that the changes in lake conditions and algal communities have been transmitted to higher trophic levels. For example, aquatic insect (chironomid) populations in three Ellesmere Island ponds greatly expanded and diversified at the same time as diatom communities changed (Quinlan et al., 2005). Changes in both algae and invertebrate species assemblages are most pronounced in areas that have warmed the most, such as parts of the Canadian High Arctic (Smol et al., 2005).
Changes in tundra plant communities
Species diversity was found to decline in warming experiments across the tundra biome after up to six years of warming (Walker et al., 2006) and remained low after more than 10 years (Elmendorf et al., 2012a). The major factor in the decline in diversity was the loss of bryophyte and lichen species, which likely were shaded by the increasing cover and height of the vascular plants (Figure 47).
As species composition changes along gradients, and as local gradients in soil moisture and exposure can be relatively short spatially, tundra plant communities can be relatively distinct (Bliss and Matveyeva, 1992). The local topography imposes exposure and moisture gradients by influencing snow depth. At the regional scale, the effects of latitude and distance from the ocean are the major gradients affecting plant community structure (Walker et al., 2005).
Snow bed tundra plant species have been identified as potentially at risk if snowfall becomes less and melt is more rapid, leading to longer growing seasons with drier soils (Bjork and Molau, 2007). Increased precipitation during the summer may also affect the species composition of tundra plant communities, especially in the dry polar semi-desert areas. The trajectory of precipitation regimes in the Arctic, however, is difficult to predict (ACIA, 2005) and there have been very few experimental studies altering both winter and summer precipitation in combination with warming. An experimental study in High Arctic tundra plant communities showed that watering with about 13 L every 10 days for three seasons did not affect net primary production (Henry et al., 1986).
Tundra plant communities are showing changes across the biome that are consistent with responses to warming (see also the section on Primary production (biomass) in tundra ecosystems on page 75). In a study of 158 plant communities across 46 sites that were resampled between 1980 and 2010, Elmendorf et al. (2012b) found increases in canopy height and height of most vascular growth forms, increased abundance of shrubs and litter, and decreased cover of bare ground (Figure 64). However, the responses varied greatly among climate zones and depended strongly on moisture and permafrost conditions. Overall, the responses were very similar to a biome-wide analysis of warming experiments (Elmendorf et al., 2012a) and provide plot-level support to the changes in reflectance measured by NDVI (Bhatt et al., 2010).
Major human stressors on ecosystem structure
As with the linked trends in ecosystem processes and functions, climate change is the main human-induced stressor on ecosystem structure in the Arctic Ecozone+.
Stressors related to industrial development, settlements, and human activity
Fragmentation and disturbance
Overall, the degree of human-induced fragmentation in the Arctic Ecozone+ is extremely low, as there are few roads and other linear features. Fragmentation, however, is a stressor of concern at the regional level, with the potential to become more widespread with increasing human population and industrial activity in the Arctic.
Disturbance from roads, pipelines, and other linear facilities, and from vehicle and aircraft traffic, has been widely hypothesized to affect wildlife. These hypotheses have been tested experimentally and by behavioural observations in many Arctic and subarctic settings (Slaney, 1975; Harding and Nagy, 1980; Gunn, 1984; James and Stuart-Smith, 2000; Gnieser, 2000; Dyer et al., 2002; Lunn et al., 2004). In general, the studies show modest impacts on animal energetics related to fleeing industrial activities, modest habitat losses from avoidance of areas with industrial facilities and activities, and little or no reluctance of migrating animals (for example, caribou) to cross linear facilities (Nellemann and Cameron, 1998; Wolfe et al., 2000b). In some cases, wild animals are attracted to such facilities. For example, caribou were attracted to the tailings pond and airstrip at Lupin Gold Mine, Contwyoto Lake (Gunn et al., 1998). It is, however, clear that some species, including caribou, grizzly bears, wolverines, and possibly muskoxen, do avoid areas of intensive activity and disturbance, at least for some life history requirements (Harding and Nagy, 1980; Dumond, 2006; Krebs et al., 2007).
A study in the Southern Arctic and Taiga Shield examined the cumulative impacts of human activities and associated infrastructure (mainly related to diamond mine development) on the distribution of caribou, wolves, grizzly bears, and wolverines over an area of 190,000 km2, 400 km northeast of Yellowknife, from 1995–2000 (Johnson et al., 2005). The study was conducted as part of the West Kitikmeot/Slave Study (see Case study on environmental governance: Kitikmeot on page 183). Mines and other major developments had the greatest effect on species occurrence, followed by exploration activities and outfitter camps, but there was not a uniform response, with some carnivores being attracted to these facilities. Modeling techniques were used to assess the reduction in habitat effectiveness (i.e., to what degree the animals selected poorer quality habitats in response to proximity to disturbances). Grizzly bears and wolves showed the strongest negative response, followed by caribou and wolverines. Seasonally, however, the biggest impact was on post-calving caribou: models suggested a 37% reduction in the area of the highest quality habitats and an 84% increase in the area of lowest quality habitats.
Effects of human infrastructure on predator habitat
Human infrastructure (roads and airstrips, often with culverts, telecommunications poles and towers, navigation beacons, and buildings) greatly enhance habitat structure for some species by providing new denning and nesting opportunities (Liebezeit et al., 2009; Reid et al., 2011; Wilson et al., 2013). These new structures generally benefit predators (foxes and raptors) and thereby change the spatial distribution of the foraging areas of these predators. This can result in changes to the reproductive success of various prey species (notably shorebirds, (Liebezeit et al., 2009)), and the nesting distribution of competing raptors (Reid et al., 2011). Mitigation measures include minimizing construction of such structures, designing structures to reduce raptor-nesting potential, and minimizing the extent of road and pipeline networks.
Effects of food supplementation on trophic interactions
Food supplementation results when garbage disposal is centralized, creating feeding opportunities in a few locations for scavengers such as foxes and gulls. The scavengers benefit, with potential spillover effects on other species. The expansion of the red fox into Arctic Canada coincided with centralized community development for Inuit in the 1930s through 1960s, and was likely related to this enhancement of food resources at certain limited sites (Marsh, 1938; Macpherson, 1964). Red fox outcompete Arctic fox, so the red fox expansion can result in loss of Arctic foxes from ecosystems. The two species co-exist on the Yukon North Slope, where there is currently no food supplementation (Gallant et al., 2012).
Case study: Cruise ship tourism, potential emerging stressor on ecosystem structure
Historically, ice conditions have prohibited most commercial shipping in the Canadian Arctic, and it was not until 1984 that cruise tourism got underway with the first cruise ship voyage through the Northwest Passage, by the MS Explorer (Jones, 1999). Increasing reduction of summer sea ice (see section on Sea ice on page 36) makes most of the Canadian Arctic increasingly accessible to cruise ship tourism (Stewart et al., 2007; Stewart et al., 2010).
From 1984 to 1991, activity was sporadic. From 1992 on, both numbers of cruises and diversity of cruise routes increased steadily. The number of planned cruises in Arctic Canada doubled between 2005 and 2006 (from 11 to 22) and continued to increase from 2007 to 2010 at a rate of approximately 10% per year (Dawson, 2012). Cruise ship activity has now begun to level off, likely due to the 2009 economic recession, limited number of appropriate ships available, and the fluctuating tourism market for Arctic cruises (Dawson, 2012). Lack of infrastructure (such as port facilities and ground transport), as well as few tourist sites and community activities for groups of visitors, also impose limitations on the growth of the cruise ship industry.
During the 2007 season, 23 separate cruises, run by six different companies, brought approximately 2,110 visitors to the Canadian Arctic (Stewart et al., 2010). Many of these ships made planned, and occasionally unplanned, stops in national parks such as Auyuittuq National Park and Sirmilik National Park on Baffin Island, and Quttinirpaaq National Park on Ellesmere Island, as well as to nearby locations (Table 6) (Stewart et al., 2008).
|Auyuittuq National Park||4||3||2|
Source: Stewart et al. (2008)
Most cruising expeditions encourage tourists to hike in the parks. Without regulated access and trails, large number of tourists hiking in parks and in other accessible coastal sites could have enduring negative impacts on the surrounding ecosystems. During investigations of abandoned settlements on three islands in the eastern Canadian Arctic Archipelago, Forbes (1996) found that flora of the Arctic has a limited number of species able to respond to disturbance and anthropogenically disturbed patches may be extremely persistent. Even relatively low-intensity, small-scale disturbances have immediate and persistent effects on Arctic vegetation and soils. Where slope is minimal, such disturbances are capable of expanding over large areas in as short a time as four years. The result is an artificial mosaic of patches of highly variable quality and quantity, compromising feeding and nesting habitats for vertebrate herbivores (Forbes, 2001). Other potential impacts include wildlife disturbance, especially of nesting seabirds (Marquez and Eagles, 2007), increased risk of introduction of invasive plants, and large local increases in human waste that, if not disposed of properly, could result in habitat degradation.
Composition refers to the species and species groups that make up the Arctic Ecozone+.
Sources of knowledge about past distribution and abundance of Arctic species include Aboriginal Traditional Knowledge and records and narratives written by early explorers, who often recorded information from their Dene and Inuit guides along with their own observations and survey results. American and Canadian governments began sponsoring biological surveys in the late 19th century. These surveys brought to the attention of the world the iconic wildlife species of the Canadian Arctic: polar bears (Ursus maritimus), barren-ground grizzly bears (Ursus arctos), vast herds of migratory barren-ground caribou (Rangifer tarandus groenlandicus), small, nearly white Peary caribou (Rangifer tarandus pearyi) of the Northern Arctic, Arctic fox (Vulpes lagopus), Arctic hares (Lepus arcticus), lemmings (Arvicolinae) and muskox (Ovibos moschatus); marine mammals such ringed (Pusa hispida) and bearded (Erignathus barbatus) seals, bowhead (Balaena mysticetus) and beluga (Delphinapterus leucas) whales and narwhals (Monodon monoceros); walruses (Odobenus rosmarus); resident birds such as the snowy owl (Bubo scandiacus) and gyrfalcon (Falco rusticolis), and the Arctic breeding grounds of migratory birds such as the now presumably extinct Eskimo curlew (Numenius borealis).
Species diversity is relatively low in the Arctic Ecozone+, declining on a gradient from south to north and west to east. The northwestern region of the ecozone+ is part of the Beringia region and its diversity is enhanced by endemic or globally rare species that survived the last glaciations (Cannings et al., 2013).
Diversity of vascular plants in Canadian tundra ecosystems declines with decreasing temperature along the latitudinal gradient (Figure 65) (Rannie, 1986). While the number of bryophyte and lichen species also declines with latitude, the decrease in richness is not as great as for vascular species. Vertebrate species richness also declines from south to north, as shown in Figure 66.
Arctic ecosystems have relatively simple food webs relative to more southern ecosystems (see section on Community and population dynamics on page 56). More complex ecosystems have multiple energy flow pathways and more species that occupy similar niches. The low species diversity and generally simple Arctic ecosystems may limit the ability of these ecosystems to resist perturbation in the first place, and to recover when damaged. However, the link between food web complexity and vulnerability to changes is not well-understood (see section on Community and population dynamics).
Nunavut, with a land area of 1.9 million km2, almost all of which is within the Arctic Ecozone+ , has 38 species of terrestrial mammals, 151 species of birds breeding in the territory, 20 species of freshwater and anadromous fishes, 8 amphibian and 1 reptile species, 47 butterfly species, and 626 species of vascular plants (Department of Environment, 2013a). By contrast, British Columbia, with half the land area, has 47 regularly occurring mammal species, 528 species of birds breeding in the province, 120 of freshwater and anadromous fishes, 22 amphibians and 22 reptile species, 275 of butterfly species, and 3,097 species of vascular plants (BC Conservation Data Centre, 2007). This comparison illustrates the relative simplicity, in terms of species richness, of Arctic ecosystems.
Trends in species of conservation concern
The biological status of species refers to their risk of extinction as determined through standard criteria and assessed at the global, national, or provincial/territorial level. Subspecies and populations can be assessed and included in the term “species”. Status at the global level is assessed by the International Union for the Conservation of Nature (IUCN). Species considered at risk in the world or of conservation concern for this overview are those assessed at a category of Near Threatened, Vulnerable, Endangered, or Critically Endangered.
The biological status of species at risk in Canada can be assessed as “Endangered” (facing imminent extirpation or extinction), “Threatened” (likely to become endangered if nothing is done to reverse threats), and “Special Concern” (at risk of becoming threatened or endangered). Species that are assessed by the Committee on the Status of Endangered Wildlife in Canada (COSEWIC) as being at risk, but are not yet legally protected under Canada’s Species at Risk Act (SARA), are undergoing consultation for listing (Table 7). Although the Arctic Ecozone+, when compared with other ecozones+, has the lowest number of species listed as at risk under SARA, several iconic or keystone species and species important to Inuit culture are under review.
Below are discussions of status and trends of selected species of conservation concern.
|Species with a Canadian distribution||Species||Subspecies or Population||COSEWIC Status||SARA Status||SARA Schedule|
|Plants||Hairy braya (Braya pilosa)||Hairy braya (Braya pilosa)||Endangered||-||-|
|Plants||Porsild’s bryum (Haplodontium macrocarpum)||Porsild’s bryum (Haplodontium macrocarpum)||Threatened||Threatened||Schedule 1|
|Plants||Spiked saxifrage (Micranthes spicata)||Spiked saxifrage (Micranthes spicata)||Threatened||-||-|
|Fishes||Atlantic cod (Gadus morhua)||Arctic Lakes population||Special Concern||No Status||No Schedule|
|Fishes||Atlantic wolffish (Anarhichas lupus)||Atlantic wolffish (Anarhichas lupus)||Special Concern||Special Concern||Schedule 1|
|Fishes||Bering wolffish (Anarhichas orientalis)||Bering wolffish (Anarhichas orientalis)||Data Deficient||Special Concern||Schedule 3|
|Fishes||Blackline prickleback (Acantholumpenus mackayi)||Blackline prickleback (Acantholumpenus mackayi)||Data Deficient||Special Concern||Schedule 3|
|Fishes||Bering cisco (Coregonus laurettae)||Bering cisco (Coregonus laurettae)||Special Concern||No Status||No Schedule|
|Fishes||Bull trout (Salvelinus confluentus)||Western Arctic populations||Special Concern|
|Fishes||Dolly varden (Salvelinus malma malma)||Western Arctic population||Special Concern||No Status||No Schedule|
|Fishes||Fourhorn sculpin (Myoxocephalus quadricornis)||Freshwater form||Data Deficient||Special Concern||Schedule 3|
|Fishes||Northern wolffish (Anarhichas denticulatus)||Northern wolffish (Anarhichas denticulatus)||Threatened||Threatened||Schedule 1|
|Fishes||Roundnose grenadier (Coryphaenoides rupestris)||Roundnose grenadier (Coryphaenoides rupestris)||Endangered||No Status||No Schedule|
|Fishes||Spotted wolffish (Anarhichas minor)||Spotted wolffish (Anarhichas minor)||Threatened||Threatened||Schedule 1|
|Fishes||Thorny skate (Amblyraja radiata)||Thorny skate (Amblyraja radiata)||Special Concern||No Status||No Schedule|
|Birds||Buff-breasted sandpiper (Tryngites subruficollis)||Buff-breasted sandpiper (Tryngites subruficollis)||Special Concern||No Status||No Schedule|
|Birds||Eskimo curlew (Numenius borealis)||Eskimo curlew (Numenius borealis)||Endangered||Endangered||Schedule 1|
|Birds||Harlequin duck (Histrionicus histrionicus)||Eastern population||Special Concern||Special Concern||Schedule 1|
|Birds||Ivory gull (Pagophila eburnea)||Ivory gull (Pagophila eburnea)||Endangered||Endangered||Schedule 1|
|Birds||Red knot (Calidris canutus)||rufa subspecies (C. canutus rufa)||Endangered||Endangered||Schedule 1|
|Birds||Red knot (Calidris canutus)||roselaari type (C. canutus roselaari type)||Threatened||Threatened||Schedule 1|
|Birds||Red knot (Calidris canutus)||islandica subspecies (C. canutus islandica)||Special Concern||Special Concern||Schedule 1|
|Birds||Ross’s gull (Rhodostethia rosea)||Ross’s gull (Rhodostethia rosea)||Threatened||Threatened||Schedule 1|
|Birds||Short-eared owl (Asio flammeus)||Short-eared owl (Asio flammeus)||Special Concern||Special Concern||Schedule 1|
|Birds||Tundra peregrine falcon (Falco peregrines tundrius)||Tundra peregrine falcon (Falco peregrines tundrius)||Non-active||Special Concern||Schedule 3|
|Land Mammals||Barren-ground caribou (Rangifer tarandus groenlandicus)||Dolphin and Union population||Special Concern||Special Concern||Schedule 1|
|Land Mammals||Grizzly bear (Ursus arctos)||Western population||Special Concern||No Status||No Schedule|
|Land Mammals||Peary caribou (Rangifer tarandus pearyi)||Peary caribou (Rangifer tarandus pearyi)||Endangered||Endangered||Schedule 1|
|Land Mammals||Polar bear (Ursus maritimus)||Polar bear (Ursus maritimus)||Special Concern||Special Concern||Schedule 1|
|Land Mammals||Wolverine (Gulo gulo)||Western population||Special Concern||No Status||No Schedule|
|Marine Mammals||Atlantic walrus (Odobenus rosmarus rosmarus)||Atlantic walrus (Odobenus rosmarus rosmarus)||Special Concern||No Status||No Schedule|
|Marine Mammals||Beluga whale (Delphinapterus leucas)||Southeast Baffin Island-Cumberland Sound population||Non-active||Endangered||Schedule 2|
|Marine Mammals||Beluga whale (Delphinapterus leucas)||Eastern Hudson Bay population||Endangered||No Status||No Schedule|
|Marine Mammals||Beluga whale (Delphinapterus leucas)||Ungava Bay population||Endangered||No Status||No Schedule|
|Marine Mammals||Beluga whale (Delphinapterus leucas)||Cumberland Sound population||Threatened||No Status||No Schedule|
|Marine Mammals||Beluga whale (Delphinapterus leucas)||Eastern High Arctic–Baffin Bay population||Special Concern||No Status||No Schedule|
|Marine Mammals||Beluga whale (Delphinapterus leucas)||Western Hudson Bay population||Special Concern||No Status||No Schedule|
|Marine Mammals||Bowhead whale (Balaena mysticetus)||Bering–Chukchi–Beaufort population||Special Concern||Special Concern||Schedule 1|
|Marine Mammals||Bowhead whale (Balaena mysticetus)||Eastern Arctic population||Non-active||Endangered||Schedule 2|
|Marine Mammals||Bowhead whale (Balaena mysticetus)||Eastern Canada–West Greenland population||Special Concern||No Status||No Schedule|
|Marine Mammals||Grey whale (Eschrichtius robustus)||Eastern North |
|Special Concern||Special Concern||Schedule 1|
|Marine Mammals||Humpback whale (Megaptera novaengliae)||Western North Atlantic population||Not at Risk||Special Concern||Schedule 3|
|Marine Mammals||Killer whale (Orcinus orca)||Northwest Atlantic/ Eastern Arctic population||Special Concern||No Status||No Schedule|
|Marine Mammals||Narwhal (Monodon monoceros)||Narwhal (Monodon monoceros)||Special Concern||No Status||No Schedule|
Those species on SARA Schedules 2 and 3--or as yet on no schedule--are at some stage of assessment for consideration for listing on Schedule 1. Note that COSEWIC assessment and SARA listing are two separate processes. COSEWIC assessment as Endangered does not automatically mean SARA listing.
Source: SARA Registry, current to June 2013 (Government of Canada, 2013)
This section is based on the ESTR technical thematic report Northern caribou population trends in Canada (Gunn et al., 2011c) and includes updates to that report and expanded discussion based on recent assessments and research results. See also the Main threats to caribou section (page 170). The classification of caribou used in this report follows the current Species at Risk Act (SARA) classification system. In 2011, COSEWIC adopted 12 designatable units for caribou in Canada that will be used in caribou assessments and subsequent listing decisions under SARA beginning in 2014 (COSEWIC, 2011).
Over the last 50 years the number of Peary caribou have declined from about 44,000 to about 11,000 to 12,000 caribou (Species at Risk Committee, 2012a) and two geographic populations appear to have disappeared. The rate of decline has varied over time and between the different geographic populations, with both reversals of some declines and absence of recovery for other populations. Survey intervals are irregular: only two of the six geographic populations (Banks and Bathurst Island complex) have been surveyed at regular intervals. Overall trends differ between the northern High Arctic islands and the larger, southern (mid-Arctic) islands and Boothia Peninsula. Although these larger mid-Arctic islands had a relatively high abundance of Peary caribou until the 1990s, declines occurred. Populations on the Boothia Peninsula and on Prince of Wales and Somerset islands had almost disappeared by the 1990s. There is no evidence for recovery over a 20-year period on Banks and northwest Victoria islands, despite severely restricted harvesting since the early 1990s (Species at Risk Committee, 2012a).
Geographic populations of Peary caribou are identified based on knowledge of movements, seasonal distribution, and genetics (Jenkins et al., 2011; Species at Risk Committee, 2012a). The Arctic islands are linked for much of the year by sea ice, and Peary caribou seasonally migrate across the ice. Sea ice coverage, however, is minimal in fall, which likely influences the rut distribution and thus the maintenance of island geographic populations. Jenkins et al. (2011) described seven geographic populations on the more northern islands: Melville and Prince Patrick, the Prime Minister group, the Bathurst Island group, Devon Island, Axel Heiberg, Ellesmere, and the Ellef group. Two of four geographic populations on the southern islands (Prince of Wales and Somerset; Boothia Peninsula) have essentially disappeared since the 1990s, leaving the two southern island populations of Banks and northwest Victoria islands.
On the High Arctic islands, weather is an overwhelming influence as periodic severe winters trigger large-scale mortality and reduction in productivity (Miller and Gunn, 2003; Harding, 2004). Although the signals of climate warming are strong in the High Arctic (Zhang et al., 2011), relating those trends in weather to changes in Peary caribou abundance is uncertain, partly because of high annual variability in climate and infrequent monitoring for most Peary caribou. The other reason is that harvest and predation also affect Peary caribou abundance. Muskox trends in abundance tend to differ from Peary caribou, although this is area-specific. Muskox increases relative to Peary caribou decreases have raised the question of competition. The role of intra- or inter-specific competition for forage is conjectural, as diet and habitat selection differ considerably between caribou and muskoxen (Gunn and Dragon, 2002). On Banks Island, however, there was overlap in some plants eaten by the two species, with, for example, willow being eaten by both Peary caribou and muskoxen (Larter and Nagy, 2004). This suggests that a competitive relationship could occur between the two species. Less emphasis has been placed on determining whether the increasing muskox abundance supported increased wolf numbers which, in turn, could increase predation rates on Peary caribou (Gunn and Dragon, 2002). Even less attention has been given to studying the relationship between caribou and muskoxen and their parasites. Hughes et al. (2009), however, discussed levels of intestinal nematode worms and warble flies in muskoxen and caribou for the range of the Dolphin and Union caribou.
