Endangered and Threatened Wildlife and Plants; 12-Month Finding on a Petition To List the Black-footed Albatross as Endangered or Threatened

DEPARTMENT OF THE INTERIOR

Fish and Wildlife Service

50 CFR Part 17

[Docket No. FWS-R1-ES-2007-0004; MO 92210-0-0008]

Endangered and Threatened Wildlife and Plants; 12-Month Finding on a Petition To List the Black-footed Albatross as Endangered or Threatened

AGENCY:

Fish and Wildlife Service, Interior.

ACTION:

Notice of 12-month petition finding.

SUMMARY:

We, the U.S. Fish and Wildlife Service (Service), announce a 12-month finding on a petition to list the black-footed albatross (

Phoebastria nigripes

) as endangered or threatened under the Endangered Species Act of 1973, as amended (Act). The petitioners provided three listing options for consideration by the Service: Listing the black-footed albatross throughout its range; listing the Hawaiian Islands breeding population of the black-footed albatross as a Distinct Population Segment (DPS); or listing the Japanese Islands breeding population of the black-footed albatross as a DPS. After a review of the best available scientific and commercial information, we find that listing the black-footed albatross rangewide is not warranted at this time. We find that the Hawaiian Islands breeding population and the Japanese Islands breeding population are separate DPSs, as defined by DPS policy. However, we further find that neither the Hawaiian Islands DPS nor the Japanese Islands DPS of the black-footed albatross warrants listing at this time. We ask the public to submit to us any new information that becomes available concerning the threats to the black-footed albatross or its habitat at any time.

DATES:

The finding announced in this document was made on October 7, 2011.

ADDRESSES:

This finding is available on the Internet at

http://www.regulations.gov

at Docket Number FWS-R1-ES-2007-0004, and

http://www.fws.gov/pacificislands/.

Supporting documentation we used in preparing this finding is available for public inspection, by appointment, during normal business hours at the U.S. Fish and Wildlife Service, Pacific Islands Fish and Wildlife Office, 300 Ala Moana Boulevard, Box 50088, Honolulu, Hawaii 96850. Please submit any new information or materials concerning this finding to the above address.

FOR FURTHER INFORMATION CONTACT:

Dr. Loyal Mehrhoff, Field Supervisor, Pacific Islands Fish and Wildlife Office (see

ADDRESSES

); by telephone at 808-792-9400; or by facsimile at 808-792-9581. If you use a telecommunications device for the deaf (TDD), call the Federal Information Relay Service (FIRS) at 800-877-8339.

SUPPLEMENTARY INFORMATION:

Background

Section 4(b)(3)(B) of the Act (16 U.S.C. 1532

et seq.

) requires us to make a finding within 12 months of the date of receipt of any petition to revise the Lists of Endangered and Threatened Wildlife and Plants, provided the petition contains substantial scientific and commercial information that listing may be warranted. In this finding, we will determine that the petitioned action is: (a) Not warranted, (b) warranted, or (c) warranted, but the immediate proposal of a regulation implementing the petitioned action is precluded by other pending proposals to determine whether species are threatened or endangered, and expeditious progress is being made to add or remove qualified species from the Federal Lists of Endangered and Threatened Wildlife and Plants. Section 4(b)(3)(C) of the Act requires that we treat a petition for which the requested action is found to be warranted but precluded as though resubmitted on the date of such finding, requiring that a subsequent finding be made within 12 months. We must publish these 12-month findings in the

Federal Register.

Previous Federal Actions

On October 1, 2004, we received a petition dated September 28, 2004, from Earthjustice on behalf of the Turtle Island Restoration Network and the Center for Biological Diversity, requesting that we list the black-footed albatross as a threatened or endangered species throughout its range, with critical habitat, or that we list either or both the Hawaiian breeding population and/or the Japanese breeding population as a DPS, and that we designate critical habitat concurrently with listing. Because the determination of critical habitat is not a petitionable action under the Act, we did not consider the designation of critical habitat in this finding. The petition included supporting information regarding the species' taxonomy and ecology, historical and current distribution, present status, potential causes of decline, and active imminent threats. In our December 3, 2004, letter to the petitioner we acknowledged the petition and provided our determination that emergency listing was not warranted. We also explained that, due to a significant number of listing rules due in 2005 under court-approved settlement agreements, we had insufficient resources to initiate a 90-day finding at that time.

In 2007 we received funding and initiated the 90-day finding. On October 9, 2007, we published a 90-day petition-finding (72 FR 57278), in which we concluded the petition presented substantial scientific or commercial information indicating listing of the black-footed albatross may be warranted, and we initiated a status review. In that notice, we announced the opening of a 60-day information collection period and invited the public to submit to us any pertinent information concerning the status of or threats to this species. We received information from 14 parties in response to this notice. We also consulted with recognized species experts and other Federal and State agencies. On August 26, 2009, we announced the reopening of the information collection period (74 FR 43092) in response to the U.S. Geological Survey-Biological Resources Discipline (USGS-BRD) publication of the

Status Assessment of the Laysan and Black-Footed Albatrosses, North Pacific Ocean, 1923-2005

(Arata

et al.

2009, entire). One additional party provided comments during the second information collection period. This notice constitutes the 12-month finding on the petition to list the black-footed albatross as endangered or threatened with critical habitat.

Outline of This Notice

In this notice, we first provide background information on the biology of the black-footed albatross. Next we analyze the threat factors facing the black-footed albatross throughout its range to determine if listing under the Act is warranted. This analysis is called a “Five Factor Analysis” because it addresses the five factors listed in section 4(a)(1) of the Act that are used in determining whether a species meets the definition of an endangered or a threatened species under the Act. For each factor, we first determine whether any stressors, or risk factors, appear to be negatively affecting black-footed albatrosses anywhere within the species' range. If we determine they are, then we evaluate whether each of these risk factors, either singly or in combination, is resulting in population-level effects. Defining a stressor to be a threat to the species does not

necessarily mean the species meets the definition of endangered or threatened. Virtually all species face some degree of threat from either natural or anthropogenic sources. Rather, for the purposes of the Act, we must consider each of the stressors and identified threats, both individually and cumulatively, and make a determination with respect to whether the species is endangered or threatened according to the statutory standard. That is, we must make a determination as to whether the threats are impacting the species to such a degree that the species is currently in danger of extinction (endangered), or likely to become so within the foreseeable future (threatened), throughout all or a significant portion of its range. Further details on this evaluation are provided below in the section

Summary of Factors Affecting the Species.

Species Information

The black-footed albatross is a migratory, open-ocean species whose current range encompasses the seas from north of the Hawaiian Islands to the Bering Sea (15° N to 60° N), eastward to the western coast of North America, and west to the northeastern coast of Japan (118° E to 112° W) (Figure 1) (Awkerman

et al.

2008, p. 4; Fischer

et al.

2009, p. 757).

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Approximately 95 percent of the breeding population nests in the Hawaiian Islands archipelago in the central Pacific; other breeding colonies are found on the Japanese Islands in the western Pacific in the Izu-Torishima Islands, the Ogasawara Islands (also known as the Bonin Islands), and the Senkaku Islands (Figure 2).

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Taxonomy and Description

The black-footed albatross is one of three north Pacific species in the seabird family Diomedeidae (albatrosses). Adults are uniformly sooty brown with a whitish ring at the base of the bill, a white patch behind the eye, and white feathers over the base of the tail and undertail coverts. Birds of all ages have a blackish bill, legs, and feet. Fledglings are uniformly dark brown and acquire a white ring at the base of the bill and around the tail as they age (Hyrenbach 2002, p. 87). The wingspan is 76 to 85 inches (in) (193 to 216 centimeters (cm)), and the average weight is 6.17 pounds (lb) (2.30 kilograms (kg)) (Cousins and Cooper 2000, p. 3). No subspecies are recognized, though significant genetic differentiation between the Hawaiian and Japanese populations has been identified (Walsh and Edwards 2005, pp. 292-294; Eda

et al.

2008, pp. 112-115), and further research may possibly indicate that taxonomic revision is warranted (Eda

et al.

2008, p. 115). At present the black-footed albatross continues to be classified by taxonomic authorities as a single species (American Ornithologists' Union 1998 and supplements; Integrated Taxonomic Information System 2011), and there does not appear to be a broad scientific consensus that this classification is incorrect; therefore, we consider it a single species in this finding.

Life History

Black-footed albatrosses range throughout the north Pacific (Cousins and Cooper 2000, p. 12). Reports of banded birds, casual observation, and studies using satellite transmitters have revealed patterns in the use of oceanic habitats by black-footed albatrosses that vary with age and breeding status, and oscillate with the breeding cycle (Cousins and Cooper 2000, p. 12). Adult birds concentrate around the colonies during egg-laying, incubation, and chick brooding. As chicks get older, breeding adults range much farther from the colony to reach productive foraging waters. Post-breeding adults forage near the western coast of North America, and south of Alaska as far west as the Aleutian Islands. Black-footed albatrosses use areas of coastal upwelling or convergence for foraging throughout the north Pacific; these highly productive areas are also used by numerous fisheries (Fernandez

et al.

2001; Hyrenbach

et al.

2002; Hyrenbach and Dotson 2003; Fischer 2007; Fischer

et al.

2009).

Black-footed albatrosses live for 40 to 50 years, and represent a classic example of a “K-selected” species (

i.e.,

the species is long-lived, has delayed reproductive maturity, produces relatively few young, and is dependent upon high annual adult survivorship). The earliest known age for first breeding by black-footed albatrosses is 4 years of age, but on average the age of first breeding is 7 years (Cousins and Cooper 2000, p. 51). Pairs mate for life, and mate loss in black-footed albatrosses can cause adults to skip up to five breeding seasons prior to forming a new pair (Committee on the Status of Endangered Wildlife in Canada (COSEWIC) 2007, p. 33). Only one egg is laid per year, and pairs do not attempt to renest if nesting failure occurs (Cousins and Cooper 2000, p. 2). Sometimes pairs will skip a breeding year. It is estimated that 75 percent of black-footed albatrosses that fledged a chick one year will go on to breed the next year, while 83 percent of pairs that experience nest failure will breed the next year (Viggiano 2001, p. 59).

Female black-footed albatrosses have a high level of affinity to the nest site. Long-term studies have shown that over 99 percent of females return to breed on the island or atoll where they hatched and fledged (known as their natal site) and establish their own nesting site nearby (Rice and Kenyon 1962a, pp. 532-533). Most have been found to return within less than 20 feet (ft) (6 meters (m)) of the same nest site season to season (Rice and Kenyon 1962a, p. 533). Such is their nest site fidelity that birds banded at a particular site in 1938 were found still nesting at that site 20 years later (Rice and Kenyon 1962a, p. 533). Since the vast majority of females nest on the island where they themselves hatched (Rice and Kenyon 1962a, pp. 532-533), recolonization of formerly occupied islands or atolls (that were abandoned or where black-footed albatrosses were extirpated due to cataclysmic or stochastic events) and colonization of new islands or atolls by dispersing breeders is relatively rare. Such events are not unknown, however. For example, black-footed albatrosses banded as nestlings on Midway Atoll were later observed breeding on Kure Atoll, and other individuals are known to have moved from their natal sites to breed between the islands of Pearl and Hermes Reef, French Frigate Shoals, and Kure Atoll as well (Woodworth 1972, p. 96). Black-footed albatrosses recolonized Torishima Island, the Ogasawara Islands, and the Senkaku Islands followed cessation of World War II military activities in the western Pacific (see

Volcanic Activity,

below), and pioneering attempts by black-footed albatrosses to breed on Mexico's Guadalupe and San Benedicto islands in the eastern Pacific have been reported recently.

Birds arrive at their nesting colonies in the central and western Pacific islands in mid- to late October (Rice and Kenyon 1962a, p. 552; Woodward 1972, p. 92). Eggs are laid between mid-November and mid-December (Rice and Kenyon 1962a, p. 540; Woodward 1972, p. 92; Awkerman

et al.

2008; Agreement on the Conservation of Albatrosses and Petrels [ACAP 2010], p. 2). Incubation lasts approximately 66 days, and most eggs hatch by early February (Rice and Kenyon 1962a, p. 546). Both adults take turns brooding the chick and attend it for approximately 1 month, after which the adults spend most of their time at sea, returning only to feed the chick (Rice and Kenyon 1962a, pp. 548-549). The chick-rearing stage lasts approximately 140 days, with fledging occurring in mid-June to mid-July (Rice and Kenyon 1962a, p. 562). Once fledged, the young birds remain at sea and do not return to land for 2 to 5 years (Rice and Kenyon 1962a, p. 520; Viggiano 2001, p. 15).

Diet and Feeding Habitats

Black-footed albatrosses are surface feeders and scavengers, generally seizing food within 3 ft (0.9 m) of the ocean's surface (Brooke 2004, p. 191). The birds take prey at the surface of the water, and occasionally partially submerge below the surface (Awkerman

et al.

2008, p. 14). Fernandez and Anderson (2000, entire) used an immersion monitor and satellite telemetry to evaluate feeding activity patterns during the chick-brooding period, when shorter foraging trips would be expected (Fernandez

et al.

2001, p. 4). The majority of time at sea was spent flying (90.8 percent), with most immersions less than 100 seconds long, indicating birds were engaged in surface foraging rather than resting (Fernandez and Anderson 2001, p. 580). Immersions (presumed feeding activity) during this study occurred primarily during the daytime, though some presumed feeding activity did occur during the night.

The diet of adult black-footed albatrosses is composed primarily of flying fish eggs, but also includes squid, fish, offal, and human refuse (Brooke 2004, p. 191). Black-footed albatrosses are known to follow fishing boats and are more aggressive than Laysan albatrosses (

Phoebastria immutabilis

) in scavenging fish discards (Fischer

et al.

2009, p. 758). Harrison

et al.

(1983, entire) and Gould

et al.

(1997, entire) studied the food habits of the black-footed albatross. Harrison

et al.

(1983, pp. 15-18) collected regurgitation samples from adult birds primarily from Laysan Island and Midway Atoll, but also collected samples during the chick-rearing stage from Kure Atoll and French Frigate Shoals, and found the contents were primarily flying fish eggs, squid, and crustaceans. Gould

et al.

(1997, p. 550) sampled birds collected from drift nets in the north Pacific during the nonbreeding season. They found the greatest percentage of stomach contents was squid species typically targeted by the squid and driftnet fisheries. In their analysis of both Laysan and black-footed albatross stomach contents, Sileo

et al.

(1990a, p. 674) found that chicks consume a variety of plastic objects. Black-footed albatrosses are especially prone to inadvertently ingesting plastic because plastic particles floating on or below the water's surface resemble flying fish eggs, a major component of their diet. In addition, flying fish eggs are often laid in floating items, including plastic refuse, thereby increasing the chances of inadvertent plastic ingestion (Cousins and Cooper 2000, p. 5).

Nesting Sites

Black-footed albatross nests are most often a depression scooped out in a sandy substrate, surrounded by a rim of sand (Arata

et al.

2009, p. 10). They are usually located on exposed sandy beaches at the beginning of the vegetation line (Cousins and Cooper 2000, p. 5; Awkerman

et al.

2008, p. 20; Arata

et al.

2009, p. 10). At Midway Atoll and Tern Island (French Frigate Shoals) in the Hawaiian Islands, nests are also located in areas with low-growing vegetation (Arata

et al.

2009, p. 10). On the volcanic islands of Torishima Island and the Ogasawara Islands, nests are not found on beaches, but are located at high elevations on sparsely to highly vegetated exposed volcanic slopes (Cousins and Cooper 2000, p. 5).

Breeding Distribution

Historically, the breeding range of the black-footed albatross likely extended from Lehua Island (offshore of Niihau Island) in the Hawaiian Islands west to the Senkaku Islands in the western Pacific. In the late nineteenth and early twentieth centuries, this range was reduced due to extirpation of the black-footed albatross from entire breeding islands by egg and feather hunters, and later by military activities on some of the nesting islands in the central and western Pacific from World War II-related military occupation and activities (Rice and Kenyon 1962b, pp. 366-367; Naughton

et al.

2007, p. 6). The likely historical breeding range of the black-footed albatross prior to these extirpation events is detailed in table 1.

