Common Murre Uria aalge Scientific name definitions
Version: 2.0 — Published August 6, 2021
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Demography and Populations
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Incompletely known, with aspects pieced together from different locations in different years, and from populations exhibiting different growth trajectories. Except for fecundity, demographic parameters very poorly known. Even fecundity, however, not really well known, as chicks reach independence at sea, achieving the ability to fly, i.e., fledge (the usual point at which avian fecundity is measured), a long time after breeding-ledge departure. Most data on demography from California, Wales, and Scotland, where long histories of marked individuals available; these populations have a similar annual cycle (minimal migration, colony visitation much of year; see Figure 1).
Measures of Breeding Activity
Age at First Breeding
Age of first colony return varies from 2–7 yr. At Isle of Canna, Scotland, 1973–1982 (expanding population; n = 100), 10% had returned by age 3 yr, 35% by 4 yr, 61% by 5 yr, and 88% by 6 yr (154); and at Skomer Island, Wales, 1974–1978 (decreasing population), 6% by 2 yr, 27% by 3 yr, 56% by 5 yr, 75% by 6 yr, and 100% by 7 yr (443, 363, 312). At Farallon Islands, California, 1986–2001 (increasing population), 0.5% returned at 2 yr, 30% by 3 yr, 55% by 5 yr, most by 6–7 yr, with all recruited by 9 yr (444).
Proportion of birds breeding, by age, is affected by food availability, with the average age being youngest when availability high, either because of favorable environmental conditions or reduced competition owing to small colony size (445). No murres <3 yr old reported to breed. Youngest first breeders: Isle of Canna and Isle of May, Scotland, 3 yr (154, 446), though upon further inspection of the data youngest breeders 4 yr at Isle of May (447), and in a separate study from Isle of May, 5 yr (Harris 1992); Skomer Island, Wales, 4 yr (443). Range in age at first breeding 4–9 yr at Isle of May, 4–15 yr at Isle of Canna (446, 447). Average age of first breeding: Isle of Canna, expanding colony, 1973–1982, 5.3 yr ± 1.2 SD (n = 26; 154). Median and modal age of first breeding 6 yr at Isle of May (1981–1992; n = 42) and 7 yr at Isle of Canna (1974–1992; n = 321); median age of successful breeding 7.67 yr at Isle of May (1983–2015, n = 55; 447). No difference between sexes (446). At Farallon Islands, 1986–2001, youngest breeders 4 yr old, reaching 100% by 7 yr (444). Lacking explanation, murres recruited to an egg-laying site that had previously been active delay first laying for at least a year longer than those occupying a new site (447).
Intervals between Breeding
"Annual," but in 3,016 bird-years among banded adults at Isle of May, Scotland, 1983–1992, nonbreeding occurred 222 (7.4%) times, about 5–10%/yr; in 21 instances, individuals were absent entire season; 6% of population accounted for 47% of nonbreeding. Nonbreeders tended to have lost a mate, did not have a breeding site, and exhibited lower breeding success during previous season (421, 448). At Farallon Islands, California, nonbreeders contributed 4.5% to total bird-years, and once recruited rarely skipped (~4% in any given year; 449, 444).
Clutch Size and Number of Clutches per Season
Annual and Lifetime Reproductive Success
Lifetime success not known, again related to problems with measuring fecundity in this species (see above). Annual success varies between and within colonies and years (see below), expressed as proportion of pairs that laid and produced a chick who departed the ledge successfully (i.e., chicks/breeding site). Breeding/departure success increases with age from 3 to 8 yr at Farallon Islands, California, and is slighter lower among oldest breeders (15+ yr; 444; see also for Isle of May, 447). In a study spanning 1983–2015 at Isle of May, among 34 birds that entered and left the breeding population, considerable variation existed in number of fledglings produced: 9 individuals (26%) failed but 1 male fledged 19 chicks from 21 attempts; 3.09 was median number of chicks fledged (447).
Proportion of pairs successful, as follows:
U. a. inornata: Chukchi Sea, Bluff (1987–1991; declining population), 0.70 (450). Bering Sea: St. Paul (1976–2019; slightly decreasing), 0.47 (range 0.0–0.7); St. George (1976–2019; decreasing), 0.48 (range 0.0–0.9); Round Island (1999–2019; increasing then decreasing), 0.16 (range 0.0–0.55; 428, 451). Gulf of Alaska: Semidi Islands (1979–1981), 0.6 (452); Barren Island (1993–1998), 0.7 (range 0.47–0.81; 378); St. Lazaria (1994–2019; stable), 0.47 (0.0–0.85); Chowiet Island (1979–2019; stable), 0.50 (range 0.0–0.75); Aiktak Island (1995–2019; greatly decreasing), 0.21 (0.0–0.8); Buldir Island (1995–2018; steeply increasing), 0.0–0.95; 451). British Columbia, Triangle Island (2003–2007; decreasing), 0.35 (range 0.26–0.45; 37).