There are limitations to describing the trends for Peary caribou due to the infrequency of surveys and the relatively brief period over which surveys have occurred. It is uncertain whether the documented high numbers of Peary caribou in the early 1960s followed by a decline and prolonged low numbers are: 1) part of regular fluctuations; 2) a period of relative stability within an unusually prolonged decline; or 3) atypically high peak numbers. Possibly Peary caribou regularly fluctuate in abundance, driven by a relationship between amounts of forage and caribou abundance (Tews et al., 2007a; Tews et al., 2007b). Alternatively, Peary caribou are in a “non-equilibrium grazing system” where sporadic, unpredictable weather events affect vital rates and population trends (Caughley and Gunn, 1993; Behinke, 2000).
Trends in abundance
Melville and Prince Patrick islands: The infrequent surveys documented steep declines between 1961 (the first range-wide aerial survey) and 1997. On Prince Patrick Island there was a 95% decline (from 1,797 to 84 year-plus old caribou). On Melville Island, Peary caribou declined by 92% (from 10,366 to 787 year-plus old caribou) (Tener, 1963; Gunn and Dragon, 2002). However, between 1997 and 2012, trends were reversed and a strong recovery was apparent for Melville, Prince Patrick, Byam Martin, Eglinton, and Emerald islands (Davison and Williams, 2012). The 2012 estimate was about 6,000 Peary caribou which, while not a complete recovery to the abundance recorded in 1961 (16,000), is the highest abundance recorded since then. See Figure 67 a and Figure 67 b. Lack of sea ice prevented the aerial survey of Mackenzie King, Brock, and Borden islands in 2012.
Prince of Wales and Somerset islands: Peary caribou seasonally crossed the sea ice between the islands in this group and some caribou also wintered on the Boothia Peninsula. Between 1974 and 1980, caribou numbers were stable in the range 4,000–6,000, one of the largest Peary caribou populations in the 1970s and 1980s (Gunn et al., 2006). There was a 15-year hiatus in surveys until 1995, when only a few caribou were found (Gunn and Dragon, 1998). In 2004, no caribou were seen during an aerial survey of the islands (McRae et al., 2010). See Figure 67 c.
Bathurst Island (and its satellite islands): Between 1961 and 1974, Peary caribou numbers declined by an order of magnitude (Miller, 1991a). Between 1974 and 1994, numbers recovered to the 1961 level (Miller, 1991a). An abrupt decline followed and, by 1997, fewer than 100 caribou remained (Gunn and Dragon, 2002). A survey in 2001 revealed the trend was for a recovering population (McRae et al., 2010). The Bathurst Island complex was resurveyed in May 2013, with preliminary estimates indicating that the herd has tripled in size since 2001 (Department of Environment, 2013a). The final population estimate based on this survey was not available at the time of report completion and is not included in the graph, Figure 67 d.
Banks Island: Peary caribou on Banks Island were one of the larger populations as they peaked at about 12,000 in the early 1970s (Gunn et al., 2000b) and remained relatively stable until 1982 (Nagy et al., 2009d). Numbers declined to about 1,000 caribou by 1992 (Nagy et al., 2009a) and an initial small recovery by 2001 was likely lost during an icing storm early winter 2003 (Nagy and Gunn, 2006). A 2010 survey led to an estimate of 1,097 ± 343 [95% confidence interval (CI)] non-calf caribou, which confirmed the persistence of low numbers on Banks Island (Davison et al., 2013). See Figure 67 e.
Northwest Victoria Island: Trends in Peary caribou on northwest Victoria Island are less clear than on Banks Island as surveys have been less frequent. Numbers were high, about 2,600 in 1987 (Gunn et al., 2000b), and declined during the late 1980s until, in 1993, only a few caribou were seen during an aerial survey (Gunn, 2005). The population then slowly recovered, based on estimates of 95 ± 60 (95% CI) in 1998 (Nagy et al., 2009b) and 204 ± 103 in 2001 (Nagy et al., 2009c). However, in 2005, the estimate was 66 ± 61 non-calf caribou, which suggested that some recovery was lost during two winters (2002/03 and 2003/04) with icing events (Nagy and Gunn, 2006). A subsequent survey in 2010 returned an estimate of 150 ± 104 non-calf caribou, which confirmed the persistence of low numbers (Davison et al., In Prep.). See Figure 67 f.
Boothia Peninsula: Peary caribou increased throughout the 1970s and early 1980s. A survey in 2006 in this region showed a decline from a 1985 estimate (Gunn and Dragon, 1998). Trends are difficult to distinguish as satellite telemetry has shown that both barren-ground caribou and Peary caribou calve and summer on the peninsula (Gunn et al., 2000a) and only three aerial surveys were flown between 1985 and 2006. During the most recent aerial survey in June 2006, only one caribou was identified as a Peary caribou and sightings of all caribou were few on the northern part of the peninsula which was used by Peary caribou for calving and summer in the mid-1980s and 1990s (Jenkins et al., 2011).
Eastern Queen Elizabeth Islands (Ellef Ringnes, Amund Ringnes, Devon, Ellesmere, Axel Heiberg Islands, Cornwall, King Christian, Graham): There is relatively little information available to assess trends as there have only been two extensive aerial surveys over a 50-year period. The islands were surveyed in 1961, although coverage was so low that the resulting figure of about 1,500 caribou was an approximation (Tener, 1963). Miller et al. (2005) re-analysed the 1961 data using updated areas for the islands and increased the estimate to 2,887± 642. The next extensive survey was between 2005 and 2008, when the Nunavut Department of Environment estimated 4,000 caribou 10 months and older based on aerial surveys (Jenkins et al., 2011):
- Ellesmere Island (including Graham Island), surveyed partly in 2005 and partly in 2006: estimate of 1,021 caribou;
- Axel Heiberg Islands, surveyed 2007: 2,291 (95% CI of 1,636 to 3,208) caribou
- Amund Ringnes, Ellef Ringnes, King Christian, Cornwall, and Meighen Islands, surveyed 2007: total of 282 (95% CI of 157 to 505) caribou
- Lougheed Island, surveyed 2007: 372 (95% CI of 205 to 672) caribou
- Devon Island, surveyed 2008: 17 caribou counted in an extensive survey
The information does not support any overall trend. The reworked 1961 estimate and the 2005–2008 estimate overlap in their confidence limits, but, with only two data points separated by almost 50 years, there is not sufficient information to interpret the trend as stable as the extent and nature of local declines and recoveries are not known.
Source: Gunn et al. (2011c); Species at Risk Committee (2012a); Jenkins et al. (2011). See also references in text.
Long description for Figure 67
This series of bar graphs show the following information:
|Year||Melville Island||Prince Patrick Island||Prince of Wales, Russell, and Somerset Islands||Bathurst Island Complex||Banks Island||Northwest Victoria Island|
|1961||12,799||2,254||-||3,509Note a of Figure 67||-||-|
|1972||2,551||-||-||-||12,098Note a of Figure 67||-|
|1974||1,679||621||4,540Note a of Figure 67||266||-||-|
|1975||-||-||3,741Note a of Figure 67||361||-||-|
|1980||-||-||5,097Note a of Figure 67||-||-||4512|
|1993||-||-||-||2,787||-||*Note b of Figure 67|
|1994||-||-||-||3,155||709||*Note b of Figure 67|
|1995||-||-||*Note b of Figure 67||-||-||-|
|1996||-||-||-||452Note a of Figure 67||-||-|
|1997||787Note a of Figure 67||84Note a of Figure 67||-||74Note a of Figure 67||-||-|
|2001||-||-||-||187Note a of Figure 67||1,142||204|
Trends in distribution
Peary caribou only occur in Canada (except occasional sightings on the northwest Greenland coast) and are restricted to the High Arctic (Queen Elizabeth Islands (404 730 km2)) and the mid-Arctic islands (129 510 km2), as well as the northern extension of the mainland (Boothia Peninsula (26 000 km2)) (Miller, 1991b).
Much of the eastern Queen Elizabeth Islands is mountainous with ice fields and glaciers. Areas above 750 m above ground level, including permanent snow and ice fields, account for 20–43% of Ellesmere, Axel Heiberg, and Devon islands. Caribou densities are much lower and so, although the eastern Queen Elizabeth Islands are 78% of High Arctic landmass, in 1961 they held only 10% of the Peary caribou (Miller et al., 2005). With the higher rate of decline on the western islands, the proportionate distribution has changed and, based on the most recent surveys, about 40% of High Arctic Peary caribou are on the eastern Queen Elizabeth Islands.
Knowledge of trends in distribution is largely based on sightings during systematic aerial surveys, which vary in timing from island to island. On a few islands (for example, northwest Victoria Island and Banks Island), tracking satellite-collared caribou has increased understanding of seasonal and annual distribution (Gunn and Fournier, 2000b; Poole et al., 2010; Gunn et al., 2012). Caribou shift their distribution during unfavourable winters, as, for example, has been recorded on Bathurst Island (F.L. Miller, unpublished data in Gunn et al., 2012) (Figure 69).
As abundance declines, distribution becomes more restricted with, for example, changes in the use of the smaller islands. In the western Queen Elizabeth Islands in 1997, Peary caribou were not seen on three islands (Brock, Eglinton, and Emerald islands) during aerial surveys (Gunn and Dragon, 2002). However, Peary caribou were consistently seen on those islands during 1961, 1972–1974 and 1987–1988 aerial surveys, and were seen there again in 2012 (Davison and Williams, 2012). Any trends in distribution within the larger islands are difficult to determine due to the variability in timing of surveys (Species at Risk Committee, 2012a).
Gaps in information due to the sporadic surveys, as well as the scale of annual variation, impede describing trends in distribution. This is especially true for the eastern Queen Elizabeth Islands, as the frequency of surveys is so low (Jenkins et al., 2011). Overall, the distribution of Peary caribou was reduced by 15% between 1980 and 2006, taking into account the decline of two populations to possibly only a few individuals (on Prince of Wales and Somerset islands and on the Boothia Peninsula).
Dolphin and Union Caribou Population
This section is based partly on Northern caribou population trends in Canada (Gunn et al., 2011c) and includes updates to trend data presented in that report.
Status and trends
Historical information and Inuit hunter reports indicate that there may have been as many as 100,000 caribou on Victoria Island in the early 1800s (Manning, 1960). By the early 1920s, numbers declined and migration across Dolphin and Union Strait halted. The causes are possibly a combination of icing storms and the introduction of rifles. The recovery was slow and caribou were rare until the 1970s. By the 1990s, numbers were increasing (Figure 70 ). In October 1997 and 2007, surveys of caribou staging along the south coast of Victoria Island led to estimates of 27,948 ± 3,367 [standard error (SE)] and 21,753 ± 2,343 (SE) caribou, respectively (uncorrected to account for caribou assumed to be outside of the census zone) (Nishi and Gunn, 2004; Dumond, 2011, pers. comm.). Those two estimates, together with variable annual pregnancy rates, relatively low cow survival (over the period 1999-2006), and high harvest rates, suggest an increased likelihood of a decline (Poole et al., 2010; Dumond, 2011, pers. comm.).
Source: Gunn et al. (2011c). See also references in text.
Long description for Figure 70
This bar graph shows the following information:
Calving is dispersed over about half of northern and central Victoria Island and, based on sightings and satellite telemetry, the summer, fall, and winter ranges have increased in size since the early 1980s into the late 1990s. A trend to an increasing size of winter range is due to fall migrations across the newly formed sea ice to the mainland coastal areas, which is a resumption of migrations observed up until the 1920s. The caribou return across the sea ice to Victoria Island in April to May. The date of freeze-up is increasingly delayed: 8 to 10 days later than in 1982, a trend which may lead to changes in the fall migration across the sea ice (Poole et al., 2010).
Approximately 55 to 65% of the world’s 20,000 to 25,000 polar bears reside in Canada (COSEWIC, 2008). Figure 71 shows the global distribution and the status and trends of all polar bear subpopulations.
Analysis of movement data from mark-recapture studies and tracking of adult female bears with satellite radio collars indicated that there are two subpopulations of polar bears in the Beaufort Sea: one that inhabits the west coast of Banks Island and Amundsen Gulf and a second that is resident along the mainland coast from about Baillie Islands in Canada to approximately Icy Cape in Alaska (Stirling, 2002). The central and Eastern Arctic has six population units: Viscount Melville Sound, Lancaster Sound, Norwegian Bay, Kane Basin, Baffin Bay and Davis Strait (Taylor et al., 2001). The Hudson Bay complex has three subpopulations: Western Hudson Bay, Foxe Basin, and Southern Hudson Bay. Although these subpopulations are sufficiently distinct for management purposes, there is enough movement among them for adequate gene flow (Obbard et al., 2010). Abundance estimate data for Canadian polar bear subpopulations are presented in Table 8.
|Subpopulation||Abundance estimate||95% confidence|
|Gulf of Boothia||1,592||870–2314||2000|
|Northern Beaufort Sea||980||670–1290||2006|
|Southern Beaufort Sea||1,526||1210–1842||2006|
|Southern Hudson Bay||969||688–1365||2011/12|
|Western Hudson Bay||1,000||715–1398||2011|
Source: Polar Bear Technical Committee (2013)
In addition to the science-based information, traditional and community knowledge forms part of decision-making about polar bear risk assessment and management. Aboriginal traditional knowledge (ATK) is considered alongside western science by wildlife management boards and government management agencies, especially in making decisions about harvest. Data brought to the management table from these different knowledge sources, reflecting different timeframes and scales as well as differences in experience, methodologies, and perspectives, are not always in agreement, though they are often complementary (e.g., see discussion in Peacock et al., 2011). As well, increasingly, shifts in distribution, timing, and habitat use make trends in overall abundance difficult to detect. Presentation of alternative information and perspectives on polar bear status and trends is beyond the scope of this primarily science-based report, but is covered elsewhere. For example, the recent Northwest Territories polar bear status assessment presents and synthesizes information from traditional and community knowledge and science studies (Species at Risk Committee, 2012b). Under the federal Species at Risk Act listing process, Environment Canada is obligated to consider ATK during assessments. Initiatives are underway to improve standardization of collection of ATK and to better integrate ATK and western science in assessments and management decisions.
Climate change and polar bears
Polar bears, adapted to hunting seals from the ice, cannot persist without seasonal sea ice (COSEWIC, 2008) and rapidly declining sea ice poses the most serious threat to polar bears (see section on Sea ice on page 36) (Peacock et al., 2011; Vongraven and Richardson, 2011; Reid et al., 2013). Polar bear subpopulations can be grouped according to ecoregions based on ice habitat (Vongraven et al., 2012). These ecoregions reflect differing types and levels of threat to polar bears from changes in ice timing and extent (Table 9). The divergent sea ice ecoregion has extensive annual ice that forms and then moves to the deep-water regions of the Arctic Ocean, while the convergent sea ice ecoregion has heavy multiyear ice. The convergent ice part of the range of the Norwegian Bay polar bear subpopulation is likely to remain the most viable for polar bear populations as seasonal ice shrinks. The Archipelago ecoregion has historically had a mixture of multiyear and seasonal ice filling the gaps between islands year-round, and polar bears in this ecoregion remain on ice all year. By contrast, in the seasonal ice ecoregion, bears are forced ashore and are deprived of food for periods when the ice melts in the summer.
|Ecoregion||Subpopulation||Current or imminent risk from climate change||Known levels of toxic contaminants|
|Divergent||Southern Beaufort Sea||High||Low|
|Convergent||Northern Beaufort Sea||Medium||Medium|
|Archipelago||Gulf of Boothia||Low||Low|
|Archipelago||Viscount Melville Sound||Low||?|
|Seasonal||Southern Hudson Bay||High||Low|
|Seasonal||Western Hudson Bay||High||Low|
Source: Vongraven et al. (2012)
Earlier sea-ice break-up around western Hudson Bay has led to poorer physical condition and poorer reproductive performance of polar bears (see next section on Declines in polar bear body condition). Sea-ice break-up has been linked with lower birth rates, lower survival in subadults and senescent bears, and lower body condition (Regehr et al., 2007). Satellite tracking has shown that polar bear movements have changed and productivity has d eclined in response to sea ice changes in both the Western and the Eastern Arctic (Stirling, 2002; Stirling et al., 2004), as has that of seals, their principal prey (Ferguson et al., 2005). Polar bear denning off Alaska has shifted landward and eastward in response to changing ice conditions (Fischbach et al., 2007). These and other observations led Derocher et al. (2004) to suggest that polar bears will likely not survive as a species, should the predicted scenarios for total disappearance of summer sea ice in the Arctic come true. A small area of the High Arctic may remain as a refuge for polar bears as their habitat shrinks (Durner et al., 2009; Vongraven et al., 2012; Eamer et al., 2013).
Declines in polar bear body condition
This section is extracted from the CAFF report Life linked to ice (Eamer et al., 2013), with an update for the Western Hudson Bay subpopulation as noted.
In areas where sea ice melts completely in the summer polar bears may be forced onto land. Earlier sea ice break-up in these areas reduces the amount of time bears have for hunting seals on the ice. In some areas where this is occurring, bears are becoming thinner, resulting in decreases in survival and reproduction.
Changes in sea ice over the past two decades have led to significant declines in physical condition of bears in the Western Hudson Bay (Stirling et al., 1999; Stirling and Parkinson, 2006), Southern Hudson Bay (Obbard et al., 2006), and Baffin Bay subpopulations (Rode et al., 2012). Regehr et al. (2007) showed that survival decreased in association with earlier sea ice break-up and that this contributed to a 22% decline in the size of the Western Hudson Bay polar bear subpopulation between 1987 and 2004. A recent aerial survey indicates that, while the population continues to decline, the rate of decline is less than anticipated based on the models used by Regehr et al. (2007). Mark-recapture studies are underway to obtain a new population estimate for the Western Hudson Bay subpopulation (R. Vallender, Environment Canada, pers. comm., 2013). Reduced survivorship in relation to sea ice conditions has also been demonstrated in the Southern Beaufort Sea polar bear population (Regehr et al., 2010).
For the western Hudson Bay population, the body condition of bears measured during the ice-free period declined from 1980 to 2007, as did the average weight of female polar bears in the fall (Figure 72). The female bears weighed were suspected to be pregnant.
For the Baffin Bay population, the decline in body condition since the early 1990s is associated with deteriorating ice conditions (Rode et al., 2012). Polar bears were in significantly worse condition in years with less summer sea ice cover, starting in the 1990s when ice in these regions began its sharp decline (Table 10).
|1978-1995 spring captures||no trend||decline||improvement|
|1992-2010 spring captures||decline||decline||insufficient data|
|1991-2006 fall captures||decline||decline||decline|
|1978-1995 spring captures||no trend||no trend||no trend|
|1992-2010 spring captures||body condition better in years with more sea ice||body condition better in years with more sea ice||insufficient data|
|1991-2006 fall captures||body condition better in years with more sea ice||body condition better in years with more sea ice||body condition better in years with more sea ice|
Other threats: harvest and contaminants
Polar bears are also very sensitive to overharvest (COSEWIC, 2008) and currently regulated harvest is the main factor limiting the size of some subpopulations (Species at Risk Public Registry, 2012). Other threats include disturbance and loss of habitat due to increased activity and opening of new areas for shipping, oil and gas development, and mining. Polar bears, as long-lived animals at the top of marine food chains, are vulnerable to the accumulation of contaminants, including mercury and persistent organic pollutants (Table 9). Mercury in polar bears can exceed threshold values for toxicological effects and there is evidence that levels are increasing (Dietz et al., 2013). In contrast, levels of some, but not all, organic contaminants in polar bear tissues have declined in recent years (Mckinney et al., 2011). Levels of contaminants, globally, in polar bears, are highest in the East Greenland subpopulation (Vongraven et al., 2012).
Grizzly bears have been extirpated across much of their former North American range, losing habitat in the southern areas and expanding in parts of the Arctic (Figure 73). Grizzly bears are a useful indicator species for revealing landscape changes in ecosystems. Body size, vital rates, and food habits correlate with habitat and seasonality of food supplies, which led Ferguson and McLoughlin (2000) to distinguish three groups of grizzlies: Pacific-coastal, interior, and barrenground. The barrenground bears typically require large home ranges: those of adult males are 1,154 to 8,171 km2 (Nagy et al., 1983; Clarkson and Liepins, 1994; McLoughlin et al., 1999), four times larger than the largest home ranges of 77 to 1,918 km2 for grizzly bears in other areas of Canada (Ross, 2002). In a study of 41 collared grizzly bears on the coastal plain and mountainous terrain of the Yukon North Slope, ranges of male grizzlies averaged 1,020 km2 with the most extensive range being over 3,000 km2 (Wildlife Management Advisory Council (North Slope), 2008b).
The most recent estimates are 3,500 to 4,000 grizzly bears in the Northwest Territories and between 1,500 and 2,000 in Nunavut, which is almost a quarter of the 26,000 grizzly bears estimated for all of Canada (COSEWIC, 2012). Trends in abundance are difficult to gauge in the Arctic, partly because there is little historic information and survey methods and jurisdictional boundaries have changed, and only few areas have been repeatably sampled over time. Estimating population sizes of grizzly bears is inherently difficult and costly, due to the low densities and difficulty in spotting bears. Information on status and trends will become more available through regional monitoring studies using mark-recapture techniques, including those based on DNA identification of hair that has been sampled by snagging on barbed wire. Studies using this technique are underway in the central barrens as a result of a collaborative effort between the mining industry and the Government of the Northwest Territories (for example NWT Department of Environment and Natural Resources, 2011) and on the Yukon North Slope (Wildlife Management Advisory Council (North Slope), 2008b). A pilot project was carried out using hair snagging around Kugluktuk, Nunavut, from 2004 to 2006, followed by a larger study of grizzly bear populations in the West Kitikmeot area (Department of Environment, 2013a). Results from these studies were not available at the time of writing. Barrenground grizzly bear density estimates from earlier work are available for several areas of the Southern Arctic. The densities vary from 3.5 bears/1,000 km2 on the central barrenlands, in 1995–1999 (McLoughlin et al., 2003) to higher densities in the western NWT: the Anderson-Horton River area, for example, had 9.1 bears/1,000 km2 in 1994 (Clarkson and Liepins, 1994).
