|Scientific Name:||Anguilla rostrata (Lesueur, 1817)|
Muraena rostrata Lesueur, 1817
|Taxonomic Source(s):||Eschmeyer, W.N. (ed.). 2014. Catalog of Fishes. Updated 27 August 2014. Available at: http://researcharchive.calacademy.org/research/ichthyology/catalog/fishcatmain.asp. (Accessed: 27 August 2014).|
|Taxonomic Notes:||Historically, a few authors have argued that European (Anguilla anguilla) and American Eels might actually be conspecific (vertebral count differences were believed to be entirely ecophenotypic), although this was never considered to be very likely (Van Den Avyle 1984) and has been clearly shown to be incorrect by a variety of types of information. Separate species status has been confirmed by recent demonstrations that American and European Eels comprise two largely separate gene pools (see Avise et al. 1990, Bastrop et al. 2000, Avise 2003, Bernatchez, St-Cyr et al. 2011). Eels in Iceland include, in low frequency, the products of hybridization between American and European eels (Avise et al. 1990, Avise 2003, Albert et al. 2006, Pujolar et al. 2014). Phylogenies of the genus Anguilla constructed based on mitochondrial DNA have shown the two species to be distinct, although more closely related than most species of the genus (Aoyama et al. 2001, Minegishi et al. 2005).
The preponderance of evidence indicates that American Eels do not exhibit significant geographic or latitudinal patterns of genetic variation (Avise et al. 1986, Avise 2003, Wirth and Bernatchez 2003, Côté et al. 2013). However, Bernatchez, Côté et al. (2011) suggest that there may be minor quantitative genetic differences between American glass eels colonizing different parts of the range despite the fact that all belong to a single genetic population. They consider that plausible explanations for this pertain either to non-random dispersal or to selective mortality based on individual genetic differences. They emphasize that on one hand, evidence for panmixia justifies the need for global co-operative action toward improved management and conservation of the species, but that on the other hand, evidence for local genetically based phenotypic differences also justify the need for local actions, citing eels of the St. Lawrence River basin as an example.
|Red List Category & Criteria:||Endangered A2bd ver 3.1|
|Assessor(s):||Jacoby, D., Casselman, J., DeLucia, M. & Gollock, M.|
|Reviewer(s):||Cairns, D., Castonguay, M., Dumont, P., Feigenbaum, M., Harrison, I.J., Jessop, B., Lee, L., Limburg, K., McCleave, J. & Miller, J, Stavinga, J.|
|Contributor(s):||Ahn, H., Bennett, L., Crook, V., Kaifu, K., Kurwie, T., Sasal, P., Silfvergrip, A., Uchida, K., Walker, M, Turnock, S. & Hammerson, G.A.|
Anguilla rostrata exhibits facultative catadromy, has multiple life stages, is semelparous and panmictic; these life history traits make application of the IUCN Red List Criteria more challenging. Anguillids are often referred to as ‘freshwater eels’, however, it is known that they can exhibit inter-habitat migration and that a proportion may stay in estuaries, lagoons and coastal waters, rarely, if ever, entering freshwater: this element of the population is particularly poorly understood in most species.
Ideally, the IUCN Red List Criteria would be applied to mature eels at their spawning grounds and in the absence of such data, the criteria would be applied to silver eels starting their spawning migration (in the case of American Eels, leaving ‘continental’ waters), as these represent the maximum estimate of spawning stock biomass, but datasets for this are very rare. The majority of available data for this species relates to yellow eel recruitment but the relationships between recruitment, yellow eel populations, silver eel escapement and spawner stock biomass are poorly understood. As such, the IUCN Red List Criteria have to be applied to an amalgamation of multiple life stages, which may not exactly mirror the mature spawning stock but can be used as the current best estimate. Finally, the American Eel is a panmictic species (i.e., all individuals come from one spawning stock); as such escapement from a specific water body/country/region is not directly correlated to the subsequent corresponding recruitment to the same area. However, as data are only available from certain parts of the species' range - data are particularly sparse for Caribbean and Central American populations, and those associated with the Gulf of Mexico - it is important that conservation initiatives and management actions are adjusted as new data become available.