U. a. californica: Farallon Islands (1972–2017; rapidly increasing), 0.72 (range 0.1–0.9; 27, 453); Castle Rock (2007–2017; increasing), 0.61 (range <0.05–0.9; 232); Yaquina Head, Oregon (2007–2017, decreasing), initially ranged 0.54–0.77, but between 2011–2017 with arrival of eagles ranged 0.00–0.27 (85); Tatoosh Island, Washington (1992–1999; increasing), 0.35 (range 0.09–0.71; 159).
U. a. aalge: Gannet Islands, Labrador: 1981–1983, average 0.80–0.85 (range 0.79–0.86 [n = 935]; 327); 1996–1997, 0.83–0.85 (n = 245; 235). At Gull Island, Witless Bay, Newfoundland (1977–1978), range 0.35–0.53 (n = 500; 359), 2007–2015, range 0.38–0.80, overall average 0.61, average for two low capelin years 0.41 (n = 310; AES). At Great Island, Witless Bay (1983–1984), range 0.57–0.63 (n = 350; DNN), and from 1998–2002, average 0.77 (range of 0.53–0.96, n = 146; AES). At Cape St. Mary's, Newfoundland (1980–1984), 0.76 (range 0.70–0.87 [n not known]; 379). Isle of May, Scotland: 1981–1985, unbridled pairs, 0.77 (range 0.76–0.82 [n = 2,513]), bridled pairs, 0.83 (range 0.69–0.94 [n = 213]; 17).
Factors Affecting Breeding Success
Included are variation in food availability, weather, predation, habitat, timing of laying, breeding population size, and age, many of which have confounding effects (see 27, 323, 419, 342, 37, 444, 232). In eastern North Pacific, including California Current, Gulf of Alaska, and Bering Sea, for locations having long time series, success varies with ocean climate, which affects food availability (454). Complete or near-complete breeding failures occurred 6 of 45 yr at Farallon Islands between 1972–2017; 2 of 11 yr at Castle Rock (2007–2017); 2 of 25 yr at Chowiet Island (1978–2019); 3 of 36 yr at St. George Island (1978–2019); and 4 of 33 yr at St. Paul Island (1976–2019). Conversely, during these time periods success was very high (≥~0.7 chick/pair) for 28 yr at Farallon Islands, 4 yr at Chowiet, 8 yr at St. George, and 7 yr at St. Paul (453, 232, 451). Breeding success was low at all these sites during strong El Niño, when food webs not well developed and appropriate prey unavailable, and high during La Niña, when food is more abundant (455, 27, 454, 101, 453, 232, 451). Extreme ocean anomalies, as in 1978 and warm event of 2015–2017, also brings breeding failure to most eastern Pacific breeding populations due to alteration of the food web (454, 456). Windy weather reduces chick-provisioning rates (368), and persistent bad weather can reduce chick survival and ultimate breeding success. At Isle of May, 1983–2005, breeding success varied between 75–85% for most of the period but then decreased to vary between 50–70% after 1998, though population continued to increase (see below: 447) without explanation, though a decrease was expected in the future (158). Indeed, after peaking in 2003, population decreased about one third by 2015 (M. P. Harris and S. Wanless, personal communication). Numbers of murres in North Sea and Barents seas have experienced a decrease owing to commercial fish extraction (see Conservation and Management: Effects of Human Activity).
Pairs on narrow ledges exhibit lower success than those breeding on wider ledges, presumably because of greater egg and chick loss due to falling (327, 192). Success higher at specific egg-laying sites regardless of what individuals are present; average site success related to physical characteristics (low slope, wall, no flooding), bird density (pairs within 3 body lengths), position on cliff (not near edge, nor top), and suitability of habitat for overwintering ticks (Ixodes spp.; 316). Falcons and eagles take more eggs, chicks, and adults from exposed ledges near cliff tops than protected ledges (340, 341); same true in case of human predators (16). Low food availability also decreases parental attentiveness, thereby increasing predator success (452). Predator success increases inversely to breeding density/group size and synchrony (308, 457, 340, 371); density and synchrony are reduced during El Niño, which makes egg and chick predators more effective (321); other environmental conditions, such as late ice breakup at high latitude breeding sites, adversely affects food supply and, in turn, murres and predators (e.g., 29).