COSEWIC (2012) describes a trend of barren-ground grizzly bears expanding their range and becoming more common in the low to mid-Arctic tundra regions of northwest Canada. McLoughlin et al. (2003) reported the grizzly bear population in Nunavut was increasing in 1999 at an annual rate of 3%. There are increasingly reports of grizzly bears in areas where where they were seldom seen or not observed until recently on the Arctic islands, including Banks, Victoria, King William and north to Melville Island (COSEWIC, 2012).
The trend toward expanding their northern distribution has had unexpected consequences. In April 2006 a sport hunter shot a hybrid polar-grizzly bear on Banks Island (Roach, 2006). His guide, Roger Kuptana of Sachs Harbor, identified it as a hybrid; DNA tests showed it had a polar bear mother and a grizzly bear father. Although polar-grizzly bear hybrids are known from zoos and are fertile (Stirling, 1999), this was the first confirmation of a wild hybrid.
Habitat trends include a large and widespread decline in caribou, which make up a large part of the diet of grizzly bears in the central barrens (Mowat and Heard, 2006). In a study in Nunavut and the NWT (Gau et al., 2002), caribou (both calves and adults) was the most prevalent item in the diet, although in early summer, when caribou were elsewhere, horsetails, sedges, and cotton grass dominated the diet, with berries also being prevalent in late summer. Some bears accompany the pre-calving migration of the caribou (Gau et al., 2002) to take advantage of the high densities of caribou on calving grounds. While caribou abundance has changed, most herds return to the same calving grounds each year, providing a predictable resource for the bears. Although the role of muskoxen in the diet of barren-ground grizzly bears has not been studied, grizzly bear predation is known to occur on muskoxen, especially on calves (Case and Stevenson, 1991; Gunn and Fournier, 2000a).
Barren-ground grizzly bears are vulnerable to the effects of industrial activities through possible displacement, but also through nuisance bear kills, and, possibly, through disturbance during denning (Harding and Nagy, 1980; Nagy et al., 1983; Clarkson and Liepins, 1994; McLoughlin et al., 1999; Edwards, 2006). Barren-ground grizzly bears have low productivity, probably as a consequence of the relatively short foraging season. In the Low Arctic tundra, average duration of den occupancy is 185 days (6.2 months) for males and 199 days (6.6 months) for females (McLoughlin et al., 2002). In the Western Arctic there may be an emerging trend toward bears entering into hibernation later in the year and emerging from hibernation earlier, based on observations of Inuvialuit hunters. This change is perhaps due to the longer growing seasons experienced in the Arctic in recent years (Wildlife Management Advisory Council (North Slope) and Aklavik Hunters and Trappers Committee, 2008; COSEWIC, 2012).
Wolverine abundance on the Arctic tundra is an indicator of both the availability of caribou for human harvest and the level of human activities. Wolverines are the largest and mostly solitary carnivore active in the winter on the tundra, as bears hibernate in winter and wolves mainly hunt in packs. Reproductive rates are relatively low and strongly influenced by food availability in winter. Typically, wolverine breed at two to three years of age and have litters averaging three kits (COSEWIC, 2003).
Wolverines scavenge carcasses from wolf and grizzly bear kills. In winter on the tundra, wolverines mainly eat caribou, most likely scavenged, as well as ptarmigan (Lagopus spp.) and small mammals. Wolverines also rely on food caches. In summer, their diet includes a wide range of small mammals, birds including geese and goose eggs, as well as berries (Mulders, 2000). Wolverines’ dependence on caribou when in the tundra zone during the winter may be an important factor in habitat selection, as well explaining their large home ranges.
When caribou numbers decline, their fall and winter ranges contract and they spend more time on tundra rather than in the boreal forests. When caribou abundance is reduced, therefore, tundra-living wolverines likely still have access, at least initially, to caribou. However, if caribou abundance continues to decline, especially for herds that winter on the tundra (for example, the Dolphin and Union herd), wolverine abundance may also change. On the Bathurst caribou herd’s summer-fall range, where caribou abundance declined 93% between 1986 and 2009 (Boulanger et al., 2011a), wolverine densities have also declined. For example, at Daring Lake, wolverine declined from an average density of 8 wolverine per 1,000 km2 to 4 wolverine per 1,000 km2 over a period of 7 years (2005–2011) (Boulanger et al., 2011b; Boulanger and Mulders, 2013).
Wolverine, with low reproductive potential, large ranges, and relatively low winter survival, are susceptible to reduction in numbers from trapping. Sustainable harvest relies on there being sufficient productive females surviving to recolonize areas following removal of animals through trapping (Slough, 2013). Monitoring, both of the harvest and of the distribution and density of wolverines, is key to ensuring that refugia are sufficient to sustain harvested populations, especially when other impacts, such as habitat loss due to resource extraction development, or decline of caribou populations, are present.
Wolverine population studies and programs to collect information from trappers are ongoing in the NWT (Slough, 2013), the Yukon North Slope (Wildllife Management Advisory Council (North Slope), 2012), and Nunavut (Department of Environment, 2013a). For example, in Nunavut, a harvest monitoring program, initiated in Kitikmeot in the 1980s, was expanded to other regions in 2009. The study covers geographic distribution, age and sex of the harvested animals as well as feeding habits. A population monitoring study that involves local hunters in the collection of snagged hair samples to identify individual wolverine using DNA is planned for the Baker Lake area. This will establish baseline information: “natural” wolverine density in areas with limited or no harvest pressure (Department of Environment, 2013a).
Trends in abundance
Wolverine densities are at moderate levels on the western tundra in the NWT and Nunavut, and low on the Arctic islands and in eastern Nunavut: at the time of the 2003 COSEWIC assessment, the population estimates were 3,500 to 4,000 wolverine for the Northwest Territories and 2,000 to 2,500 for Nunavut (COSEWIC, 2003). The trends were considered stable for the NWT and Nunavut, but sensitive to changes in harvesting, as wolverine fur is highly regarded (COSEWIC, 2003). The estimate for numbers of adult wolverine resident in the Northwest Territories has since been revised to 5,100 (Slough, 2013).
Wolverine in northern Quebec and Labrador have been reduced to unconfirmed sightings since 1978 and the 1950s, respectively (COSEWIC, 2003). The declines are attributed to trapping and hunting and the extreme reduction in the caribou herds early in the 20th century, as well as human encroachment, reduced wolf numbers and poison baiting (COSEWIC, 2003). Despite the increase in caribou populations up until the 1990s and restrictions on hunting and trapping of wolverine, populations of wolverine in these locations have not recovered (COSEWIC, 2003).
Further information on wolverine, based both on science studies and ATK, will soon be available. A new COSEWIC assessment of wolverine is due to be released early in 2014. The NWT’s Species at Risk Committee is conducting an assessment of the western population, also to be released in 2014.
Trends in species of special interest
Species and groups of species of special interest for ecological, cultural, and economic reasons are discussed in this section. See the ESTR report Ecosystem status and trends report: Arctic marine ecozones (Niemi et al., 2010) for discussion of anadromous fish species of special interest (dolly varden and Arctic charr).
Migratory tundra caribou
This section is mainly excerpted and summarized from the ESTR national thematic report Northern caribou population trends in Canada (Gunn et al., 2011c) and includes updates as noted. Caribou groupings are based on that report while acknowledging that reviews of historical data and new information provide varying interpretations of geographical units relevent to caribou. Note that COSEWIC has published a report defining “designatable units” for caribou that is now in effect for the COSEWIC assessment process (COSEWIC, 2011).
Caribou, with their central role in tundra and taiga ecology and their inter-connection with the culture of many Aboriginal people, have parallels with the role of salmon on Canada’s Pacific West Coast. Ecosystem aspects of caribou are discussed in previous sections and the importance of caribou to Aboriginal people is discussed in the section on Ecosystem goods and services (page 167).
Northern caribou include migratory caribou of three subspecies plus Peary caribou (Banfield, 1961; Rothfels and Russell, 2005). The three subspecies included in migratory tundra caribou are: 1) barren-ground caribou (Rangifer tarandus groenlandicus), ranging east of the Mackenzie River; 2) Grant’s caribou (R. t. granti), ranging west of the Mackenzie River,; and 3) the two large herds in Ungava of woodland caribou plus two smaller herds in the Hudson Plains Ecozone+ (R. t. caribou).Peary caribou (R. t. pearyi) range on the High and southern (mid-Arctic) Arctic islands and the Boothia Peninsula (Figure 74 and Figure 75).
Trends in abundance
Aboriginal elders recall periods of abundance and scarcity. Other indicators of past caribou abundance and distribution include traditional place names (Legat et al., 2002). Highs and lows in historic abundance since the 1800s have been reconstructed from the frequency of hoof scars on spruce roots, at least for the Bathurst and George River herds (Payette et al., 2004a; Zalatan et al., 2006). Current ranges and recent trends are presented in Figure 76 , based on information summarized in Gunn et al. (2011c). Note that the Beverly and Ahiak herds are shown here as they have been historically defined, consistent with Gunn et al. (2011c). Other authors using re-evaluation of satellite-collared caribou on the north-eastern mainland have reached different interpretations leading to uncertainty about the trends, distribution, and structure of subpopulations in that area (Nagy et al., 2011; Gunn et al., 2013b). Detailed discussion of this topic is beyond the scope of this report.
On the mainland, numbers were low from the 1950s to the 1970s, when the major herds began to increase (Kelsall, 1968; Gunn et al., 2011c). The increases continued into the 1980s. All eight major mainland caribou herds from the Western Arctic east to Hudson Bay declined following their peak abundance in the mid-1980s to mid-1990s (the exact timing depends on the herd). Two herds considered to be still in decline are the Leaf River and George River herds.
By 2012, the Cape Bathurst and Bluenose-West herds had stabilized at extremely low numbers following a period of sharp declines. A 2012 census for the Bathurst Herd revealed a slight decrease in the number of breeding females and a slight increase in the number of younger caribou--essentially the herd is stable at a very reduced abundance, despite the harvest being reduced from over 3,000 to 300 (B. Croft, pers. comm., 2012). At such low numbers it is difficult to detect whether “stability” is a slow decline, a slow recovery, or no trend. After a calving ground photographic census in 2008, which was the first census since 1994, the trend for the Qamanirjuaq Herd was determined to be a statistically insignificant decline.
The George River Herd declined following the mid-1980s, based on the census results for 2010. The neighboring Leaf River Herd, which increased from the mid-1980s at least until the most recent census (2001), is now considered to be declining based on information on demographic rates.
The status of the Ahiak and several herds on the northeast mainland (Wager Bay, Lorillard, Melville Peninsula, and other smaller herds on Boothia Peninsula and Simpson Peninsula), Baffin Island, and the smaller islands in Hudson Bay are currently unknown. The exception is the Southampton Island Herd whose abundance is tracked during aerial surveys at relatively regular intervals. By 2007, the herd had declined to half the peak size estimated in 1997 (30,000 caribou) (Gunn et al., 2011c) and the decline continued to about 7,900 caribou by 2011 (Greer, 2013). The 2013 survey resulted in an estimate of 7,000 caribou, with a higher proportion of calves considered to indicate that the population is stabilizing (based on interview with M. Campbell reported in Greer, 2013). On Baffin Island, Inuit observations and science-based studies indicate that caribou numbers are at a low in the cycle of abundance (Baffinland Iron Mines Corporation, 2012; Department of Environment, 2013b).
The trends in abundance are based on one indicator--the number of caribou in the herd, estimated either through calving-ground or-post calving counts (Gunn and Russell, 2008). In a few herds, such as the Bathurst and George River herds, the trends in total numbers are supported by measured trends in demographic indicators such as adult or calf survival. In other herds, especially the Beverly Herd, monitoring of herd size was infrequent and supporting data on demographic rates were not collected.
The rates of increase and decrease of individual herds vary greatly as can be seen when the rates of change for herds are plotted for periods when they were increasing (after 1970) and periods when they were decreasing (generally after the 1990s) (Figure 77). The herds with the greatest rates of increase were the Southampton and Bathurst, while the Bluenose-West and Porcupine herds showed the lowest rates of increase among herds for which there are sufficient data. During the decline phase, the Cape Bathurst Herd had the greatest rate of decline, although, with only a few breeding females on the Beverly traditional calving grounds in recent surveys, the rate of decline of the Beverly Herd may have been greater. Data are insufficient for the Beverly Herd to calculate this rate.
The chart shows the annual rate of change during increase and decline phases, based on conversion of the population estimates to natural logarithms. The years used vary among herds depending on when herds were increasing and decreasing and when population estimates were made. For the Porcupine Herd, where a change in trend direction was detected in the 2010 survey, the rate of increase is the average of 0.033 (1972–1989) and 0.035 (2001–2010).
Source: Gunn et al. (2011c)
Long description for Figure 77
This graphic represents a bar graph showing the following information:
|Herd||Exponential rate of decline||Exponential rate of incres|
The current declining trends for some caribou herds, as well as the recent declining trends with current indications of stabilization or recovery for other herds, are likely a reflection of natural cycles in caribou abundance accentuated by the cumulative effects of increasing human presence on the caribou ranges (Gunn et al., 2013c). More conjectural is to what degree climate warming and attendant broad-scale habitat changes are factors in the natural cycles.
The causes of declines are complex, with the roles of the various contributing factors changing as the declines continue. Caribou are similar to other northern herbivorous mammals (voles, lemmings, and hares) in that their abundance is cyclic (Morneau and Payette, 2000; Gunn, 2003b; Zalatan et al., 2006) and, overall, the cycles are likely driven by climate interacting with forage availability, predation, and pathogens. Weather tends to have a decadal pattern, influenced by major patterns, such as the Arctic Oscillation, switching between negative and positive phases (Bonsal and Shabbar, 2011). Winter temperatures and snowfall patterns interact with forage growth and availability. Winter conditions and forage availability influence caribou condition, which determines birth rates and calf survival (Couturier et al., 2009a; Couturier et al., 2009b). Trends in annual calf survival and fecundity also play a role in changing herd abundance.
Weather also interacts with parasites such as warble flies, whose activity depends on summer weather. Weather affects the transmission of internal parasites, which in turn influences forage intake as caribou try to reduce their exposure to the parasites (Van der Wal et al., 2000). Predation and harvest by humans have a pivotal role in declines as even small annual reductions in adult female survival strongly influence population trends (Gaillard et al., 1998). Without mandatory harvest reporting, however, it is not possible to assess the impact of harvesting on the caribou populations.
Combining population estimate data on migratory herd numbers (including the Dolphin and Union population) since 1970 and scaling herd size relative to maximum estimates for each herd indicates that, on average, caribou numbers have increased from lows around 1975 to a peak around 1995, followed by a decline with some indication of a recent levelling off or reversal of the decline (Figure 78). The timing and magnitude of the changes vary.
The line represents the six-year running average. Other symbols represent individual herds. Relative population size is calculated as the population estimate for the year as a proportion of the maximum recorded estimate. Note that the maximum recorded estimate is not necessarily the peak population over this timeframe, as surveys usually did not cover the entire period and were not conducted every year.
Source: Gunn et al. (2011c)
Long description for Figure 78
This scatterplot and line graph shows the following information:
|Year||Porcu-pine||Cape Bathurst||Bluenose-West||Bluenose-East||Bathurst||Dolphin and Union||Beverly||Qamanir-juaq||South-ampton Island||George River||Leaf River||6 year running average|
The cumulative effects of increasing human presence on caribou ranges (number of people as well as non-renewable resource exploration and extraction and infrastructure development) are largely unknown. However, tools are being developed to examine how responses of the individual caribou can be scaled up to measure population-level effects (Gunn et al., 2011b). Some recently constructed mine projects monitored effects on caribou and reported changes in caribou distribution and time spent foraging (Gartner Lee Limited, 2002). In response to large open-pit mines on the tundra summer range of the Bathurst Herd, caribou distribution was reduced in a 10 to 15 km zone of influence around the mines (Boulanger et al., 2004). Changes in the atmospheric transport of contaminants on individual caribou body burdens are monitored for some herds (Gamberg, 2009) and the results evaluated in relation to potential impacts on human health. These evaluations conclude that nutritional benefits of consuming caribou far outweigh any risks from the low levels of contaminants (Van Oostdam et al., 2005; Donaldson et al., 2010).
See Gunn et al. (2011c) for herd-specific assessments. These include summaries of status and trends and notes on each herd’s ecology.
Trends in distribution
Herd distributions will change through time; shifts in distribution, however, are not well documented and are uncertain. Information from aerial surveys and satellite-collared individuals generally has not been analyzed to describe trends in distribution. Migratory tundra caribou characteristically shift their winter distribution among years and winter ranges often overlap between neighbouring herds (Schmelzer and Otto, 2003; Bergerud et al., 2008). Additionally, as herd abundance rises and falls, distribution--especially winter distribution--can shift (Bergerud et al., 2008). Maps of historical distribution (Banfield, 1961) and winter distribution since the 1970s, at least for the Beverly, Qamanirjuaq, and Bathurst herds (Gunn et al., 2001; BQCMB, 2004), hint at a contraction in the southern boundary of the winter distribution in northern Manitoba, Saskatchewan, and Alberta. During the 1996 to 2010 decline of the Bathurst Herd, the winter distribution of the satellite-collared cows showed a trend towards wintering further north of the 60th parallel (Gunn et al., 2011b).
The muskox (Ovibos moschatus) is important as a symbol of the Arctic tundra and its long association with people, extending back to Paleolithic times (Lent, 1988). Its ecological role stems from its function as a large-bodied social herbivore, well-adapted for the pulse of summer productivity and the long winters that characterize tundra ecology. In many tundra areas, muskoxen are the only large-bodied prey available to wolves and humans during winter, as muskoxen are rarely long-distant migrants.
Muskoxen are distributed through northeast Canada and Greenland and have been introduced or re-introduced to Alaska, western Greenland, Scandinavia and Russia (Figure 79 ). Muskoxen reintroduced to Alaska have spread, in small numbers, to the Yukon North Slope in the 1980s (Wildlife Management Advisory Council North Slope, 2008). By the early 1900s, muskox abundance in Canada had collapsed on the Arctic mainland and on some Arctic islands. Since then, numbers have built up, through natural recovery and range extension, aided by 30 to 50 years of almost no harvesting. Combining the most recent estimates for the islands and mainland, Canada has about 114,300 muskoxen, which is about three-quarters of the world’s muskoxen (for references, see below). Most muskoxen occur on the Arctic islands (85%), and 77% of Northwest Territories and Nunavut muskoxen occur on the large islands (Banks and Victoria), even though these two islands form only 40% of the landmass of the Arctic islands’ muskox range.
Circumpolar muskox status and trends
The world distribution of muskoxen, including introduced and re-introduced populations, is shown in Figure 79. Muskoxen have spread widely from their release sites, perhaps in search of new range as populations expanded (Reynolds, 1998; Gunn and Adamczewski, 2003). Alaska muskox numbers have stabilized and, in 2013, numbered over 4,200 in five regions (Harper, P., 2011; Gunn et al., 2013a). In Greenland, status and trends in muskoxen are unknown. The most recent survey, in northeast Greenland about 20 years ago, estimated 9,500 to 12,000 muskoxen (Boertmann et al., 1992). More is known about the introduced population at Kangerlussuaq, where the initial 27 muskoxen introduced in the 1960s increased rapidly. Rough estimates are for a current population of 10,000 to 25,000. Other populations in West Greenland have also shown good growth, and quota-based harvesting has been implemented. Based on the above, the total numbers of muskoxen in Greenland may be from 20,000 to 40,000 (Greenland Institute of Natural Resources, 2012). In Russia, re-introductions started in the mid-1970s in several locations. The total muskox population in Russia was estimated at over 10,000 in 2013, with the great majority (8,700) being on the Taimyr Peninsula (Gunn et al., 2013a).
Since the Pleistocene, muskoxen have varied little genetically and in appearance. Skull and dental characteristics from contemporary and Pleistocene muskoxen are similar (Harington, 1970) and a comparison of ancient and modern DNA also suggests muskoxen have changed little over thousands of years and their variation has not been sufficient to recognize subspeciation (Tener, 1965; Groves, 1995). This is not to say that mainland and Arctic island muskoxen do not differ, but that the differences were not enough to warrant sub-specific designation. Muskoxen have notably low genetic variability, though mainland muskoxen are slightly more genetically variable than island muskoxen, based on microsatellite DNA analysis (Van Coeverden de Groot, 2001).
Information on changes in the number or distribution of distinct muskox populations is not available, as Canadian muskoxen are managed for harvesting based on management units (rather than natural populations). Often, for the Arctic Islands, the management units are islands or groups of islands. On the mainland, the units are based on hunting patterns and on changing information about distribution, and the managment unit boundaries are modified at intervals. Nunavut’s management units were recently revised as part of the development of new muskox management plans (Dumond, 2006; Kivalliq Wildlife Board, 2010; Government of Nunavut, 2012). The NWT has seven muskox management units (Environment and Natural Resources, 2012a). The boundaries of aerial survey areas do not always coincide with management unit boundaries, or they may cover several units, which complicates analysis of historical trends (Fournier and Gunn, 1998; Dumond, 2006).
Trends in abundance
The record of European exploration and settlement explains why muskoxen on the mainland were reduced to isolated remnants by the early 1900s, as traders encouraged commercial hunting for sales of hides in European markets (Barr, 1991a). The muskox hides were marketed to replace the loss of bison robes as the bison numbers were collapsing, also in part due to unregulated commercial harvesting (Barr, 1991a). From 1862 to 1916, the Hudson’s Bay Company purchased more than 15,000 muskox hides (Tener, 1965). The company continued to purchase hides as the number brought in fell from more than 1,000/year in 1891 to only a few score in 1908, reflecting the declining population. In 1916, the last year before the Canadian government provided legislative protection, just one muskox skin was sold (Figure 80).
Source: Tener (1957)
Long description for Figure 80
This bar graph shows the following information:
On the Arctic islands, European explorers hired Inuit to provide meat for their expeditions, and Hone (1934) calculated that, between 1852 and 1916, British and Canadian expeditions to Melville Island killed over 600 muskoxen for meat and to capture calves for zoos. Over 1,000 muskoxen were killed on Ellesmere Island by Norwegian and American explorers between 1880 and 1917 (Barr, 1991b). However, how those harvests affected long-term trends is essentially unknown. On Banks and western Victoria islands, muskoxen had essentially disappeared by the late 1800s. This drastic decline was attributed to ice storms (Gunn, 1990).