Available data for this species indicates that, overall, there have been declines in recruitment, population and escapement over the period of three generations (36 years). However, the majority of data are for young yellow eels and the data available for older yellow eel populations and, most importantly, escaping silver eels, are limited and do not encompass the entire range of the species. These issues have been taken into account during the assessment process, and it would seem that applying criteria A2bd, as the limited available data indicates, the American Eel has seen a ~50% decline in silver eel escapement over three generations and thus is on the cusp of the Vulnerable and Endangered categories. Data relating to yellow eel recruitment and yellow eel populations would indicate that these metrics have declined to a slightly greater degree (50-60%) over three generations.
There are a suite of threats that have been implicated in causing the decline of American Eel recruitment and stocks; barriers to migration – including damage by hydropower turbines; poor body condition; climate change and/or changes in oceanic currents; disease and parasites (particularly Anguillicola crassus); exploitation and trade of glass, yellow and silver eels; hydrology; habitat loss; pollutants; and predation. The impacts of these threats individually or synergistically, are likely regionally specific, however, more broadly, climate and ocean currents have been suggested to play an important role in the survival of the leptocephalus larvae and recruitment of glass eels to coastal, brackish and freshwater habitat. Further research is required to fully understand the complexities of this particular aspect of the eel's life history but there are conflicting opinions as to the degree, if any, to which oceanic factors contribute to declines in eel numbers.
To date, the American Eel is not protected under any global treaties or conventions (e.g., CITES). However, in Canada, COSEWIC reviewed the status of the American Eel and revised the Canadian status from “Special Concern” in 2006, to “Threatened” in 2012. The Canadian government has commenced a response by initiating a Recovery Potential Assessment Review. In 2007, eels in Ontario were classified as “Endangered” under the Ontario Endangered Species Act and a Recovery Strategy was developed with an emphasis on restoring, protecting, and diversifying suitable habitat. The Ontario government is currently preparing a formal Response Statement addressing the recommendations in the Recovery Strategy. In the U.S, a review by the U.S Fish and Wildlife Service and the U.S. National Marine Fisheries Service is due to be completed in 2015 that will determine whether a listing of the species as Threatened, under the Endangered Species Act (ESA), is warranted.
Utilising the available historical data from range states, and taking into account the continued threats to this species, it was deemed appropriate to assign an IUCN listing of Endangered, using IUCN Red List Criteria, under past and current observations of population decline (A2bd). However, it is imperative to highlight that recruitment is increasing in some sites and that should management actions relating to anthropogenic threats prove effective, and/or there are positive effects of natural influences on the various life stages of this species, a listing of Vulnerable would be achievable. Further, a drive to fill data gaps – particularly in the southern range of this species - would increase the robustness of the assessment of this species. Another point to make is that it was not felt that the species was in immediate risk of extinction due to the wide geographic ranges they inhabit. However, this is not to say there wasn’t concern as to the species status being ‘outside of safe biological limits’- for this assessment, ‘risk’ is determined based on past and present population levels.
Assessment of the population data was carried out during a workshop held at the Zoological Society of London from July 1st-5th 2013.