Late egg layers, which include a high proportion of young or first-time breeders, exhibit lower breeding success (458, 27, 457, 447, 232); among experienced breeders late layers not less successful (309). Most of temporal decline related to egg loss (458). Predators' presence early in breeding season causes temporary desertion of ledges, thereby delaying egg-laying and subsequently lowering success (192, 339, 340).
Ingestion of oil can affect reproductive cycles and lower hatchability of eggs (see Conservation and Management: Effects of Human Activity).
The increasing number of eagles (Haliaeetus spp.) have recently become a serious threat to murre populations in the eastern North Pacific, the western Atlantic, and northern Europe (37, 351, 342). As eagle populations recovered from the effects of pesticide pollution, they increasingly cause serious disturbances of surface-breeding seabirds, including Common Murre. Effects come from direct predation as well as from causing murres to flush so that gulls have access to eggs and chicks. Massive breeding failures also occur by natural (e.g, Gannet Islands, Labrador; 339) and unnatural (e.g., Alaskan islands; 346, 347, 48) introductions of foxes on murre colonies. Colony abandonment, possibly involving hundreds of thousands of murres, has been observed in the Pacific coast colonies, as well as in northwestern Atlantic and northern Europe (351, 342). In northern Norway, predation from a growing population of White-tailed Eagles (Haliaeetus albicilla) has exacerbated the decline in the populations of Common Murre. Birds in several colonies (e.g., Røst, Bleiksøya, Hjelmsøya) have been forced away from the open ledges to breeding under cover, for example in large cracks or stone screes, to avoid predation (459). At present, birds breeding in such habitats are much more productive than those on exposed cliff ledges.
Life Span and Survivorship
Oldest known Common Murre western Atlantic: 34 yr 8 mo (banded as chick on Gannet Islands, Labrador, 6 August 1981, and shot by murre hunter on 11 February 2016 in eastern Newfoundland waters; CWS-ECCC National Banding Office; DNN, unpublished data). Oldest known Common Murre was determined to be 44 yr of age, at Stora Karlsö, Sweden (T.R. Birkhead, personal communication).
U. a. californica: Farallon Islands, 1986–1991 (increasing), 93.9%; survival differed within a subcolony that experienced variable falcon predation, with 58% survival between 1989–1990 with high falcon predation, and 92.6% survival in years with lower falcon predation (1988–1989, 1990–1991; 449). Survival (recapture probability) among breeders increases with age, from ~80% among 3–6 yr olds, to near 100% among 12–14 yr olds; among non-breeding adults, survival increases from 30–55% among 3–6 yr olds, to near 100% among birds ≥7 yr (444).
U. a. aalge: Gannet Islands, 1981–1983 (declining), 94.5% annual (460); Great Island, 1997–2000 (expanding), 97% (AES).
Elsewhere: British Isles (band recoveries from dead birds), 1960–1972, 87.9% uncorrected, 93.7% corrected (461, 462); studies of banded birds at breeding sites: Skomer Island, Wales, 1963–1980 (declining population), 91.5% (443) but, 1980–2011 (increasing), 93.0% (463); Isle of May, Scotland, 1973–1981, 96.9% (expanding population), compared to 1982–1993, 88.1% (stable), and 1982–1993, 94.9% (expanding; 464, 421); Sweden, Stora Karlsö (declining), 87.1% (Hedgren 1980); Baltic Sea, 1962–1989 (increasing), 87–90%, 1989–1997 (perhaps decreasing), 78% (465); Hornøya, Northern Norway, 1988–2019 (expanding), 97.4% (466).
Colony departure to recruitment: U. a. aalge – Gannet Islands, Labrador (declining population), 29.5%; Green Island and Funk Island, Newfoundland (expanding), 41.1% and 37.1%, respectively (443); Skomer Island, Wales (declining), 24.6% (443), but more recently when population increasing 43.0% (to age 3 yr; 463); Stora Karlsö (declining), 36.0% (309); Baltic Sea (increasing), 44–68% (465). U. a. californica – Farallon Islands (increasing), 30–55% (444).