Harvesting of muskoxen was suspended from 1924 to 1969, when a small quota was established for southern Ellesmere Island. The Thelon Game Sanctuary (later to become the Thelon Wildlife Sanctuary) was established in 1927, in large measure to protect muskoxen (Taylor, 2006). During the 1970s and 1980s, conservative quotas for harvesting became more widespread as hunters’ reports and surveys revealed increasing muskox numbers (Barr, 1991b). Recovery appeared to be relatively slow over the decades, partly as a consequence of lack of surveys to track the recovery, and partly because the muskoxen were dispersing and recolonizing their previously occupied ranges (Barr, 1991b). Muskoxen have a potential doubling rate of three years, although that is relatively rare. They are similar to other large-bodied herbivores, having high adult survival, low fecundity, late onset of breeding, a single birth every one to three years, and females having a lifespan often exceeding 15 years (Gunn and Adamczewski, 2003). Muskox numbers reached an estimated 134,000 in 2001 (summarized in Dumond, 2006) but, since then, have declined, mostly because of the decline in muskoxen on Banks Island. Although the mid-Arctic islands of Banks and Victoria are only 9% of the landmass of the Arctic tundra, they are home to 72% of Canada’s muskoxen. Only 14% of muskoxen are on the mainland. Estimates of Canadian muskox abundance from 1967 to 2012 are shown in Figure 81 . The breakdown of the most recent estimates by location across the Arctic Ecozone+ is shown in Table 11.
Source: Dumond (2006) and references therein; references in Table 11
Long description for Figure 81
This bar graph shows the following information:
|Region||Area||Muskox numbers||Error estimate|
|Nunavut||Bathurst Island Complex||82||-||2001||Jenkins et al. (2011)|
|Nunavut||Cornwallis||18||-||2002||Jenkins et al. (2011)|
|Nunavut||Devon Island||513||302–864 (95% CI)||2008||Jenkins et al. (2011)|
|Nunavut||Prince of Wales||2,086||1,582–2,746 (95% CI)||2004||Jenkins et al. (2011)|
|Nunavut||Somerset||1,910||962–3,792 (95% CI)||2004||Jenkins et al. (2011)|
|Nunavut||North Ellesmere Island||8,115||6,632–9,930 (95% CI)||2006||Jenkins et al. (2011)|
|Nunavut||South Ellesmere Island||456||312–670 (95% CI)||2005||Jenkins et al. (2011)|
|Nunavut||Axel Heiberg Island||4,237||3,371–5,325 (95% CI)||2007||Jenkins et al. (2011)|
|Nunavut||Amund Ringnes, Ellef Ringnes, King Christen, Cornwall, Meighen, and Lougheed islands||21||-||2007||Jenkins et al. (2011)|
|Nunavut||King William Island||300||-||2007||White (2002)|
|Nunavut||West of Bathurst Inlet (MX 19)Note * of Table 11||2,141||±586 SE||2005||Dumond (2007a)|
|Nunavut||(MX 14, western part)||434||±168 SE||2005||Dumond (2007a)|
|Nunavut||West of Coppermine R.||132||±71 SE||2007||Dumond (2007b)|
|Nunavut||Boothia Peninsula |
High density area
|348||±62 SE||2006||Dumond (2007c)|
|Nunavut||Medium density area||645||±119 SE||2006||Dumond (2007c)|
|Nunavut||Low density area||104||±72 SE||2006||Dumond (2007c)|
|Nunavut||East Queen Maud Gulf||2,621||379 SE||2000||Campbell and Setterington (2001)|
|Nunavut||Northeast Kivalliq||165||-||2000||Campbell and Setterington (2001)|
|Nunavut||North of Baker Lake||4,736||554 SE||2010||Campbell and Setterington (2001)|
|Nunavut||Southeast Victoria Island (NU)||25,000-30,000||-||1992-1999||Dumond (2006)|
|Nunavut||Thelon Sanctury (NU and NT)||1,095||±281 SE||1994||Gunn et al. (2009)|
|Northwest Territories||Northwest Victoria Island (NT)||11,442||1,637 (95% CI)||2010||Davison et al. (2013)|
|Northwest Territories||Banks Island||36,676||4,031 (95% CI)||2010||Davison et al. (2013)|
|Northwest Territories||Melville Island||3,033||852 (95% CI)||2012||Davison and Williams (2012)|
|Northwest Territories||Prince Patrick Island||507||320 (95% CI)||2012||Davison and Williams (2012)|
|Northwest Territories||Eglinton Island||213||211 (95% CI)||2012||Davison and Williams (2012)|
|Northwest Territories||South Paulatuk||1,215||±525 SE||2002||Community of Paulatuk et al. (2008)|
|Northwest Territories||North Great Bear||1,457||±919 SE||1997||Veitch (2013)|
|Northwest Territories||Alymer Lake||161||±39 SE||1991||Shank and Graf (2013)|
|Northwest Territories||Artillery Lake||1,606||±277 SE||1998||Bradley et al. (2001)|
|Northwest Territories||Beaverhill Lake||532||±149 SE||2000||Gunn et al. (Gunn et al., 2009)|
|Yukon||Yukon North Slope||101||-||2011||Wildlife Management Advisory Council (North Slope) (2012)|
|Yukon||Quebec (Nunavik)||1,400||-||2003||Chubbs (2007)|
Surveys are conducted in the spring and are for one-year-plus-old muskoxen. In general, when there is no range or error estimate provided, the number respresents a mimimum count (i.e., the actual number seen in the survey).
Notes of Table 11
- Note * of Table 11
MX indicates Nunavut muskox management units
Trends in regional populations
By the 1930s, remnants of muskox populations (with a total of perhaps about 500 muskoxen) were scattered across the Arctic mainland, with a cluster of herds north of Great Bear Lake, south of Bathurst Inlet, in the Thelon Game Sanctuary, and south of Boothia Peninsula (Anderson, 1934; Barr, 1991b). As hunters reported sightings of muskoxen, aerials surveys were undertaken for several mainland regions. North of Great Bear Lake, muskoxen recolonized, and their numbers increased during the 1990s (Fournier and Gunn, 1998; Dumond, 2006).
While the muskoxen continued to spread west toward the Mackenzie River, their numbers declined behind the “colonizing front}. For example, north of Great Bear Lake and west of Coronation Gulf, muskoxen increased from scattered remnants in the 1900s to 425 in 1967 (see review in Dumond (2007b)). In a small portion of this area, the Rae-Richardson watersheds, there were 869 ± 300 (SE) in 1980, 1295 ± 279 (SE) in 1983, and 1,805 ± 289 (SE) in 1988 (Fournier and Gunn, 1998; Dumond, 2007b). However, between 1989 and 1994, the muskox numbers declined, possibly due to a parasitic lungworm (Umingmakstrongylus pallikuukensis) which may have increased muskox vulnerability to grizzly bear predation (see section on Wildlife diseases and parasites on page 62). Between 1994 and 2007, numbers appeared to stabilize. The 1994 estimate of 540 ± 139 (SE) is similar to a 2007 estimate of 509 adult muskoxen (Dumond, 2007b).
East of the Coppermine River to Bathurst Inlet, a 2005 survey found 2,141 ± 586 (SE) muskoxen in the western part of the study area, indicating an increase in muskox abundance since a 1991 survey. However, the abundance had declined to 434 ± 168 (SE) in the eastern part of the study area, compared to a 1986 survey (Dumond, 2007a).
Few muskoxen were reported in the Queen Maud Gulf and Adelaide Peninsula area until the 1960s. By August 1982, however, numbers were estimated at 8,494 ± 2,673 (SE) (Gunn et al., 1984). The trend between 1982 and 1996 was a 50% decline in numbers, as the 1996 estimate was 4255 ± 680 (SE) (Fournier and Gunn, 1998). It is unclear to what extent the decline is a distributional shift, as hunters in Kivalliq and northeast Kitikmeot areas report increasing numbers and expanding distribution through the 1990s (Campbell and Setterington, 2001). South of the Queen Maud Gulf area, in the Kivalliq Region, muskoxen were surveyed in 1986, 1991, 1999, and 2010. Preliminary analyses of the 2010 findings indicate both population and range expansion. The 2010 estimate was 4736 ± 554 (SE) muskoxen with a colonizing edge moving east toward Hudson Bay coast and a decline in densities in the survey area (Ford et al., 2012).
Further north and east of the Queen Maud Gulf, Gunn and Dragon (1998) estimated 554 muskoxen in 1995 on the Boothia Peninsula where they were scarce or absent through the 1980s (Gunn and Dragon, 1998). This number had approximately doubled to an estimated 1,097 by 2006 (Dumond, 2006).
South and west of the Thelon Game Sanctuary, muskoxen were re-colonizing their former ranges toward the treeline. The estimated number of adult muskoxen in the sanctuary in 1994 was 1,095 ± 281 (SE) (Gunn et al., 2009). Hunters in Lutsel K’e also reported increasing numbers of muskoxen east of Artillery Lake and surveys in 1989 and 1998 revealed that muskox numbers had doubled and that muskoxen had spread into the taiga to the west and southwest (Bradley et al., 2001). A survey documented the increased spread of muskoxen south toward the treeline when, in 2000, 1,320 ± 183 (SE) muskoxen were estimated (Gunn et al., 2009).
Muskoxen disappeared from the Yukon North Slope between 1858 and 1865 and were re-established in the Arctic National Wildlife Refuge in Alaska in 1969–1970 (Reynolds, 1998). From there a few animals moved into the Yukon in 1972 (L. Harding, pers. obs.) but it was not until the mid-1980s that repeated sightings of cow muskoxen were reported; mixed groups with calves were first sighted in 1987 (Johnson et al., 2005). The range continued to expand east to the Mackenzie River and south into the Taiga Cordillera Ecozone+. Because the North Slope muskoxen move between Alaska and the Yukon, population estimates include animals on the Alaska and Yukon North Slope and areas to the south. The North Slope population was estimated in 2006 to be 400 animals, a decline from 700 animals estimated in the mid-1990s, based on extensive aerial surveys. A total of 291 muskoxen were spotted in the survey conducted in 2011--of these, 101 were in the Yukon (Wildlife Management Advisory Council (North Slope), 2012). Reports of sightings by Aboriginal people in the region indicate a decline since the 1990s (Arctic Borderlands Ecological Knowledge Co-op, 2006).
Muskoxen from Ellesmere Island were initially introduced to a farm in 1967 near Kuujjuaq and then released between 1973 and 1983 at three locations in Nunavik (Lehenaff and Crete, 1989a; Chubbs and Brazil, 2007). The muskoxen rapidly increased to 1,400 by 2003 and have been sighted in Labrador (Chubbs and Brazil, 2007).
The most recent muskox numbers available for the Arctic Islands indicate a total of 94,130 (non-calves): these estimates span the period 1992 to 2012. This total includes 13,300 muskoxen on the eastern Queen Elizabeth Islands, based on aerial surveys between 2004 and 2008 (Jenkins et al., 2011). No trend for these islands is discernible as the previous estimate was in 1961 (Tener, 1963).
On the western Queen Elizabeth Islands, the more frequent surveys in the early 1970s, mid-1980s, and 1990s indicate that there have been two sharp declines and one or two periods of recovery. The total number of estimated muskoxen in 2012 on the western Queen Elizabeth Islands was 3,750, an increase of 48% over 15 years (Davison and Williams, 2012).
The trends for the western Queen Elizabeth Islands can be related to a pattern of increases and sharp decreases during winters with above-average snow depths and icing (while acknowledging that the weather data are sparse). For the islands in the Western Arctic with sufficient number of estimates to describe trends, there is very little regional consistency. This can be seen for Prince Patrick, Melville and Bathurst islands, which are stretched over 600 km across a progressively decreasing Beaufort maritime influence (Maxwell, 1981). Muskox numbers on Prince Patrick Island increased from none to 507 ± 320 (95% confident interval) between 1961 and 2012, while, during the same period, muskox numbers on Melville Island peaked in 1986 and declined by the mid-1990s before recovering to 3,033 ± 852 (95% confident interval) (Davison and Williams, 2012). Based on aerial surveys, muskoxen on Bathurst Island have been through two declines and one recovery since 1961 (Figure 82 ). In 1961, muskox numbers were relatively high: Tener (1963) estimated 1,136 muskoxen. When the population was next estimated in 1973, it had declined 59% and, by August 1974, the continued decline had reduced the estimated number of muskoxen to 164 ± 70 (SE) after a severe winter (Miller et al., 1977). Muskox numbers recovered and increased between 1974 and 1993 to an estimated 1,200 (Ferguson, 1987; Miller, 1987; Miller, 1998). A series of three winters, 1994–1997, with rain on snow and unusually deep snow reduced the muskox abundance and, in summer 1996, Miller (1998) estimated that there were 425 ± 136 (SE) muskoxen alive. By 1997, only an estimated 124 ± 45 (SE) one-plus-year-old muskoxen were alive, which indicated a 90% decline since 1994 (Gunn and Dragon, 2002). During an aerial survey in 2001, only 94 muskoxen were observed and no population estimate was derived (Jenkins et al., 2011).
Error bars are ± standard error. See text for more information on trends.
Source: Tener (1963); Miller et al. (1977); Miller (1998); Gunn and Dragon (2002); Jenkins et al. (2011)
Long description for Figure 82
This bar graph shows the following information:
The muskox trends on the larger, more southern mid‐Arctic islands are quite different from those on the High Arctic islands (Queen Elizabeth Islands). A relatively well‐documented trend is that of Banks Island, where the population grew from hundreds in the 1960s, based on Inuvialuit accounts to a peak around 2001, followed by a decline to the most recent survey in 2010 (Table 11, Figure 83) (Gunn et al., 1991b; Inuvialuit Game Council, 2002; Lyberth, 2003; Davison et al., 2013).
Similar to the trends on Banks Island, muskox numbers increased on Victoria Island during the 1970s to 1990s: the Nunavut part of Victoria Island had an estimated 25,000 to 30,000 muskoxen in 1992–1999 (Dumond, 2006; Gunn and Patterson, 2012) and the Northwest Territories portion had an estimated 11,442 in 2010 (Davison and Williams, 2013). In 2004 surveys of the large islands east of Victoria Island, 2,086 muskoxen were estimated on Prince of Wales, and 1,910 muskoxen were estimated on Somerset Island (Jenkins et al., 2011). The muskox populations on those two islands have been increasing since at least the 1970s (Nagy et al., 1996; Fournier and Gunn, 1998; Jenkins et al., 2011; Gunn and Patterson, 2012; Davison and Williams, 2013).
Trends in distribution
The Canadian population is almost entirely within the Arctic Ecozone+ except for in the Thelon Wildlife Sanctuary and, recently, where muskoxen have spread along the treeline from northeast of Great Slave Lake toward the Saskatchewan border. They are distributed on the tundra in areas with either shallow snow (20 to 40 cm) or sufficient relief for the wind to maintain areas with shallow snow. Summer habitats include low-lying river valleys or coastal plains with sedge meadows, riparian willows, and gravel bars. Muskoxen are widely distributed and occur in the Low, Middle, and High Arctic tundra ecoregions (an area of 2,293,372 km2).
There have been no overall analyses yet of the trends in distribution. Since the early 1900s, muskoxen have recolonized most of the mainland tundra except northeastern areas, including the Melville Peninsula. Likewise, muskoxen occur on most Arctic islands except the northeast islands (Baffin and the islands in Hudson Bay). Muskoxen disappeared from Baffin Island during the fifteenth century, with only ocassional records since then, such as a herd of eight muskoxen observed south of Clyde River in 1968 (Barr, 1991b). The recolonization of historical ranges has been natural, except for the Yukon North Slope and the successful introduction to northern Quebec. Muskoxen apparently did not naturally colonize Quebec after the glacial retreat (Lehenaff and Crete, 1989b). The rate of recolonization is slow--perhaps less than 10 km/yr (Fournier and Gunn, 1998).
Muskoxen occasionally appear on and disappear from smaller and medium-sized islands. Tener (1963) did not find muskoxen on Prince Patrick, Eglinton, and Mackenzie King islands in 1961, although they were recorded there during the aerial surveys in 1972–1974, 1986, 1997, and 2012 (Miller et al., 1977; Gunn and Dragon, 2002; Jenkins et al., 2011; Davison and Williams, 2012). Within the islands, relatively little analysis has been undertaken for trends in distribution. Even so, information does suggest some changes. Between 1972 and 1980, exceptionally high densities of muskoxen were found on Bailey Point on southwest Melville Island (Thomas et al., 2013). The densities were from 0.6 to 1.1 muskoxen/km2 over the entire peninsula, reaching 2.6/km2 below 100 m, which was comparable to other “hotspots” of muskox abundance, including the Thomsen River, Banks Island. Bailey Point was considered to have a particularly favourable microclimate to serve as a refugium for muskoxen. The high densities persisted at least to 1987 (Miller, 1988), but in 1997, when numbers had declined by about 50% since 1987, almost no muskoxen were seen at Bailey Point, a situation which persisted in 2005 and 2012 (Gunn and Dragon, 2002; Davison and Williams, 2012).
Trends in harvest
Muskoxen are harvested under a quota system introduced in 1969. Quotas have increased and, by 2011, in Nunavut had reached a Total Allowable Harvest of 2,303 muskoxen (Giroux et al., 2012b). This includes harvest for domestic use, commercial harvest, and sport hunting by non-residents (Wildlife Research Section, 2011). In the Northwest Territories, the total harvest quota is 1,112 in the Inuvialuit Settlement Region, which is mainly two quotas that cover Banks and northwest Victoria islands (NWT Environment and Natural Resources, 2011). These two quotas were assigned to influence the rate of increase in the 1990s and to encourage commercial harvesting. Annual use of the quotas varies, especially in the commercial harvesting for meat and qiviut (wool) sales.
Source: NWT Environment and Natural Resources (2011)
Long description for Figure 84
This bar graph shows the following information:
Threats that cause or contribute to population declines may lie in the unexpected, which may be related to shortcomings in data collection or predictive models, or unanticipated environmental changes accumulating in ways not encompassed by traditional knowledge. The extent of effective management (including habitat and landscape management planning) is the deciding factor for anticipating and dealing with threats. The status of muskox management planning varies between Nunavut and Northwest Territories. Within Nunavut, regional management plans with a focus on harvest management have recently been completed (Dumond, 2006; Kivalliq Wildlife Board, 2010; Government of Nunavut, 2012) which include adaptive management to adjust harvest levels if popuation levels change. The Inuvialuit Settlement Area has community conservation plans and harvest reporting (Community of Paulatuk et al., 2008; NWT Environment and Natural Resources, 2011); governance, however, is more complex elsewhere in the Northwest Territories. The harvest, except on Banks and northwest Victoria islands, is relatively low and an overall approach to muskox management is absent.
Predation is a natural driving force in muskox evolutionary ecology. However, almost nothing is known about how muskoxen trade off the risks of predation (and parasitism) relative to foraging. Wolf predation on muskoxen is likely common, with packs or single wolves observed killing adult as well as younger muskoxen (Gray, 1987; Mech and Adams, 1999). On northern Ellesmere Island, and northern and eastern Greenland, muskox occurrence in wolf scats ranges from 65–98%, with lemmings and Arctic hares as the next most frequent item (Marquard-Petersen, 1998). On Banks Island, where Peary caribou are less numerous, muskox remains were found in 103 out of 115 stomachs and 34 out of 38 scat samples (1992–2001), with lemmings being the next most frequent item (Larter, 2013). Trends in wolf predation on muskoxen are unknown.
Although unmeasured, an increase in grizzly bear sightings on muskox ranges is being reported (Ford et al., 2012). Newborn calves that cannot keep up with a herd are easily killed by pursuing bears (Clarkson and Liepins, 1994). Deep, crusted snow increased the vulnerability of muskoxen in at least one instance of a grizzly bear killing a bull (Case and Stevenson, 1991).
Predation in some areas, such as west of Kugluktuk, may be accentuated by infection with a lungworm (see section on Wildlife diseases and parasites on page 62 and below). Muskox abundance increased and peaked at 1,800 in 1987, but then declined by about 50% in 1994 and remained stable in 2007 (Dumond, 2007b). The muskoxen were infected with the lungworm Umingmakstronylus pallikuukensis, which may increase muskox vulnerability to predation, especially from grizzly bears, as the lungworm forms cysts in the lungs, causing difficulties breathing.
Health and trends in health
It is difficult generally to detect trends in health, as knowledge of muskox parasites has been increasing since the 1990s. However, it likely that the spread of at least one parasitic lungworm may be associated with a warmer climate. In the 1990s the lungworm Umingmakstrongylus pallikuukensis was described (as a previously unknown species) with a range restricted to the western Canadian Arctic mainland, extending west of Kugluktuk (Hoberg et al., 1995). Climate models predicted range expansion of the parasite and it has now been found on southernVictoria Island (Kutz et al., 2009) (see section on Wildlife diseases and parasites on page 62).
Muskoxen may have a low resistance to parasites even when in good body condition (Alendal and Helle, 1983; Korsholm and Olesen, 1993). This characteristic may be related to their low genetic variability. Muskoxen in some areas, notably Banks Island, are subject to outbreaks of the bacterial disease Yersinia pseudotuberculosis,which may have an environmental trigger such as unusually warm weather. Exposure to Yersinia is high (Larter and Nagy, 1999) and periodic outbreaks include deaths. A previously unknown disease in muskoxen that led to deaths on Banks Island in 2012 was erysipelas, a bacterial disease that is common in livestock (M. Branigan. pers. comm., 2012).
Most weather-related deaths are indirect, as they are caused by malnutrition. The general projections for a warmer climate include more incidents of rain on snow, which are associated with muskox deaths, at least on Banks Island (Rennert et al., 2009; Nagy and Gunn, 2009). Periods of rain when calves are young can also increase mortality (P. Hale, pers. comm., 2013).
The trend is for increasing industrial exploration and development, including construction of new roads. Predicting how muskoxen and their habitat will be influenced is uncertain as relatively little is known about cumulative effects on muskoxen. Earlier studies focused on behavioural responses of muskoxen to seismic and aircraft disturbance (e.g., Miller and Gunn, 1980). Studies in Alaska related to the National Petroleum Reserve provide more information on potential disturbance and mitigation measures associated with development (Bureau of Land Management, 2012).
Wolves are the major large carnivore in Arctic tundra systems, with grizzly bears being important in much of the Southern Arctic. Medium-sized predators include Arctic and red foxes and wolverines, and avian predators include snowy owls and jaegers. Grizzly bears are discussed above in the section on Trends in species of conservation concern (page 119), and wolves and wolverines are discussed below. See also the section on Community and population dynamics (page 56) and the Main threats to caribou (page 170) for discussion of predator-prey dynamics.