|Previously published Red List assessments:|
|Range Description:||The American Eel is a facultatively catadromous anguillid; they are migratory, spending most of their lives in freshwater, estuaries or coastal waters before migrating to the open ocean to spawn – although it is possible that there are elements of the population that exclusively occupy offshore marine habitats. The continental distribution of the American Eel ranges from the coasts and streams of West Greenland in the north (Nielsen and Bertelsen 1992), Labrador (Hamilton Inlet-Lake Melville Estuary), and Newfoundland southward along the Atlantic coast of Canada and the eastern USA to southern Florida along the coast of the Gulf of Mexico to the northern tip of the Yucatan Peninsula, the Caribbean coast of Central America, the Caribbean and Atlantic coasts of the Greater and Lesser Antilles and other West Indies islands, and occasionally the northern portion of the Atlantic coast of South America (Venezuela and Columbia) (Ogden et al. 1975, Wenner 1978, Erdman 1984, Van Den Avyle 1984, Böhlke and Chaplin 1993, Claro 1994, Kenny 1995, Bussing 1998, Lim et al. 2002). This species is also present in Puerto Rico, Bermuda and the US Virgin Islands. Despite this large range there is strong evidence that this species forms one panmictic population with all mature individuals migrating to the Sargasso Sea area (Bonhommeau et al. 2008, Miller et al. 2009, Gagnaire et al. 2012, Côté et al. 2013).|
Historically, the American Eel was widely distributed in the continental watersheds associated with this extensive range, including the Mississippi River basin and Great Lakes-St. Lawrence River basin below Niagara Falls. After construction of the Welland Canal to bypass Niagara Falls, eels, which were abundant in Lake Ontario, were frequently reported in the waters of Lake Erie (Trautman 1957). Bensley (1915) reported that eels were occasionally seen at the mouth of the Severn River in southeastern Georgian Bay, Lake Huron, possibly gaining access through the Trent-Severn Waterway. American Eels found on the Canadian side of Lake Superior likely were ballast-water introductions (Mandrak and Crossman 1992). In fact, construction of many large and small dams throughout the eastern half of the North American Continent has caused the range of the American Eel to be greatly reduced (USFWS 2007). Anecdotal evidence indicates that it has disappeared in much of the Mississippi River drainage due to the construction of dams or locks, and no longer reaches many upstream areas in other river systems along the U.S. east coast.
American Eels have been introduced (stocked, released and/or escaped) in several inland areas and several areas in the western United States. Goode (1881) reported that live eels had been privately imported and released into a tributary in the south end of Lake Michigan and dispersed widely in the lake. In 1873, the United States Fisheries Commission released eels in the Calumet and Dead rivers in Illinois and in the Fox River, Wisconsin. They were even introduced into California. In 1878, the Michigan Fish Commission introduced eels from the upper Hudson River, New York, into southern Michigan waters, and in 1882, with wild eel numbers decreasing, the Ohio Fish Commission began the same practice throughout Ohio, stocking from the same source (Trautman 1957). No doubt some of the records of rare eels in Ohio waters of Lake Erie above Niagara Falls around that time came from these stockings. As an experiment, American Eels were introduced into Lac Pelletier (1951), a saline lake in southern Saskatchewan. Some escaped into the South Saskatchewan River (Scott and Crossman 1973). The last record of the species in Saskatchewan was in 1981 (Atton and Merkowsky 1983).
Native:Anguilla; Antigua and Barbuda; Aruba; Bahamas; Barbados; Belize; Bermuda; Bonaire, Sint Eustatius and Saba; Canada; Cayman Islands; Colombia; Costa Rica; Cuba; Curaçao; Dominica; Dominican Republic; Greenland; Grenada; Guadeloupe; Haiti; Honduras; Jamaica; Martinique; Mexico; Montserrat; Nicaragua; Panama; Puerto Rico; Saint Barthélemy; Saint Kitts and Nevis; Saint Lucia; Saint Martin (French part); Saint Pierre and Miquelon; Saint Vincent and the Grenadines; Sint Maarten (Dutch part); Trinidad and Tobago; Turks and Caicos Islands; United States; Venezuela, Bolivarian Republic of; Virgin Islands, British; Virgin Islands, U.S.