Table 1—Summary of the Historical Distribution of Black-Footed Albatross Breeding Colonies and Their Current Status as Extant E; Extinct X; or Prospecting P (Occasional Breeders Scouting Out New Nest Sites; Considered a Possible Early Sign of Range Expansion)

Breeding colony

Year cited

First known reference

Status

Central Pacific Islands

Northwestern Hawaiian Islands:

Nihoa Island

1923

Wetmore

1

E

Necker Island

1923

Wetmore

1

E

French Frigate Shoals

1923

Wetmore

1

E

Laysan Island

1923

Wetmore

1

E

Lisianski Island

1923

Wetmore

1

E

Pearl and Hermes Reef

1923

Wetmore

1

E

Midway Atoll

1923

Wetmore

1

E

Kure Atoll

1923

Wetmore

1

E

Main Hawaiian Islands:

Kaula (Kauai)

1923

Wetmore

1

E

Lehua (Kauai)

1923

Wetmore

1

E

Other:

Taongi Atoll (Marshall Islands)

1874

Dall

1

X

Wake Atoll or Wake Island

1841

Peale

1

P

Minami-Torishima (Marcus Island)

1902

Bryon

1

X

Johnston Atoll

1923

Wetmore

1

X

Western Pacific Islands

Iwo Jima (Volcano Islands)

1891

Seebohm

1

X

Izu Shoto (Torishima Island)

1889

Hattori

1

E

Ogasawara Gunto (Bonin Islands)

1890

Seebohm

1

E

Senkaku Retto (Ryukyu Shoto)

unknown

unknown

E

Eastern Pacific Islands

Isla Guadalupe

1998

Pitman & Ballance

2

P

San Benedicto

2000

Pitman & Ballance

2

P

1

Referenced in Rice & Kenyon, 1962a, p.21

2

Referenced in Pitman & Ballance, 2002, p. 13.

Wake Island or Wake Atoll was first reported as a breeding colony for black-footed albatross in December 1841 by Titian R. Peale while on a U.S. Exploring Expedition. During this expedition, an egg and a black-footed albatross skin were collected; however, the egg was later judged, by size and shape, to be that of a Laysan and not a black-footed albatross (Rice and Kenyon 1962b, p. 379). Thus, because a single collected skin of a black-footed albatross does not denote nesting or breeding, we cannot conclude that these birds historically nested or bred on Wake Atoll.

Present breeding populations of black-footed albatross occur as follows (table 2): (1) Hawaiian Islands (central Pacific, Hawaii archipelago) (1a) Northwestern Hawaiian Islands—Nihoa Island, Necker Island, French Frigate Shoals, Laysan Island, Lisianski Island, Pearl and Hermes Reef, Midway Atoll, and Kure Atoll; (1b) Main Hawaiian Islands—Lehua Island, Kaula Island; (2) Japanese Islands (western Pacific) (2a) Izu Islands—Torishima Island; (2b) Ogasawara Islands (also known as the Bonin Islands)—nine islets; (2c) Senkaku Islands — three islets (Kawakami

et al.

2006, p. 187; Chiba

et al.

2007, p. 5; Eda

et al.

2008, p. 109).

Table 2—Black-Footed Albatross Population Counts or Estimates of Breeding Pairs From All Known Breeding Sites 1993-2010 (United States, Japan) (ACAP 2010, Table 3, p. 4; Flint 2011a, pers. comm.)

Breeding site

Jurisdiction

Last year surveyed

Number of breeding pairs

Hawaiian Islands (Central Pacific)

Northwestern Hawaiian Islands:

Nihoa Island

United States

2007

1

Necker Island

United States

1995

112

French Frigate Shoals

United States

2009

4,309

Laysan Island

United States

2010

22,272

Lisianski Island

United States

2006

2,126

Pearl and Hermes Reef

United States

2003

6,116

Midway Atoll

United States

2010

25,581

Kure Atoll

United States

2010

3,486

Main Hawaiian Islands:

Kaula Island (Kauai)

United States

1993

3

1

Lehua Island

United States

2007

25

Total Central Pacific

64,031

Japanese Islands (Western Pacific)

Torishima Island (Izu Islands)

Japan

2003

2,150

Ogasawara (Bonin) Islands (Muko-jima Island)

Japan

2006

967

Ogasawara (Bonin) Islands (Haha-jima Island)

Japan

2006

11

Senkaku Islands

Japan

2002

56

Total Western Pacific

3,184

Total Rangewide

67,215

1

Survey at Kaula was done 16-17 November, 1998, which is early for nesting. Nine birds were present on the island.

As of 2010, there are no established breeding colonies in the Marshall Islands or on Wake Atoll. While black-footed albatrosses have attempted to breed at Wake Atoll on occasion, most nests, both with and without eggs, were subsequently abandoned, and none have ever successfully fledged young. Birds are likely prospecting the atoll for potential nesting sites (Rauzon

et al.

2008, pp. 14-15) (see Marshall Islands in “Current Population Status” below). Isolated attempts by black-footed albatrosses to breed on the Revillagigedo Islands of Mexico have been reported on Guadalupe and San Benedicto islands (Pitman and Ballance 2002, p. 13), but there is no record of a breeding population ever being established (Henry 2007, pers. comm.; Hebshi 2010, pers. comm.). Other than one unsubstantiated report of a “fully-feathered chick” on Guadalupe Island in 1998, there is no evidence that any young have been fledged (see Mexican Islands in “Current Population Status” below).

Foraging Distribution During the Breeding Season

Satellite telemetry data collected in 1988 and 1989 indicate black-footed albatrosses forage north and northeast of breeding colonies in the Hawaiian Islands. They tend to forage in pelagic (open ocean) oligotrophic (low in dissolved nutrients and high in oxygen) waters within the vicinity of the nest (maximum range 188 miles (mi) (303 kilometers) (km)) during the nest-guard phase (when chicks are less than 18 days old) (Fernandez

et al.

2001, pp. 4-5; Hyrenbach

et al.

2002, p. 288). When feeding older nestlings, black-footed albatrosses breeding on Tern Island mixed short trips near nest sites with long trips to the highly productive waters along the continental shelf of North America (Fernandez

et al.

2001, pp. 4-7; Hyrenbach

et al.

2002, pp. 288-294). They foraged along the North Pacific Transition Zone, which separates the Subarctic Domain (defined as a water mass with temperature less than 50 °F (10 °C)) from the North Pacific Subtropical Gyre (a large-scale circular feature made up of ocean currents that spiral around a central point; it is made up of four large, clockwise-rotating currents—North Pacific, California, North Equatorial, and Kuroshio)), and is characterized by convergence fronts and high productivity (Hyrenbach

et al.

2002, p. 296). Overall, the adults ranged from 18° N to 48° N latitude in the north Pacific and over a large area in the eastern Pacific (121° W to 172° W longitude) (Fernandez

et al.

2001, p. 4). Similar results have been reported using Geographic Positioning Systems (GPS) tracking of breeding birds in the Bonin Islands (Kawakami

et al.

2006, p. 189). Adults incubating eggs or brooding young chicks foraged within 252 mi (405 km) of the breeding site; over 90 percent of the observations were within 124 mi (200 km) of the colony.

Foraging Distribution During the Nonbreeding Season

During summer months (postbreeding), female black-footed albatrosses captured off the coast of California foraged largely along the transition zone between the California Current (a cold current originating in the northern part of the Pacific Ocean, flowing southeast along the coast of western North America) and the North Pacific Gyre, and spent 39, 43, and 18 percent of their time at sea in tropical waters, subtropical frontal zones, and subtropical waters, respectively (Hyrenbach and Dotson 2003, p. 397). Likewise, they spent 25, 24, and 51 percent of their time foraging in the exclusive economic zones (EEZ) of the United States, Mexico, and the high seas, respectively (Hyrenbach and Dotson 2003, p. 397).

Postbreeding black-footed albatrosses captured off the coast of Alaska ranged from 60° N to 36° N, and 125° W to 180° W (Fischer

et al.

2009, p. 757). Within this range, they spent more time in continental margin waters versus oceanic waters; within the continental margin waters they spent equal time in the continental shelf, shelf break, and slope waters (Fischer

et al.

2009, pp. 755-756).

Demography and Population Resiliency

Certain intrinsic aspects of black-footed albatross ecology and demography are relevant to the species' status. Stable populations of K-selected species, such as the black-footed albatross, generally live in relatively constant (

i.e.,

not highly variable) environments and are characterized by low annual productivity rates balanced with high annual survival rates, meaning that individuals must live many years to replace themselves with offspring that survive to recruit into the breeding population. (The letter “K” represents the carrying capacity of a given environment, and is also used to represent a species whose reproductive strategy is to keep a stable population close to the carrying capacity.) Cousins and Cooper (2000, pp. 53-54) found that black-footed albatross population trends

were more sensitive to changes in survival than to changes in fecundity.

Although factors that compromise productivity can cause populations to decline, adult survival is often the more important determinant of population size and persistence for a K-selected species (Cousins and Cooper 2000, p. 53). Annual adult death rates for the black-footed albatross are normally very low, on the order of 3 to 8 percent (in other words, annual adult survivorship is about 92 to 97 percent (Cousins and Cooper 2000, p. 50; Veran

et al.

2007, p. 7; Arata

et al.

2009, p. 47)). If a sufficient number of adults are removed from the population prior to replacing themselves (

i.e.,

adult survival is decreased beyond a certain threshold), the population will decline. Additionally, reduced juvenile survivorship will also affect the population; Cousins and Cooper (2000, p. 53) estimated that juvenile survival of black-footed albatrosses has to be 86 percent or higher to prevent a population decrease. Estimates of juvenile survivorship for the black-footed albatross have been more varied over the years; Arata

et al.

(2009, p. 47) report a rate as low as 0.688 for the period 1963-1982, but estimate juvenile survivorship of 0.993 over the period 1994-2002. For French Frigate Shoals, juvenile survivorship was estimated at 0.79 for the years 1994-2000 (ACAP 2010, Table 5, p. 8). All of the characteristics of the black-footed albatross—its longevity, low reproductive rates, delayed sexual maturity, irregularity in annual breeding, and life-long pair bonding (with consequent delays in subsequent breeding if a mate is lost)—make it difficult to detect changes in population structure, particularly the recruitment of juveniles into the population. Species with such characteristics are slow to exhibit population declines and are inherently more vulnerable to extinction (Primack 1993, p. 102; Meffe and Carroll 1994, p. 128). These intrinsic aspects of black-footed albatross ecology and demography signal the continuing need to monitor their populations, despite the fact that numbers are presently stable and the species continues to be widely distributed across its range (Arata

et al.

2009 p. 2; see “Current Population Status” below).

Current Population Status

Rangewide

Feather and egg hunters decimated black-footed albatross populations until the 1920s, and an estimate of population size prior to this period is not known. In 1923, the estimated breeding population was 17,800 pairs in Hawaii, and 200 in Japan (Arata

et al.

2009, p. 35). The current black-footed albatross worldwide population estimate, with most recent counts from the 2010 nesting season, is approximately 67,215 breeding pairs (ACAP 2010, p. 4; Flint 2011a, pers. comm.). Based on a Leslie matrix model, roughly 60,000 breeding pairs were estimated to represent a total world population of approximately 300,000 black-footed albatrosses, including both breeding and nonbreeding individuals (Cousins and Cooper 2000, p. 19; Niel and Lebreton 2005, p. 833); the most recent counts of more than 67,000 nesting pairs therefore puts the estimated world population of black-footed albatrosses at well over 300,000 individuals.

Cousins and Cooper (2000) present data on the number of breeding black-footed albatrosses from Midway Atoll, Laysan Island, and French Frigate Shoals as well as the available information for all other sites throughout the world. An examination of their data indicates a stable or increasing global trend in the number of breeding black-footed albatross in the years 1992 through 1999 (Cousins and Cooper 2000, p. 19 and Figure 19). More recently, data presented by Arata

et al.

(2009, Figure 22) indicate an increasing world population of the black-footed albatross between 1923 and 2005. In addition, survey data indicate populations in the Japanese Islands have been steadily increasing (Cousins and Cooper 2000, p. 23; Hasegawa 2010, pers. comm.; see Figure 4 of this document). All of these population data are based on counts of active nests at breeding sites. It should be noted that because only the breeding component of the species' population is counted, changes in population demographics that could affect the population in the long term cannot be detected with this method (Viggiano 2001, p. 5). For example, any significant increase in juvenile mortality would not be detected until years later, when these birds would normally be entering the breeding population that is counted. In the absence of more precise data, however, these counts are generally used as a rough index of population numbers, and represent the best scientific information available to us.

Hawaiian Islands

Roughly 95 percent of the world population of black-footed albatrosses breed in the Hawaiian Islands. Black-footed albatrosses currently nest on Lehua Island and Kaula Island off of Kauai in the main Hawaiian Islands, and in the Northwestern Hawaiian Islands on Nihoa Island, Necker Island, French Frigate Shoals, Laysan Island, Lisianski Island, Pearl and Hermes Reef, Midway Atoll, and Kure Atoll. Many of the smaller breeding populations of black-footed albatross are not regularly monitored, but standardized counts and estimates of active nests have been conducted in the Northwestern Hawaiian Islands since 1980 at French Frigate Shoals and since 1991 at Midway Atoll and Laysan Island (Naughton

et al.

2007, p. 6). These three colonies collectively comprise 77 percent of the global breeding population of the black-footed albatross as of 2010 (ACAP 2010, p. 4).

Based on the latest nest count data as of 2010, the largest colony of black-footed albatrosses at 25,581 breeding pairs is on Midway Atoll, representing approximately 40 percent of the world's breeding population. Laysan Island has the second largest colony with 22,272 breeding pairs (approximately 35 percent of the global breeding population), and French Frigate Shoals is the smallest of the three with 4,309 breeding pairs, or roughly 7 percent of the world's breeding pairs (Flint 2011a, pers. comm.). Prior to 1997, instead of direct nest counts on Laysan Island, nesting estimates were derived from counts on plots from a portion of the island that were then extrapolated to represent total nesting area. Beginning in 1997, the direct count method (counts of all nests) used at French Frigate Shoals and Midway Atoll was adopted on Laysan Island as well. An analysis of the nest count data from these three regularly monitored colonies at Laysan Island, French Frigate Shoals, and Midway Atoll for the years 1998 to 2009 demonstrates an increasing trend on the order of 0.93 percent per year for the three islands combined (ACAP 2010, p. 5, Fig. 2A). Individually, the breeding population at Midway increased at an average annual rate of 1.3 percent between the years 1992 and 2009 (ACAP 2010, p. 7, Table 4). At French Frigate Shoals, the colony for which the longest time series of data is available, the number of breeding pairs has fluctuated between the years 1980 and 2009, but overall is increasing at an average rate of 0.43 percent annually (ACAP 2010, p. 7, Table 4). Laysan Island, however, has shown a negative trend over the years 1998 to 2009, decreasing at an average annual rate of 1.1 percent (ACAP 2010, p. 7, Table 4). Laysan Island formerly supported the largest breeding population of black-footed albatrosses, until it was surpassed by Midway Atoll in 2004 (ACAP 2010, p. 6). Figure 3 shows the linear trend between 1998

and 2009 for the number of pairs nesting at French Frigate Shoals, Laysan Island and Midway Atoll, individually and combined (taken from ACAP 2010, p. 6, Figure 2).

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These trends are consistent with those reported in a recent status assessment of the black-footed albatross conducted by the U.S. Geological Survey (Arata

et al.

2009, entire). The linear regression analysis in that report indicates a significant increasing trend between the years 1923 and 2005 for black-footed albatrosses at Midway Atoll, Laysan Island, and French Frigate Shoals combined, and no trend (stable population) for the more recent time periods examined, from 1957 to 2005 and 1998 to 2005 (Arata

et al.

2009, p. 29, Table 6). The divisions in time steps represent the earliest thorough surveys of the population in 1923 and 1957, and the beginning of standardized surveys at Midway Atoll and Laysan Island in 1998. The authors attribute the positive growth in the black-footed albatross population, since 1923, to the cessation of poaching at nesting colonies. In addition, they state that only the time-series data from French Frigate Shoals are long enough to show a potential change over time, and note that this population shows positive annual population growth rates with a median trend for growth over the next 60 years. However, they also point out that French Frigate Shoals represents only a small fraction of the global population and advise caution in extrapolating these numbers (Arata

et al.

2009, p. 50), and we note further that the projected growth trend is based on an implicit assumption of no changes in conditions.

Arata

et al.

(2009) also used matrix models to examine population data for the black-footed albatross over the time period 1955 through 2003. These results, summed across all three colonies at Midway Atoll, Laysan Island, and French Frigate Shoals in the Northwestern Hawaiian Islands, suggest the black-footed albatross population overall was stable or slightly increasing during that time period, with an annual population growth rate of 0.3 percent a year (Arata

et al.

2009, p. 46). Although positive, the authors note the observed growth rate of 1.003 is less than the natural annual growth rate estimate of 1.035 for the species. They attribute this difference of 3.2 percent in potential population growth to fishery mortality (Arata

et al.