With changes in land-use patterns and the expansion of human populations in southern Canada, carnivores have lost much of their former North American ranges, making the northern regions of the continent increasingly important for species conservation (Figure 85).
A well-known species and top predator, the continued presence of wolves is an important indicator of Arctic ecosystem integrity. Wolf populations are not monitored regularly. Moreover, since some follow caribou and some do not, interpreting abundance and distribution data would be problematic for some areas. Wolves in highly productive environments can reproduce rapidly, making them resilient to various types of disturbance, including hunting.
Gray wolves, Canis lupus, in the Arctic Ecozone+ were formerly considered to include up to eight subspecies (Mech, 1974). Although subspecies status is currently in doubt, genetic analyses suggest that wolves in the Northern Arctic (sometimes called “Arctic” wolves) are separable from Southern Arctic (“tundra”) wolves (Carmichael et al., 2007), and that these, in turn, are separable from boreal forest (“gray”) wolves (Musiani et al., 2007). Aside from their genetic structures, the three groups have different prey specialization patterns (Carmichael et al., 2001; Musiani et al., 2007).
Wolf populations of the Arctic Archipelago form a metapopulation, some populations of which have endured severe bottlenecks that probably occurred during catastrophic declines in their principal prey, Peary caribou and muskoxen. Arctic hares are also important prey species. From 2000 to 2006, Arctic wolves, Arctic hares, and muskoxen were counted in a study area on Ellesmere Island. Wolf numbers were strongly correlated with hare numbers, but not with muskox numbers (Mech, 2007).
Wolves on Banks, Ellesmere, and Devon islands show genetic signatures of recent population declines (Carmichael et al., 2008). Island wolf populations in general have significantly less genetic variability than mainland populations, suggesting that there is movement among the islands and that populations recover after a decline, mainly through recolonization from other islands (Carmichael et al., 2008). The wolves’ genetic diversity appears to have been augmented, however, by periodic migration from mainland populations, occurring primarily through two corridors: Baffin Island and Victoria Island. This gene flow could be compromised by the loss of sea ice due to climatic warming and increased human activity (Carmichael et al., 2008). In particular, year-round marine traffic would keep an open water route through certain marine passages, blocking movement of wolves in the winter.
Tundra wolves in the Southern Arctic make heavy use of eskers, which are also a source of gravel for industrial uses such as mine site and road construction, a resource often in short supply in Arctic infrastructure construction (McLoughlin et al., 2004). Some tundra wolves are caribou specialists and follow migrating caribou for long distances (Frame et al., 2004).
Trends in wolves
There is some information on trends in wolf abundance, based on community and traditional knowledge and on observations of wolves during systematic aerial surveys to count caribou and muskoxen. There is no direct information on predation rates.
On Banks Island, wolves were poisoned in the 1950s and recovered slowly, increasing during the 1980s and 1990s (Figure 86 ) and often spotted in areas with high muskoxen density.
Group size includes adults and pups. Values for 1994 and 1998 are averages of two surveys.
Source: data from Species at Risk Committee (2012a)
Long description for Figure 86
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|Year||Groups observed||Average group size|
|1994Note * of Figure 86||12.5||2.7|
|1998Note * of Figure 86||12||3.1|
Note of Figure 86
- Note 1 of Figure 86
Values for 1994 and 1998 are averages of two surveys.
For the Bathurst Herd, wolf sightings between 1987 and 2008 during late winter aerial surveys to estimate caribou calf survival suggest no consistent trend in either wolf sightings or mean pack size (Williams and Fournier, 1996; Gunn, 2013) (B. Croft, pers. comm., 2010). During that time, the numbers of caribou declined and the number of wolves seen at their dens, as well as the number of occupied dens, declined (D. Cluff, unpublished data). The rate of sightings in the vicinity of caribou, however, suggests that predation rates were maintained.
Arctic-nesting birds winter in many parts of the world where they may be vulnerable to stressors including loss of food supplies and habitat, pollution, disturbance, and overharvesting during winter and during migration. In the Arctic, they are vulnerable to changes in their habitat and food supplies and, in some cases, to overharvest. Data are lacking or sparse for many species, and it is often difficult to determine the causes of trends. Many Arctic-nesting shorebird and landbird species are known to be declining, as are some sea ducks. Other Arctic bird groups, such as geese and swans, have mainly stable populations or have increased over the past few decades. Trends in Arctic seabirds (not including sea ducks) are discussed in the ESTR report Ecosystem status and trends report: Arctic marine ecozones (Niemi et al., 2010). An overview of seabird trends for all of Canada is presented below (text box and Figure 87 ). For more information on the status of bird populations in Canada see The State of Canada’s Birds 2012 (NABCI, 2012).
Trends of seabirds in Canada
Seabirds nesting in the Arctic are covered in more detail in the ESTR report on Arctic marine ecosystems (Niemi et al., 2010). This text box, based on the ESTR review of status and trends of seabirds in Canada (Gaston et al., 2009a), is extracted from the report Canadian biodiversity: ecosystem status and trends 2010 (Federal, Provincial and Territorial Governments of Canada, 2010).
Note: only populations with significant breeding populations, long-term datasets, and those unaffected by terrestrial human activities are included.
Source: adapted from Gaston et al. (2009a)
Long description for Figure 87
This graph shows the following information:
Worldwide, the status of seabirds is deteriorating faster than any other bird group (BirdLife International, 2010). In Canada, trends are regional in nature and result from a variety of factors, including climate change, fishing by-catch, resource extraction, transportation, and pollution (Coe and Rogers, 1997; Melvin and Parrish, 2001; Schreiber and Burger, 2002; Stenseth et al., 2004; ACIA, 2005; Gaston et al., 2009a). A trend to an earlier breeding date has been found in several populations (Bertram et al., 2001; Gjerdrum et al., 2003; Hipfner and Greenwood, 2008), as have changes in diet and condition (Parsons et al., 2008).
With the exception of ivory gulls (Pagophila eburnea), which are declining rapidly, change in Arctic seabird populations is slow and possibly the result of events on wintering grounds in the Northwest Atlantic (Gaston, 2002; Gaston et al., 2003). Changes in seabird diet and growth have been found to be related to reduction of Hudson Bay sea ice. This may have negative consequences for populations in the long term (Gaston et al., 2003). Conversely, in the High Arctic, less sea ice may benefit the birds (Gaston et al., 2005; Gaston et al., 2009b).
Waterfowl (ducks, geese and swans) are among the most important wildlife species for Inuit for food and ceremonial uses and large concentrations of migrating and nesting waterfowl are a prominent feature of the ecozone+. Status and trends of waterfowl are comparatively well known as federal agencies in the USA and Canada monitor populations annually in order to set hunting limits and guide conservation programs. There is limited monitoring on Arctic nesting grounds, but for many species regular or periodic monitoring is conducted on wintering grounds or at points where the birds congregate during migration. This section describes trends for selected Arctic tundra nesting waterfowl species with substantive parts of their global populations nesting in the Canadian Arctic. In addition, the ecozone+ provides important habitat for waterfowl species, such as Canada geese (Branta canadensis) and scoters (Melanitta spp.) that also nest in more southerly ecozones+.
King eiders (Somateria spectabilis) nesting in the Canadian Arctic overwinter both in the eastern and western part of the continent. There is growing evidence that the western population of King Eiders has declined over past few decades. The total population of king eiders nesting in the western and central Canadian Arctic was estimated in 1960 to be 900,000; in the early 1990s the population had dropped to between 200,000 and 260,000 (Canadian Wildlife Service Waterfowl Committee, 2008). Counts are conducted during spring and fall migration at Point Barrow, Alaska, about every 10 years (Peacock et al., 2013; Quakenbush et al., 2013). The most recent estimate indicates that numbers remained stable between 1996 and 2003 (Figure 88). Surveys on nesting grounds in western Victoria Island conducted in the early 1990s were repeated in 2004 and 2005 (Raven and Dickson, 2006), showing a 56% decline in abundance (Figure 89).
Error bars are 95% confidence intervals.
Source: Raven and Dickson (2006)
Long description for Figure 89
This bar graph shows the following information:
|Year||NW Victoria Island||SW Victoria Island|
Data on wintering grounds in Greenland indicate that the eastern population of king eiders is also declining, though this decline may be related to a shift in distribution related to human disturbance (Canadian Wildlife Service Waterfowl Committee, 2008). Surveys in 2010 confirmed that large numbers of king eiders were wintering at the northern tip of Labrador and southern tip of Baffin Island (Canadian Wildlife Service Waterfowl Committee, 2012). King eiders nesting in the Rasmussen Lowlands, Nunavut, declined 86% between surveys conducted in 1974–1975 and repeated in 1994–1995 (Gratto-Trevor et al., 1998).
Three subspecies of common eiders (S. mollissima)nest in the Canadian Arctic: Pacific (nesting in the western and central Arctic), northern (nesting in the Eastern Arctic), and Hudson Bay common eider.
Numbers of Pacific common eiders in migration counts at Point Barrow, Alaska, declined by 50% between 1976 and 1996, then increased again by 2002 (Figure 90). Areas of particular importance for nesting are the Dolphin and Union Strait, Coronation Gulf, and Queen Maud Gulf. The breeding population for the central Arctic was estimated at about 37,000 in surveys conducted during 1995–1998 (Canadian Wildlife Service Waterfowl Committee, 2012). Additional surveys, conducted at nesting areas in Bathurst Inlet to establish a baseline for detection of trends, indicate a decline of about 50% since 1995 (Canadian Wildlife Service Waterfowl Committee, 2012).
95% confidence intervals are shown where available.
Source: Suydam et al. (2000) and Quakenbush et al. (2013)
Long description for Figure 90
This line graph shows the following information:
Source: Raven and Dickson (2009)
Long description for Figure 91
This bar graph shows the following information:
Northern common eiders winter along the southern coasts of Newfoundland and Labrador, Quebec, and Greenland. Surveys in Greenland, where most overwintering occurs, indicate severe declines since the 1970s (Canadian Wildlife Service Waterfowl Committee, 2012). No clear trends in nesting populations were shown in surveys in Ungava Bay (historical data and surveys in 2000), while surveys along the Labrador coast showed increases between the 1980s and early 2000s. Surveys have been conducted on the wintering range in eastern Canada every third year since 2003. Results to 2009 indicate a stable population of about 200,000 (Canadian Wildlife Service Waterfowl Committee, 2012).
Hudson Bay eiders, wintering entirely in Arctic waters near Belcher Islands and off the western coast of Quebec, are subject to periodic weather-related population crashes (Canadian Wildlife Service Waterfowl Committee, 2008). Belcher Islands breeding eiders declined 70% between 1985–1988 and 1997 (Erickson and Meegan, 2007), probably as a result of a large die-off during the winter of 1991–1992 that occurred when areas of open water froze (Robertson and Gilchrist, 1998). There have been no significant winter kill events since then, and the population appears to be recovering (Canadian Wildlife Service Waterfowl Committee, 2012).
Threats: harvest and disease
Declines in some eider populations may be related to harvest. An estimated 115,000 eiders were harvested in 2001 in Chukotka, but the proportion of these that nest in Canada is not known (Canadian Wildlife Service Waterfowl Committee, 2012). The commercial and subsistence harvest of northern common eiders in southwest Greenland was estimated at over 100,000 birds, a large proportion of which are Canadian Arctic nesters (Canadian Wildlife Service Waterfowl Committee, 2008). Demographic modelling indicated that this was not sustainable, and more restrictive harvest regulations were put in place in Greenland from 2002 to 2004 (Canadian Wildlife Service Waterfowl Committee, 2012). Subsistence harvest in Canada is relatively low, but better information on harvest is needed, especially for the northern common eider.
Avian cholera may also be an emerging issue for northern common eiders--the first Arctic outbreak was recorded in 2004 in northern Quebec, and it has recurred there and in the vicinity of Southampton Island (Canadian Wildlife Service Waterfowl Committee, 2012). A small island off Southampton Island, the largest breeding colony of northern common eiders in the Canadian Eastern Arctic, has been studied since 1996. Colony size was stable at around 5,000 pairs between 2001 and 2005, then nearly doubled to an estimated 9,800 pairs in 2006. During an outbreak of avian cholera mortality, colony size dropped to 4,700 pairs in 2007 (Buttler, 2009). Further studies have demonstrated that outbreaks of avian cholera in common eider colonies at a frequency of more than one outbreak per decade could drive colonies to extinction (Descamps et al., 2012).
The western population of tundra swans (Cygnus columbianus) breeds along the Alaska coast and winters in the western United States, while the eastern population breeds from Alaska to the northeast shore of Hudson Bay and Baffin Island, wintering in the eastern United States.
Estimates of the eastern population of tundra swans acquired through midwinter surveys on the U.S. wintering ground (Figure 92) show fluctuations around a mean of about 90,000 to 100,000 birds.
Source: U.S. Fish and Wildlife Service (2012)
Long description for Figure 92
This line graph shows the following information:
|Year||Tundra swan (eastern population) abundance|
(Estimated number, thousands)
Snow geese breed only in the Arctic, including along the Hudson Bay coast. Northernmost eastern breeding populations, greater snow geese (Chen caerulescens atlanticus), breed in the Canadian Eastern Arctic and northern Greenland and winter along the U.S. Atlantic coast, staging in southern Quebec during migration. More southerly and westerly breeding populations, lesser snow geese (C. c. caerulescens), nest in colonies in coastal and inland areas across the Canadian Arctic and winter in southern Canada and all North American flyways.
Greater snow geese increased dramatically from a few thousand birds in the 1930s to over 500,000 in the early 1990s, counted on the main staging grounds in southern Quebec (Canadian Wildlife Service Waterfowl Committee, 2012). The increase is related to changes in agricultural practices in the U.S. wintering grounds leading to a change of feeding habits from marshes to cultivated fields (Canadian Wildlife Service, 2003). The spring 2012 population estimate was about one million geese (Canadian Wildlife Service Waterfowl Committee, 2012). The population has remained fairly stable since 1999 when special conservation measures were implemented (Reed and Calvert, 2007). The dramatic increase in this population has had significant impacts on staging and breeding areas in Canada through overgrazing (Environment Canada, 2007a).
Lesser snow geese also show population growth, with a levelling off of the population, as estimated through the midwinter counts, after the late 1990s and a recent increase (Figure 93). This trend is also evident in inventories on major nesting grounds, with the exception of West Hudson Bay (Figure 93). Inventories provide only a sampling, and banding and harvest studies indicate that the population of lesser snow geese may be much higher than has been estimated--likely exceeding 15 million in 2010 (Canadian Wildlife Service Waterfowl Committee, 2012).
Midwinter counts are primarily of lesser snow geese that breed in the central and Eastern Arctic.
Sources: Canadian Wildlife Service Waterfowl Committee (2012) (breeding colony data) and U.S. Fish and Wildlife Service (2012) (midwinter population counts)
Long description for Figure 93
This figure is comprised of a bar graph (A) and a line graph (B) showing the following information:
|Major breeding colonies||1970s||1980s||1990s||2000s|
|West Hudson Bay||317,000||436,300||212,000||261,000|
|Queen Maud Gulf||56,000||317,100||740,600||-|
|Western Canadian Arctic||169,600||207,500||486,100||580,000|
|Year||Midwinter population: abundance index|
This section is extracted from the ESTR technical thematic report prepared for this ecozone+, Shorebird trends Arctic Ecozone+ (Gratto-Trevor et al., 2011) and has not been comprehensively updated.
The Arctic Ecozone+ is of great importance globally for shorebird production. Sixty percent of North American shorebirds breed in the Arctic. The Canadian Arctic alone provides 75% of the North American breeding range for 15 of the 49 species of shorebirds that are common to North America (Donaldson et al., 2000).
Globally, 44% of estimated population trends for Arctic-breeding shorebirds are declining (Figure 94) making the problem more widespread than was originally thought (Morrison et al., 2001). Overall, the Arctic breeders as a group are declining 1.9% per year (Bart et al., 2007).
Globally, population trends have been estimated for 52% of Arctic-breeding shorebirds (100 biogeographical populations of 37 species). Of these, 12% are increasing, 42% are stable, 44% are decreasing and 2% are possibly extinct.
Source: Delany and Scott (2006)
Long description for Figure 94
This bar graph shows the following information:
|Population trend||% of population for which trends are known|
An analysis of fall migration count data was undertaken to determine if the declining numbers of birds recorded on migration counts could be explained by changes in migration routes or timing or by changes in detection rates (Bart et al., 2007). The authors concluded that migration counts most likely reflected a true reduction in population size. They found no evidence of major shifts in the number of birds migrating along specific routes and no major changes in variables related to detection. Annual rates of change were calculated over the period 1974 to 1998 in this study--results are shown in Figure 95 for Arctic-breeding shorebirds with sufficient survey counts in fall migration surveys conducted in the Canadian-United States North Atlantic or United States Midwest regions.
NA = North Atlantic migration survey; MW = Midwestern migration survey.
Source: data from Bart et al. (2007)
Long description for Figure 95
This bar graph shows the following information:
|Trends||Annual percent change||P|
|Black-bellied plover (North Atlantic migration survey)||-5||<0.01|
|American golden-plover (North Atlantic migration survey)||-7.2||<0.01|
|Sempalmated plover (North Atlantic migration survey)||0.4||-|
|Whimbrel (North Atlantic migration survey)||-3.3||-|
|Hudsonian godwit (North Atlantic migration survey)||-3.5||0.01<p<0.05|
|Ruddy turnstone (North Atlantic migration survey)||-12.3||-|
|Red knot (North Atlantic migration survey)||-3.3||-|
|Sanderling (North Atlantic migration survey)||-3.3||-|
|Semipalmated sandpiper (North Atlantic migration survey)||-4||0.05<p<0.1|
|White-rumped sandpiper (Midwestern migration survey)||-5.2||-|
|Baird's sandpiper (Midwestern migration survey)||-1||-|
|Pectoral sandpiper (North Atlantic migration survey)||-4.5||<0.01|
|Dunlin (North Atlantic migration survey)||-2.6||-|
|Red-necked phalarope (Midwestern migration survey)||-7.6||0.01<p<0.05|
Two major shorebird trend reviews by the U.S. Shorebird Conservation Plan Committee (in 2001 and 2004) and Canadian Wildlife Service Shorebird Committee (in 2001) assessed 18 species of Arctic-breeding shorebirds with very similar results (Table 12). Eight species were listed in both assessments as having significant population declines (Brown et al., 2001; Morrison et al., 2001; U.S. Shorebird Conservation Plan, 2004).
|Species||Trend Summary||U.S. Shorebird Conservation Plan (2004)||CWS Shorebird Committee|
|Black-bellied plover (Pluvialis squatarola)||Significant declining population trend||Significant decline||Significant decline|
|American golden-plover (Pluvialis dominica)||Significant declining population trend||Significant decline||Significant decline|
|Semipalmated plover (Charadrius semipalmatus)||Conflicting information||Not enough information||Significant decline|
|Eskimo curlew (Numenius borealis)||Significant declining population trend||Significant decline||Likely extinct|
|Whimbrel (Numenius phaeopus)||Conflicting information||Significant decline||Mixed trends|
|Hudsonian godwit (Limosa haemastica)||Probable or declining population trend, not statistically significant||Not enough information||Decline|
|Ruddy turnstone (Arenaria interpres)||Significant declining population trend||Decline||Significant decline|
|Red knot (Calidrus canutus)||Significant declining population trend||Significant decline||Significant decline|
|Sanderling (Calidrus alba)||Significant declining population trend||Significant decline||Significant decline|
|Semipalmated sandpiper (Calidris pusilla)||Significant declining population trend||Significant decline||Significant decline|
|White-rumped sandpiper (Calidris fuscicollis)||Not enough information to conclusively determine population trend (mixed trends)||Not enough information||Mixed trends|
|Baird’s sandpiper (Calidris bairdii)||Conflicting information||Not enough information||Decline|
|Pectoral sandpiper (Calidris melanotos)||Not enough information to conclusively determine population trend (mixed trends)||Not enough information||Mixed trend|
|Purple sandpiper (Calidris maritima)||Conflicting information||Stable||Significant decline|
|Dunlin (Calidris alpina)||Significant declining population trend||Significant decline||Significant decline|
|Buff-breasted sandpiper (Tryngites subruficollis)||Probable or declining population trend, not statistically significant||Decline||Decline|
|Red-necked phalarope (Phaloropus lobatus)||Significant declining population trend||Decline||Significant decline|
|Red phalaropen (Phalaropus fulicarius)||Significant declining population trend||Significant decline||Significant decline|
What is of most concern is that over the past 30 years many species trends have changed from slightly declining to significantly declining, indicating that the decline is persistent and ongoing (Morrison et al., 2001; Delany and Scott, 2006). The declines are observed in species with a range of migration, habitat, and breeding strategies and needs. Preliminary investigations by Thomas et al. (2006a) and Bart et al. (2007) found no common factors among declining species.
In the U.S. Shorebird Conservation Plan (Brown et al., 2001), population trend information was combined with five other variables (relative abundance, threats during breeding season, threats during non-breeding season, breeding distribution, and non-breeding distribution) to create a conservation prioritization scheme. The scheme, adopted in the Canadian Shorebird Conservation Plan (Donaldson et al., 2000), is useful because species with stable or slightly downward-trending populations with threats on their wintering grounds and very specific breeding ground habitat requirements may be more at risk than species with significant population declines. The highest priority species were those designated ‘highly imperiled’. Using this prioritization scheme, the only Arctic species listed in 2001 as ‘highly imperiled’, Eskimo curlew (Numenius borealis), is believed to be extinct (Environment Canada, 2007b).
In 2004, species were re-evaluated (U.S. Shorebird Conservation Plan, 2004) and the status of several species was upgraded (Table 13).
|Highly imperiled (first priority)||Species of high concern (second priority)|
Source: U.S. Shorebird Conservation Plan (2004)
Notes of Table 13
- Note a of Table 13
Upgraded species are denoted with an asterisk (*)
Local studies have recorded population declines over a range of periods. Analysis of the Atlantic coastal migration stop-overs from 1972 to 1983 (Howe et al., 1989) found significant declines for black-bellied plover (Pluvialis squatarola) (decreasing by 5.4% per year), whimbrel (Numenius phaeopus) (–8.3% per year) and sanderling (Calidrus alba) (–13.7 per year). Breeding populations of red phalarope (Phalaropus fulicarius), black-bellied plover, and American golden-plover (Pluvialis dominica) decreased significantly, by 76, 87, and 79% respectively, in the Rasmussen Lowlands (Central Arctic) over a 20-year period (Gratto-Trevor et al., 1998). Given the long time interval between studies, natural fluctuation as a result of a series of poor breeding seasons rather than a persistent and continuous population decline could explain the differences between the two study periods, but it may represent a true decline in these species (Gratto-Trevor et al., 1998).