|FAO Marine Fishing Areas:|
Atlantic – western central; Atlantic – northwest; Atlantic – northeast
|Range Map:||Click here to open the map viewer and explore range.|
|Population:||Data were primarily collated from the Atlantic States Marine Fisheries Commission (ASMFC) American Eel Benchmark Stock Assessment (ASMFC 2012) – the body that manages American Eel stocks along the Atlantic Coast of the U.S. - and from the DFO (Canada’s Department of Fisheries and Oceans) Recovery Potential Assessment (RPA) report (DFO 2014). There were very few data that related to the Southern extent of the species’ range – Central America / Caribbean – and this was taken into account during discussions. Data that were eel-specific (i.e., not part of a multi-species fishery) and/or collected as part of fisheries independent scientific monitoring or fisheries dependent data with an associated metric of effort (e.g., catch per unit effort) were given greater weight than data sets that did not meet these criteria. Catch effort can be variable in fishing data and under-reporting, and in some cases, an absence of reporting of landings is a serious problem; thus while landings data are discussed, in the absence of other data, it cannot be accepted as a precise measure of stock status. Data were analysed to assess trends in recruitment, population and spawner escapement, however, they were primarily used to guide further discussions that included threats and associated management measures. These and the associated review inputs resulted in the consensus of the ‘Endangered’ listing by assessors, contributors and reviewers.|
According to a regional assessment and status report by the Committee on the Status of Endangered Wildlife in Canada (COSEWIC), eel abundance has seen significant declines in the last 50 years, and COSEWIC assessed the Canadian population that receives recruitment at the northernmost part of the species' range as ‘Threatened’ (COSEWIC 2012). During the ASMFC Baseline Stock Assessment for American Eel, nearly 100 fishery-dependent and independent U.S. data sources representing several life stages and geographical and temporal scales were evaluated. Ultimately, fifty-two fishery-dependent and independent data sources were selected for use in this assessment because they were considered adequate for describing life history characteristics and abundance trends of eels on either a coast-wide or regional basis. The assessment concluded that in the US the American Eel stock is depleted with the stock being at or near historically low levels (ASMFC 2012). To provide a note of historical context, Haro et al. (2000) analysed commercial and non-commercial data to determine population trends of A. rostrata along the east coast American/Canadian range from Ontario and Nova Scotia down to Virginia. Of the 16 localities analysed between the years 1984 and 1995, seven areas showed significant declines in population numbers using Mann-Kendall trend analysis, while the remaining nine were non- significant.
Further to this, time series data from the National Marine Fisheries Service (NMFS) and the FAO provide details of fisheries declines in the American Eel from the southern part of its range. In Florida, a state that has sustained an eel fishery on record since 1950, numbers were highest between 1975 and 1980, with average annual catches peaking at 213 t. The last reported figures for this fishery were from 2007 with an average of 2.11 t in the five preceding years (NMFS NOAA 2012, FAO 2012). Further south, Cuba reported 0.5–1.0 t annually during the 1980s and early 1990s, but no eel catches have been reported to FAO since 1995. Conversely, in the Dominican Republic, fishing for eels is relatively recent with first records appearing in 1995 (five year average = 4 t). From 2006 to 2011 this figure has increased to 22 t. Figures from the western Atlantic Coast in Mexico began in 1984 and are characterised by long inter-annual periods of no eels caught. The most consistent periods were 1995 to 2000 with an average catch of 18 t per year before falling back to zero (NMFS NOAA 2012, FAO 2012). In 2011, an anomaly in this term saw a remarkable 140 t caught suggesting a sudden surge in effort. These fluctuating values from the southern part of the American Eel range perhaps reflect changes and surges in demand from abroad, which can quickly drive up the price of eels (see the Use and Trade section below). According to a 50-year average, the southern states of the U.S. contribute just 10 t per year to a commercial fishery with a combined total catch between Canada and the US averaging 1,564 t per year (Casselman 2003). Thus, fisheries-dependent data suggest that eels in the Mid-Atlantic U.S. and eastern Canada contribute disproportionally to the overall panmictic population (Barbin and McCleave 1997).