2009, p. 46). In other words, the data indicate that the black-footed albatross population was stable or slightly increasing between 1955 and 2003, but that it was increasing at less than its potential annual growth rate. Wiese and Smith (2003, pp. 34-35) similarly concluded that the world population of black-footed albatross was stable, with an observed annual growth rate of 1.005 (based on demographic rates as published in Cousins and Cooper 2000 and Lewison and Crowder 2003), but also noted the population was growing at less than its estimated

potential annual growth rate of 1.04 (Wiese and Smith 2003, p. 33). The authors cautioned that, although the black-footed albatross population appeared to be stable, this reduced annual growth rate renders the population vulnerable to changes in their environment, especially in conjunction with sustained anthropogenic impacts (Wiese and Smith 2003, p. 35).

Japanese Islands

Breeding populations of black-footed albatross currently occur on Izu-Torishima (Torishima) Island in the Izu Islands, on nine islets in the Ogasawara islands within the Bonin Island complex, and on three islets in the Senkaku Islands (Kawakami

et al.

2006, p. 187; Chiba

et al.

2007, p. 5; Eda

et al.

2008, p. 109). Few data are available specific to the breeding population of the black-footed albatross in Japan. The Western Pacific Regional Fishery Management Council (Council) provided us with fledging success estimates for the Ogasawara Islands for 2009. The Council reported 801 chicks fledged, which is not directly comparable to the 967 nesting pairs in 2006 shown in table 2. They extrapolated these fledgling count data to estimate the number of nesting pairs, and concluded approximately 1,070 black-footed albatross nesting pairs were present on the Ogasawara Islands in 2009, which they interpreted as representative of an increase in the population. Because of the documented annual variability in nesting activity in black-footed albatross breeding colonies and lack of other supporting information, we believe extrapolation from a single year of fledging success data to an increase in the black-footed albatross population trend is inappropriate. However, Dr. Hiroshi Hasegawa of Toho University in Japan has additionally reported that the number of black-footed albatross chicks reared on Torishima Island has increased steadily between 1957 and 2010 (Figure 4) and that the populations on the Ogasawara and Senkaku Islands have also increased (Hasegawa 2010, pers. comm.).

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Marshall Islands and Wake Atoll

Black-footed albatrosses have infrequently been reported on Wake Island, a U.S. territory in the Marshall Islands archipelago in the central Pacific, an area from which they had been extirpated by feather hunters prior to World War II (Rice and Kenyon, 1962a, pp. 379-380; Rauzon

et al.

2008, pp. 15-16). Although a few birds have occasionally been observed nesting on Wake Island, any eggs laid were subsequently abandoned, and there have been no reports of black-footed albatross fledging here (Rauzon

et al.

2008, p. 15). These birds are attempting to breed and may be prospecting for future nesting sites on this island, but based on the available information, we conclude that at present there is no established breeding population of black-footed albatrosses on Wake Island or on any island, atoll, or reef in the nearby Marshall Islands (see Tables 2 and 3).

Mexican Islands

There have been a handful of reports recording intermittent nesting activity by black-footed albatrosses on Guadalupe and San Benedicto islands in the Revillagigedo Island archipelago off the tip of Baja California, Mexico (Pitman and Ballance 2002, p. 13). In 1999, Pitman and Ballance (2002, p. 13) recorded a single black-footed albatross standing among a group of Laysan albatrosses on Albatross Beach on San Benedicto Island. In 2000, they recorded an adult black-footed albatross sitting on an egg on the rim of Herrera Crater on San Benedicto Island (Pitman and Ballance 2002, p. 13). Also in 2000, but on Guadalupe Island, military personnel identified a black-footed albatross nesting area that was set apart from the Laysan albatross nesting area. They also reported seeing a “fully-feathered chick” at this same site in 1998, no breeding in 1999, and no breeding in 2000, although one pair of birds was reported “visiting” the site every afternoon in 2000 (Pitman and Ballance 2002, p. 13). Apart from these reports, no black-footed albatross have been recorded nesting on either San Benedicto or Guadalupe islands in the last 10 years, although they have been recorded visiting and possibly prospecting for nesting sites on both islands during breeding seasons (Henry 2007, pers. comm.; Hebshi 2010, pers. comm.). The International Union for Conservation of Nature (IUCN) Red List reports a population of 400 black-footed albatross exists on Guadalupe Island (IUCN Red List,

http://www.iucnredlist.org

). We note this report appears to be in error, as there is a known population of approximately 400 Laysan albatross on Guadalupe Island, but there are no black-footed albatross. In conclusion, at this time, there is no established breeding population of black-footed albatross on either San Benedicto Island or Guadalupe Island in Mexico, but birds may sporadically nest there and appear to be prospecting the islands for potential nesting sites (Naughton 2010, pers. comm.).

Population Trends and Projections

Noticeable declines in nesting activity during the 1990s generated concern for the health of the black-footed albatross population, and several population modeling efforts were undertaken to evaluate the cause of the decline and to estimate the worldwide status of the black-footed albatross. Cousins and Cooper (2000, entire), Lewison and Crowder (2003, entire), Wiese and Smith (2003, entire), Niel and LeBreton (2005, entire), Veran

et al.

(2007, entire), and Arata

et al.

(2009, entire) used the nest count data collected by the Service on French Frigate Shoals, Midway Atoll, and Laysan Island, estimates of bycatch rates from the domestic and international fisheries, estimates of adult survival, and other population parameters to analyze and project black-footed albatross population trends. Population projections specific to the Japanese breeding colonies of black-footed albatross are not available.

The conclusions regarding future black-footed albatross population trends based on these different modeling efforts are not easily comparable because of limited or nonexistent empirical data. The various researchers consequently had to rely on various assumptions, and these assumptions often varied between models, as did the methods. In part due to these differences in assumptions, the conclusions reached by the various models are not consistent, making it difficult to project the future population condition of the black-footed albatross with certainty. Here we briefly summarize and evaluate each of these efforts.

Cousins and Cooper (2000, entire) investigated the population parameter values available at the time of their analysis, evaluated changes in demographic rates such as adult and juvenile survival, and modeled effects of longline fishing activity on the black-footed albatross. They reported a mean adult survivorship rate of 0.923 (range 0.81-0.994) over the years 1961 to 1966, based on data from Midway Atoll, and stated that this estimate of adult survival was based on data collected when the Hawaii-based longline fishing fleet represented only a small fraction of the north Pacific fishing effort (Cousins and Cooper 2000, p. iv). They also noted that this adult survivorship rate may be an underestimate (Cousins and Cooper 2000, p. 50). They estimated more recent adult survivorship, based on the years 1991-1997, as in the range of 0.90-0.94 (Cousins and Cooper 2000, p. 50).

According to a predictive model that estimated the annual population growth rate based upon varying levels of mortality and adult and juvenile survival rates, Cousins and Cooper (p. 53) found black-footed albatross population trends were more sensitive to changes in survival than fecundity, and reported juvenile survival has to be 86 percent or higher to prevent a population decrease, assuming adult survivorship of 0.93 and fecundity of 0.25 fledglings per adult (note that this model utilized a combination of experimental rates from black-footed albatrosses and Laysan albatrosses, since data for black-footed albatrosses were limited at the time). However, the most recent values for black-footed albatross survivorship (adult survivorship 0.967 and juvenile survivorship 0.993; Arata

et al.

2009, p. 47) are higher than those for Laysan albatrosses, which were used in their models (adult survivorship 0.947 and juvenile survivorship 0.57; Cousins and Cooper 2000, p. 49).

Their models indicated the potential annual growth rate of the black-footed albatross population, without any bycatch loss, is in the range of 0 to 4 percent (annual growth rate, or lambda (λ) of 1.0 to 1.04) (Cousins and Cooper 2000, p. 56). In addition, they developed an estimate of potential biological removal—the maximum mortality that can be sustained before declines are observed—as 10,000 birds per year (Cousins and Cooper 2000, p. 57). Based on anecdotal evidence, they report the interactions of Japanese fisheries with black-footed albatross as insignificant (H. Hasegawa, Toho Univ., pers. comm., as cited in Cousins and Cooper 2000, p. 67). The demographic parameters and modeling efforts presented by Cousins and Cooper (2000, entire)) serve as the basis for some of the predictive models developed by several later researchers.

Lewison and Crowder (2003, entire) developed an age-structured matrix model. They based their longline fishing bycatch rates on published rates for the Hawaii and Alaska fisheries, and estimated annual fishing effort by international longline fleets (Lewison and Crowder 2003, pp. 774-746). Since their baseline population model was based on the demographic parameters reported by Cousins and Cooper (2000), the authors state that “double-dipping” (adding estimated fisheries bycatch to a demographic rate that already reflects mortality from fisheries) was not likely, based on their stated assumption that significant fisheries mortality was not occurring during the time period when the data used by Cousins and Cooper were collected (mid-1970s; Lewison and Crowder 2003, p. 747). The authors assigned three levels of mortality and age-based survival probabilities to evaluate the effect of longline fishing on the black-footed albatross. Population trajectories under all mortality levels resulted in projected declines over a 20-year period (Lewison and Crowder 2003, p. 748). According to these models, mortality from longline fishing exceeded the potential biological removal value developed by Cousins and Cooper (2000) (Lewison and Crowder 2003, p. 748).

The authors stated their estimates are likely conservative, since the reported bycatch estimates do not include the estimated 30 percent of birds caught in fishery operations that are scavenged or dislodged from the hooks prior to observation, and are, therefore, not counted as bycatch (Lewison and Crowder 2003, p. 751). In addition, they pointed out that due to the life-history characteristics of the black-footed albatross—longevity, delayed maturity, low fecundity—there is a lag in population response, and the impact of threats that may cause declines in adult survival may not be detectable for many years (Lewison and Crowder 2003, p. 751). The authors concluded that although declines had not been observed, the bycatch rates for black-footed albatross suggested population-level effects were likely (Lewison and Crowder 2003, p. 751).

Wiese and Smith (2003, pp. 29-31) also estimated black-footed albatross annual growth rates using an age-structured matrix model based on the published demographic parameters of Cousins and Cooper (2000) and Lewison and Crowder (2003, Table 1). However, unlike Lewison and Crowder (2003), they assumed incidental fishing mortality was already incorporated in the adult survival rate, based on their observation that longline fishing has occurred in the north Pacific since the mid-1900s, and thus would have been in place when the data serving as the basis for calculating that adult survival rate were collected (Wiese and Smith 2003, p. 30). Wiese and Smith's estimate of a potential annual growth rate of 1.04 in the absence of fisheries mortality is identical to the estimate presented by Cousins and Cooper (2000, p. 56). Wiese and Smith's results showed the population was stable with a stochastic annual intrinsic growth rate of 1.005 (range 0.990-1.018), and projected annual population growth rates of 0.98-1.04 percent over a period of 20 years based on known demographic values at the time of their analysis (Wiese and Smith 2003, p. 33 and Figure 4), indicating a stable population.

In addition, the authors found their model successfully fit real data (COSEWIC 2007, p. 29). Wiese and Smith (2003, p. 35) pointed out data collected during breeding bird censuses since 1992 and subsequent population projections do not support the projected decline that served as the basis for the IUCN designation of black-footed albatross as a vulnerable species (upgraded to “endangered” by the IUCN in 2003). However, they also emphasized the decreased annual population growth rate of the black-footed albatross, reduced below its maximum potential, renders the species vulnerable to additional stressors, even if the species is currently abundant, and they stressed the need for careful monitoring of colonies and the use of bycatch reduction measures in Canadian and international longline fisheries.

Niel and Lebreton (2005, entire) developed a model to estimate the annual maximal growth rate of a species from incomplete demographic data and used the black-footed albatross as a case study. They applied the population parameters developed by Cousins and Cooper (2000, entire) in their model and calculated a maximal annual growth rate of 1.059 (Niel and Lebreton 2005, p. 833). Additionally, they calculated the potential excess growth (used as an estimate of the maximum additional mortality the population could sustain on an annual basis without declining) of the population as 8,850 individuals. (It should be noted that Niel and Lebreton (2005) utilized the population parameters for the Laysan albatross presented in Cousins and Cooper (2003, p. 49; breeding age of 8.6 years and adult survivorship of 0.947) rather than those specific to the black-footed albatross, since Cousins and Cooper used the parameters for the Laysan albatross in their initial modeling efforts in the absence of data for the black-footed albatross (Cousins and Cooper 2000, p. 49)). It is not clear why they did so, since Cousins and Cooper (2000, p. 47) did provide an adult survivorship estimate specific to black-footed albatross, but it may be because Cousins and Cooper (2000, p. 50) believed their data likely underestimated adult survivorship of black-footed albatross. More recent estimates of black-footed albatross adult survivorship are 0.967 for the time period 1994-2002 (Arata

et al.

2009, p. 47), slightly greater than the estimate of 0.947 for Laysan albatross used by Niel and LeBreton (2005)). Based on their calculations, Niel and LeBreton (2005, p. 833) concluded the additional mortality associated with the longline fishery, based on an estimated mortality of 12,000 individuals a year during the 1990s, has a biologically significant impact on the growth potential of the black-footed albatross population.

Lacking reliable estimates of bycatch rates, Veran

et al.

(2007, entire) developed a model to quantify the relationship between albatross populations and longline fishing by using capture-recapture data to develop survival estimates, and investigated the relationship between fishing effort and black-footed albatross adult survival using principal components analysis. One of the key assumptions of their model was that the level of bycatch is proportional to fishing pressure; thus, they assumed mitigation measures were not in place to reduce incidental mortality from fisheries (Veran

et al.

2007, p. 4). Their adult survivorship estimates were based on capture-recapture data gathered between the years 1992-2003 on Tern Island in the Northwestern Hawaiian Islands (Veran

et al.

2007, p. 3). Their results suggested a significant negative relationship between adult survival and fishing effort (Veran

et al.

2007, p. 1). When fishing effort was high, adult survival was estimated to be 92 percent, which the authors described as low compared to other albatross species, and adult survival was related to fishing effort in a nonlinear fashion (Veran

et al.

2007, pp. 5-7). Inspection of the adult survivorship data presented for 17 albatross species shows that Veran

et al.'

s estimated 0.92 survivorship of the black-footed albatross is on the borderline between those albatross species that were categorized as being impacted by fisheries (range 0.84 to 0.91) and those not impacted by fisheries (range 0.926 to 0.98) (Veran

et al.

2007, Appendix S2). The authors estimated annual adult survival of black-footed albatross would be approximately 95 percent in the absence of fishing mortality (Veran

et al.

2007, p. 8).

Veran

et al.

(2007, p. 9) concluded the low adult survival probability during the study period, combined with the significant correlation with longline fishing, suggests an anthropogenically induced decline for the black-footed albatross population. However, their only reference to evidence of any decline in the breeding population is a citation to unpublished data from the Service for the years 1992 to 2004 (Veran

et al.

2007, p. 2); we note that more recent Service data for 1998 to 2009 indicate the black-footed albatross population is not in decline, but is stable or increasing at a rate of 0.93 percent a year (95 percent confidence interval (CI) 0.85 to 1.00; ACAP 2010, p. 5). (The Service used data from 1998 through 2009 because it reflects direct counts of breeding black-footed albatross on Laysan; we considered data from 1992 through 1998 less reliable as it reflects only estimates of breeding numbers, with resulting wide margins of error). In conclusion, Veran

et al.

(2007, p. 9) stressed the importance of efficient mitigation measures to reduce incidental mortality and maintain a sustainable survival probability for the black-footed albatross.

Arata

et al.

(2009) conducted a status assessment of the black-footed albatross, evaluated current population trends using linear regression and matrix models (both discussed above under “Current Population Status”), and projected future trends using population viability analyses (PVA), assuming current conditions but incorporating environmental and demographic stochasticity. The authors based their analyses on counts of nesting birds from Midway Atoll, Laysan Island, and French Frigate Shoals in the Northwestern Hawaiian Islands; counts were available for 11 years: 1923, 1957, 1992, and 1998-2005 (Arata

et al.

2009, p. 77). The survivorship rates presented and utilized by Arata

et al.

(2009, p. 47) were higher than those reported in earlier studies; for the years 1994 to 2002, they calculated an adult survivorship rate of 0.967 (compared to 0.926 for the years 1963 to 1982, and 0.892 for the years 1983 to 1993) and a juvenile survivorship rate of 0.993 (compared to 0.688 for 1963 to 1982 and 0.668 for 1983 to 1993). These rates suggest that both adult and juvenile survivorship may have increased from the mid-1990s to 2002, the last year covered in the survivorship estimates.

Arata

et al.

(2009, p. 46) estimated total fishery bycatch, including international fisheries, at 5,228 birds per year in 2005 and found this was within the mortality level that can be sustained by the black-footed albatross population without causing a decrease (Arata

et al.

2009, p. 46). Their calculated maximum potential biological removal rate was 11,980 birds per year (range 10,579-12,796) (Arata

et al.