A study in the Foxe Basin (Prince Charles and Air Force islands) found significant population declines for white-rumped sandpiper (Calidris fuscicollis) (–61%) and red phalarope (–43%) over an eight-year time span (1989–1997) (Johnston and Pepper, 2009). For red phalarope the decline was even more pronounced at East Bay, Southampton Island, where there was a 93% decline over six years (1999–2005) (Pirie et al., 2012). All shorebird species (n=5) at East Bay declined by more than 90% over the same interval. In 2007 there was a small rebound to about 33% of the original 1999 values. This coincided with a high lemming (and therefore low predation) year (Pirie et al., 2012).
Near Churchill, Manitoba, a comparison of six qualitative bird abundance studies between 1930 and the 1990s found that semipalmated sandpiper (Calidris pusilla), stilt sandpiper (Calidris himantopus) and red-necked phalarope (Phalaropus lobatus) experienced a ‘great decrease’, and dunlin (Calidris alpina) a ‘decrease’ (Jehl and Lin, 2001). Huge declines were also noted at La Perouse Bay, Manitoba (40 km east of Churchill), for semipalmated sandpiper and red-necked phalarope (Gratto-Trevor, 1994).
One of the current major limitations to determining population trends for Arctic-breeding shorebird species is the lack of reliable population estimates. In many cases intensive surveys of shorebirds on the Arctic breeding grounds have led to increases to the world population estimate for a given species (Johnston et al., 2000; Latour et al., 2005; Johnston and Pepper, 2009). This does not reflect an increase in world population size but instead is an indication that initial population estimates were probably low (Brouwer et al., 2003; Morrison et al., 2006). The large-scale Program for Regional and International Shorebird Monitoring (PRISM), which has an Arctic component, is partway through a multi-year survey program that will produce continental population estimates for 19 species of shorebirds that breed in the North American Arctic. Once the first pass of surveys is complete, a second set is planned to assess species-specific as well as North American Arctic-wide population trends (Skagen et al., 2003; Bart and Earnst, 2004; Bart et al., 2005; Bart and Johnston, 2012).
Proposed causes of shorebird population declines include: loss of migration stop-over sites, loss of wintering habitat, and life history characteristics (migratory behaviour, life history, and biogeography) which may predispose shorebirds to population decline. Future population decline is expected to be accelerated by habitat changes on the Arctic breeding grounds.
Since many shorebirds are long-distance migrants that tend to gather in very large numbers at relatively few sites, loss of one or two major stop-over sites could have a huge effect on shorebird populations. Declining food availability at existing stop-over sites can also have a large impact on populations because birds may not be able to take in enough fuel to move to the next stop-over site, or may not be able to acquire the body stores essential for survival and successful reproduction (Senner and Howe, 1984; Donaldson et al., 2000; Morrison et al., 2001; Baker et al., 2004; Morrison et al., 2004; Morrison et al., 2007). Analysis of population trends of North American shorebirds found species that followed continental migration routes (as opposed to coastal or oceanic migration routes) were at higher risk of population decline because of ecosystem loss and alteration (Thomas et al., 2006a; Bart et al., 2007). Continental migrants use small, ephemeral ponds and wetlands that are scattered over a large area. These ponds and wetlands are difficult to delineate for conservation initiatives making it harder to preserve them as compared to larger stop-over sites (Thomas et al., 2006a). Little is known about Arctic stop-over sites because of their remoteness. Observations along a 200 km stretch of coast line in the Kivalliq Region (northwestern Hudson Bay) during the 2008 spring migration found hundreds of High Arctic nesting migrants feeding on insects in the wrack lines on their journey north to the breeding grounds (Johnston and Rausch, unpublished data). The importance of sites such as these to migration and subsequent breeding success is not known.
Loss or degradation of habitat on the non-breeding grounds from human activities such as oil pollution (Harrington and Morrison, 1980), mechanical dredging or fishing (Piersma et al., 2001), conversion of native grasslands and wetlands to agriculture (Isacch and Martinez, 2003; Shepherd et al., 2003), and tourism and development on marine beaches (Blanco et al., 2006) may be a cause of population decline (Thomas et al., 2006a). Complicating the assessment of the importance of wintering habitat is that little is known about food resources on the wintering grounds (Morrison et al., 2004). Threats on the wintering grounds, however, have been found to have a weak influence on the likelihood of a species being in population decline (Thomas et al., 2006a).
The intrinsic biology of shorebird species may make them more susceptible to population decline. Migratory behaviour (such as distance and routes) is suspected to be the most influential intrinsic factor, with more continental migrants in population decline than coastal or oceanic migrants (Thomas et al., 2006a). Phylogenetic characteristics such as body and clutch size, lifespan, and relatedness were found to be unimportant to population decline, but limited clutch sizes means that recovery following a decline is likely to be slow (Myers et al., 1987). Sexual selection may have an influence on declining populations since most socially polygamous species have declining populations while socially monogamous species have stable or increasing population trends--but the data are not conclusive. There are no clear intrinsic factors held in common by shorebird species with declining population trends and extrinsic factors are more likely to be the primary cause of decline (Thomas et al., 2006a; Thomas et al., 2006b; Bart et al., 2007).
Habitat changes in the Arctic caused by climate change are expected to have an exacerbating effect on the declining population trends of Arctic-breeding shorebirds (Bart et al., 2007). Arctic-breeding shorebirds are adapted to the annually variable weather conditions of the Arctic during the breeding season. However, their conservative life-history strategy (low reproduction and long lifespan) makes it difficult for them to adapt to climate change. This puts Arctic-breeding shorebirds more at risk of population decline than other groups (Donaldson et al., 2000; Meltofte et al., 2007). Effects of climate change on breeding habitat include: drying of tundra ponds (Walsh et al., 2005; Smol and Douglas, 2007a), shrub encroachment (Callaghan et al., 2005a), and asynchrony of insect-chick hatch (Tulp and Schekkerman, 2006).
The synchrony of shorebird chick hatch with the peak of insect emergence is not as critical as hatch occurring when there is sufficient food supply. The availability of the food supply is strongly influenced by weather and a sufficient supply is only available for 40% of the insect season (Tulp and Schekkerman, 2008). The peak date of insect emergence fell between 8 July and 23 July for 75% of the 33-year study period. These earliest and latest peak emergence dates were recorded in consecutive years, showing that the date of peak emergence is not advancing linearly with time. Overall, however, the date of peak insect emergence as well as the range of dates with sufficient food available for the normal growth of chicks is getting earlier in the season (Tulp and Schekkerman, 2008).
Since Arctic shorebirds time nest initiation to occur as soon as the snow melts, the advancement in the timing of insect emergence may not be as critical for the survival of chicks hatched from the earliest nests. It could be a serious problem for chicks from late nests, or from re-nests (clutches laid late to replace an earlier nest that was unsuccessful) because they will hatch too late in the season to obtain sufficient food resources (Meltofte et al., 2007). Studies indicate that, while shorebirds exhibit considerable flexibility in nest initiation, the accommodation in timing may not completely match the change in date of snowmelt, especially in years with exceptionally early springs (Smith et al., 2010; McKinnon et al., 2012; Grabowski et al., 2013). Further analysis is needed to determine if snowmelt is advancing at the same rate as the timing of insect emergence, which would permit birds to nest earlier. It is not known whether shorebirds will be able to adjust their migration strategies to arrive on the breeding grounds sooner in response to an earlier snow-free season. Species which make the final jump to the breeding ground from latitudes closer to the Arctic may be more successful than species that use internal length-of-day cues to initiate migration from very distant wintering grounds (Tulp and Schekkerman, 2008).
This report has been extracted from the ESTR thematic technical report Landbird trends in Canada, 1968-2006 (Downes et al., 2011) and has not been updated with new information.
The Arctic Ecozone+ is relatively pristine and there are few immediate threats to landbirds from human activity, although species are affected by climate change, contaminants, and other wide-ranging factors. All birds listed in Table 14 overwinter in more populated areas of Canada and the United States where development pressures are more intense in both their wintering ranges and along their migration routes. Canada has a high stewardship responsibility for these species because large portions of their Western Hemisphere breeding populations are concentrated in the Arctic Ecozone+.
There are relatively few landbird species in the Arctic Ecozone+ and few data on their population trends. The lack of information on population status and trends has been highlighted as the most pressing conservation need in relation to landbirds for this region (Rich et al., 2004). Because of the remoteness and lack of roads, the Breeding Bird Survey (BBS) has not been carried out in the Arctic and there are few other surveys of breeding birds. However, many birds that breed in the Arctic spend their winters in the United States and more southerly parts of Canada, where their populations can be monitored by the Christmas Bird Count (CBC). The CBC, now over 100 years old, monitors the status and trends of winter bird populations through an all-day, annual census conducted by groups of volunteers throughout North America. Data from the CBC complement the BBS by providing results for some species that cannot be monitored on their breeding grounds. Results presented below are preliminary findings based on CBC data from Canada and the United States combined (Butcher and Niven, 2007).
CBC trends (Table 14 and Figure 96) suggest that several species such as Harris’s sparrow (Zonotrichia querula) and snowy owl, have been undergoing long-term, statistically significant declines since the 1960s. Other species, such as the rough-legged hawk (Buteo lagopus) and Lapland longspur (Calcarius lapponicus), have shown relatively stable overall population trends.
|Species||Population Trend (%/yr)||P||CBC Abundance Index|
|CBC Abundance Index|
|CBC Abundance Index|
|CBC Abundance Index|
|CBC Abundance Index|
|Hoary redpoll (Acanthis hornemanni)||–4.97%||significant trends (P<0.05)||0.29||0.18||0.14||0.09||-68%|
|American tree sparrow (Spizella arborea)||–2.16%||significant trends (P<0.05)||62.8||56.3||42.4||34.4||-45%|
|Harris's sparrow (Zonotrichia querula)||–2.13%||significant trends (P<0.05)||9.6||7.5||6.2||5.3||-45%|
|Snowy owlNote a of Table 14 (Bubo scandiacus)||–2.12%||significant trends (P<0.05)||0.24||0.17||0.14||0.11||-53%|
|American pipit (Anthus rubescens)||–0.97%||significant trends (P<0.05)||5.9||4.8||4.4||4.7||-19%|
|Snow bunting (Plectrophenax nivalis)||–0.93%||-||15.8||14.4||11.6||9.3||-41%|
|Rough-legged hawk (Buteo lagopus)||–0.06%||-||1.8||1.6||1.6||1.7||-7%|
|Lapland longspur (Calcarius lapponicus)||0.40%||-||0.9||0.9||0.9||1||12%|
|Common redpoll (Acanthis flammea)||0.60%||-||19||17.8||18.1||17.9||-6%|
"Change" is the percent change in the average index abundance between the first decade for which there are results (1970s) and the 2000s decade (2000–2006).
Source: Downes et al. (2011)
Notes of Table 14
- Note a of Table 14
This reflects sightings of snowy owls in southern latitudes. Note, however, that many snowy owls remain the in the Arctic or northern taiga throughout the winter, far from human settlements.
The rough-legged hawk and snow bunting show no significant trends, though the latter may be in decline; the snowy owl and Harris’s sparrow have declined significantly (P<0.05). See the note on snowy owl for Table 14.
Source: Downes et al. (2011)
Long description for Figure 96
These four line graphs show the following information:
|Year||Christmas Bird Count|
|Year||Christmas Bird Count|
|Year||Christmas Bird Count|
|Year||Christmas Bird Count|
Harris’s sparrow, a species with its entire breeding range in Canada, is classified by Partners in Flight as a Continental Watch List species (Rich et al., 2004). The species has apparently experienced a long-term decline over the last 40 years (Figure 96b). Because of its isolated breeding range, direct influence of human activity on the breeding range is unlikely to be a factor in its decline. Harris’s sparrows, however, are susceptible to predation, especially by merlins (Falco columbarius), whose populations are increasing. The influence of factors such as climate change is unknown (Niven et al., 2004; Norment and Shackleton, 2008).
Population indices for the snow bunting (Plectrophenax nivalis) vary annually but this species has apparently experienced a large decline in its population over the long term (Audubon Society, 2007) (Figure 96d). The Arctic has a very high stewardship responsibility for snow buntings (Rich et al., 2004), which breed throughout the Arctic Cordillera and Northern Arctic and the northern portions of the Southern Arctic. Causes of the apparent snow bunting decline are not known and further demographic studies are needed. Potential factors affecting populations on the breeding grounds include availability of nest sites (rock talus or cavities, driftwood cavities, or sometimes buildings), predation on nests and incubating females, and availability of food (Lyon and Montgomerie, 2011). Reduction in snow bunting populations may be related to earlier thawing of the tundra, which allows more woody plants to grow, converting the open foraging habitats preferred by snow bunting to more shrub-dominated communities (Audubon Society, 2007).
Insects and pathogens
Warmer temperatures will benefit free-living bacteria and parasites whose survival and development is limited by temperature and there are indications that this is occurring in some Arctic muskox and caribou populations (see section on Wildlife diseases and parasites on page 62). Arthropods such as ticks that transmit disease agents may also benefit from climate change, and the diseases they spread may consequently become more prevalent or widespread. Other abundant flying insects in the Arctic, especially oestrids (warble flies) and mosquitoes, are a significant factor in caribou ecology and their abundance is closely related to climate conditions. Inuit from the Queen Elizabeth Islands report that mosquitoes and black flies have increased along with warmer weather in certain areas (e.g., Nunavut Tusaavut Inc., 1997).
Major range shifts of species native to Canada
Arctic residents have reported changes in animal behaviour and distribution. Indigenous people of the Arctic especially are familiar with the long-term trends, variability, and extent of species distribution, and habitat use and behaviour. As they travel and work in their lands, they observe changes that cannot be easily detected through other monitoring and research. These observations have been documented through several studies using interviews, questionnaires, and group discussions, but there is not a systematic methodology or repository for this information. See Table 15 for examples of reported observations of range shifts.
|Location||Observations that may indicate species range shifts||Reference|
|Nunatsiavut||See new insects and strange kinds of birds: 2001 small yellow and red birds came in large groups that ate|
anything including seal skin and seal fat. Hummingbirds seen (new species).
|Nickels et al. (2005) Nunatsiavut workshop|
|Nunavik||Deer were seen in the summer of 2002 for the first time around the community of Kangiqsujuaq.||Nickels et al. (2005) Nunavik workshop|
|Chesterfield Inlet||Insects, birds and even grizzly bear were reported to be appearing farther north than usual.||Nunavut Research Institute (2004)|
|Kitikmeot||New birds seen for the first time such as the robin and an unidentified yellow songbird. More abundant and new species of shrubs and lichens.||Thorpe (2000)|
|Baker Lake||At least ten new kinds of insects were reported to be seen in the area, all winged insects, some recognized from the treeline area.||Fox (2004)|
|Banks Island||New bird species observed include robins and barn swallows. Shorter winters, longer summers and more water were thought to have caused an increase in the number of insects and led to the arrival of new species of beetles and sand flies. Changes in behaviour include overwintering of birds that normally migrate.||Ashford and Castleden (2001)|
|Beaufort coast||Ravens and eagles more numerous further north including Tuktoyaktuk.||Gordon et al. (2008)|
|Western Inuvialuit Settlement Region||Cougars are new to the treeline areas.||Nickels et al. (2005) Inuvialuit workshop|
Root et al. (2003) reviewed 143 journal articles related to changes in species distribution (globally) and found that 80% of species show changes that can be explained by physiological constraints. These included changes in density at a particular location, changes in range either poleward or up in elevation, changes in timing of migration, changes in phenology (plant growth and flowering), change in the timing of egg laying, changes in morphology (body size), and shifting frequencies of genetic markers. Many of the species with documented range changes were birds and butterflies, which can more easily adjust their distribution than other taxa (for example, small mammals) that may be constrained by geological features such as rivers (Root et al., 2003).
Climate-related changes in tundra ecosystems--including reduction of lichens, increase in shrubs, and increased canopy height (see section on Changes in tundra plant communities on page 98)--underlie observed and anticipated changes in distribution of flora and fauna in the Arctic Ecozone+ (Gilg et al., 2012; Reid et al., 2013). More rapid changes may be caused by human introductions-- either inadvertently or as deliberate moves to establish populations--for example, the introduction and re-introduction of muskoxen to some regions (see section on Muskoxen on page 129). Land-use change and disturbance from industrial activity and other human activities may also lead to changes in animal distribution.
Range extensions due to climate change could show up as previously vagrant/accidental species that appear more and more frequently and finally become established. Such extensions could involve boreal species moving into the Southern Arctic, or species from that zone moving into the Northern Arctic. Detecting such changes in status requires significant effort in well-designed monitoring programs.
Mammals that have experienced relatively recent range shifts include grizzly bears, which have moved northward in some areas (see the section on Trends in species of conservation concern, on page 104) and moose, which have also expanded their range northward (see text box below). Red foxes have expanded their distribution in parts of the Canadian Arctic and not in others. The current, best-supported hypothesis is that red fox expansion in the 20th century in Arctic Canada was promoted by human food supplementation, and that such food supplementation could lead to similar effects elsewhere if not checked (see the discussion on this topic under Major human stressors on ecosystem structure on page 100).
Expansion of moose northward
Moose (Alces alces) are found primarily in shrub habitats south of the Arctic Ecozone+ but they have been increasing in recent decades in much of the Arctic Ecozone+. Expansion of moose ranges can have an impact on populations of predators and of other ungulates. Moose can provide alternate prey for predators, and alternate food sources in Arctic subsistence economies, especially when caribou are scarce. Wolf populations in caribou winter ranges (in the taiga ecozones+) can increase in response to higher levels of other prey such as moose. When caribou return to the winter range they are preyed on more heavily by the increased number of wolves (e.g., Basille et al., 2013).
Evidence of changes in moose distribution includes:
- Expansion of moose range northward into suitable habitat has been documented for the Northwest Territories and Nunavut, with more frequent sightings of moose in shrub-rich tundra regions, especially since the 1970s. In the NWT, moose have been sighted as far north as near Bathurst Inlet, Coronation Gulf, and the east side of Victoria Island (NWT Department of Environment and Natural Resources, 2012).
- Although not historically resident in Nunatsiavut, moose have increased in numbers and range over the past four decades. Moose hunting commenced there in 1977. The northward range expansion of moose reached the treeline in Nunatsiavut in the 1990s (Chubbs and Schaefer, 1997). Inuit report that moose have moved north of Nain and to the Voisey’s Bay area, not far south of the Arctic Ecozone+ (Nickels et al., 2005 Nunatsiavut workshop).
- Moose from the British–Richardson Mountains of the Yukon (in the Taiga Cordillera Ecozone+) probably began frequenting the Yukon North Slope coastal plain within the last 100 years (Wildlife Management Advisory Council (North Slope), 2008a), utilizing mainly riparian shrub zones. Surveys in the 1980s showed that most of the moose summering along the coastal plain migrated about 100 km south during the winter--but a survey in 2000 (Smits, 2000) as well as observations by Inuvialuit (Arctic Borderlands Ecological Knowledge Co-op, 2003) indicate that increasing numbers are wintering in the Arctic Ecozone+ section of their range . Moose in the northern part of their range in the Yukon are increasing: counts in the Richardson Mountains/coastal plains surveys increased by 67% from 1989 to 2000 (Smits, 2000).
Human stressors on ecosystem composition
Climate change is currently a human-induced stressor for some species, such as polar bears, and is thought to contribute to changes in others, including shorebirds and caribou. Implications of climate change are dealt with throughout this report.
Human population increase
Despite the low density of human population in the Arctic Ecozone+, human settlements and associated infrastructure affect ecosystem composition through habitat change, disturbance, and harvest. Increasing human populations place stress especially on large mammals. For example, disturbance to female polar bears during denning, or prior to denning if it alters their choice of dens, may affect growth of cubs (Lunn et al., 2004). Disturbance at den sites may reduce productivity and alter the distribution of tundra wolves (Walton et al., 2001).
The reverse is also true: decreasing intensity of human use may result in increasing wildlife populations. For example, muskox populations in the Queen Maud Gulf mainland and islands increased following withdrawal of Inuit families to more distant settlements in the 1950s. Barren-ground caribou, lesser snow goose, and Ross's goose (Chen rossii) populations have also rapidly increased in the area since the late 1960s (Gunn et al., 1984).
Contaminants and pollution
Sources of Arctic wildlife contamination include long-range transport of persistent organic pollutants (POPs) and mercury, mines, DEW line sites, transportation-related emissions, and at least one military base in addition to the DEW line sites. An emerging issue is potential future contamination related to the anticipated increase in marine traffic and oil and gas activity in the Arctic Ocean as the ice melts.
A local source of pollution, with the potential for transfer of contaminants to wildlife, is liquid and solid waste disposal in Arctic communities. For example, Nunavut dumps are seldom contained and waste is often burned in an uncontrolled manner (ARKTIS Solutions, 2011). Wastewater from small communities is stored in open lagoons that may lack liners. In summer some communities decant the liquids from the lagoons onto the surrounding ground, into lakes, and into the sea (Jamieson and Krkosek, 2013).
Contaminants in wildlife
The concentration of toxic contaminants in wildlife has been a concern in Canada since the 1970s and monitoring has been ongoing since then for some species and some compounds. In the late 1980s concerns about elevated levels of contaminants in wildlife species that were important to the traditional diets of northern Aboriginal peoples led to increased monitoring and research in northern Canada. Early results found a wide variety of substances, many of which had no Arctic or Canadian sources, but which were, nevertheless, reaching unexpectedly high levels in Arctic ecosystems primarily through long range atmospheric transport (INAC, 2008).
Air currents transport several types of contaminants, for example, PCBs and mercury, to the Arctic from southern, more populated and industrial parts of the world. When the airborne pollutants arrive in the Arctic, cold temperatures cause the air, and the pollutants, to fall lower in the atmosphere and to be deposited on the land and into aquatic environments. The contaminants are transported through soil and water, incorporated into biota, and often concentrated as they move up the food chain. Many of these substances become concentrated in particular tissue types, especially fat, or in particular organs such as livers. In general, the highest contaminant levels tend to build up in marine fish-eating animals or their predators, especially those that are long-lived such as whales and polar bears.
Concentrations of all contaminants in wildlife vary depending on the individual, the species, and the location. Trends for some animal populations are shown in Table 16. Three classes of contaminants are of concern in the Arctic Ecozone+ (Stow, 2008):
- Concentrations of persistent organic pollutants (POPs), such as the pesticide dichlorodiphenyltrichloroethane (DDT), polychlorinated biphenyls (PCBs) and toxaphene, have generally declined (Riget et al., 2010). In long-lived species such as whales, decline of contaminant levels has been slower.