Glass eels/young yellow eels:
The data sets that were the main focus of analysis and framed further discussions were as follows: East River Chester (NS, Canada), Little Egg Inlet (NJ, United States), Beaufort Inlet (NC, United States). Moses Saunders, upper St. Lawrence River-Lake Ontario recruitment (ON and NY, Canada/United States), Richelieu River/Chambly Dam (QC, Canada), Highlands River and Northeast Brook (NL, Canada), Trepassey (NL, Canada), Rivière du Sud-Ouest (QC, Canada), Long Island (NY, United States), Hudson Estuary (NY, United States), Delaware River (NJ, United States), Delaware River (DE, United States), Chesapeake Bay (MD, United States), Lower Chesapeake Bay and Tributaries (VA, United States), North Carolina trawl survey (NC, United States).
Recruitment of young eels to areas of historically high abundance, such as the St. Lawrence River watershed, has been in decline for some time (Casselman et al. 1997, Casselman 2003). More recently some of these declines have slowed and appear to have stabilised and some watersheds have recently exhibited increasing trends in glass eels and elvers, but these have yet to manifest themselves as increases in continental populations (DFO 2014). This means that the population at this locality is potentially more vulnerable to future threats because of very low abundance. A few long-term data sets from eel ladders are available, and these have shown a major decline in recruitment. A severe decline has occurred in the number of eels ascending the eel ladder at the Moses-Saunders Dam (Ontario-New York) during the peak spring migration in the St. Lawrence River-Lake Ontario segment of the range (Castonguay, Hodson et al. 1994, Casselman et al. 1997, Casselman 2003, COSEWIC, 2006). The number of juvenile eels climbing the eel ladder declined from more than one million per year in the early 1980s to fewer than 4,000 per year in the late 1990s and to a level approaching zero by 2001. Larger numbers of eels have begun climbing the ladders again starting in 2004 indicating a reversing trend (MacGregor et al. 2013).
In the US, there has been monitoring in the Hudson Estuary since the mid-1970s – this is fisheries independent but not species specific. Catches of glass eels/elvers indicate that there has been a significant decline in recruitment over this period. Further south, the situation in the Chesapeake and Delaware bays are less clear cut, over three generations, with some indices showing declines, others stable and some exhibiting increased recruitment. However, the general trend is that of a recruitment decline in this part of the US.
The American Eel exists in the Gulf of Mexico, Caribbean islands, and some Atlantic coastal areas of North and South America, however data for these regions is scant. Commercial fisheries have existed in Mexico and the Caribbean islands, specifically the Dominican Republic. FAO and FishStatJ indicate that harvest in Mexico and the Caribbean islands increased dramatically in the mid-1990s correlating with the decline in the Canadian harvest. In 2011 harvest in both jurisdictions reached a record-high, approaching almost 20% of total harvest of the American Eel (Casselman unpub. data). Evidence is that this harvest was exported to Asia for culture. This increase could be attributed to increasing abundance, but price and demand for American glass eels and elvers has increased substantially in recent years, likely due to the export of the European Eel outside of the EU being banned in 2010. As such it is more than likely that this demand for young American Eels for export is driving this increased harvest.
Overall, despite regional variations, and in some cases, signs of stabilisation and/or recovery in the recent past, the general picture across the data sets used, was that over the period of three generations analysed there had been a significant decline in recruitment across the species' range.
Older yellow eels:
The data sets that were the main focus of analysis and framed further discussions were as follows: Bay of Quinte (ON, Canada) St Lawrence River (ON, Canada), Miramichi River, (NB Canada), Gulf of St Lawrence (Canada), Upper Mississippi River (United States).
All five of the data sets indicated that there had been declining trend in continental yellow eel populations over the period of three generations. Fisheries-dependent data indicate that eel harvest declines first occurred along the Atlantic Seaboard of the U.S. in the mid to late-1980s, followed by a decline in the mid-1990s in Canadian harvest. One hypothesis for this is that eels harvested in Canada are older than those harvested in the United States - mean age of silver eels is 14.1 compared to 7.6 years respectively. Hence these decreases are virtually synchronous when age is taken into consideration, reflecting an overall declining abundance of the species (Casselman 2003).
The data sets that were the main focus of analysis and framed further discussions were as follows: Lower St. Lawrence River x2 (QC, Canada), Rivière du Sud-Ouest (QC, Canada).