2009, p. 47). All of their model scenarios indicated that when both the pelagic longline and pelagic driftnet fisheries were active during the 1980s the incidental mortality of black-footed albatross exceeded the potential growth capacity for the species (Arata

et al.

2009, Figure 4, p. 15), and they concluded that the closure of the high seas pelagic driftnet fishery in 1992 was critical to preventing further population declines for the black-footed albatross (Arata

et al.

2009, p. 46). In terms of the current conditions, the authors advised caution in interpreting results because there is such great uncertainty in the bycatch estimates and suggested that if the estimated bycatch level is doubled as a conservative safeguard for potentially underestimating bycatch, the resulting value approaches the potential biological removal maximum, and the upper 95-percent confidence limit exceeds that value (Arata

et al.

2009, pp. 46, 51).

Although Arata

et al.

(2009, p. 51) stated that fishery bycatch “may be causing a decrease in black-footed albatross populations,” it is not clear how they arrived at that conclusion since they offer no evidence of a population decrease, and their conclusions point to rangewide populations being stable or increasing for their period of analysis. We assume the authors meant that, given the uncertainty in bycatch estimates, a population decline might be expected if the worst-case scenario were realized and bycatch was actually twice as much as the estimate they used (see, for example, the discussion regarding the uncertainty of bycatch estimates, particularly with regard to international longline fisheries, on p. 67 of Arata

et al.

2009). Individual PVAs showed breeding colonies on Midway Atoll and French Frigate Shoals are stable or increasing, with projected annual population growth rates of 1.5 percent (95 percent CI 1.1 to 1.9) and 1 percent (CI 0.8 to 1.2) a year, respectively (Arata

et al.

2009, pp. 39, 41). In contrast, the population on Laysan Island is declining, with a negative annual growth rate of 1.3 percent (CI -1.7 to -0.9) per year (Arata

et al.

2009, p. 41).

Projections of future trends for all three colonies showed a high degree of uncertainty, with high probabilities of colonies both increasing and decreasing in the future, although in most cases the probability of future increases is greater than the probability of future decreases (Arata

et al.

pp. 39-45, 51). The authors concluded that, under conditions present in 2005, the black-footed albatross population is not at risk of a substantial decrease over the next 60 years (Arata

et al.

2009, p. 50). Overall, the decreases at Laysan Island appear to be offset by the positive growth observed at Midway Atoll and French Frigate Shoals, resulting in the overall stable or positive trend.

Arata

et al.

(2009, p. 50) reported that the assumption of zero bycatch prior to 1970, as assumed by the previous analyses of Cousins and Cooper (2000) and Lewison and Crowder (2003), is not supported by their model (Arata

et al.

2009, p. 46). They questioned the key assumption in the Lewison and Crowder (2003) model that led to the prediction of a rapid population decline for the black-footed albatross over the 60 years following their analysis, namely the assumption that there was no effect of fishery bycatch on adult survivorship at the time the data were gathered that served as the basis for demographic parameter estimates, in the 1960s and 1970s. Arata

et al.

(2009, p. 50) report that fishery effort data from the Ocean Fisheries Program indicate that fishery bycatch was in fact most likely significant during this time period (see Arata

et al.

2009, Figure 4, p. 15). If so, this would result in analyses such as those of Cousins and Cooper (2000) and Lewison and Crowder (2003) having inadvertently doubled the impact of fisheries bycatch, since bycatch effects would already be reflected in the survivorship rates used, but mortality from bycatch was then additionally imposed on the population in the models. This inadvertent doubling of bycatch effects would account for the prediction of particularly rapid population decreases. Arata

et al.

(2009, p. 79) point out the nonindependence between survival estimates and fishery bycatch levels since the pelagic fishery started in 1952, and caution that survival estimates affected by fishery mortality used in previous population assessments may have significantly influenced results.

We additionally received comments during the information solicitation period indicating this possible double-counting of fisheries mortality in the Lewison and Crowder (2003) model, and pointing out that it would have resulted in erroneous predictions of sharp population decline (

e.g.,

Harrison 2008, pers. comm., p. 9). More importantly, perhaps, we received a communication from the senior author of the Lewison and Crowder (2003) analysis, in which Dr. Rebecca Lewison points out that their paper had illustrated population-level trajectories for the black-footed albatross

if bycatch levels remained constant and bycatch was unmitigated,

and assuming the bycatch levels observed in the Hawaii fishery from 1994 to 2000 would continue over the 60 year time period of the projection. “This assumption has already been shown to be false,” Dr. Lewison stated, “There have been several NOAA regulations from 2001-2004 which have included initial and revised mitigation device requirements, improved performance specifications of mitigation devices, and spatial/temporal fishing closures. It is clear that mortality levels have dropped dramatically as a result” (Lewison 2007, pers. comm., p. 2). An accurate understanding of the Lewison and Crowder (2003) model has important consequences, as the severe population declines projected by that particular model led, at least in part, to the IUCN changing the status of the black-footed albatross from “vulnerable” to “endangered” in 2003 (IUCN 2011), a change which further played a key role in spurring the original petition to list the species under

the Act (EarthJustice 2004, p 2). The IUCN change in classification also apparently served as the basis for NatureServe to change the ranking of the black-footed albatross from G5 (globally secure) to G3/G4 (vulnerable) (NatureServe 2011).

Subsequent modeling efforts have produced different results. The modeling by Arata

et al.

(2009, pp. 50-51), which accounted for bycatch impacts in the observed demographic rates utilized in the models (Arata

et al.

2009, p. 79), did not project future declines in the black-footed albatross population. Wiese and Smith (2003, p. 30) likewise considered that the black-footed albatross had sustained mortality in the northeastern Pacific fisheries since the 1970s, and, therefore, considered the survival rate data collected during that time to represent a population already affected by incidental mortality due to fisheries; their model also did not support projections of a population decline (Wiese and Smith 2003, p. 35). We consider these models to provide more accurate projections of future population trends in the black-footed albatross since they avoid the issue of double-counting mortality from fisheries bycatch. However, it appears the conservation status of the black-footed albatross has not yet been updated in light of this new information by either NatureServe (2011; population trend information cites to IUCN 2000) or the IUCN, although the IUCN does note that its current categorization of the black-footed albatross is likely to be revisited pending the outcome of a review of the species' population status (IUCN 2011).

Population-level estimation of demographic parameters in black-footed albatrosses has proved difficult because of multiple factors, including band loss and variation in capture-recapture efforts (Doherty

et al.

2006, pp. 175-176). Until recently, the population monitoring program in the Hawaiian Islands consisted only of annual counts of breeding birds in three colonies at French Frigate Shoals, Midway Atoll, and Laysan Island. The program did not account for the proportion of nonbreeding birds in a year. A change in the count data from year to year could, therefore, reflect either a change in the total breeding population size or a change in the proportion of birds returning to breed in a given year (Naughton

et al.

2007, p. 15).

The Service has contracted with USGS's Patuxent Wildlife Research Center to review the black-footed albatross monitoring program implemented by the Service in the Northwestern Hawaiian Islands. A pilot study has been undertaken that is expected to provide information on adult survival, probability of a breeder skipping a year, and reproductive success (Arata

et al.

2009, p. 21; Naughton 2009, pers. comm.). These parameters are important for refining demographic models and determining population trends.

Summary Evaluation of Population Status and Trend Data

Following the end of feather hunting at nesting colonies, the world population of the black-footed albatross recovered from an estimated low of 17,800 breeding pairs in Hawaii and 200 breeding pairs in Japan in the early 1920s (Arata

et al.

2009, p. 35) to an estimate of 64,031 breeding pairs in Hawaii and 3,184 breeding pairs in Japan as of 2010 (ACAP 2010, Table 3, p. 4; Flint 2011a, pers. comm.). Our evaluation of the best available scientific data indicates the world population of the black-footed albatross is currently stable or slightly increasing, although population growth is below its potential maximum, likely due to the impact of incidental bycatch in fishery operations (Wiese and Smith 2003, p. 35; Niel and Lebreton 2005, p. 833; Arata

et al.

2009, p. 46). In the Hawaiian Islands, home to an estimated 95 percent of the breeding population of the black-footed albatross, a decrease in the number of breeding pairs on Laysan Island appears to be offset by increases at Midway Atoll and French Frigate Shoals, resulting in an overall positive trend and an increase of 0.93 percent annually for these three areas combined for the years 1998 through 2009 (ACAP 2010, p. 5). The nearly 40 percent reduction in the size of the colony on Laysan Island since the late 1950s (ACAP 2010, p. 7), however, does indicate cause for concern, as well as the need for further research to determine the underlying cause of this decline. In Japan, indications are that the number of breeding pairs has steadily increased over time (Cousins and Cooper 2000, p. 23; Arata

et al.

2009, p. 39; Hasegawa 2010, pers. comm.).

There is little doubt that incidental mortality from fisheries had a significant negative impact on black-footed albatross populations in the past (Niel and Lebreton 2005, p. 833; Arata

et al.

2009, p. 46), and recent analyses demonstrate a significant negative relationship between black-footed albatross survivorship and fisheries effort (Veran

et al.

2007, p. 1). Examination of estimated bycatch data over the past 50 years shows high numbers of black-footed albatrosses killed in the pelagic driftnet and longline fisheries, peaking with 15,290 birds in 1961 and again with 16,215 birds in 1988 (Arata

et al.

2009, p. 14). Past bycatch estimates ranged generally between approximately 6,000 and 10,000 birds a year, often exceeding the maximum potential biological removal value estimated for the black-footed albatross (Arata

et al.

2009, Figure 4, p. 15; p. 46). However, mortality of black-footed albatrosses was greatly reduced following the closure of the high seas driftnet fishery by a United Nations resolution in 1992 (ACAP 2010, p. 12) and implementation of regulatory bycatch measures in U.S. longline fleets in 1997 and 2002 (Arata

et al.

2009, p. 14, Figure 4; Moore

et al.

2009, p. 444, Figs. 3A and 3B). Bycatch of black-footed albatrosses in the Hawaii-based pelagic longline fishery has decreased from over 1,300 birds taken annually in 1999 and 2000 to less than 100 in 2007 (annual report on seabird interactions and mitigation efforts in the Hawaii longline fishery for 2007, Administrative Report, U.S. Dept. of Commerce, NOAA, NMFS, PIRO, April 2008). The increased survivorship probabilities observed for both adult (0.967) and juvenile (0.993) black-footed albatross since these measures have been in place, for the years 1994-2002, may reflect this significant reduction in mortality (Arata

et al.

2009, p. 47).

Attempts to project the future condition of the black-footed albatross population have produced inconsistent results. However, some of the past models that suggested incidental mortality from fisheries bycatch may exceed the level that can be sustained by the black-footed albatross population were based on demographic data gathered prior to both the high-seas driftnet moratorium (1992) and to regulatory bycatch reduction measures implemented in U.S. fisheries (1997, 2002) (

e.g.,

Cousins and Cooper 2000). It is not known what these models might project under current conditions, as these bycatch reduction measures have resulted in a significant decrease in incidental mortality of albatrosses (American Bird Conservancy 2008, pp. 7-9; Awkerman

et al.

2008; Arata

et al.

2009, pp. 14, 46; Moore

et al.

2009, p. 444; ACAP 2010, p. 12).

The model of Lewison and Crowder (2003) assumed bycatch mortality was constant, and the model of Veran

et al.

(2007) assumed no bycatch mitigation measures were in place; neither of these assumptions are met under present conditions since effective bycatch reduction measures have been put in place in the U.S. fleets (acknowledging the level of bycatch in international

fleets remains unknown and knowledge of bycatch in the U.S. North Pacific fleets is imperfect). That these assumptions are now known to be false has been acknowledged (Lewison 2007, pers. comm., p. 2). In addition, the studies of Cousins and Cooper (2000) and Lewison and Crowder (2003) appear to have used demographic parameters based on a potentially erroneous assumption of zero bycatch at the time the data on survivorship values were collected, resulting in likely exaggerated predictions of rapid population declines when mortality from bycatch was added to demographic rates that already reflected ongoing bycatch at the time the data were collected (Arata

et al.

2009, p. 46). When bycatch mortality is considered to be already reflected in the survivorship parameters utilized, models project stable or slightly increasing populations of the black-footed albatross (Wiese and Smith 2003, p. 24; Arata

et al.

2009, pp. 50-51).

We have evaluated the various predictive models for the black-footed albatross, and agree with Arata

et al.

(2009, p. 50) that the model of Lewison and Crowder (2003, entire) most likely overestimated bycatch impacts by adding mortality from fisheries bycatch on demographic parameters that already reflected bycatch impacts. This inadvertent doubling of the mortality rate from bycatch would have resulted in the projection of precipitous population declines for the species. We base our conclusion on the data from the Ocean Fisheries Program presented in Figure 4 of the report of Arata

et al.

(2009, p. 15), which shows significant levels of bycatch mortality of black-footed albatrosses from commercial fisheries occurring from the mid-1950s through the early 1990s. It follows that demographic parameters based on data collected during the mid-1970s, used by Lewison and Crowder (2003, p. 747) in their efforts, would have reflected ongoing levels of bycatch at that time.

Other models based on the assumption that bycatch mortality is already reflected in demographic data collected during this time period (and, therefore, did not incorporate further bycatch effects into simulations) project future black-footed albatross populations to be relatively stable or even slightly increasing in size under conditions present at the time of the analyses (Arata

et al.

2009, pp. 46, 50-51; Wiese and Smith 2003, p. 35). Although stable, the results of these models also show that black-footed albatross populations are growing at less than their potential growth rate, most likely due to bycatch mortality (Arata

et al.

2009, pp. 46, 50-51; Wiese and Smith 2003, p. 35). Because the models of Wiese and Smith (2003, entire) and Arata

et al.

(2009, entire) avoid double-counting mortality from fisheries bycatch, we consider them to provide the most reliable projections of population trends for the black-footed albatross.

All studies we examined acknowledged the vulnerability of the black-footed albatross to bycatch mortality, and all indicated that declines may occur in the future if bycatch levels are greater than estimated (

e.g.,

Arata

et al.

2009, p. 47). At this point in time, however, we do not see any evidence that the black-footed albatross population is in decline, and current data suggest recent bycatch reduction measures have been effective in increasing survivorship (Arata

et al.

2009, p. 65). Advances in avoiding seabird bycatch include methods such as the use of streamer lines, which are found to reduce incidental mortality of albatrosses by nearly 100 percent (Melvin

et al.

2006, p. 4). Other seabird avoidance measures under evaluation include, but are not limited to, side setting, night setting, underwater setting, towing buoys, using heavier branch line weights, and dying bait (

e.g.,

Gilman

et al.

2005, Table 1, pp. 40-41; Gilman

et al.

2008, p. 12). Such measures are now required in most U.S. fisheries (some smaller vessels are exempted; for details, see the discussion under Factor D “The Inadequacy of Existing Regulatory Mechanisms,” below).

Although the conservation measures implemented thus far have been highly effective in reducing the incidental mortality of black-footed albatrosses (Arata

et al.

2009, pp. 14, 46; Moore

et al.

2009, p. 444; ACAP 2010, p. 12), great uncertainty surrounds the actual level of bycatch from international longline fisheries, and the true impact of those fisheries is currently unknown (Arata

et al.

2009, p. 47). On the whole, however, the demonstrated effectiveness of current bycatch mitigation measures, where mandated, in conjunction with: (1) Indications that past models predicting severe declines may have inadvertently overestimated the impacts of fishery bycatch or operated under assumptions that are now known to be false; (2) analyses that show populations are collectively stable or increasing; and (3) recent modeling that projects no substantial decreases over the next 60 years if current mitigation measures remain in place (and assuming continuation of other conditions present in recent years), all lead us to the conclusion that black-footed albatross numbers are stable at present rangewide, in the Hawaiian Islands and in the Japanese Islands.

Summary of Factors Affecting the Species Throughout Its Range

Section 4 of the Act (16 U.S.C. 1533) and implementing regulations (50 CFR part 424) set forth procedures for adding species to the Federal Lists of Endangered and Threatened Wildlife and Plants. A species may be determined to be an endangered or threatened species due to one or more of the five factors described in section 4(a)(1) of the Act:

(A) The present or threatened destruction, modification, or curtailment of its habitat or range;

(B) Overutilization for commercial, recreational, scientific, or educational purposes;

(C) Disease or predation;

(D) The inadequacy of existing regulatory mechanisms; or

(E) Other natural or manmade factors affecting its continued existence.

Listing actions may be warranted based on any of the above threat factors, singly or in combination.