Survival by Sex
Disease and Body Parasites
Avian pox known (467); large pox growths seen on bills of some chicks in Barkley Sound, British Columbia (H. R. Carter and S. G. Sealy, unpublished data). Adenovirus observed (468); avian malaria develops among captive birds (469). Among 7 oiled birds in California, some exhibited hepatocellular dissociation and hemosiderosis, renal tubular necrosis, pancreatitis, lymphoid hyperplasia, and hemolytic anemia; causes and consequences unknown (470). Basic hematological and plasma biochemical references established among healthy individuals at Shumagin Islands, Alaska; significance of inconsistencies with other diving species unknown (271).
In California (n = 7), 1 had lung flukes, 3 had nematodes (Entracecum sp.), 3 had tapeworms in proventriculus, 1 had a peritoneal nematode, 1 had coccidial gametes and oocysts in intestinal wall, and 2 had parasitic ova in eretreal and cloacal mucosa (470). Infestation apparent in Newfoundland, including at least 13 species of 8 genera: Ornithobilharzia, Cyptocotyle, Tetrabothrius, Anonmotaenia, Eustrongylides, Contracaecum, Cosmocephalus, Anisakis; extent of burdens varied annually, with no differences related to sex or age (471).
Phthiraptera Lice. In British Columbia, birds were infested with Cummingsiella (Quadraceps) obliquus (25 of 25), Saemundssonia calva (9 of 25), and/or Austromenopon uriae (23 of 25); 62.5% infested with 2 species, 25% with 3 species, and all juveniles had at least 2 species (472). Same species, in same order of abundance, on murres (n = 29) from Pribilof Islands, Alaska (473). In Newfoundland, 23 of 28 adults and 5 of 6 chicks infested with at least one of the same 3 (above) species, 35% with 2, and 48% with 3 (474). A heavy louse infestation (722 individuals) can contribute to chick death (474).
Mites. Weak birds captured by hand in British Columbia were infested with Cymbaeremaeus sp. and Rhinonyssus sp. (472). On Pribilof Islands, Alaska, 29 of 29 had latter nasal mite, but chiggers (Neotrombicula) in just 1 bird; most numerous ectoparasites were feather mites (Alloptes sp., Calamicoptes sp.; 473).
Ticks. In Newfoundland, 15 of 28 breeding adults (53.6%) infested with Ixodes uriae (all life stages) during the breeding season, mostly from back of neck, with lower numbers on head and body and none on wings, legs, or tail; inter-year variation low (1972–1973) and no significant differences in burden between sex and body mass of host (475). Manner and extent of infestation similar elsewhere where examined in Atlantic colonies (for details, see 475). Deaths recorded in British Columbia attributed to heavy infestation by I. uriae (472); tick life cycle is 2–4 yr, and most feeding occurs during murre incubation period, not during fall and winter (476). Ixodes uriae and I. signatus found on murres (n = 29) at Pribilof Islands, Alaska (473). Prebreeding (immature) murres can be important carriers of tick-borne viruses at breeding colonies (see 477).
Nonparasitic Arthropods. Oribatid mites, Ameronothrus nidicola (n = 29) and Svalbardia paludicula at Pribilof Islands (473).
Causes of Mortality
Natural causes include predation (see Behavior: Predation), parasite infestation (see Demography and Populations: Body Parasites and Disease), and starvation due to inadequate food supplies (478, 479, 439), perhaps at times related to storms (335).
Ocean anomalies negatively affecting food availability can also lead to higher mortality, as during El Niño (27, 82; see also below, Exposure). Piatt et al. (456) reported extensive murre die-offs in eastern North Pacific (estimated as high as 1 million birds) related to less food associated with anomalous ocean conditions; more southerly regions also affected, but not as badly. They report that “The 2015–2016 Common Murre die-off in the northeast Pacific is unprecedented globally in magnitude, spatial extent and duration resulting from record-high and prolonged water temperatures that greatly reduced the abundance of forage fish. Necropsies indicated starvation.”
Timing of mortality differs by age. In waters off Oregon in 1930s, 52% of banded murres recovered in first yr of life, 76% in July–September; recoveries of other age classes scattered throughout year (35, 162). In California, a huge peak in dead chicks occurs after colony departure, July–September (14-yr average: 1.2 carcasses/km in northern and central California, but higher in isolated areas); peak among subadults and adults occurs July–October and February–April. Mortality related to poor food supply (El Niño and other warm-water periods) during winter and spring, but not summer/fall (data compromised by huge gill-net kills; 480, 481; see Conservation and Management). Number of adult carcasses/km in August (immediate postbreeding) 10 times higher during anomalous warm ocean temperature; numbers of chick carcasses high then, too, except in years when few murres breed (455, 480).