The three data sets showed varying changes in silver eel abundance over the period of three generations, though all were in decline. The 2012 COSEWIC assessment initiated an official review of the recovery potential of the species in Canada (DFO 2014). The number of mature individuals within the Great Lakes has been estimated from previous spawning escapement data by Verreault and Dumont (2003), who predicted silver eel departures from the upper St. Lawrence River and Lake Ontario from a model based on eel passage, commercial landings, percentage of migrating eels in the commercial catch and turbine survival rate. Recent estimates (fisheries independent data including electrofishing CPUE and tailwater surveys at dams) indicate that total spawning escapement has decreased by 65% over the period of 14 years despite a reduction in mortality from commercial fisheries (50% of fishing effort between 2002 and 2009) (G. Verreault, MRNF pers. comm. 2011; COSEWIC 2012). A number of datasets exist for the St. Lawrence River system and as declines were more extreme and occurred earlier upstream, there is value in their use in demonstrating the depth and breadth of the decline and its progression.
|Current Population Trend:||Decreasing|
|Habitat and Ecology:||Habitat: |
The species is found in a range of habitats from small streams to large rivers and lakes, and in estuaries, lagoons, coastal waters and possibly further offshore. Under natural conditions, it only occurs in water bodies that are connected to the sea, although it is stocked elsewhere. American eels are sometimes found in landlocked lakes, particularly in the northeast U.S. (Facey and Van Den Avyle 1987). The latitudinal range of the American Eel is extremely broad, from 5°N to 60°N (Bertin 1956), and their range covers approximately 30,000 km of coastline. The American Eel is thought to occupy the broadest range of habitats of any fish in the world (Helfman et al. 1987).
As both larvae and adults, the American Eel migrates through and spawns in the ocean. Research has indicated that coastal and estuarine waters could be appropriate habitat for eels within the species' range (Chaput et al. 2014). It has further been suggested that, as a result of this fact, only a small proportion of available habitat is subject to the threat of exploitation (Cairns et al. 2012). Indeed, the yellow/silver eel fishery in the southern Gulf of St Lawrence was estimated to harvest about 7–8% of the standing stock biomass over the period of 1997–2008 (Chaput et al. 2014). However, it was felt that to make the assumption that this habitat is populated simply because it is theoretically appropriate is a risky strategy in terms of conservation. As such, and as was also the case for the other anguillid assessments, a precautionary approach has been employed and only areas where data on the presence of eels was available has been included in analysis.
As a catadromous species, American Eels use marine habitats for migration to their spawning area, spawning, and egg and larval phases. There is evidence that American Eels are facultatively catadromous, and yellow eels use estuarine—and possibly marine—habitats during the growth phase. Little is known about the precise spawning habitat of the American Eel. Spawning location is inferred based on size of the leptocephali - the larval stage of anguillid eels - with the smallest leptocephali captured over a relatively large 550 km arc in the southern Sargasso Sea within the frontal area of the Subtropical Convergence Zone (ASMFC 2012, DFO 2014). The northern limit of spawning appears to be defined by temperature fronts separating surface water masses with high temperature and higher-saline water to the south and seasonally cooled, lower-saline water to the north (McCleave et al. 1987, Kleckner and McCleave 1988, Munk et al. 2010). These highly dynamic fronts vary in shape and locations within months and between years.
Breeding in the wild has never been observed, however, based on size of the leptocephali that have been captured, spawning activity occurs in late winter and early spring, with the smallest leptocephali being present in February to April. The Sargasso Sea is a high-salinity (~36.6 ppt) region with warm surface temperatures ranging from 18 to 24°C (Kleckner and McCleave 1985, Kleckner and McCleave 1988). Some evidence suggests that American Eels may spawn in the upper 150–200 metres of the water column (Kleckner et al. 1983, Kleckner and McCleave 1985). Leptocephali, drift and swim with prevailing currents (Antilles Current, Florida Current, and Gulf Stream), which take them to areas near continental coasts or continental-slope waters. The leptocephali undergo metamorphosis and develop into glass eels. Eventually glass eels metamorphose into elvers, a smaller version of yellow eels. It is believed that metamorphosis from leptocephali to glass eels occurs in the first year primarily at the edge of the continental shelf; active transport is thought to be necessary for leptocephali and glass eels to exit the Gulf Stream. In the Gulf of St. Lawrence, migration of glass eels is 10–15 km/day primarily at night (Dutil et al. 2009). Leptocephali and glass eels appear to be tolerant of a wide range of temperatures and salinities.