- Brominated flame retardants [e.g., polybrominated diphenyl ethers (PBDEs) and fluorinated surfactants, perfluorooctane sulfonate (PFOS)] are examples of toxic contaminants that have been increasing since the mid-1980s in most locations and some species [e.g., ringed seals (Pusa hispida); see the ESTR report Ecosystem status and trends report: Arctic marine ecozones (Niemi et al., 2010)].
- Increases in mercury have been observed since the mid-1970s in some Arctic marine mammals, seabirds, and fish around the circumpolar Arctic, with the highest proportion of increases being in Canada and Greenland (Riget et al., 2011). There were far fewer records for terrestrial mammals, and few are of adequate length to determine trends (Riget et al., 2011). While some of the mercury found in wildlife is from natural sources, much of the mercury found in marine and aquatic systems is from industrial sources.
|Trend||Measurement taken from…||Contaminant||Time Frame||Reference|
|Significant increasing trend||Caribou (Bluenose East Herd)||Mercury||1994–2002 increase, decline 2002-2006||Gamberg (2008)|
|Significant increasing trend||Caribou (Porcupine Caribou Herd)||Mercury||1993 to 2007 (slight increasing trend)||Gamberg (2008)|
|Significant increasing trend||Caribou (Bathurst Herd)||Mercury||1992–2006||Gamberg (2008)|
|No significant trend but indications of possible change – more data needed to draw conclusions||Charr muscle (Amituk Lake, Resolute)||Mercury||1989–2007||Muir (2008)|
|Significant decreasing trend||Charr muscle (Amituk Lake, Resolute)||PCBs, DDTs, toxaphene||1989–2007||Muir (2008)|
|No significant trend but indications of possible change – more data needed to draw conclusions||Charr muscle (Amituk Lake, Resolute)||Mercury||1993–2007||Muir (2008)|
|No significant trend||Charr muscle (Amituk Lake, Resolute)||PCBs||1993–2007||Muir (2008)|
|Significant decreasing trend||Charr muscle (Amituk Lake, Resolute)||PCBs, DDTs, toxaphene||1993–2007||Muir (2008)|
|Significant decreasing trend||Charr muscle (Lake Hazen, Quttinirpaaq National Park)||PCBs, DDTs, toxaphene||1990–2007||Muir (2008)|
|No significant trend but indications of possible change – more data needed to draw conclusions||Charr muscle (Lake Hazen, Quttinirpaaq National Park)||Mercury||1990 - 2007||Muir (2008)|
|No significant trend||Charr muscle (Amituk Lake, Resolute)||Mercury||1993–2007||Muir (2008)|
|No significant trend||Charr muscle (Amituk Lake, Resolute)||PCBs||1993–2007||Muir (2008)|
|Significant decreasing trend||Polar bear (Western Hudson Bay)||DDTs||1991–2008||Letcher (2008)|
|No significant trend||Polar bear (Western Hudson Bay)||PCBs||1991–2008||Letcher (2008)|
Based on a review of research on ecological effects of POPs and metals, Fisk et al. (2005) found that concentrations in Arctic terrestrial wildlife, fish, and seabirds are generally below effects thresholds, with a few possible exceptions. These were PCBs in burbot (Lota lota) in some Yukon lakes (in the Boreal Cordillera Ecozone+), Greenland shark (Somniosus microcephalus), glaucous and great black-backed gulls (Larus hyperboreus and L. marinus), and dioxin-like chemicals in seabird eggs. PCB and DDT concentrations in several Arctic marine mammal species exceed effects thresholds, although there was no evidence of stress in these populations. Some polar bears and beluga whales have high enough levels of contaminants that they may experience sublethal effects such as immune suppression. Papers reviewed found weak relationships between cadmium, mercury, and selenium burdens and health biomarkers in common eiders in Nunavut, although these metals were probably not influencing the health of these birds. The authors concluded that there is little evidence that contaminants are having widespread effects on the health of Canadian Arctic organisms, with the possible exception of polar bears. However, they recommended that further research and better understanding of contaminant exposure in Arctic biota is needed considering factors such as tissue levels that exceed effects thresholds, exposure to “new” organic contaminants of concern (such as brominated flame retardants), contaminated regions, and climate change (Fisk et al., 2005).
Sources of local and regional pollution
Current and past activities in the Arctic result in potential for both contamination of wildlife, affecting their safety as human food, and effects on the animals themselves. Overall, the Arctic has relatively few sources of local pollution, with little industry and a very low population density. Sources of pollutants include mining activity, especially abandoned mines, military activities, especially DEW line sites (see text box The Distant Early Warning (DEW) Line in the introductory section on page 12), and oil and gas activities.
An example of localized contamination from a still-active radar station is Saglek Bay, Nunatsiavut (Labrador), which had been the site of a military radar station since the late 1950s. PCB contamination in soils and leaching into the marine environment was discovered in 1986 and cleanup was initiated in 1997–1999. Sediments, fish, and seabirds (black guillemots, Cepphus grylle) have accumulated high levels of PCB concentrations due to marine sediment contamination (Kuzyk et al., 2003; Kuzyk et al., 2005).
In 2005, two sites in the Arctic Ecozone+ had been designated as “contaminated” by the federal contaminated sites program: Kittigazuit Military Site and Atkinson Point Military Site (INAC, 2005), both within the Inuvialuit Settlement Region.
Ecosystem Goods and Services
More so than perhaps any other group in Canada, modern Inuit and Inuvialuit occupy and use huge tracts of land as they travel to neighbouring communities to visit and conduct business and to remote (often traditional) camp sites to hunt, fish, and trap (Damas, 2002). “Travel” in this context means not only travel by dogsled, snow machine, and boat, but also the charter of aircraft to visit, confer with, and exchange goods with friends, relatives, and colleagues in remote communities. Many communities and individual families regularly occupy several traditional hunting, trapping, and fishing camps far from their primary homes for long periods at different seasons. These annual events are important sources of social cohesion. As with the capture and consumption of country foods (see below), use of space is integral to Inuit and Inuvialuit culture and tradition (Condon et al., 1995).
Ice and snow form an important part of this living space. Travel over river, lake, and sea ice provides access to the land for hunting and fishing and the main corridors for regular transport of goods and people over much of the year.
Permafrost is another important component of living space. It provides an unquantifiable service by supporting building structures; when it unexpectedly thaws, structures built on it collapse. Structures such as homes must be insulated to prevent them from thawing the permafrost beneath them, and this is an additional (quantifiable) cost; its presence, however, allows structures to be built on land that otherwise would be unusable for habitation, a definite benefit. The reduction in extent of permafrost and the increase in the depth of annual thawing that may accompany climate change are present risk factors and future threats.
Detailed, quantitative studies since the early 1970s (Usher, 1976; Berger, 1977) through the 1980s (Gunn et al., 1986) and to the present (Wein et al., 1996; Helander-Renvall, 2005; Inuuvik Community Corporation et al., 2006) show that Inuit and Inuvialuit rely heavily on traditional foods (also called country foods): wild-caught fish and wildlife and collected plants such as berries. Harvest of country foods is not merely a matter of calories and nutrients in a subsistence economy, but a central feature of Inuit and Inuvialuit cultural identity.
Extent and patterns of subsistence activities were included in the Survey of Living Conditions in the Arctic (SLiCA), an interview-based study of Inuit around the Arctic, conducted in Canada in 2001 in cooperation with Statistics Canada, including all four major Inuit groups, and with a total sample size of 4,700 interviews (Kruse et al., 2009). Results were very consistent across Canadian Inuit groups and show the high rates of participation in fishing and hunting, with far fewer Inuit being active trappers (Figure 97). As shown in Figure 98, participation rates in subsistence activities are higher for people who live outside of major population centres of the Arctic, especially in Western Arctic.
Based on the Survey of Living Conditions in the Arctic, a survey of circumpolar Inuit, a project of the Arctic Council. Percent is of respondents who engaged in the subsistence activity in the 12 month period before the interview (2001). Sample size is 4,700.
Source: Kruse et al. (2009)
Long description for Figure 97
This bar graph shows the following information:
|Hunt caribou, moose or sheep||62||47||63||59|
Based on the Survey of Living Conditions in the Arctic, a survey of circumpolar Inuit, a project of the Arctic Council. Percent is of respondents who engaged in the subsistence activity in the 12 month period before the interview (2001).
Source: Kruse et al. (2009)
Long description for Figure 98
These two bar graphs shows the following information:
|hunt caribou, moose or sheep||34||68|
|Hunt caribou, moose or sheep||53||60|
Caribou are an important part of Canadian Arctic cultures and still play a central role in people's lives. A measure of the importance is the annual harvest, which in Nunavut (1996–2001) averaged 24,522 caribou (Priest and Usher, 2004). In the Northwest Territories, Dene, Inuvialuit, Metis, and non-Aboriginal people from almost all communities hunt the migratory herds and the minimum annual harvest is 11,000 caribou, valued as at least $17 million dollars, including meat replacement and outfitting (Department of Environment and Natural Resources, 2006). Assuming an average carcass weight of 45 kg, the Northwest Territories and Nunavut harvest is about 1.6 million kg caribou. At a beef replacement value of $20/kg, the annual average harvest of caribou just for meat replacement is $35 million. This excludes any commercial harvesting or any value for hides, let alone the intangible cultural strength and value of hunting. A study commissioned by the Beverly and Qamanirjuaq Caribou Management Board estimated the total net economic value of the caribou harvest of the Beverly and Qamanirjuaq herds at $20.0 million annually (InterGroup Consultants Ltd., 2013). This includes domestic and commercial harvest, less production costs.
Other wild meat, such as muskox (Campbell and Setterington, 2001), beluga whales, waterfowl, and, especially, fish, are important sources of protein and are essential connections to community heritage and traditions.
- Stressors/cumulative impacts
Stressors are discussed throughout this report, with a common theme of growing evidence of impacts of climate change. Some of the impacts, such as those from melting permafrost, are only beginning to be observed and monitored, though there is a good body of research linking climate variability and change with ecological processes, structure, and composition. As all climate model projections show continued above global average increases in temperatures in the Arctic, climate change will continue to be the major large-scale stressor for the foreseeable future. Other stressors are also widespread, including contaminants from long-range atmospheric transport, but appear to be of lesser magnitude in terms of impacts on ecosystems. At the local and regional level, stressors that lead to habitat fragmentation and disturbance, overharvest, and localized contamination can also be significant for ecosystems.
These stressors interact, often in complex and unpredictable ways. Freshwater ecosystems in the Arctic Ecozone+, for example, are undergoing changes related to climate change, but they also are affected by other human activities, including depletion of stratospheric ozone, deposition of persistent organic pollutants and mercury transported through the atmosphere, and land and watercourse disturbance associated with development activities (Schindler and Smol, 2006). Projected warming and changes in precipitation will result in higher contaminant loads and biomagnifications, while changes in ice cover are predicted to increase ultraviolet radiation levels, resulting in cumulative and/or synergistic effects on aquatic ecosystem structure and function (Wrona et al., 2006). The range of human stressors of concern for terrestrial systems is illustrated below in the section on Main threats to caribou. There are, of course, bound to be unexpected and poorly understood stressors affecting Arctic ecosystems--for example, long-term transport of nitrogen compounds (pollutants) may increase nitrogen availability in tundra soils, interacting with warming temperatures to alter habitat for Arctic species.
Main threats to caribou
This section is mainly summarized and updated from the ESTR technical thematic report Northern caribou population trends in Canada (Gunn et al., 2011c). Because migratory tundra caribou mainly winter to the south of the Arctic Ecozone+, some of this discussion also focuses on the taiga ecozones+.
In northern Canada, migratory tundra caribou typically have large annual ranges with annual migrations of hundreds to thousands of kilometres linking their seasonal ranges. While caribou are exposed over the spatial scale of their annual range, they are also accumulating responses to environmental influences over their lifetime (typically about 15 years for an adult female). While much attention has focused on the effects of industrial exploration and development on caribou, the extent and severity of effects depend on the vulnerability of caribou. Vulnerability changes during the cycles of caribou abundance as, during declines and the phase of low numbers, factors that influence caribou births and deaths can have stronger effects. The challenge for hunters, biologists, managers, and co-management boards, therefore, is to understand the vulnerability of their herds, keeping in mind all these factors and assessing not only what the impact might be on the herds, but also what the impact will be on those communities that depend on the abundance of caribou. In the following sections, we present some examples of how ecological influences such as predation, harvesting, and parasites change caribou vulnerability and how threats from development, contaminants, and climate change affect caribou.
Predation, parasites, disease
Although predation, diseases, and parasites are part of the ecology of migratory tundra caribou, they are listed here as threats because their role in influencing trends in caribou abundance interacts with stressors related to human activities. The interactions work in a number of ways, both at the individual and herd scales, and include variables such as whether predation is additive or compensatory to harvesting (meaning whether interactions cause the overall effect of predation plus harvesting to be more or less than the sum of the individual effects). The major predators of migratory tundra caribou are wolves and grizzly bears, but wolverines, lynx, and eagles all take caribou as well. As caribou abundance declines, the role of wolves and bears has greater effect on the caribou until the numbers of predators themselves decline (Bergerud et al., 2008).
Information on trends in predation rates or numbers of predators is patchy, but some comparisons over time are available using a relatively simple index. For example, Heard (1992) reported an average of eight wolves seen per 100 hours of flying surveys in the Queen Maud Gulf area during the 1980s, lower than the 24 to 33 wolves/100 hours observed during surveys of the Ahiak Herd in 2007 and 2008 (Poole et al., 2013). For the Bathurst Herd, wolf sightings observed for 16 years between 1987 and 2008 during late winter aerial surveys to estimate caribou calf survival suggest no consistent trend in either wolf sightings or mean pack size (Figure 99). During that time, the numbers of caribou declined 90% and the number of wolves seen at their dens and the number of occupied dens declined (D. Cluff, pers. comm., 2012). Thus, while caribou were declining and wolf populations were probably declining in the region, the sightings of wolves in the vicinity of caribou did not decline. This observation suggests that predation rates were maintained, which would have increased the vulnerability of the caribou herd.
Information on the status and trends of diseases and parasites in migratory tundra caribou is fragmentary among herds. Knowledge about if and how disease rates change when caribou abundance changes is particularly lacking. We do, however, know that in at least one instance, a disease has played a role in increasing the vulnerability of caribou. On Southampton Island, caribou abundance and condition are monitored and a high incidence of brucellosis in both sexes is implicated in the herd’s decline (see section on Wildlife diseases and parasites on page 62).
A recurring theme in wildlife and fisheries management over the centuries is that numerical abundance is not always a hedge against declines to the point of local extinctions. One has only to think of salmon, northern cod, and plains bison to remember that numerical abundance carries the risk of over-confidence--“there’s still lots”. What determines persistence is rate of change, not the size of the starting population. Another contribution to confidence among users is that the caribou, being cyclic in their abundance, have been low in number before and have come back. Given changing environmental conditions, however, the past may not be a secure guide to the future.
Changes that have occurred on caribou ranges since the 1970s that influence caribou vulnerability include an overall increase in the number of people and shifting socio-economic patterns (such as wage-earning), both of which may influence harvest levels. The human populations of the Arctic and the three taiga ecozones+ have all increased, the combined population almost doubling, from 59,390 people in 1971 to 107,213 people in 2006 (Environment Canada, 2009b). The increase in number of people is reflected in the increasing size of larger communities (centralization) and increased seasonal and all-year road developments, especially in the Northwest Territories and northern Saskatchewan, on the southern edges of the winter ranges in the taiga and boreal ecozones+ (BQCMB, 2011; Trottier, 2011, pers. comm.).
The increasing number of people, a shift to wage-earning, and changing technologies for hunting (snowmobiles, winter roads, and rapid communications) have likely altered hunting effort and made finding and harvesting caribou more efficient. The relationship between hunting effort and harvest levels, however, is largely unknown and this limits understanding of the effects of hunting. Most of the understanding about the importance of measuring hunting effort is from a range of exploited species other than caribou (Ludwig, 2001). Schooling fish have parallels with gregarious caribou in relation to harvest vulnerability. For pelagic fisheries, constant yield harvests can lead to population collapses if harvesting effort is not directly related to local abundance (Mullon et al., 2005).
Understanding how harvest can add or subtract from herd vulnerability is complicated as there are few measures of hunting effort, and data on harvest levels are sporadic over time. Levels of, or trends in, wounding loss are unknown, although increased effort in management planning includes education and help from Aboriginal elders about respectful hunting practices (e.g., PCMB, 2009; Tlicho Government and Department of Environment and Natural Resources, Government of the Northwest Territories, 2011). Harvest levels at the community scale vary; this is related to annual variability of the distribution of caribou on their seasonal ranges. In Canada’s territories, most hunters are Aboriginal and, under Canada’s constitution and land claims settlement acts, their hunting rights are respected. Generally, Aboriginal hunters can harvest unlimited numbers of caribou unless there is a conservation issue.
A second category of hunters is resident hunters. The trend is to fewer resident hunters in the Northwest Territories (Government of the Northwest Territories and NWT Biodiversity Team, 2010) as a result of restrictions following the decline of herds. The third category of hunters is under the heading “commercial”, which can include harvesting for commercial meat sales and guided, outfitted hunts by non-residents. The level of commercial harvesting has varied among herds. Commercial use has been progressively reduced as the herds have declined and currently there is no commercial harvesting of any Northwest Territories barren-ground herd (Department of Environment and Natural Resources, 2006) and commercial harvesting has been sharply reduced in Nunavut (Coral Harbour, 2011; Dumond, 2011, pers. comm.).
Declines in caribou numbers have led to management measures that consider all user groups and that are tied to herd trend and herd abundance (see Figure 100, for example).
Along with increasing numbers of people on the caribou ranges, there are trends toward more exploration and resource development activity. Most attention has focused on activities associated with mineral and hydrocarbon exploration and development, and caribou behavioural responses to these human activities are quite well known (Wolfe et al., 2000a; Stankowich, 2008). Impacts from power lines and shipping have received less attention. The lack of information of impacts from railway development became evident in the Mary River Project (Baffinland) Environmental Impact Assessment (Baffinland Iron Mines Corporation, 2012), leading to conditions of approval that included adaptive management measures to be enacted if unanticipated negative effects on caribou are observed.
As the rate of human activities increases, our lack of understanding about the cumulative effects on caribou at the individual and the herd level becomes a more worrisome gap (Cameron et al., 2005). Limited progress has been made in measuring and managing these cumulative effects (Festa-Bianchet et al., 2011; Gunn et al., 2011a). Examples of development activities and concerns for caribou populations are described below.
Mineral exploration activities boomed in the 1990s on the range of the Bathurst Herd and over the period 2003–2008 for the Beverly Herd’s calving and summer ranges. The timing of the herd’s decline coincided with an increase in mining exploration and development. While there had been mining exploration for decades, the discovery of diamonds in 1991 triggered a surge in exploration with an increase in camps, aircraft, and helicopters on the tundra ranges of the Bathurst herd (pre-calving to fall ranges). At the peak, 118,124 km2 of new claims had been staked in 1993 in the NWT (G. Bouchard, Natural Resources Canada., pers. comm.). Since 1996, four diamond mines have been built and are operational on the Bathurst Herd's summer range. Three of the diamond mines are large open-pit mines and their activities have reduced the occurrence of caribou in their vicinity over a greater distance than expected (Johnson et al., 2005).
Elsewhere on caribou ranges, exploration and mining of uranium has been the greatest concern in the past among communities that harvest Beverly caribou. Uranium exploration and development has occurred for decades on the Beverly winter range in northern Saskatchewan. Mineral exploration has been increasing over the past 10 years on the ranges of both the Beverly and Qamanirjuaq herds in the Northwest Territories and Nunavut (BQCMB, 2010). As of May 2010, there were many active prospecting permits, mineral claims, and mineral leases on the Beverly and Qamanirjuaq traditional calving ground (BQCMB, 2010).
The possibility of opening the Arctic National Wildlife Refuge to oil and gas development (a proposal that has been under consideration in the United States since the late 1970s) has the potential to constitute a major threat to the Porcupine Caribou Herd. The "1002" area on the Alaskan coastal plain contains both high potential for hydrocarbon deposits and contains the most sensitive habitat for the herd during the calving and post-calving periods (Griffith et al., 2002).
Roads, mineral exploration, mines, and hydrocarbon production are land use activities whose effects on caribou may accumulate over time. Although there are concerns about those cumulative effects, progress on measuring them and managing them is barely underway. In particular, roads can increase human access and act as partial barriers to caribou movement. Table 17 indicates main caribou issues related to three Nunavut development prospects.
|Mary River Iron Mine||Baffin Island, eastern Nunavut||North Baffin caribou: low in 70-year cycle; railway effects on movement|
|Kiggavik Uranium Mine||Baker Lake, central Nunavut||Multiple herds: new road access through summer habitat; increased harvest pressure|
|Izok Corridor Zinc Mine||Western mainland Nunavut||Bathurst caribou: road at edge of traditional calving ground |
Dolphin and Union caribou: road in winter range
Source: Introduction by M. Setterington, in Eamer et al. (2013)
It is uncertain as to how longer-distance and local sources of contaminants and pollutants influence the vulnerability of caribou. While levels are monitored and levels are mostly low, synergistic effects are uncertain and how contaminants may interact with health and condition is unknown. The Canadian Northern Contaminants Program has been active in monitoring Persistent Organic Pollutants (POPs), heavy metals, and radionucleotides for the last three decades. The following account is derived from the NCP summary report (Northern Contaminants Program, 2003), except where noted. Some 15 different caribou herds across Nunavut, the Northwest Territories, and the Yukon were monitored during the 1990s through two large monitoring programs; additional monitoring has been undertaken since then for some herds (Gamberg, 2009). Assessments of risk to human health from contaminants show that caribou is a safe and nutritious food choice across northern Canada (Donaldson et al., 2010).
Persistent organic pollutants such as DDT, PCBs, dioxins, and furans were found at only very low levels in caribou (often too low to be detected at all) and are not of concern for either caribou or human health (see also Gamberg et al., 2005a).