In considering those factors that might constitute threats, we must look beyond mere exposure of the species to the factor to determine whether the species responds in a way that causes actual impacts to the species. If there is exposure to the factor, but no response, or only a positive response, that factor is not a threat. If there is exposure and the species responds negatively, the factor may be a threat, and we then attempt to determine how significant that threat may be. All species face some degree or source of threat. We consider a threat to be “significant” if that threat may drive or contribute to the risk of extinction of the species such that the species warrants listing as threatened or endangered as those terms are defined by the Act. The mere identification of factors that could impact a species negatively is not sufficient to compel a finding that listing is appropriate. We require evidence that these factors are operative threats that act on the species to the point that the species meets the definition of endangered or threatened under the Act; that is, the species is presently in danger of extinction throughout all or a significant portion of its range (endangered), or is likely to become endangered within the foreseeable future (threatened).

In making this finding, we have considered and evaluated the best available scientific and commercial information, including information received in response to our 90-day finding (72 FR 57278, October 9, 2007) and received or acquired in response to

our August 26, 2009, notice (74 FR 43092) reopening the information collection period. Below we summarize the information regarding the status and threats to the black-footed albatross across the range of the species in relation to the five factors in section 4(a)(1) of the Act.

Factor A. The Present or Threatened Destruction, Modification, or Curtailment of Its Habitat or Range

As with other members of the family Diomedeidae, black-footed albatrosses feed offshore or pelagically, and return to land only to breed. In this section, we describe and evaluate various conditions in relation to the present or threatened destruction, modification, or curtailment of the marine and terrestrial habitats and range of the black-footed albatross, including: Military activities; volcanic activity; natural gas development; invasive plant species; and conditions related to climate change, including sea level rise and coastal inundation, tropical storm frequency and intensity, impacts to marine productivity, and ambient temperature. Each of these topics is discussed in relation to the two breeding populations (Hawaiian Islands and Japanese Islands) that collectively constitute the entire breeding range of the species.

Military Activities

Historical occupation by armed forces on islands important to black-footed albatross breeding populations occurred during much of the twentieth century, mostly associated with World War II. Activities associated with warfare and development of military infrastructure throughout black-footed albatross breeding habitat, including the intentional modification of breeding habitat to reduce albatross nesting activity, negatively impacted albatross colony size in the past (Rice and Kenyon 1962b, p. 384). However, little information exists with which to deduce the original size of the black-footed albatross colonies on these islands because there were very few early quantitative studies.

Northwestern Hawaiian Islands.

French Frigate Shoals and Midway, Kure and Johnston atolls all supported armed forces stations or sustained military activities during World War II (Rice and Kenyon 1962b, pp. 366-378). In addition to the obvious disruptive impact of active warfare during that time, black-footed albatross populations were severely diminished by the development of military bases that led to loss and degradation of nesting habitat and large-scale albatross eradication programs intended to reduce interference of the birds with aircraft operations (Arata

et al.

2009, p. 17; ACAP 2010, p. 6). By 1996, management of nearly all of the Northwestern Hawaiian Islands was transferred to the jurisdiction of the Service, and active military impacts had ceased. The black-footed albatross' breeding sites on Midway Atoll National Wildlife Refuge (NWR), as well as Pearl and Hermes Reef; Lisianski, Laysan, Necker, and Nihoa islands; and French Frigate Shoals, which are part of the Hawaiian Islands NWR, are now all protected from human-related habitat modification or destruction because these islands are under the jurisdiction of the Service's NWR system.

The mission of the NWR System is to administer a national network of lands and waters for the conservation, management, and where appropriate, restoration, of the fish, wildlife, and plant resources and their habitats within the United States for the benefit of present and future generations of Americans (U.S. Fish and Wildlife Service (USFWS) 2009b). Management of Kure Atoll was transferred from the U.S. Coast Guard to the State of Hawaii in 1993. Breeding sites for the black-footed albatross on Kure Atoll are protected from human-related habitat modification or destruction because this atoll now is a State wildlife sanctuary and is managed by the Hawaii Department of Land and Natural Resources (HDLNR) for the conservation and protection of indigenous wildlife, including seabirds (Hawaii Administrative Rules Title 13, Subtitle 5, Part 2, Chapter 125, sections 1-7). Further military impacts to black-footed albatross breeding habitat are unlikely in light of the transfer of the military lands to the Service and State, as described above.

Future military activity on these lands is further constrained by the 2006 establishment of the Northwestern Hawaiian Islands Marine National Monument (renamed Papahanaumokuakea Marine National Monument (PMNM) in 2007), which encompasses all of the islands, atolls, reefs, shoals, banks, and seamounts from 50 mi (80 km) east of Nihoa Island to 50 mi (80 km) west of Kure Atoll, and waters 50 mi (80 km) on either side of the lands. The co-trustees of the area are the Department of the Interior through the Service; the Department of Commerce through the National Oceanic and Atmospheric Administration (NOAA); and the State of Hawaii through the HDLNR. PMNM management is also accomplished in coordination with the State Office of Hawaiian Affairs. Within the boundary of the PMNM are two National Wildlife Refuges: Hawaiian Islands NWR and Midway Atoll NWR; the State Seabird Sanctuary at Kure Atoll; the State Marine Refuge; and the Northwestern Hawaiian Islands Coral Reef Ecosystem Reserve. Current uses are limited primarily to management activities by jurisdictional agencies, research, education, Native Hawaiian practices, a small-scale commercial bottomfishing and pelagic trolling operation, and a small number of recreational trips and visits to historical sites at Midway Atoll. Although military activities are not expressly prohibited within PMNM, the management regulations do require that all activities and exercises of the Armed Forces shall be carried out in a manner that avoids adverse impacts on monument resources and qualities, to the extent practicable and consistent with operational requirements (71 FR 51138; August 29, 2006). We have no reason to anticipate any active military operations within the PMNM.

Kaula Island and Lehua Island.

Kaula Island has been under U.S. Navy control since 1965 and is still used for munitions training. In 1977, Kaula Island was designated a State Seabird Sanctuary by the State of Hawaii (U.S. Navy 2009, unpubl.). Currently the Navy uses the southeastern portion of the Kaula Island for inert ordnance and gunnery activities, and it was previously used as a practice range for air-to-surface and surface-to-surface weapons delivery. Black-footed albatrosses have been observed on Kaula Island as recently as 1998 (USFWS 2009a; U.S. Navy 2009, unpubl.), but the last breeding data collected from a 1993 survey reported a breeding population of only three pairs (ACAP 2010, p. 4). Because of concerns regarding bird-aircraft hazards and unexploded ordnance, access to the island for bird surveys or management has been denied (U.S. Navy 2009, unpubl.). Lehua Island is administered by the U.S. Coast Guard and managed by the State of Hawaii as a State Seabird Sanctuary. No current military activities occur on this island, and none are anticipated.

Japanese Islands.

Torishima Island has been a protected national natural monument since 1965, when it was still under U.S. authority, and can be visited only by research scientists with special permission. In the Ogasawara Islands, Muko-jima is known to have been occupied during World War II by a Japanese garrison that presumably “wiped out” whole bird colonies for use as a food source (Austin 1949, pp. 290-291). The Senkaku Islands were used by the U.S. Navy as maneuver areas.

Following World War II, all of the western Pacific islands were controlled by the United States. However, in 1972, all of the islands were returned to Japan, although Taiwan and the People's Republic of China (China) both claimed sovereignty to the Senkaku Islands, and this is still under dispute (Senkaku Islands 2009).

Natural reoccupation or recolonization by black-footed albatrosses since World War II has occurred on Torishima Island, the Ogasawara Islands (Muko-jima Island and Haha-jima Island), and the Senkaku Islands (Arata

et al.

2009, p. 39). The Ogasawara Islands are now part of Japan's Ogasawara National Park, and current protective management of the islands likely precludes future military activities.

In summary, significant military activity is not currently taking place anywhere within the range of the black-footed albatross, and we have no reason to anticipate any increase in future military activity. Therefore, military activity does not pose a threat to the black-footed albatross in relation to the present or threatened destruction, modification, or curtailment of its habitat or range rangewide, in the Hawaiian Islands, or in the Japanese Islands.

Volcanic Activity

Within the nesting range of the black-footed albatross, volcanic activity in historical times is recorded only from Torishima Island, where a 1903 volcanic eruption occurred during the nonbreeding season for several species of albatrosses, so that the only apparent effect was to destroy part of their nesting habitat. By 1930, it was apparent that many birds had returned and were breeding on the island, as human harvesting of all the albatross species was resumed by settlers. The volcano erupted again in 1939, burying most of the former breeding grounds and making them uninhabitable for the birds. The main crater overflowed once more in 1941, closing the natural anchorage that had allowed free access to human hunters in the past. When visited in 1949, the island was described as “birdless” (Austin 1949, p. 289). The island was again naturally reoccupied by black-footed albatrosses subsequent to this eruption, growing from a count of 6 chicks in 1957 to 914 chicks by 1998 (H. Hasegawa, unpublished data, as cited in Cousins and Cooper 2000, p. 23). Volcanic activity on Torishima Island was last recorded in 2002, with volcanic ash and rock blanketing the central portion of the island. The following year, surveys resulted in an estimate of 2,150 breeding pairs of black-footed albatross on Torishima (ACAP 2010, p. 4), demonstrating that the breeding population was largely unaffected by this most recent event.

Given this history, it is likely that Torishima Island will continue to experience volcanic activity. The evidence from past events suggests that black-footed albatrosses may survive such an event, as they have in the past, since at any given time approximately 75 percent of the birds are at sea and, therefore, are likely to be absent at the time of a volcanic eruption or other catastrophic event (Finkelstein

et al.

2010, p. 328). Past reoccupation of Japanese islands by black-footed albatrosses has occurred subsequent to volcanic events as well as recolonization following extirpation of colonies due to military activities during World War II. Therefore, if the nesting population should be eliminated from the island due to volcanic activity impacts on nesting habitat, as has apparently occurred in the past, the historical evidence suggests that natural reoccupation of the island is probable assuming no other substantial changes in present conditions. In addition, Torishima might also be recolonized by birds from the nearby Ogasawara Islands.

Some researchers have suggested this scenario to be unlikely as movement of black-footed albatrosses between colonies is typically low (

e.g.,

Finkelstein

et al.

2010, p. 323). However, we believe natural reoccupation is likely, based on past evidence of several separate reoccupation events, although we acknowledge the population would likely suffer reduced productivity for several years following a catastrophic volcanic event. Torishima provides nesting habitat for 3.5 percent of the rangewide population and is the only nesting island for black-footed albatrosses with an active volcano. Most birds nesting on Torishima likely would be at sea if there were an eruption, and based on past history it is reasonable to assume the island would be reoccupied over time following any such event. Therefore, we conclude that volcanic activity does not pose a threat to the black-footed albatross in relation to the present or threatened destruction, modification, or curtailment of its habitat or range in the Hawaiian Islands, the Japanese Islands, or rangewide.

Natural Gas Exploration

Exploration for natural gas has recently become a potential issue for birds on the Senkaku Islands. A dispute, primarily between Japan and China but also including Taiwan, over the territorial jurisdiction of the Senkaku Islands has been ongoing since the summer of 1970 (Cheng 1973-1974, p. 221; Downs and Saunders 1999, p. 124). Although this dispute originated in 1945 following World War II, it escalated in the 1970s when potential undersea natural gas reserves off the continental shelf near the Senkaku Islands became an economic issue. However, there is no firm evidence that commercially exploitable petroleum reserves exist in the area (Downs and Saunders 1999, p. 124). Furthermore, it has been suggested that multinational petroleum companies have little interest in drilling near the Senkaku Islands because of difficult terrain, political uncertainty, existence of unexploded ordnance from use of the islands as a target range, and doubts about whether any reserves that might exist can be commercially exploited in viable terms (Downs and Saunders 1999, p. 124). Regardless of the outcome of the territorial dispute and the unlikely progress of gas exploration, we have no information to indicate that such development of natural gas resources in the area of the Senkaku Islands would potentially modify or destroy black-footed albatross nesting or foraging habitat. The black-footed albatross population of the Senkaku Islands comprises less than 0.1 percent of the rangewide population, and less than 2 percent of the breeding population of black-footed albatross in the Japanese Islands (56 breeding pairs; ACAP 2010, p. 4). Thus, even if such development were to occur and impact habitat on the Senkaku Islands, it would likely not pose a significant threat to the Japanese Islands population. It appears unlikely that gas exploration will occur in the Japanese Islands because: (1) There is no strong evidence that such resources exist; (2) commercial interest to develop these resources is weak, even if they were found to exist; and (3) sovereignty of the Senkaku Islands continues to be in dispute. Therefore, we conclude that natural gas exploration off the Senkaku Islands does not pose a threat to the black-tailed albatross in relation to the present or threatened destruction, modification, or curtailment of its habitat or range, including across its entire range, in the Hawaiian Islands, or in the Japanese Islands.

Invasive Plant Species

Many plant species have been introduced to the Hawaiian Islands, and of these

Verbesina encelioides

(golden crown-beard) has been identified as the greatest threat to black-footed albatross

nesting habitat (Naughton

et al.

2007, p. 12).

Verbesina encelioides

is well established on Kure Atoll, Midway Atoll, and Pearl and Hermes Reef, where it inhibits native plant growth (Shluker 1999, p. 4; Naughton

et al.

2007, p. 17). It is a woody herb that forms tall, dense stands, which can reduce access to nesting habitat of ground-nesting birds, including the black-footed albatross. Dense growth of

V. encelioides

can entangle black-footed albatross chicks or prevent parents from locating and feeding chicks (Shluker 1999, p. 4; Flint 2010, pers. comm.). It also restricts windspeed at the nest sites, potentially reducing the ability of adult and juvenile birds to thermoregulate using convective cooling (Flint 2010, pers. comm.). The Service and the HDLNR have implemented programs to control and eradicate

V. encelioides

on Midway Atoll and Kure Atoll (Shluker 1999, pp. 4-7; Flint 2010, pers. comm.), where approximately 43 percent of the rangewide black-footed albatross population breeds (see Table 2). In 2003, the Service and the HDLNR increased efforts to reduce the extent and spread of this invasive plant on Midway Atoll, including hand-pulling, mowing, and herbicide application.

In addition to

Verbesina encelioides,

other nonnative plant species occur in the Northwestern Hawaiian Islands, including

Casuarina equisetifolia

(common ironwood), a nonnative tree that has been identified as a threat to ground-nesting seabirds on Midway Atoll (Naughton

et al.

2007, p. 12). Like

V. encelioides,

the dense growth of

C. equisetifolia

around black-footed albatross nest sites can block the wind and thereby reduce the potential for convective cooling. Growing as they do in an area normally devoid of tall vegetation, these trees can potentially interfere with the flight of long-winged birds such as albatrosses. The trees also may break off or fall onto ground-nesting birds during wind storms; nesting Laysan albatrosses and chicks were killed on Midway in January 2011 by falling ironwood trees and flooding (ACAP 2011).

Casuarina equisetifolia

is also subject to a control program (Flint 2010, pers. comm.). Furthermore, the Papahanaumokuakea Marine National Monument Plan (PMNM Plan) has incorporated a nonnative species action plan to identify, control, eradicate, and avoid the introduction of new nonnative species to the PMNM (NOAA

et al.

2008, pp. 201-214).

The number of birds nesting on Midway Atoll has been relatively constant since 1992 (USFWS, unpubl. data) and has increased each year between 1999 and 2005 (Arata

et al.

2009, p. 36), so

V. encelioides

and

C. equisetifolia

as currently controlled do not appear to have significant negative impacts on the availability of black-footed albatross nesting habitat. Also, while standardized annual nest counts are not conducted on Kure Atoll and Pearl and Hermes Reef, a program to control

Verbesina

has been initiated on Kure Atoll (Flint 2010, pers. comm.). While uncontrolled growth of

V. encelioides

and

C. equisetifolia

would likely have negative impacts on habitat and thus possibly on the black-footed albatross population, based on the evidence from current control efforts, we anticipate these and expected future levels of control will continue to reduce and limit these impacts to the extent that these nonnative plants do not pose a significant threat to the black-footed albatross.

We found no information regarding nonnative plants within the nesting range of the black-footed albatross on the Japanese Islands, and have no evidence indicating that nonnative plants pose any threat to the black-footed albatross or its breeding habitat on the Japanese Islands.

Therefore, based on our evaluation of the best available scientific and commercial data, we conclude that invasive plants do not pose a significant threat to the black-footed albatross in relation to the destruction, modification, or curtailment of habitat or range of the species in the Hawaiian Islands, the Japanese Islands, or rangewide. In the section below, we further consider the potential spread of invasive plants in relation to conditions related to climate change.

Effects Related to Climate Change

The anticipated impact of climate change on black-footed albatross habitat, ecology, and life history in tropical and subtropical terrestrial and marine ecosystems is complex. In this section we begin with a general overview of climate change projections, followed by our evaluation of the potential response of the black-footed albatross to possible changes in their nesting and foraging habitat related to climate-related changes in sea level, coastal inundation, and storm events. We then consider changes in foraging habitat related to altered marine productivity that could occur in relation to climate change, and possible physical effects to the black-footed albatross related to changes in ambient temperatures.