Unnatural causes of mortality also occur. The most important are oil spills and entanglement in fishing gear (480, 482, 483, 481, 82, 249); see Conservation and Management. Fishery extraction of forage fish can exacerbate negative impact of ocean climate fluctuation, e.g., El Niño (82, 83). There is a limited hunt for murres in waters off eastern Canada and Greenland, and before 1978 in waters off Denmark (484; see Effects of Human Activity).
Thousands, practically in "windrows," found dead on beaches, many emaciated, and often after storms, so-called "seabird wrecks" (Alaska: 485, 486, 456; Washington: 124, 456; Canada: 38; Britain: 487); such cases may be initiated by food shortage (see above), sometimes exacerbated by contaminant loads (488, 489). Numbers found on California Current beaches much higher during El Niño and similar ocean anomalies, indicative of food shortage but complicated by gill-net mortalities in 1979–1987; in many cases, adults emaciated (455, 480, 439, 490). Death from paralytic shellfish poisoning also occurs, possibly related to coastal pollution (491, 478, 481).
Chicks killed from falling during colony departure and from getting separated from adult; return to shore and die of starvation or predation (38, 16). Late heavy rains lead to smothering or trampling of large chicks (38). Large waves known to wash eggs and chicks from low-elevation ledges (492). Rockfalls also kill cliff-breeders: 439 adults and 2,962 chicks killed among 208,000 breeding adults (the two murres species combined) at Cape Thompson, Alaska, in 1960 (16). Intense storms at time of chick cliff-leaving can also take a toll (493). See also Conservation and Management: Effects of human Activity.
Population Spatial Metrics
Breeders normally remain <1 m from egg site; failed breeders may visit other ledges toward end of breeding season. Pre- and nonbreeders may have loafing areas at colony peripheries.
Home Range Size
Away from colony (up to 80 km), breeders concentrate where food availability is high; locations variable within and between breeding seasons (see Habitat).
Only rough estimates possible unless rigorous protocols followed, with years of most recent counts varying widely by region (Appendices 2, 3, 4, 5, 6, 7, 8, and 9); survey effort needed for management related to oil pollution, gill-netting, and population changes (494, 495, 496, 101, 109, 328; see also Conservation and Management).
Estimated world breeding population based on breeding site data is 16.6 million birds (plus 1.2 million more murres in mixed colonies, in which species ratios could not be determined, in western Arctic/sub-Arctic). The approximate regional break down in the total Common Murre breeding population is as follows: eastern Asia, 2.6 million (Appendix 2); western North America, 7.4 million (Appendix 3 and Appendix 4); eastern North America, 1.8 million (Appendix 5, Appendix 6); Greenland and Iceland, 1.5 million (Appendix 7); Eurasia, 0.5 million (Appendix 8); United Kingdom, Ireland, and Faeroe Islands, 2.8 million (Appendix 9). Other estimates without data sources include: 9 million pairs, with 4 million (range: 3.0–4.5) in Atlantic and 5 million (range: 4.0–5.5) in the Pacific, or ca. 18 million breeding birds overall (48); 12–15 million birds (8). About 70% breed at locations in low/sub-Arctic waters, between ca. 50–65o N. Given difficulties in estimation (497, 364) or distinguishing between murre species, use or nonuse of correction factors, and rapid switches or population changes in both species (see below), all totals above remarkably close. Harris et al. (164) recommended the need for annual estimation of correction factors, but this is done at only a few locations, e.g., Isle of May, Scotland, and Farallon Islands, California (see above). However, since censusing/monitoring breeding populations is difficult, the estimates are given above to generate a sense of perspective.
Past at-sea counts often problematic owing to lack of logistical and statistical precision. For example, eastern Bering Sea population based on counts at sea totals 3.4 million (147) to 3.5 million (498), or roughly half the adjusted colony-based estimate (275, 499). Techniques, based on flux-corrected strip transects (500) and state-of-the-art statistical analysis (501), allow estimates within 95% confidence intervals of rigorous colony counts (e.g., central California populations).