Some heavy metals, however, are found at elevated levels in caribou, though not to the same extent as in some marine mammals. There are wide variations in the levels of metals from herd to herd, probably due to the variation of levels in the underlying geology. Cadmium levels tend to be higher in the kidneys and livers of the Beverly caribou in the Northwest Territories and Nunavut, compared to the levels in other herds. Natural sources of cadmium in the underlying rocks in the area are likely responsible. This cadmium accumulates in lichen which is then eaten by the caribou. Mercury levels show no clear pattern (Gamberg et al., 2005a), with the highest levels found in the Beverly Herd and in Meta Incognita Peninsula caribou (part of the South Baffin population). In the central and northeastern parts of northern Canada, levels of mercury in caribou follow the same geographic pattern as levels found in sediments. Scientists consider that much of this mercury has been transported from human-made sources in other parts of the world. An exception is mercury in caribou from the Yukon, where local geology may be more important (Northern Contaminants Program, 2003).
The importance of trends in forest fires to caribou is related to winter forage, dominated by fruticose lichens which are most abundant in older successional forests. The effect of forest fires on caribou abundance has been controversial but most studies do document that caribou tend to avoid recently burned areas (Joly et al., 2003).
Climate changes will affect the movements and distribution of the herd as well as the ability of hunters to get on the land to be able to continue their traditional hunting practices. Table 18 provides a general treatment of observed and potential climate impacts on caribou, their ranges, and the communities that depend on them. The table was developed for the Porcupine Caribou Herd and as such some of the impacts described (for example, changes to calving grounds on coastal plains) do not apply to all herds.
|Climate change condition||Impact on habitat||Impact on movement||Impact on body condition||Impact on productivity||Management implications|
|Earlier snowmelt on coastal plain|
|Warmer, drier summer|
|Warmer, wetter fall||-||-|
|Warmer, wetter winters|
Overall effect: In very general terms: the calving range improves but with movement and reliance on more northern portions of the calving range; animals leave calving range earlier; cows and calves suffer reduced summer and fall body reserves due to increase in oestrid fly harassment; mosquito harassment may be reduced if summers drier; more frequent icing in fall, winter, and spring ranges, which depend on the location of these ranges; may have moderate to severe implications to body condition and survival.
Source: Gunn et al. (2011c), update of Chapter 10 of the Arctic Climate Impact Assessment (2005) by the author (Russell)
It is difficult to attribute a single event such as an icing storm as being within the “normal” range or as an indication of a warming climate. Examples of this type of event occurred in the fall of 2003, when coastal areas from Alaska to Kugluktuk, Nunavut, experienced icing conditions that forced caribou to move in search of accessible forage. Ice on the land formed a barrier between the caribou and their food (Nagy, 2007).
Another factor that adds to the complexity of predicting impacts of climate change is that all herds have evolved and adapted to a unique suite of environmental factors within their ranges --some herds cope with winter ranges characterized by deep, persistent snow, while others enjoy mild winter conditions; some herds occupy excellent summer ranges with an abundance of fresh green vegetation; others have to replenish fat and protein reserves depleted over winter with vegetation that is limited by a brief, intense summer growing season. Changes that result in more severe winter range conditions, for example, would have a different effect on different herds--even neighbouring herds. For example, under a warmer climate, the annual range for the Leaf River Herd may increase while that of the George River Herd may contract (Sharma et al., 2009).
Further, at the population level, some herds have exhibited a high rate of increase, over 15% annually, while others have increased at rates of less than 5% annually, primarily reflecting higher adult female mortality rates (Figure 77 ). Environmental changes that result in an increase in adult female mortality would have a greater impact on herds that demonstrate a low rate of increase.
Within the range of the Porcupine Herd, for example, the trends of climate change are marked. Spring, in particular, has warmed over the last three decades. During late spring, after calving, this has resulted in early snowmelt and more food available for nursing mothers. As a consequence, early calf survival has improved (Griffith et al., 2002). In early spring, however, when the herd is on migration, warmer weather has resulted in more freeze-thaw cycles as temperatures get above freezing during the day and below freezing at night. Specifically, the number of days where the temperature has risen above zero during spring migration doubled during the population decrease phase (1989–2001) compared to the population increase phase (1975–1988) (Griffith et al., 2002). The greater difficulty in traveling and feeding through ice crusts would result in higher energetic costs, and moving onto wind-blown ridges during migration would result in potential increased mortality from wolves, as wolves are at an advantage in shallow snow (Griffith et al., 2002).
It is clear from this example that specific information on how changes in temperature and precipitation alter environmental conditions for caribou throughout during their annual cycle is needed to understand how each herd is adapted to its range and to be able to predict impacts of climate change on caribou abundance and distribution. Work is underway through the CircumArctic Rangifer Monitoring and Assessment Network (CARMA) to produce datasets of relevant climate metrics at the scale of the herd ranges (Russell et al., 2013). Climate data and other data (such as measures of biomass) from satellite monitoring are used to develop daily median values of “caribou-relevant” climate variables for defined regions within the range of each herd, for each of five seasons (calving, summer, fall, winter, and spring). Variables, aside from temperature and precipitation, include: accumulated growing degree days above 0°C and above 5°C, leaf area index and vegetation greenness fraction; measures related to mosquito and oestrid (warble and bot fly) activity; snow depth, density, cover, and melt rate; and, accumulated number of days with freeze/thaw events.
Cumulative energy costs to caribou
Cumulative effects assessment (CEA) is part of environmental impact assessment (EIA) and focuses on the combined impact of the individual effects of industrial development projects. However, the context is also set by multiple stresses from all human activities additional to natural environmental effects at both the individual and population scales. To scale up the individual animal’s behavioural responses to the population requires being able to estimate the energy costs to the individual and whether those costs will affect its reproduction and survival. Estimating the costs of a behavioural response is not straightforward: as well as the energy costs of movement and interruption in foraging time, there may also be an effect on diet (energy protein intake) if a displacement puts the individual in a different habitat. Understanding and integrating the relationships between behaviour, habitat selection, energy, and protein intake relative to reproduction and survival is data intensive and requires interdisciplinary collaboration, as the understanding is based on ecology, nutritional ecology, and modeling. Russell (2012) describes modelling to integrate environmental influences to project caribou vulnerability to developments.
Addressing climate change
A major human stressor for the Arctic Ecozone+ is climate change as demonstrated throughout this report. This stressor cannot be overcome through stewardship and conservation measures within the ecozone+, although measures can be taken to reduce greenhouse gas emissions within the ecozone+ and to adapt to coming changes. This is a global problem, and the most significant measure of actions taken to conserve Arctic ecosystems is the rate of reduction of releases of greenhouse gases worldwide. Globally, however, emissions continue to grow, though the pace of growth appears to be slowing (Environment Canada, 2013a).
Historical and projected trends in greenhouse gas emissions for Canada are graphed in Figure 101, showing that additional greenhouse gas reduction measures will be needed to achieve Canada’s 2020 target for emissions. The decrease in emissions around 2009 due to the global recession reflects a global trend for that time period and shows that actual emissions are closely tied to economic factors (Environment Canada, 2013a).
Projections with current measures include the compliance contribution of the Land Use, Land-Use Change and Forestry (LULUCF) sector for each year; actual emission trends are projected to be 28 Mt higher than the projections shown in 2020.
Source: Environment Canada (2013b; 2013c)
Long description for Figure 101
This line graph shows the following information:
|Year||Total greenhouse gas emissions (megatonnes of carbon dioxide equivalent)|
The global scale of climate change does not mean that regional and local stewardship are not important in the Arctic. As plant and animal ranges shift and ecosystem function and processes are affected by climate change, it becomes even more important to focus on conservation of important lands and waters such as breeding areas for migratory birds and caribou, building into conservation plans the flexibility needed to adapt to the changing landscape. An important stewardship tool is the creation and management of protected areas. In addition, pressure is increasing in some areas and likely to increase further for development and expansion of human use of the sea, coastal zone, and land. Protected areas are an important management tool for conserving the ecosystems needed to support Arctic biodiversity and the ecosystem services they provide. Arctic ecosystems operate on huge scales, often crossing the terrestrial/marine divide. No one protected area is sufficiently large to maintain the integrity of ecosystems and stewardship of the entire land base is crucial, taking into consideration ecosystem elements and processes that routinely move in and out of protected areas, such as migratory birds and caribou.
Analyses for this section were conducted for ESTR based on data provided by federal, territorial, and provincial jurisdictions and have not been updated (see figures for references). Location of protected areas is shown for all of Canada in Figure 102 and the growth of protected areas in the Arctic Ecozone+ is shown in Figure 103. In addition to federal parks and wildlife sanctuaries, the ecozone+ has (generally smaller) territorial and Quebec parks and reserves, often protecting cultural heritage sites and areas such as estuaries with particularly high value for biodiversity.
While the overall percentage of land protected is 11.3% for the ecozone+, this is not evenly distributed. The Arctic Archipelago has (as of May 2009) the highest proportion of protected area at 24.0% in 10 protected areas, followed by the Southern Arctic with 15.9% protected through 21 protected areas. The Northern Arctic has 6.7% of its land protected through 22 protected areas.
Much of the growth of protected areas in the Arctic has been related to the settlement of land claims. Many Arctic parks are managed through regimes created through land claim settlements, protecting the harvesting rights of Inuit within the parks, while working to conserve species and areas of special cultural and ecological significance.
Data provided by federal, provincial, and territorial jurisdictions, updated to May 2009. Only legally protected areas are included. All areas shown are protected as IUCN (International Union for Conservation of Nature) categories I–III. IUCN categories are based on primary management objectives (see text for more information). Selected protected areas are shown on the figure, along with their dates of establishment.
Source: Source: Environment Canada (2009a) using data from the Conservation Areas Reporting and Tracking System (CARTS), v.2009.05, 2009 (CCEA, 2009); data provided by federal, provincial, and territorial jurisdictions
Long description for Figure 103
This bar graph shows the following information:
|Year||Northern Arctic||Southern Arctic||Arctic Cordillera|
National Wildlife Areas in Nunavut
National Wildlife Areas protect nationally significant habitat for migratory birds, support species or ecosystems at risk, or protect rare or unusual habitat. Critical natural features are conserved and activities considered harmful to species or habitats are prohibited. Three new National Wildlife Areas were created in Nunavut in June 2010 to protect critical habitat for Arctic seabirds, bowhead whales, and other species. They will be co-managed by local and federal governments, and were chosen based on advocacy and involvement from the communities of Qikiqtarjuak and Clyde River.
Akpait National Wildlife Area (774 km2) is an important area for migratory birds. It provides breeding habitat for one of Canada’s largest thick-billed murre (Uria lomvia) colonies, black-legged kittiwakes (Rissa tridactyla), glaucous gulls (Larus hyperboreus), and black guillemots. It is also home to polar bears, walruses, and several species of seals.
Qaqulluit National Wildlife Area (398 km2) is home to Canada’s largest colony of northern fulmars (Fulmarus glacialis), representing an estimated 22% of the total Canadian population. Marine animals, including walrus and ringed seals, also use the waters of this National Wildlife Area.
Ninginganiq National Wildlife Area (Isabella Bay) (336 km2) protects critical summer habitat for the Eastern Arctic population of bowhead whales, a Threatened species.
Source: Latour et al. (2008)
Beginning in the 1970s and accelerating since the close of the Cold War, “the North” is emerging as a new circumpolar geopolitical entity (Heininen and Nicol, 2007). Arctic countries and their indigenous peoples’ organizations are entering new discourses concerning the scale and nature of circumpolar regionalism and critical new environmental, human security, and economic development challenges. These northern discourses stress regional co-operation, human security, and sustainable development (Heininen and Nicol, 2007).
Several features set Arctic governance apart from the rest of Canada:
- There is only one Aboriginal language group and culture (with important regional differences) across the entire region--a quarter of Canada--and its members remain numerically dominant.
- The Inuit have forged alliances with related cultures/linguistic groups around the circumpolar region, for example, through the Inuit Circumpolar Council (ICC), an international non-government organization representing approximately 150,000 Inuit of Alaska, Canada, Greenland, and Chukotka in Russia.
- The ICC and other Arctic indigenous organizations are Permanent Participants at the Arctic Council, an intergovernmental organization of the eight circumpolar countries that provides a forum for collaboration and oversight of many international initiatives related to ecological science and sustainable development.
- Resource management boards established pursuant to the land claims settlements have become the dominant forces in management of land and natural resources. The networks of resource management boards, councils, and local hunters’ and trappers’ associations function as “bottom-up”, co-operative decision-making systems that are mandated to make use of science and Aboriginal traditional knowledge in their decision-making.
Wildlife in the Arctic Ecozone+ portion of the Yukon, NWT, and Nunavut is co-managed by governments and Inuvialuit pursuant to the Inuvialuit Final Agreement Indian Affairs and Northern Development Canada, 1984) and by governments and Inuit pursuant to the Nunavut Land Claims Agreement (Indian Affairs and Northern Development Canada, 1993). While the terms of these agreements differ, in general they recognize the Aboriginal rights of Inuit and Inuvialuit to manage the harvest of wildlife, subject only to the need for conservation and public safety. The primary management bodies are two Wildlife Management Advisory Councils (WMAC) for the Inuvialuit Settlement Region in the Northwest Territories and the Yukon North Slope, and the Nunavut Wildlife Management Board (NWMB) for Nunavut. Although the mandates of these boards differ, both organizations bring together Inuit and government representatives.
In Nunavut, the Nunavut Department of Environment and Environment Canada (Canadian Wildlife Service) both appoint members to sit on the NWMB, along with Inuit appointed by their regional organizations. NWMB membership also includes other federal departments and Nunavut Tunngavik Incorporated. However, as an institution of public government, all NWMB members represent the public interest and not necessarily the interests or opinions of their appointing bodies. These boards are supported by hunters’ and trappers’ associations from each community and other community committees.
A comprehensive review of how these new governance structures influence environmental status and trends, and knowledge thereof, is beyond the scope of this report. A case study will, however, serve as an example of evolving Arctic environmental governance.
Case study on environmental governance: Kitikmeot and the West Kitikmeot/Slave Study
In addition to the Nunavut Wildlife Management Board, three regional Inuit associations were established under the Nunavut Land Claims Agreement to manage Inuit-owned lands in the regions. The Kitikmeot Inuit Association (KIA) has a mandate “To represent the interests of Kitikmeot Inuit by protecting and promoting our social, cultural, political, environmental and economic well-being” (Kitikmeot Inuit Association, 2013). Under the KIA, the Kitikmeot Corporation and the Kitikmeot Economic Development Commission are delegated the responsibility of promoting economic development in the region. The KIA is responsible for administering surface lands in the Kitikmeot region. There is also a Kitikmeot Heritage Society that promotes Kitikmeot culture and history.
Even before its formal establishment, Kitikmeot communities were cooperating to organize the generation of information needed to inform environmental management decisions. For example, more than 20 years ago they completed a major study that quantified the country foods that they harvested from the land (Gunn et al., 1986).
The discovery of diamonds in 1991 at Lac de Gras initiated one of the largest mineral staking rushes in the history of the world (Environment and Natural Resources, 2012b). Out of concern about the impacts of rapid development on the environment, Kitikmeot communities formed a partnership with other Aboriginal (Dene) organizations, environmental organizations, government, and industry. The West Kitikmeot/Slave Study (WKSS) Society was founded in 1996 by nine partner organizations. The purpose of the Society was to collect environmental and socioeconomic information to enable better-informed planning and contribute to a baseline for assessing and mitigating cumulative effects of development (Environment and Natural Resources, 2012b). Community capacity and traditional knowledge were priorities.
The WKSS region encompassed the western part of Kitikmeot (Arctic Ecozone+, in Nunavut) plus the area between Great Slave Lake and Contwoyto Lake (Taiga Shield Ecozone+, in the NWT), with the treeline dividing the area approximately in half (Figure 104) (West Kitikmeot Slave Study Society, 2001). The calving, migrating, and wintering ranges of the Bathurst Caribou Herd were largely within the WKSS region. The Inuit communities of Kugluktuk, Bathurst Inlet, Umingmaktok, and Cambridge Bay, and the Dene communities of Gameti, Wha Ti, Rae Edzo, Wekweti, Yellowknife, Dettah, and Lutselk'e were included in the study area. The partners developed an initial five-year research program to provide information necessary to examine the long- and short-term effects of development in the WKSS area. The initial WKSS research program ended in 2001, but the society continued to fund projects until 2009, when it was dissolved (Environment and Natural Resources, 2012b).
The WKSS sponsored a number of wildlife-related studies, including studies of traditional ecological knowledge, specific habitats (such as the use of eskers by wildlife), place names as indicators of biogeographical knowledge, use of native plants for mine reclamation and revegetation, water quality, and ecology of important wildlife species, including caribou, moose, muskoxen, grizzly bears, wolves, and wolverines. Results of scientific studies were often presented at workshops attended by people from villages throughout the region so that results could be integrated with traditional ecological knowledge and priorities for action could be developed (Dumond et al., 2009). Conclusions were fed into the management process, to inform decisions, for example, on harvest quotas and on monitoring and research needs.
Collaborative programs and projects such as those undertaken through the WKSS can be effective because they provide a structure for making effective use of both scientific and traditional ecological knowledge (Gunn et al., 1988; Freeman, 1992; Stevenson, 1996; Zamparo, 1996; Berkes, 1998; Duerden and Kuhn, 1998; Usher, 2000). A side benefit is the promotion of mutual understanding, mutual respect, and mutual learning among local communities, regional management authorities, researchers, and other stakeholders, including industry.
Appendix 1: Descriptions of surficial materials
Alluvial Deposits: stratified silt, sand, clay, and gravel; floodplain, delta, and fan deposits; in places overlies and includes glaciofluvial deposits
Alpine Complexes: rock, colluvium, and till; rock and Quaternary deposits complex in an area, characterized by alpine and glacial landforms
Coarse grained (Glacio)Lacustrine: sand, silt, and gravel; deposited as deltas, sheet sands, and lag deposits
Coarse grained (Glacio)Marine: sand and gravel; deposited as sheet sands, deltas, and extensive flights of beaches
Colluvial Blocks: Blocks, and rubble with sand and silt; derived from crystalline bedrock, medium grade metamorphic substrate, and cemented sandstone
Colluvial Fines: silt, clay, and fine sand; derived from substrate weakly consolidated shale and siltstone substrate
Colluvial Rubble: rubble and silt; derived from carbonate and consolidated fine clastic sedimentary rock substrate
Colluvial Sand: Sand and gravel; derived from poorly lithified sandstone and conglomerate substrate
Eolian Deposits: sand and minor silt; dunes, blowouts, and undulating plains; in most places overlies deltaic sediments, coarse lacustrine sediments, or glaciofluvial deposits
Fine grained (Glacio)Lacustrine: silt, and clay, locally containing stones; deposited as quiet water sediments
Fine grained (Glacio)Marine: dominantly silt and clay, locally containing stones; deposited as quiet water sediments
Glaciers: ice and minor morainal debris
Glaciofluvial Complex: sand and gravel and locally diamicton; undifferentiated ice contact stratified drift, and outwash; locally includes till and rock
Glaciofluvial Plain: sand and gravel; deposited as outwash sheets, valley trains, and terrace deposits
Lacustrine Mud: fluid silty clay and clayey silt; deposited as quiet water sediments'
Lacustrine Sand: sand and locally gravel; deposited as sheet sands, lags, and beaches
Lag (Glacio)Marine: sand, gravel, and pockets of finer sediment; thin to discontinuous sediment veneer and residual lag developed during marine submergence; includes areas of washed till and rock
Marine Mud: fluid silty clay and clayey silt; deposited as quiet water sediments
Marine Sand: sand and locally gravel; deposited as sheet sands, lags, and beaches
Organic Deposits: peat, muck and minor inorganic sediments; large bog, fen, and swamp areas where organic fill masks underlying surficial materials; geneally >2 m thick
Quaternary Volcanics: consolidated lava, breccia and tephra; dominantly basaltic and andesitic in composition; includes flows, volcanic piles, and cinder cones
Till Blanket: thick and continuous till
Till Veneer: thin and discontinuous till; may include extensive areas of rock outcrop
Undivided: rock with minor Quaternary deposits
Source: Geological Survey of Canada (1994)
Appendix 2: Detailed land cover classes
|Class||Land Cover Type||Abbrev.||ESTR Land Cover Class||Southern Arctic|
|1||Tussock graminoid tundra||Tgt||Barren||107,242||18.62%||65,659||5.0%||1,461||0.7%|
|3||Moist to dry non-tussock graminoid/dwarf shrub tundra||Tntg/ds||polar tundra||51,247||8.9%||81236||6.2%||3,660||1.7%|
|4||Dry graminoid prostrate dwarf shrub tundra||Tg/pds||polar tundra||9,821||1.7%||18,959||1.4%||146||0.1%|
|5||Low shrub||Sl||polar tundra||87,461||15.2%||4,315||0.3%||203||0.1%|
|6||Tall shrub||St||polar tundra||24,699||4.3%||167||0.0%||32||0.0%|
|7||Prostrate dwarf shrub||Spd||polar tundra||109,871||19.1%||110,034||8.4%||4,785||2.2%|
|8||Sparsely vegetated bedrock||Rsv||Barren||52,243||9.1%||323,284||24.5%||18,575||8.7%|
|9||Sparsely vegetated till colluvium||TCsv||Barren||11,539||2.0%||177,724||13.5%||14,853||7.0%|
|10||Bare soil with cryptogam crust - frost boils||TScc||Barren||18,823||3.3%||155,281||11.8%||2,935||1.4%|
|13||Snow and/or Ice||S/I||snow/ice/glacier||303||0.1%||33,633||2.6%||115,736||54.4%|
|Class||Land Cover Type||Abbrev.||ESTR Land Cover Class||Total Arctic Ecozone+|
|Total Arctic Ecozone+|
|1||Tussock graminoid tundra||Tgt||Barren||174,362||8.3%|
|3||Moist to dry non-tussock graminoid/dwarf shrub tundra||Tntg/ds||polar tundra||136,143||6.5%|
|4||Dry graminoid prostrate dwarf shrub tundra||Tg/pds||polar tundra||28,926||1.4%|
|5||Low shrub||Sl||polar tundra||91,979||4.4%|
|6||Tall shrub||St||polar tundra||24,898||1.2%|
|7||Prostrate dwarf shrub||Spd||polar tundra||224,690||10.7%|
|8||Sparsely vegetated bedrock||Rsv||Barren||394,102||18.7%|
|9||Sparsely vegetated till colluvium||TCsv||Barren||204,116||9.7%|
|10||Bare soil with cryptogam crust - frost boils||TScc||Barren||177,039||8.4%|
|13||Snow and/or Ice||S/I||snow/ice/glacier||149,672||7.1%|
The Southern Arctic is dominated by graminoid and shrub tundra; the Northern Arctic is dominated by sparsely vegetated bedrock, till colluvium, and barren lands; the Arctic Cordillera is dominated by snow and ice and mountains.
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