Climate Change Overview

Consideration of the effects of climate change is a component of our analyses of species under the Act. Here we provide a brief overview of the general topic of climate change as a way of providing a broad context for the more detailed consideration that follows with respect to the black-footed albatross.

Described in general terms, “climate” refers to average weather conditions, as well as associated variability, over a long period of time (

e.g.

decades, centuries, or thousands of years). Climate variables most often described are temperature and precipitation, and the typical period for calculating the mean of these properties is 20 or 30 years. The term “climate change” thus refers to a change in the state of the climate (whether due to natural variability, human activity, or both) that can be identified by changes in the mean or variability of its properties and that persists for an extended period—typically decades or longer. (See Intergovernmental Panel on Climate Change (IPCC), 2007, pp. 30, 78, for technical definitions that are the basis for our description of these terms.)

Analyses of observed trends in climate demonstrate that climate change is occurring, as illustrated by examples such as an increase in the global mean surface air temperature (SAT) (“global warming”), substantial increases in precipitation in some regions of the world and decreases in other regions, and increases in tropical cyclone activity in some oceanic areas (IPCC 2007, p. 30). Because relatively small but sustained changes in temperature can have substantial direct and indirect effects on natural processes and human populations, temperature is one of the most widely used indicators of climate change. Based on extensive analyses, the IPCC concluded that warming of the global climate system over the past several decades is “unequivocal” (IPCC 2007, p. 2). These changes in global climate are affecting many natural systems (see IPCC 2007, pp. 2-4, 30-33 for global and regional examples, and Global Climate Change Impacts in the United States (GCCUS) 2009, pp. 27, 79-88, for examples in the United States).

Analyses of natural variability in climate conditions and the effects of human activities led the IPCC to conclude that most of the increase in global mean surface air temperature that has been observed since the mid-20th century is very likely due to the observed increase in greenhouse gas (GHG) concentrations related to human activities, particularly emissions of CO

2

from fossil fuel use (IPCC 2007, p. 5 and Figure SPM.3). Extensive analyses point to continued changes in climate and considerable efforts are occurring to make projections of the magnitude, rate,

and variability of future changes and to understand the mechanisms underlying them, including the role of greenhouse gases.

Projections by the IPCC in 2007 for climate change for the earth as a whole and for broad regions were based on simulations from more than 20 Atmospheric-Ocean General Circulation Models used in conjunction with various scenarios of different levels and timing of greenhouse gas emissions (Christensen

et al.

2007, pp. 847-917; Meehl

et al.

2007, pp. 753-796; Randall

et al.

2007, pp. 596-599). The emissions scenarios were developed in the late 1990s and described in the Special Report on Emissions Scenarios (SRES) published in 2000 (Carter

et al.

2007, p. 160, and references therein). The scenarios span a broad range of potential GHG emissions over the coming decades based on a wide spectrum of economic, technological, and human demographic possibilities for the planet; the SRES made no judgment as to which of the scenarios are more likely to occur, and although they cover a very broad range it is possible that emissions could be higher or lower than the range covered by the scenarios.

The IPCC's projections of change in global mean warming (global annual mean surface air temperature (SAT)) and how they differ over time across emissions scenarios as compared to the observed SAT from1980-1999, are described by Meehl

et al.

(2007, pp. 760-764). Several key points emerge from their projections. First, the projected changes in magnitude of warming are similar under all emissions scenarios to about 2030 and to some degree even to about mid-Century although more divergence is evident then, and the divergence continues to increase over time,

i.e.,

in the near-term the projections differ by only 0.05 °C (0.09 °F), but by the last decade of the century the difference across scenarios is 1.6° C (0.9 ° F); as noted by Cox and Stephenson (2007, p. 208), total uncertainty in projected decadal mean temperature is lowest 30 to 50 years in the future. Second, the magnitude of projected warming increases across each scenario, including the lowest emission scenario. Under the lowest emission scenario, annual man SAT change is 1.19 ° F (0.66 °C) for 2011-2030 and 2.32 ° F (1.29 ° C) for 2046-2065 (See Meehl

et al.

2007, p. 763, Table 10.5). Third, the pattern of projected increases is relatively consistent whether considering the average across all models for a given scenario or the projections from the individual models, including consideration of ± one standard deviation around the mean projection for each scenario (see Meehl

et al.

2007, pp. 762-763, Figures 10.4 and 10.5, and Table 10.5). Thus although differences in projections reflect some uncertainty about the precise magnitude of warming, we conclude there is little uncertainty that warming will continue through the end of century, even under the lower emissions scenario. We note also that more recent analyses using additional global models and comparing other emissions scenarios have resulted in projections of global temperature change that are similar to those reported in 2007 by the IPCC (Prinn

et al.

2011, pp. 527, 529).

While projections from global climate model simulations are informative, their resolution is coarse and it is helpful to have higher-resolution projections that are more relevant to the spatial scales used for various assessments involving climate change. Various methods to “downscale” climate information have been developed to generate projections that are more specific to regional or relatively local areas (see Glick

et al.

2011, pp. 58-61 for a summary description of downscaling). In conducting status assessments of species, the Service uses downscaled projections when they are the best scientific information available regarding future climate change.

In the case of marine areas, however, adequate procedures for downscaling are still under development, thus global projections for various conditions related to climate change (

e.g.,

sea and land surface temperatures, precipitation, storm frequency and intensity, marine productivity, and ocean acidification) are used for marine areas and small islands within them, including the Northwest Pacific Islands. Efforts are currently underway by the Pacific Islands Climate Change Cooperative and climate modelers at the University of Hawaii to develop regional models that will increase our understanding of climate change effects specific to the Pacific Islands. However, this information is not yet available to us. In most cases, therefore, global projections of future climate conditions constitute the best available scientific information available for purposes of our analyses for this finding.

Projections of Sea Level Rise

On a global (eustatic) scale, the main factors currently contributing to sea level rise are thermal expansion of warming ocean water, water input to oceans from the melting of ice sheets, glaciers, and ice caps, and the addition of water from terrestrial systems (United Nations (UN) 2009a, p. 26). The IPCC's model-based projections of global average sea level rise for the last decade of this century, as compared to the average for 1980-1999, ranged from 0.59 ft to 1.94 ft (0.18 m to 0.59 m) across various emissions scenarios (Meehl

et al.

2007, p. 812). This projection includes contributions from ocean thermal expansion, melting of glaciers and ice caps, and limited contributions from ice sheets; however, it did not include the possible contribution from relatively rapid melting of the Greenland and West Antarctic Ice Sheets. Several recent scientific publications have addressed problems that the IPCC's approach had in accounting for the observed level of sea level rise in the late 20th and early 21st centuries, and yielded new projections which reflect the possibility of rapid contributions from ice sheet dynamics beyond surface melting (see summaries by Church

et al.

2010, Rahmstorf 2010, and Nicholls

et al.

2011). Table 3 gives the ranges from these recent projections, along with the range given by the IPCC for purposes of comparison.

Table 3—Projected Ranges of Global Average Sea Level Rise for the 21st Century, Including the IPCC Projection (Meehl

et al.

2007) for Comparison

Projected range of global mean sea level rise feet (meters)

Source

0.59-1.94 ft, (0.18-0.59 m)

Meehl

et al.,

2007 (IPCC), pp. 820-822,Table 10.7.

2.6-6.6 ft, (0.8-2.0 m)

Pfeffer

et al.,

2008, p. 1340.

2.46-6.23 ft, (0.75-1.90 m)

Vermeer & Rahmstorf 2009, p. 21530.

2.36-5.25 ft, (0.72-1.60 m)

Grinsted

et al.,

2010, pp. 469-470.

2.0-5.3 ft, (0.6-1.6 m)

Jevrejeva

et al.,

2010, L07703, p. 4.

3-4 ft (0.9-1.2 m)

(GCCUS) 2009, p. 25.

As shown in Table 3, the ranges of recent projections of sea level rise all indicate substantially higher levels than the projection by the IPCC in 2007. They also show a much larger difference (approximately 3 to 4 ft (0.9 to 1.2 m)) from the low to the high ends of the ranges, which indicates the magnitude of global mean sea level rise at the end of this century is still quite uncertain.

In their review of sea level rise projections, Nicholls

et al.

noted that the earlier acceleration of some of the southeast Greenland glaciers had reversed by 2006, adding to uncertainty about whether the recent rates of mass loss are temporary and the extent to which they should be extrapolated into the future; they concluded that the upper part of the projected ranges of global sea level rise are possible but not likely to occur (Nicholls

et al.

2011, pp. 165, 168). Lowe and Gregory (2010, p. 4) similarly concluded that global mean sea level rise by the end of the century is “almost certain to be below two metres and that there is currently very little evidence to suggest that increases at the top of this range are likely.” Church

et al.

(2010, p. 411) reported that new information from satellite-based data for 2002-2009 indicates an accelerating contribution to sea level rise from both the Greenland and Antarctic ice sheets, but that “improved understanding of the processes responsible for ice-sheet changes are urgently required to improve estimates of the rate and timing of 21st-century and longer-term sea-level projections.” Similarly, Nicholls and Cazenave (2010, p. 1519) state “The extent of future SLR [sea level rise] remains highly uncertain—more so than in 2007, when the IPCC AR4 was published” and they call for additional analyses to focus on understanding ice sheet instabilities and other processes drive sea level rise.

Viewed from broad regional and particularly more local perspectives, the picture is further complicated by the fact that sea level rise is not uniform around the world and deviations from the observed global mean of sea level rise have been substantial in some areas. The fact that future sea level change will not be the same everywhere has been characterized by Milne

et al.

(2009, p. 471) as “one of the few statements that can be made with certainty.”

The considerable uncertainty about the magnitude of global average sea level rise by the end of the century is additionally complicated by the variability in sea level change observed in different parts of the world. This includes differences in open oceans, such as non-uniform changes in temperature and salinity and differences in ocean circulation patterns; the contributions of various factors to relative sea level change at regional scales are not fully understood and different contributions may dominate depending on the geographic location (Bindoff

et al.

2007, p. 409). A recent analysis aimed at providing a better understanding of sea level change at regional scales indicates that the entire range of the black-footed albatross is within a very broad ocean region where sea level rise by the end of this century is projected (under each of three emissions scenarios) to be higher than the global mean, and Hawaii is expected to have slightly higher rise than the global average (Slangen

et al.

2011, pp. 9-15). This analysis included numerous assumptions (including assumptions about changes in ice mass dynamics) and the authors made the point that the absolute values presented in their study required careful interpretation (Slangen

et al.

2011, p. 16).

Different rates of sea level rise observed locally add further complexity to the evaluation of this factor. Specifically, Honolulu, on the island of Oahu, and Hilo, on the island of Hawaii, have had different observed trends in sea level rise since the mid-1940s, although the relative differences in the rate of sea level rise between these Hawaiian islands have been more limited since the mid-1970s; these differences may be related to variations in both space and time in land motion (subsidence, uplift), and it may be related to interdecadal variations in upper ocean temperatures (Caccamise

et al.

2005, L03607, entire). Regardless of the cause(s) of the difference, this information adds to our caution in interpreting global sea level rise projections in our analysis of potential effects on the black-footed albatross and its habitat at a more localized scale.

In addition to reporting a projected range of sea level rise for the end of the century, Jevrejeva

et al.

(2010) also reported projections for the mid-century. In contrast to the relatively divergent range projected for 2100 (2.0-5.3 ft (0.6-1.6 m)), they found relatively close agreement in projected sea level rise across various emissions scenarios until about 2050 using the six emissions scenarios used by the IPCC, with projections ranging from a low of approximately 0.98 ft (0.3 m) to a high of 1.8 ft (0.55 m) (Jevrejeva

et al.

2010, p. 3, Figure 2).

As discussed above, results for models projecting sea level rise further than mid-century become increasingly divergent, and this is particularly true with regard to the maximum bounds of projected sea level rise. Furthermore, with regard to evaluating the possible upper bounds of projected sea level rise over the next century, we considered the statements of both Nicholls

et al.

(2010, p. 168) and Lowe and Gregory (2010, p. 43) that the probability of rises at the high end of the spectrum are very low. Nicholls

et al.

(2010, p. 174) concluded that, although a sea level rise between 1.6 and 6.6 ft (0.5 and 2.0 m) is not an implausible range, “owing to our poor understanding of the underlying processes driving climate-induced sea-level rise, we cannot associate any likelihood with this range, and we conclude that rises above 0.5 m and especially 1 m by 2100 are possible, rather than inevitable.”

As there is so much uncertainty surrounding global sea level projections, particularly at the upper bounds at the end of the century, and this is further complicated by uncertainty about regional and local divergences from the global mean, we believe it is more appropriate to focus our analysis on less variable projections over a somewhat shorter timeframe. Therefore, we evaluated what we consider to be reasonable approximate projected levels of sea level rise for the habitat of the black-footed albatross, based on consideration of the global estimates described above, over three time intervals: For the next 10-20 years we use an estimate of 0.5-1.0 ft (0.1-0.3 m); for 30-40 years we use 1.4-1.9 ft (0.4-0.6 m); and for 50 years we use 2.4 ft (0.7 m) (see USFWS 2011b, unpubl., for additional details). While we recognize that several models project an accelerated rate of sea level increase later in the century (

e.g.,

Vermeer and Rahmstorf 2009, Figure 6, p. 21531), we determined that, in light of the significant variability in projections following mid-century, for the purposes of this status evaluation using a linear projection of sea level rise (see Baker

et al.

2006, pp. 5-6) and time-intervals up to mid-century is a reasonable approach. Note also that the level we use for 50 years from now, 2.4 ft (0.7 m), is conservative in that it is higher than the mid-century projection by Jevrejeva

et al.

(see above), and in fact is very close to the end of century level (0.8 m) described as “plausible” by Pfeffer

et al.

(2008, p. 1342), and even closer to the low ends of the ranges projected at 2100 by Vermeer and Rahmstorf (2009, p. 21530) and Grinsted

et al.

2010, pp. 469-470. We believe this approach is reasonable and provides a reliable basis for our analysis.

Sea Level Rise and Coastal Inundation

There is very little existing information in the scientific literature

on how projected sea level rise will affect the islands currently used by black-footed albatross for nesting, as topographical information for these islands in most cases is extremely limited and, as noted above, regionally specific models of sea level rise for the area are still under development (

e.g.,

Klavitter 2010, pers. comm.). A rigorous geomorphological coastal analysis is needed to fill this information gap. In a limited study of several of the Northwestern Hawaiian Islands, Baker

et al.

(2006, p. 2) noted this lack of spatial data, and developed models that can be used to estimate the proportional rate at which land area may disappear in the Northwestern Hawaiian Islands, based on cumulative elevation data (Baker

et al.

2006, p. 6, Figure 3).

As only maximum elevation data are available for most of these islands, these researchers collected elevation data from three locations: Lisianski Island, Pearl and Hermes Reef (the islets of Southeast, Seal-Kittery, Grass, North, and Little North), and French Frigate Shoals (including the islands of East, Gin, Little Gin, and Trig). We did not use the projections of surface area lost presented by Baker

et al.

(2006) in their Table 1 since those estimates were developed using the older IPCC 2001 projections of sea level rise. However, based on their cumulative elevation models (Baker

et al.

2006, Figure 3), we estimated the effects of the projected sea level rise on each of the islands over the three time intervals (Table 4).

Note that detailed topographical information is not available for the island of Midway Atoll or Laysan Island, which support the two largest colonies of black-footed albatrosses in the world, and these islands were not included in the analysis of Baker

et al.

(2006). However, results for Laysan Island are likely to be similar to those for Lisianski Island, as detailed below.

Table 4—Projected Effects of Sea Level Rise on the Land Area of Islands Supporting Nesting Black-Footed Albatross at 10-20, 30-40, and 50 Years in the Future, Based on Passive Flooding and the Cumulative Elevation Models of Baker

et al.

2006

[We assumed islands greater than 165 ft (50 m) in maximum elevation would retain at least 95% of their land area above sea level]

Island

Maximum

elevation

Island area

Number of breeding pairs of black-

footed

albatross

(survey year)

Percent of northwestern

Hawaiian

islands

breeding

population

Percent of Japanese islands breeding population

Percent of world

breeding

population

10-20 years—proportion of land area remaining above 0.5-1.0 ft (0.1-0.3 m); range for

individual

islets is in

parentheses

30-40 years—proportion of land area remaining above 1.4-1.9 ft (0.4-0.6 m); range for

individual

islets is in

parentheses

50 years—proportion of land area remaining above 2.4 ft (0.7 m); range for

individual

islets is in

parentheses

Northwestern Hawaiian Islands

Kure Atoll

8-20 ft (2.4-6.1 m)

213 ac (87 ha)

3,486 (2010)

5.4

NA

5.2

Data not available

Data not available

Data not available.