While many American Eels appear to migrate into freshwater, some settle in estuaries or near-shore coastal waters and may complete their yellow phase without ever entering and remaining in freshwater (MacGregor et al. 2008, Greene et al. 2009). Jessop et al. (2009) reviewed studies which used otolith microchemistry to investigate lifetime use of salinity zones by American Eels; eels can be divided into residents of freshwater, residents of brackish or salt water, and shifters between these habitats. Secor et al. (2002) found these three modes of habitat use by yellow-phase eels in the Hudson River. Despite knowledge of these modes, there is still little known about the American eels’ use of the nearshore marine environment, but there are some reports of small numbers of eels being captured in Canadian brackish waters in early summer, suggesting that they are not emigrating individuals. These captures have occurred at a wide range of depths, including at locations >100 m deep (DFO 2014).
Movement into freshwater may be the result of a density-dependent process (Haro and Krueger 1991, Feunteun et al. 2003), with the largest immigrating pulses migrating farthest (Ibbotson et al. 2002, Casselman 2003). Glass eels enter estuaries and a proportion of these move upstream to spend the majority of their life growing as yellow eels in rivers, streams, ponds, freshwater marshes and the shallow, more productive areas of lakes (DFO 2014). Environmental predictors of glass eel runs are variable, but increased temperature and reduced flow may trigger upstream movement (Greene et al. 2009). This immigration begins when water temperatures reach 10°C and continue until temperatures exceed 20°C, though lower temperatures can decrease migratory activity (Martin 1995). Glass eels appear to be attracted to flowing water with a current velocity <25-40 cm/sec (Greene et al. 2009).
Significant environmental, temporal and spatial factors affect the distribution of the American Eel, including the distribution of habitat areas (Greene et al. 2009). Greene et al. emphasized that much of the American Eel habitat has not been well quantified and has been considerably affected by human actions. Yellow eels have a primarily benthic orientation, use substrate (rock, sand, silt, and mud) and other structures such as woody debris and submerged vegetation (well-documented Vallisneria, eelgrass) for protection and cover. Casselman (2003) reports that historically in the upper St. Lawrence River and Lake Ontario, eels constituted half the inshore (<10 m) fish biomass.
Yellow-phase American Eels spend 3 to 30 or more years inland or in coastal areas before becoming mature, entering the silver phase (Daverat et al. 2006, MacGregor et al. 2009). Silvering eels emigrate from freshwater estuaries and near-shore coastal areas back to the open ocean and migrate to their spawning area in the Sargasso Sea. The timing of silver eel downstream migration occurs on a latitudinal cline, with eels leaving the Canadian provinces in summer through fall and from fall to early winter in the southern U.S. depending upon location. Emigration appears to occur over a narrow timeframe, usually a few months and appears to be initiated by environmental factors such as water temperature, river or stream discharge, water level and/or precipitation events, and light intensity, including moon phase (DFO 2014). Silver eels emigrate from estuaries at water temperatures ranging from 10 to 18°C, with a laboratory preference of 17°C (Haro 1991). During this migration, silver eels appear to remain relatively near the ocean surface; they have been captured in November and December in depths ranging from 9 to 82 m and at water temperatures from 8 to 12°C (DFO 2014). Physiological changes associated with silvering have been reviewed by Facey and Van Den Avyle (1987) and include a change in colour to a metallic bronze-black sheen, pectoral fin change from yellow-green to black, increased body fat content, thickening of the skin, increased length of the capillaries in the rete gland of the swim bladder and degeneration of the digestive tract. Additionally, the eyes become larger and the visual pigments in the eye are altered (Vladykov 1973, Beatty 1975). These changes are thought to better suit the American Eel for migration at depth.