The actual size and trends of Common Murre populations are a moving target, as regional populations rise and fall on a decadal scale due to climate oscillations, and ultimately effects on food supply, such as Pacific Decadal Oscillation and North Atlantic Oscillation (77). Therefore, no real valid assessment of global trend (see below). Trends in annual standardized whole-breeding site counts from aerial photographs have served to monitor population changes along United States west coast since 1979 (101, 502, 503, 504). Annual plot monitoring possible at few colonies owing to logistics, but conducted regularly at 5-yr intervals at seabird sanctuaries along the north shore of the Gulf of St. Lawrence (initiated in 1925; see 505) and annually at Farallon Islands using correction factors (27, 506, 453). Aerial photographs and direct ground counts used for annual whole-breeding site counts in Labrador, North Shore of Gulf St. Lawrence, and eastern Newfoundland (see 507). Direct counts during short seasonal interval from land and sea, then corrected, made in England, Wales, Scotland, Northern Ireland, and Ireland (328). In Alaska, British Columbia, and many colonies in eastern Canada, whole-colony counts conducted irregularly using standardized methods: trends determined from counts of breeding birds and/or plot changes (e.g., see 497, 364, 365, 451).
U. a. inornata: Overall, despite numbers in Gulf of Alaska decreasing, a 72% increase in Alaska populations overall occurred during 1976–2013, but with some colonies increasing and others decreasing (508, 451). At Walrus Island, Alaska, predominance of one versus the other murre species shifted several times since early 1900s. When North Pacific Pressure Index (PNI) anomalies are consistently positive for a period, Common Murre is abundant, and conversely when it is negative. In accord, during 1870s, 1911–1914, and 1949–1954, this species predominated; in 1900 and 1940s Thick-billed Murre (Uria lomvia) predominated (references in 509, 251); since then, all murres gone due to invasion of sea lions (329; see Behavior: Social and Interspecific Behavior). Oscillating pattern typical of many other (but not all) Bering Sea and Aleutian Islands colonies (139, 409). With increasingly negative PNI (see 251), Common Murres at Bluff, Norton Sound, decreased markedly during 1970s, then stabilized in 1980s; conditions on wintering grounds in southern Bering Sea likely responsible (450); a decline, too, observed at Cape Thompson, Chukchi Sea, Alaska, stabilizing by early 1980s (409). PNI shifts bring food-web changes as indicated by declines in the diets of murres on Pribilof Islands as lipid-poor juvenile walleye pollock/Alaska pollock (Gadus chalcogrammus) switched with lipid-rich sandlance and capelin (250). Overall, however, 1976–2013, Common Murre numbers in Alaska colonies have increased 1.5%/yr and 72.4% overall, despite major, temporary decrease, 1989–1992, owing to oil spill (508).
Otherwise, population changes prove to be complex, and regional concordance problematic, e.g., comparing early 1900s with 1970s–1980s in southeastern Alaska, numbers increased from 300 to 2,500 pairs on St. Lazaria Island and decreased from 20,000 to 5,000 pairs on nearby Forrester Island; numbers on Middleton Island increased from very few in 1956 to 6,000–8,000 in 2001. In many cases, introduced predators and their subsequent removal have caused some changes (348). Apparent switches in presence/absence among colonies on Shumagin, Kodiak, and Barren islands during 1940s-1950s remain a puzzle (509).
In British Columbia, 2003–2007, murre numbers have been decreasing, in part owing to recovery of raptor populations (37). In Japan, numbers have decreased dramatically during latter part of 1900s owing to fishery interactions (110; see below).
U. a. californica: At Farallon Islands, California, numbers decreased from 500,000–1.5 million pairs in early 1800s to <17,000 pairs by 1911, due to commercial egging and human disturbance, exacerbated by an extended period of anomalous warm ocean conditions; population nearly extirpated by 1930s, likely due to past egging, oil pollution, and continued disturbance (99, 101). With increasing protection and reduced oil spills, partial recovery to 6,000 pairs in 1959 and 51,000 pairs by 1982; in 1983–1985, colony crashed to 26,000 pairs, in accord with all other colonies in central California (99, 510, 101). Between 1979 and 1989 (when colonies reached lowest levels), central California population (including Farallon Islands) declined by 9.9%/yr. Since then, increase has been steady, to 275,000 by 2017 (453). In the region, murre numbers increased from 60,000–90,000 in early-1990s to >460,000 by 2015 (453). Decline during 1980s precipitated by extensive mortality from gill-netting and oil spills in 1979–1987, exacerbated by El Niño 1982–1983 (see Conservation and Management: Effects of Human Activity). Earlier, slow recovery likely influenced by shift to a warmer, less productive ocean that affected both the California and Humboldt upwelling systems, with analogous population trends in other diving species (511, 103), as well as prey depletion due to fisheries (249). Recent spectacular increase due to few oil spills and effective fishery management increasing prey availability (82, 249).