Midway Atoll

12 ft (3.6 m)

1,532 ac (624 ha)

25,581 (2010)

40.0

NA

38.1

Data not available

Data not available

Data not available.

Pearl and Hermes Reef

≉ 9.8 ft (≉ 3 m)

88 ac (36 ha)

6,116 (2003)

9.6

NA

9.1

99-88% (range 99-55%)

82-72% (range 82-30%)

67% (range 79-25%).

Lisianski

≉ 40 ft (≉ 12.9 m)

391 ac (159 ha)

2,126 (2006)

3.3

NA

3.2

99%

98-99%

97-98%.

Laysan

1

≉ 40 ft (≉ 12.9 m)

1,000 ac (407 ha)

22,272 (2010)

34.8

NA

33.1

99%

98-99%

97-98%.

French Frigate Shoals

8-12 ft (2.4-3.6 m)

67 ac (27 ha)

4,309 (2009)

6.7

NA

6.4

98-86% (range 98-75%)

82-74% (range 87-50%)

69% (range 80-40%).

Necker

276 ft (84 m)

45 ac (18 ha)

112 (1995)

0.2

NA

0.2

> 95%

> 95%

> 95%.

Nihoa

903 ft (275 m)

171 ac (70 ha)

1 (2007)

0.0

NA

0.0

> 95%

> 95%

> 95%.

Offshore Main Hawaiian Islands

Kaula

165 m

64 ha (158 ac)

3 (1993)

0.0

NA

0.0

> 95%

> 95%

> 95%.

Lehua

214 m

116 ha (284 ac)

25 (2007)

0.0

NA

0.0

> 95%

> 95%

> 95%.

Japanese Islands

Torishima Island

1,293 ft (394 m)

1,184 ac (479 ha)

2,150 (2003)

NA

67.5

3.2

> 95%

> 95%

> 95%.

Senkaku Islands

1,257 ft (383 m)

1,446 ac (633 m)

56 (2002)

NA

1.8

0.1

> 95%

> 95%

> 95%.

Ogasawara

Muko-jima Retto

Not available

1,631 ac (664 ac)

967 (2006)

NA

30.4

1.4

> 95%

> 95%

> 95%.

Haha-jima Retto

1,525 ft (462 m)

6,805 ac (2,770 ha)

11 (2006)

NA

0.3

0

> 95%

> 95%

> 95%.

1

Land area estimates assume similar conditions to Lisianski, based on similar elevation and topography.

Tern Island was estimated to comprise about 66 percent (57 ac (23 ha)) of the terrestrial area of French Frigate Shoals (Arata

et al.

2009, p. 76). Originally the island was only about 10 ac (4 ha) in size but was expanded in 1942 to 57 ac (23 ha) (Amerson 1971, p. 12). Sand and Eastern islands, the two main islands at Midway Atoll, have also

undergone extensive human modifications, and are approximately 12 ft (3.6 m) above sea level. All three islands (Tern, Sand, and Eastern) transition from sea level to maximum elevation over a few meters and are relatively flat across their full expanse to accommodate aircraft runways on each island. How much projected levels of sea level rise over the next 10 to 20 years (0.5-1.0 ft (0.1-0.3 m)), 30 to 40 years (1.4-1.9 ft (0.4-0.6m)), and 50 years (2.4 ft (0.7 m)) will impact black-footed albatross nesting habitat on these islands is unknown in the absence of more detailed geomorphological information, but given their relatively low elevation, sea level rise may result in some loss of nesting habitat for black-footed albatrosses.

It is also possible, however, that there will be no net loss of land area depending on relative rates of beach erosion in some (seaward) areas and beach deposition in other (lagoon-side) areas that may occur, as has been observed in other Pacific atoll islands in response to rising sea level (Webb and Kench 2010, p. 234). Webb and Kench (2010, entire) studied 27 Central Pacific islands using a combination of historical aerial photography and remote sensing imagery from years spanning from 1943 through 2006 (the timeframe of analysis for each island differed, depending on the availability of imagery, but ranged from 19 to 61 years). Despite the expectation that such islands would diminish in size due to ongoing and future sea-level rise, they found that with a historical sea level rise of 0.08 in (2 mm) per year over the period studied (roughly 4.8 in (12 cm) maximum), the terrestrial area of 43 percent of the 27 atoll islands studied remained stable while another 43 percent actually increased in size by 3 to 30 percent (Webb and Kench 2010, p. 241). Only 14 percent of the atoll islands showed a loss of 3 to 10 percent of area. The observed adjustment for 65 percent of these atoll islands was a net lagoon-ward migration, but also included island migration along the atoll reef. Overall, these atoll adjustments added 156 ac (63 ha) of coastal land area to these islands.

In the Northwestern Hawaiian Islands, sediment transport has resulted in the submersion of Whale-Skate Island in French Frigate Shoals and has accreted island area at Spit Island (Midway Atoll), Seal-Kittery Island (previously 2 islets), and North Island at Pearl and Hermes Reef (Amerson

et al.

1974, pp. 8 and 11, comparing reported islet area to current estimates). These data, and taking into consideration the results reported by Webb and Kench (2010, see above) for atoll islands elsewhere, indicate projected sea level rise will likely change the physical shape and position of Tern, Sand, and Eastern islands and may reduce or possibly increase the size of these atoll islands. However, it is also important to note that we do not have information to indicate how these processes may work under potentially accelerated rates of sea level rise. Any such changes, however, whether positive or negative in terms of total land area, are likely to occur gradually over many years, giving black-footed albatrosses a long period of time to potentially adjust their breeding locations. Tern and East Islands each support just under half of the black-footed albatross breeding pairs at French Frigate Shoals (Arata

et al.

2009, p. 38, Figure 14).

Based on the cumulative elevation model developed by Baker

et al.

(2006, p. 6, Figure 3) East Island will lose about 2 to 10 percent of its land area to 0.5-1.0 ft (0.1-0.3 m) sea level rise in 10 to 20 years, 12 to 19 percent of its land area to 1.4-1.9 ft (0.4-0.6 m) sea level rise in 30 to 40 years, and roughly 20 percent of its current land area to a 2.4 ft (0.7 m) rise in sea level in 50 years. Potential losses of land area at the smaller islets of French Frigate Shoals are projected to be greater (Table 4, this document), but no estimates were available for Tern Island, where relatively large numbers of black-footed albatrosses breed. We estimated that, collectively, the islets of French Frigate Shoals will have roughly 86 to 98 percent of terrestrial area remaining after 10 to 20 years, 74 to 82 percent remaining after 30 to 40 years, and 69 percent after 50 years (Table 4, this document).

We note Baker

et al.

(2006) do not take into account geomorphological features that can alter sea level rise impacts, as shown by Webb and Kench (2010, p. 241). All of these islands may change shape, size and position through erosion and accretion, such that future land areas may be larger or smaller than projected due to sea level rise alone. The islets of Pearl and Hermes Reef support 10 percent of the world black-footed albatross breeding pairs and comprise some of the lowest elevation areas used for nesting by the species. Collectively, we estimate that these islets will retain roughly 88 to 99 percent of their land area in 10 to 20 years, 72 to 82 percent of their land area over 30 to 40 years, and 67 percent of their land area in 50 years (Table 4, this document). This does not take into account potential changes in shape, size, or position that may occur due to erosion and accretion, as demonstrated by Webb and Kench (2010, p. 241) for island atolls elsewhere, but due to their small size and low elevation we consider these islets to be some of the most vulnerable to sea level rise and may be a potential loss of nesting habitat for the black-footed albatross.

Lisianski Island (currently supporting 3.2 percent of world black-footed albatross breeding pairs) is one of the larger Northwestern Hawaiian islands at 391 acres (159 ha) in size. We estimated that Lisianski would still have 99 percent of its terrestrial area over the next 10 to 20 years, 98 to 99 percent over 30 to 40 years, and about 97 percent of its terrestrial area in the face of a 2.4-ft (0.7 m) rise in sea level in 50 years (based on Baker

et al.

2006, p. 6, Figure 3; see Table 4, this document). Laysan Island (currently supporting 35 percent of world black-footed albatross breeding pairs) has a maximum elevation that is the same as Lisianski Island (about 40 ft (13 m)) and, like Lisianski, has a large central depression (a lake on Laysan but not on Lisianski) surrounded by higher elevation sandy ridges (Macdonald

et al.

1990, pp. 480-481). In addition, at approximately 1,000 ac (407 ha) in size, Laysan is substantially larger than Lisianski (391 ac; 159 ha). Presuming a similar island atoll geomorphology, sea level rise will affect a limited area of Laysan Island, most likely similar to the projections for Lisianski. As discussed above, this analysis does not consider geomorphological features that can alter early sea level rise impacts, as shown by Webb and Kench (2010, p. 241). Their information indicates that levels of sea level rise expected over the next 50 years will likely change the shape and position of Lisianski and Laysan Islands, and that processes of erosion and accretion may either reduce or even increase the size of these islands. All of these changes are likely to occur gradually over many years.

Kure Atoll (which supports 5.2 percent of world black-footed albatross breeding pairs) was not included in Baker

et al.

's projections of sea level rise impacts on the Northwestern Hawaiian Islands. Kure Atoll has a maximum elevation of approximately 24 ft (7.5 m; Arata

et al.

2009, p. 75). Impacts from sea level rise at Kure Atoll are likely to be similar to those discussed for these other atoll areas, although Kure Atoll has greater land area and maximum elevation than the islets of Pearl and Hermes Reef.

While black-footed albatrosses are typically characterized as nesting on the sandy beaches of low atoll islands, there are several colonies that currently nest upslope on high-elevation islands and

do not utilize shoreline nesting sites: Toroshima (2,150 breeding pairs), Senkaku (56 breeding pairs), Ogasawara (Hahajima Island) (11 breeding pairs), Necker (112 breeding pairs), Nihoa (1 breeding pair), Kaula (3 breeding pairs) and Lehua (25 breeding pairs) (Arata

et al.

2009, p. 3, Figure 1). Nesting on these islands occurs well above sea level in volcanic substrates or on the top of hill and upland slopes (Clapp and Kridler 1977, p. 36; Clapp

et al.

1977, p. 44; Cousins and Cooper 2000, p. 5; Pitman and Ballance 2002, p. 13). Due to their topography and elevation, we do not expect these islands and their breeding populations of black-footed albatross to be affected by anticipated levels of sea level rise.

As noted earlier, detailed, spatially-explicit data specific to the breeding islands of the black-footed albatross are limited or nonexistent. Although the USGS is currently studying the potential impacts of sea level rise on the Northwestern Hawaiian Islands, the results of this research was not available in time for our status assessment. Based on the best scientific information available to us, we can make rough approximations of the land area that may remain under various sea level rise scenarios on these islands, but we do not have detailed spatial information that would enable us to determine how much of the land area that would be lost currently serves as nesting habitat for the black-footed albatross. However, given that black-footed albatrosses on the low-lying islands and atolls of the Northwestern Hawaiian Islands select sites in sandy habitats generally close to the shoreline for nesting, it is reasonable to assume that much of the initial losses of land area would constitute potential or current nesting habitat. This assumption does not apply to black-footed albatrosses that nest upslope on steep, high islands, such as Necker, Nihoa, or the Japanese Islands. Therefore, we must consider the potential effects of the loss of an unknown amount of current shoreline nesting habitat on the black-footed albatross, based on estimated losses of land area and related considerations.

For those black-footed albatrosses that do nest near the shoreline, inundation by high surf currently destroys some nests, and high winds bury nests and kill eggs or chicks and sometimes incubating adults, although the proportion of nests affected each year has not been quantified (Flint 2009a, pers. comm.). Winter storms and the associated high tides and high winds were identified as a major cause of black-footed albatross nest failure on Kure Atoll in the 1960s (Woodward 1972, p. 93). Recently on French Frigate Shoals, the smaller islands of Little Gin and Trig were washed over while adult black-footed albatrosses were incubating eggs (Flint 2009a, pers. comm.). Also on the larger islands of Tern and Eastern, black-footed albatross nests on the islands' northern sides that were exposed to the larger winter swells were often inundated or washed away (Flint 2009a, pers. comm.). During the 2008 breeding season, all of the nests, eggs, and chicks on Tern Island were washed away by high surf (Flint 2009a, pers. comm.). In addition, severe events may happen on occasion, as in the estimated loss of more than 20,000 black-footed albatross chicks from the Northwestern Hawaiian Islands in the aftermath of the March 2011 tsunami generated off the coast of Japan (Flint 2011b, pers. comm.). Such events, although random and unpredictable in occurrence, are not unexpected, and have presumably occurred throughout the history of the species (

e.g.,

see Cousins and Cooper 2000, pp. 115-117). Whether such events may potentially increase in frequency as a potential effect of climate change is an important consideration; however, at this point in time we do not have sufficient information to quantify the probability of such occurrences for this region (see “Storm Frequency and Intensity,” below). That most adults survive such events, and population viability in this species is more dependent on adult than juvenile survivorship, enables the species to persist despite occasional severe impacts to productivity or recruitment.

Reproductive success may also be affected in the event birds are forced to relocate their nesting sites due to high surf or winds. For example, black-footed albatrosses whose nest sites were lost on Midway Atoll because of habitat modification related to military activity, both during and immediately following World War II, were found in later years breeding at a different location on the atoll, though it is likely that they lost at least 1 year of breeding due to the displacement (Cousins and Cooper 2000, p. 44). More recently, black-footed albatrosses forced to relocate due to construction activities on Midway were later found nesting elsewhere in the atoll, although they similarly likely lost a year of breeding as a consequence (Flint 2009a, pers. comm.) If a nest site is destroyed, the birds may have difficulty in pairing up with the same mate. In general, mate loss in black-footed albatrosses can cause adults to miss up to 5 years of breeding before forming a new pair (COSEWIC 2007, p. 33). Increased storm surges or other events due to anticipated climate change may therefore result in some decreased productivity for black-footed albatrosses, especially those nesting on very low-lying islands; however, the actual potential extent of this impact would be purely speculative at this time.

A key uncertainty in our evaluation of the effects of sea level rise is the behavioral response of breeding black-footed albatrosses to the possible future inundation of their current nesting sites. The strong nest site fidelity of black-footed albatrosses is an important consideration in this regard. As described in the

Life History

section, above, more than 99 percent of black-footed albatrosses breed on the island where they hatched (Rice and Kenyon 1962a, p. 532), and they construct their nests every year on almost the same site. On Tern Island, black-footed albatrosses were found to nest within 16 ft (5 m) of the previous year's nest (Cousins and Cooper 2000, p. 44). Data from a 2-year study of the closely related Laysan albatross on Midway Atoll showed nests to be within 20 ft (6 m) of the previous year's nest site, and over 50 percent of nests were within 4 ft (1.3 m) (Rice and Kenyon 1962a, p. 533).

In an experimental study, adults of the closely-related Laysan albatross generally responded to displacement of their chick from the nest site by not feeding their chick unless it was within 7 ft (2 m) of the nest site (Rice and Kenyon 1962a, pp. 534-536). That is, adults are oriented on the location of the nest, not on the location or identity of the chick. Based upon this information, an unknown number of black-footed albatross nest sites may be lost each breeding season due to increasing high tides or storm surge from sea level rise, and chicks that get displaced from their nest site may die because their parents do not feed them. Offsetting this potential impact, however, is the availability of additional nesting habitat for black-footed albatrosses and the possibility that birds will relocate their nest sites to more suitable (higher-elevation, inland) habitat over time.

Although black-footed albatrosses do predominantly nest on sandy beaches near the shoreline, there is apparently some behavioral flexibility in nest site selection by the species, as they are found nesting further inland in vegetated areas on Midway and French Frigate Shoals, including amongst bushes, in clearings among introduced ironwood trees, and in grassy areas (Awkerman

et al.

2008). On steep, volcanic high-elevation islands, such as Necker, Nihoa, and Japanese Islands such as Torishima, black-footed

albatrosses nest high upslope in grassy or rocky areas (Cousins and Cooper 2000, p. 32; see, for example,

http://www.mnc.toho-u.ac.jp/v-lab/ahoudori/Photo/photo03/68.html

).

Although in some cases black-footed albatrosses have exhibited a reluctance to move, despite repeated nest failures, there are other examples of breeding pairs relocating, as for example cited above at Midway Atoll in response to displacement from military activities or construction (Arata

et al.

2009, p. 39; Flint 2009a, pers. comm.). On Torishima Island, black-footed albatrosses established new breeding colonies following volcanic eruptions in 1903, 1941, and 2002 (see “Volcanic Activity,” above). In addition, anecdotal evidence suggests that black-footed albatrosses have moved to other islands as smaller islands have disappeared or become overwashed, as suggested at Tern Island by Cousins and Cooper (2000, p. 32) and at French Frigate Shoals (ACAP 2010, p. 7). The recent increase in breeding birds at French Frigate Shoals may be due to the redistribution of black-footed albatrosses that once nested on the island of Whale-Skate, which was lost entirely to erosion from winter storms and sea level rise in 1997 (ACAP 2010, p. 7); however, this supposition is apparently based on the circumstantial timing of the increase on French Frigate Shoals following the disappearance of Whale-Skate, and is not supported by observations of banded birds.

Whether established breeders would move to new nest sites is a major source of uncertainty in our evaluation. The question of whether birds just coming into breeding age would establish new colonies, assuming their natal sites may be lost, is less uncertain. Despite their normally high degree of philopatry, we do have evidence that some black-footed albatrosses banded as nestlings have become breeders on other than their natal islands (Woodworth 1972, p. 96). For example, of 124 banded nestlings, mostly from Midway Atoll, 22 were later observed breeding on Kure Atoll (Woodworth 1972, p. 96). Other movements of smaller numbers of black-footed albatrosses between their natal and breeding sites were observed between Pearl and Hermes Reef, French Frigate Shoals, and Kure Atoll as well (Woodworth 1972, p. 96). Although most movements of black-footed albatrosses between breeding colonies have been over a relatively small range (Woodworth 1972, pp. 96, 109), there is evidence of quite long-range movements from the recent observations of black-footed albatrosses prospecting for nesting sites on the islands of Guadalupe and San Benedicto off the coast of Mexico (Awkerman

et al.

2008). Colonization of new islands and range expansion, including the establishment of breeding colonies in the eastern Pacific on the islands of Guadalupe and San Benedicto, has also been observed in the related Laysan albatross (Young

et al.

2009, p. 722), a bird that exhibits a similarly high degree of natal philopatry, suggesting it is not unreasonable to anticipate that black-footed albatrosses are capable of colonizing new areas if their current nesting habitat is lost.

In general, gradual shifts from the loss of old habitat to the availability of new habitat, as would occur under a scenario of gradual sea level rise, are considered most conducive to the establishment of new colonies (as opposed to the abrupt loss of all breeding sites) (Schippers

et al.

2009, p. 469). The availability of nest sites is only rarely limiting for seabirds (Kildaw

et al.

2005, p. 55), and we have no evidence to suggest that suitable nest sites are a limited resource for black-footed albatrosses in the Hawaiian Islands (COSEWIC 2007, p. 20). There are, however, some new challenges that black-footed albatrosses may face as a result of relocating their nest sites. For one, if the birds attempt to relocate to some of the higher-elevation Hawaiian islands in response to sea level rise, they will encounter predators that are currently not a threat to the species (

e.g.,

mongooses, cats, dogs, pigs, rats) (Naughton

et al.

2007, p. 10). Whether such an option may be feasible for black-footed albatrosses in the future may rely on the implementation and success of current management efforts to restore habitat and eradicate nonnative predators on other nearby, higher elevation islands (Naughton

et al.

2007, p. 19). There are no introduced predators on the islands of San Benedicto or the small islets off of Isla Guadalupe in the eastern Pacific (Naughton

et al.

2007, p. 12). In addition, reduced habitat area will in turn mean increased competition with other nesting seabirds, such as the Laysan albatross, which often nests in the same habitat as the black-footed albatross. However, the evidence from historical photographs indicates that great numbers of seabirds can successfully nest at very high densities on these islands, suggesting that the same number of black-footed albatrosses may be able to continue nesting into the future on islands that have diminished in size, despite the presence of other potential competitors. The maximum density of nesting seabirds on these islands is unknown, and although available habitat does not presently appear to be restricted, it is unknown at what point in time it may potentially become a limiting factor.

There will undoubtedly be some short-term impacts to productivity of nesting black-footed albatrosses due to displacement from sea level rise; based on the elevation and topography of the islands, we anticipate such impacts would be concentrated in the Northwestern Hawaiian Islands and would not affect the Japanese Islands populations (see Table 4). In the Northwestern Hawaiian Islands, our assessment of the projected levels of terrestrial area lost over the next 10 to 20, 30 to 40, and up to 50 years suggests that the loss of terrestrial area on islands used for nesting by black-footed albatrosses will be relatively gradual. Moreover, the remaining land area for some of the larger colonies at Laysan Island, Pearl and Hermes Reef, and French Frigate Shoals will still be relatively substantial at the end of that time period (estimated as 97 percent terrestrial area remaining at Laysan with 34.8 percent of the Northwestern Hawaiian Islands breeding population, 67 percent terrestrial area remaining at Pearl and Hermes Reef with 9.6 percent of the breeding population, and 69 percent terrestrial area remaining at French Frigate Shoals with 6.7 percent of the breeding population).

We note that information was not available for the largest breeding colony of black-footed albatrosses at Midway Atoll. Lost land area may disproportionately affect black-footed albatross nesting habitat, since many individuals select nesting sites on beaches near the shoreline, which will in many cases represent the first land area lost. In addition, diminished land area will not be the only effect of sea level rise, as the remaining land will consequently become increasingly vulnerable to overwash events. However, based on the relatively gradual nature of sea level rise over time, the amount of land area projected to remain, the ability of black-footed albatrosses to nest in habitats other than sandy beaches, the apparent capacity of these islands to support high densities of nesting seabirds, and the evidence suggesting that black-footed albatrosses will breed on other than their natal islands and colonize new sites, albeit in low numbers, we believe it is reasonable to conclude that the black-footed albatross may shift to new nest sites over time in response to sea level rise in the Northwestern Hawaiian Islands.

In summary, many uncertainties remain with regard to the potential impacts of future sea level rise on the

black-footed albatross. As mentioned previously, at present we have no regional models of sea level rise specific to the islands used for nesting by black-footed albatross, but must instead rely primarily on global projections of sea level rise. Yet we know that sea level rise is likely to vary considerably in different locations across the globe, as described above. As also noted above, although we have some rough projections of how much terrestrial area may be lost on a limited number of the islands used for breeding, at present we do not have the data to inform us as to how much of the land area that may be lost currently serves as nesting habitat. In addition, projected losses of land area above sea level using a simple passive inundation or “bathtub” model do not account for other potential consequences of climate change that may impact the suitability of remaining terrestrial areas for nesting, such as storm surge.

The greatest uncertainty in evaluating the threat of sea level rise and potential loss of nesting habitat is the behavioral response of the birds over time. The biggest question in this regard is whether established adult breeders would eventually shift their nesting locations in response to habitat loss as a consequence of inundation; there is some evidence that supports such a potential shift, and some evidence that suggests such a shift would more likely require waiting for birds hatched on the islands to attain reproductive age and establish new nest sites elsewhere. Whether suitable, predator-free habitat would be available for these birds in the future is another uncertainty. In any case, we anticipate some unknown level of reduced productivity and likely diminished population sizes will be realized as a consequence of smaller habitat area. However, based on the land area projected to remain and the relatively large breeding population of black-footed albatrosses (Table 4), we do not anticipate that these interim losses will be so great as to pose a significant threat to the black-footed albatross.

We conclude, based on this assessment, that there will likely be some short-term impacts to black-footed albatross nesting success due to sea level rise and coastal inundation, and that future population sizes in the Hawaiian Islands may be smaller due to a reduced area of available nesting habitat. However, we do not have evidence to suggest the projected changes will be so great as to pose a significant threat to the breeding populations of the species rangewide, in the Hawaiian Islands, or in the Japanese Islands.

Climate Change and Wave Inundation

The central Pacific location of the Northwestern Hawaiian Islands exposes the atoll islands to wind and ocean swells from all directions but mostly from the northeast and northwest (Vitousek and Fletcher 2008, p. 541). The northeastern trade winds predominate during three quarters of the year, and generate average wave heights of 6.6 ft (2 m) (Fletcher and Feirstein 2009, pp. 3-4). During winter, when black-footed albatrosses are nesting on the Northwestern Hawaiian Islands, northwestern Pacific storms generate much larger waves with an average height of 25.3 ft (7.7 m) (Fletcher and Feirstein 2009, p. 3). Wave inundation of coastal atoll island areas or overwash of entire atoll islands is known to occur, but information specific to this issue in the Northwestern Hawaiian Islands is limited. Two major features will affect future wave inundation: sea level rise and storm frequency and intensity. These are discussed below, based on the best scientific information available.

Winter (November through April) mid-latitude (30 to 60° N latitude) storms (extra-tropical cyclones) can produce waves that may impact black-footed albatross breeding. The southern cold fronts of these winter storms bring rain to the Northwestern Hawaiian Islands (Juvic

et al.

1998, p. 54). The low-pressure centers of these mid-latitude storms generate ocean waves that can propagate to the Hawaiian Islands. Approximately 20 strong mid-latitude storms occur each year in the north Pacific (Graham and Diaz 2001, p. 1,874). Large waves generated by these storms are known to periodically overwash small islets (

e.g.,

Sand and Bird islets at Pearl and Hermes Reef; Gin and Little Gin islets at French Frigate Shoals) and inundate coastal sites in the Northwestern Hawaiian Islands and destroy near-shore black-footed albatross nests (Arata

et al.

2009, p. 11). Most recently, a large wave event destroyed approximately 40 percent of black-footed albatross nests on Laysan Island in February 2011, resulting in the loss of an estimated 9,000 chicks, and more than 20,000 black-footed albatross chicks are estimated to have been lost when the Northwestern Hawaiian Islands were overwashed by a tsunami following the March 11, 2011, earthquake off Sendai, Japan (Flint 2011b, pers. comm.). The reported mortality of chicks from the tsunami is likely an underestimate, as counts were not available for all islands affected. There are no estimates as to the number of adults that may have been lost, but in general it is expected that chicks make up the vast majority of mortalities in such events.

Current climate models indicate that mid- and high- (60° to 90° N latitude; too far north to generate Pacific waves) latitude Pacific storms will shift to the north with a decrease in storm frequency in the mid-latitudes, an increase in frequency in the north latitudes (USCCSP 2008, p. 64), and an increase in the intensity of mid- and high-latitude storms (USCCSP 2008, p. 115). These model results are supported by observations from 1959 through 1997 that show similar trends (USCCSP 2008, pp. 64, 115). Winter (November through March) wave heights generated from climate models show significant increases in the northwestern and northeastern Pacific, but in the vicinity of the major black-footed albatross breeding areas (Northwestern Hawaiian Islands, Torishima Island, and the Ogasawara Islands), winter wave heights are predicted to remain relatively unchanged for the period 1990-2080 (Wang and Swail 2006, p. 116). Reduced future storm frequency in the mid-latitudes combined with no significant change in wave heights suggests that black-footed albatross may likely not be negatively affected to a degree beyond historical and current impacts, if these predictions generally hold. As in the past, wave surge and occasional overwash events will occasionally impact black-footed albatrosses breeding at localized areas. Although such events may have a large short-term impact on productivity in a single year, as with the significant wave events and tsunami observed in early 2011, most adult breeders generally survive these events, and the long-term impact on the species is limited. Therefore, based on the best available data, we have no information to indicate that the impact of wave or storm events will be so great as to pose a significant threat to the breeding populations of the species rangewide, in the Hawaiian Islands, or in the Japanese Islands.

Climate Change and Tropical Cyclone Storm Frequency and Intensity

The Pacific tropical cyclone (

e.g.,

typhoon and hurricane) storm season conservatively starts in May or June, with the core storm season running from July through November in the eastern and central Pacific, and through December in the western Pacific. Black-footed albatrosses arrive at their nesting sites in mid- to late October and do not begin to lay eggs until mid-November. Thus, the overlap between adult birds arriving at nesting sites and the end of the tropical cyclone storm season is likely only a few weeks. There are no

climate model predictions for a change in the duration of Pacific tropical cyclone storm season.

Climate modeling has projected changes in tropical cyclone frequency and intensity due to global warming over the next 100 to 200 years (Vecchi and Soden 2007, pp. 1068-1069, Figures 2 and 3; Emanuel

et al.

2008, p. 360, Figure 8; Yu

et al.

2010, p. 1,371, Figure 14). The frequency of hurricanes generated by these tropical cyclones is projected to decrease in the central and eastern Pacific (

e.g.,

the main and Northwest Hawaiian Islands and the islands off Mexico where black-footed albatrosses have recently attempted to breed) while storm intensity (strength) is projected to increase by a few percent over this period (Vecchi and Soden 2007, pp. 1,068-1,069, Figures 2 and 3; Emanuel

et al.

2008, p. 360, Figure 8; Yu

et al.

2010, p. 1,371, Figure 14). In the western Pacific (

e.g.,

the Mariana Islands and the Japanese Islands that currently, or in the past, supported black-footed albatross populations), the frequency and intensity of typhoons are projected to increase by a few percent over the next 100 to 200 years (Vecchi and Soden 2007, pp. 1,068-1,069, Figures 2 and 3; Emanuel

et al.

2008, p. 360, Figure 8; Yu

et al.

2010, p. 1,371, Figure 14). Although there is some indication that the impacts of tropical cyclones are expected to increase in general as a result of projected sea level rise (Knutson

et al.

2010, p. 157), we do not have any modeling available specific to the regions used by nesting black-footed albatross, and we do not have sufficient data to quantify or evaluate the potential impacts of such events on the species or to assess the possible population-level response over the extended timeframes of the projections, except to note that the timing of such events does not usually coincide with the nesting season of the black-footed albatross, when potential impacts from such events would be expected.

In summary, based on the limited information available to us and the climate model analyses described above, the anticipated increases in cyclone intensity or frequency are minimal. This is especially true toward the end of the storm season when albatross begin to arrive at the breeding grounds and cyclone intensity and frequency is normally decreasing. Furthermore, we believe it is highly unlikely that multiple nesting sites would be impacted in a single storm season, given the wide geographic spread of the nesting sites used by black-footed albatrosses. We further note that the frequency of hurricanes in the Northwestern Hawaiian Islands, where the majority of black-footed albatrosses nest, is currently low and is predicted to decrease with climate change.

We conclude, based on this assessment, that while there may be some short-term impacts to black-footed albatross nesting success due to the potential overlap between the arrival of birds at nesting sites and the end of the tropical storm season, we do not have evidence to suggest that projected changes in storm frequency or intensity will be so great as to pose a significant threat to the breeding populations of the species rangewide, in the Hawaiian Islands, or in the Japanese Islands.

Climate Change and Marine Productivity

The link between marine productivity and climate is not well understood (McGowan

et al.

1998, p. 210; Polovina 2005, p. 233). The potential impacts of climate change on the food supply of the black-footed albatross (mainly flying fish eggs and squid (Arata

et al.

2009, p. 11)), and thus survival and reproduction, has not been well studied. There are, however, two major natural climate oscillations associated with major changes in marine ecosystems in the Pacific: El Niño-Southern Oscillation (ENSO; (McPhaden

et al.

2006, p. 1,741) and the Pacific Decadal Oscillation (PDO) (Miller

et al.

2004, p. 163).

The ENSO can influence productivity in the tropical Pacific (Fiedler 2002, p. 270; McPhaden

et al.

2006, p. 1,741) and the west coast of Central and North America (McGowan

et al.

1998, p. 214). El Niño-Southern Oscillation is a 2- to 7-year fluctuation of unusually warm (El Niño) and cool (La Niña) conditions in the tropical Pacific associated with an unstable interaction between sea surface temperature and atmospheric pressure. It results in variations in wind, rainfall, ocean thermocline depth, circulation, and ultimately oceanic biological productivity (McGowan

et al.

1998, p. 214; Fiedler 2002, p. 267). At present, the relationship between the future frequency and intensity of ENSO events related to global climate change is not yet determined and may be unchanged, increasing or decreasing (Guilyardi

et al.

2010, p. 325; Vecchi and Wittenberg 2010, p. 260). ENSO affects areas used by the black-footed albatross mainly along the west coast of the United States and Canada (McGowan

et al.

1998, p. 214; McPhaden

et al.

2006, p. 1,741; Arata

et al.

2009, p. 6). In this area, ENSO can affect plankton biomass, the distribution of fishes and invertebrates, and the breeding success of seabirds, sea lions, and seals (McGowan

et al.

1998, p. 214).

A qualitative analysis of black-footed albatross reproductive success (1980 through 2008 on Tern Island, French Frigate Shoals) and number of breeding birds (Laysan Island, Midway Atoll, and French Frigate Shoals) showed no relationship with El Niño or La Niña events (USFWS 2009a, unpubl.). Although there have been references to “dramatic breeding failures” of black-footed albatrosses in years following El Niño events, inspection of the underlying data sugges

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