Farther north in California Current, at northern Washington colonies, numbers have been increasing, from 4,000–5,000 in mid-1990s to >20,000 currently (Thomas and Lyons 2015, unpublished data); emigration might be involved (340). A major decline has occurred at southern Washington colonies, from continual occupancy of a couple thousand birds in mid-1990s to sporadic occurrence of none to a few hundred birds in recent years (496, 100, Thomas and Lyons 2017, unpublished data). Trend complicated by continuing high levels of oil spill and gill-net mortalities, human disturbance, and eagle predation, exacerbated by marine climate change (496, 340, 103, 101, 159). Largest colonies at Split and Willoughby Rocks almost completely abandoned; decline precipitated by warm-water event in 1981 and El Niño in 1983 (103). Without as much anthropogenic influence, Oregon and northern California populations more resilient (495, 101), though recovery of eagle populations have become problematic for murres in Oregon (502, 351, 342; see Conservation and Management: Effects of Human Activity). Colonies in northern California have been increasing during last two decades (503).
Annual at-sea and colony counts in central California inversely correlated during 1980s–1990s (501), including but not confined to large-scale colony abandonment during El Niño (27, 496). Thus, proportion of birds breeding an important variable to monitor (364). Owing to effective fishery management and no recent oil spills (see Conservation and Management), murres in California, especially central California, have been dramatically increasing since the mid-1990s, now two orders of magnitude more abundant (82, 453, 83). Recolonization recently occurred in northern Channel Islands (98).
U. a. aalge: Quebec: Establishment of federal seabird sanctuaries along North Shore of the Gulf of St. Lawrence, Quebec, in 1925 resulted in annual increases of 5.4% during 1920s–1930s (from 7,200 pairs; 512, 513); growth during 1977–1982 reached 7.5%/yr, and during 1982–1988 reached 10.1%/yr (population in 1993 ~20,000 pairs; 514, 515, 516), and by 2010 increased to 24,700 pairs (505).
Newfoundland and Labrador: At Funk Island, Newfoundland, population increased from 10,000 pairs in 1936 to 400,000 by 1972 (364, 517), owing to relaxed exploitation (518). The colony, the largest in the Atlantic (519, 47), remained at constant size through 1980s (47, 520, DNN), and reached 470,000 pairs in 2009 (S. Wilhelm, ECCC-CWS). Annual rate of increase to 1970s was 10.8% (510). Green Island, Witless Bay, Newfoundland, colonized (or recolonized; DNN) in 1936; reached 50,000 pairs by 1959 (38), and 74,000 pairs by 1980 (517, 47). Otherwise, the three Witless Bay colonies have been increasing and reached 250,000 pairs in censuses from 2007 to 2019 (521, S. Wilhelm ECCC-CWS), though recently there has been a great deal of colony re-distribution following an increase in the number of juvenile eagles (351). Cape St. Mary's, Newfoundland, colony remained at 2,500 pairs from late 1800s to late 1950s (38); thereafter it grew to 10,000 pairs (517), but then declined owing to chronic oil pollution (522); reached 15,000 pairs in 2007 (S. Wilhelm ECCC-CWS, unpublished data) with breeding sites on the adjacent mainland, as well as the traditional rock stack. Numbers at Baccalieu Island, Newfoundland, increased from 500–1,000 pairs in 1934 to 2,500 pairs in 1959 (38), and increased again to 4,000 pairs by early 1980s (523, 520, DNN), but then decreased to 1,000 by 2012 (S. Wilhelm ECCC-CWS, unpublished data). All increases likely the result of release from persecution, including egging and hunting (see Conservation and Management: Effects of Human Activity) and decreased use of gill-nets. Recovery slowed during early 1900s perhaps due to changed prey availability related to alteration of Labrador Current (38). Approximately 11,000 additional pairs were reported breeding on islands adjacent to Newfoundland (S. Wilhelm, ECCC-CWS, unpublished data).
Recent declines have been observed for major colonies in coastal Labrador: breeding population at Gannet Islands, which was >60,000 pairs in 1983 (29) declined to ~37,000 pairs in 1998 (524) and 18,650 pairs in 2015 (S. Wilhelm ECCC-CWS, unpublished data). Unusual occurrence of arctic foxes, known to have impacted breeding (339), plus visits by polar bears are believed to be responsible for the declines; a shifting prey base may also be a factor (235), as well as egging and hunting (DNN). Approximately 6,000 pairs breed on other islands adjacent to Labrador (S. Wilhelm ECCC-CWS, unpublished data).
United Kingdom and Ireland: At least during the 1900s into the early 2000s, Common Murres appear to have been resilient to a number of changing environmental factors, showing minimal alteration of regional populations (104). At a smaller scale, however, trends are evident, as for example at Skomer Island, Wales, where numbers decreased from 100,000 individuals during the 1930s to 5,000 individuals by the 1940s (due oil pollution). Since about 1980, however, numbers have been increasing at ~5% per annum to reach ~26,000 individuals by 2015 (105). At Isle of May, Scotland, numbers initially decreased slightly from ~15,000 females in 1983 to ~11,000 in 1987, but then increased gradually to ~21,000 by 2005 (158); numbers then decreased to ~13,000 by 2015, then varying at about that level through to 2020 (M. P. Harris and S. Wanless, personal communication).
Norway: Numbers in northern Norway declined during the 1960s to the 1980s from around 150,000 pairs in the 1960s (459). A steep decline was observed in the winter 1986/1987 with a 70–90% decline in the population on Bjørnøya (Svalbard) and northern Norway (107). During this die-off, the Bjørnøya population declined from 245,000 to 36,000 pairs (525). The numbers have increased slowly since 1987 (109), and the population in northern Norway counts 46,000 pairs and Bjørnøya 156,000 pairs in 2019.
Iceland: On the basis of a comparison of counts during 1983–1986 with those during 2005–2009, numbers of Common Murre have decreased by 30%, especially in the southwest portion, now being 698,000 pairs (526).
Common Murre was almost extinct in the Baltic Sea by the end of the 1800s. However, with protection their numbers grew during the latter part of the 1900s to reach ~45,000 birds by 2000 (15,000 breeding pairs; 465).
As noted above, somewhat relaxed natal philopatry, compared to many other colonial seabirds, makes definition of a ‘colony’ problematic. Trends among meta-colonies or in sub-regions likely more indicative of population changes (see Priorities for Future Research), and as noted below, depending on the situation, large numbers at one location can inhibit or facilitate growth at neighboring locations within foraging range of a large population. Along west coast of North America, populations at relatively few, broadly spaced breeding sites tend to be very large, in contrast to east coast or Europe, where breeding sites are numerous, and at closely spaced individual sites numbers often small (cf. Appendices 2, 3, 4, 5, 6, 7, 8, and 9).
Simulation models indicate population growth sensitive to adult survivorship, secondarily to subadult survivorship (527, 528, 312, 450, 444, 158, 463). Assuming 90% adult survival, 29% postfledging (i.e., ledge departure) survival to breeding age, and 0.7 chicks/breeding pair annual production, a 5%/yr population growth rate would require one of the following: increase adult survival to 97% (70% decrease in mortality), increase subadult survival to 51% (31% decrease in survival), or (the impossible) increase production to 1.24 chicks/pair (312). Using a more extensive data set than that available to Hudson (312), Reynolds et al. (158) indicated that adult survival varied by year, 90.4–99.1%, first-year survival varied 13.0–87.9%, and that age 0 to age 2 survival was 54.0%. As reviewed above, survival and annual production is higher in many populations growing in size. Age structure stabilizes slowly after a perturbation; beginning at 0 in yr 1, after 20 yr the last survivors of the first generation would be still present, as would representatives of 6 other generations. Assuming 85–88% survival, a 3-yr-old has a life expectancy of 7–8 yr (527).
Ultimately, availability of food within foraging range during breeding limits growth of a breeding site; size is related to amount of foraging habitat available (35, 529, 530, 82, 83). At some point, owing to competition, smaller populations establish nearby if breeding habitat available (161), as usually occurs in newly formed colonies in northern California and Oregon (101, 160), and possibly in Gulf of St. Lawrence (J.-F. Rail, personal communication), and elsewhere including Bay of Fundy and Gulf of Maine (DNN). Smaller breeding site populations, however, if in range of a large population, may not be able to grow once asymptote of foraging density reached (530, 531). On the other hand, a large ‘source’ population nearby can facilitate growth or even founding of peripheral populations (160; see Conservation and Management).