The American Eel is tolerant of a wide range of temperatures. In Chesapeake Bay, it has been captured at temperatures ranging from 3 to 31°C, with the greatest occurrence at 16–28°C (DFO 2014). Water temperature is important in the seasonal activity of eels; in winter, when temperatures are <5°C, eels bury themselves in soft substrate and are believed to enter a state of torpor (Greene et al. 2009). In Atlantic Canada, some eels that are resident in brackish and salt water remain there in winter, but wintering migrations from salt to fresh water are also known (Jessop et al. 2009, Tomie et al. 2013). Although American Eels lack plasma antifreeze, they commonly overwinter in areas where water temperatures fall below the freezing point of fish blood, apparently by taking advantage of warmer temperatures found in bottom mud (Tomie 2011). The occurrence of eels may also be affected by dissolved oxygen levels. Catches are most abundant in waters where dissolved oxygen is >4 mg/L. Any decrease in oxygen concentration will negatively affect growth and production (Greene et al. 2009). Barriers that restrict or impede upstream movement lead to increased concentrations of eels immediately downstream. It has been suggested that higher densities of eels in these regions can increase cannibalism, predation, and competition for food and have a negative effect on growth and ultimately size and fecundity (DFO 2014). However, small eels, <10 cm in length, can traverse vertical or nearly vertical wet barriers (Greene et al. 2009).
Sex determination in anguillids is, at least to some extent, environmentally determined and appears to be a function of density and growth rate, with males occurring at higher local population densities (Krueger and Oliveira 1999, Davey and Jellyman 2005). Consequently and like other species of freshwater eels, regional sex bias is common in the American Eel, potentially exacerbating the impact of localised threats. The once large and abundant eel stock of the upper St. Lawrence River–Lake Ontario system was exclusively females that were large, old, and potentially highly fecund (Casselman 2003). As such, it has been argued that the northern portion of the geographic range of the American Eel may contribute a great proportion of the reproductive potential to this panmictic species because of apparent increases in average female size and female bias with latitude (Barbin and McCleave 1997). With this apparent increase in length and weight comes an increase in fecundity for eels residing at higher latitudes with a single large female caught in a river in Maine reported to contain 21.95 million eggs (Barbin and McCleave 1997). Counter to these arguments, however, is the contention that individuals at lower latitudes grow more quickly, reaching maturity at a younger age, and are more abundant and have a shorter generation time than at more northern latitudes (ASMFC 2012). Even though the individuals are smaller, this may result in a greater overall reproductive capacity.
To assess decline in a species with a complex life history such as anguillid eels, it is important to be able to estimate generation time for the species. Generation lengths for American Eels can vary considerably. American Eels living in brackish habitats, on average, grow 2.2 times faster than freshwater residents, possibly because of increased food resources and better osmotic conditions (Jessop et al. 2002). Consequently eels in freshwater habitats have extended times to maturity and are therefore more vulnerable (freshwater, 22 yr; estuaries 9 yr – Lamson et al. 2009). This is complicated by the fact that the species inhabits a broad range of latitudes, producing widely varied growth rates and ages at silvering (Jessop 2010).
Comparisons of eels commercially harvested at inland areas (St. Lawrence River-Lake Ontario) with those harvested along Atlantic coast provinces revealed that average lengths range from 80.3 to 46.3 cm and average ages from 17.5 to 10.6 between freshwater and estuarine or brackish habitats, respectively (Casselman 2003). The temperature-size rule (increase in body size at lower temperatures) applies to American Eel females, but not males (Jessop 2010). Upper and lower bound estimates for silver eel age are provided from a broad study by Jessop (2010), who suggests minimum mean ages of 5.8 years for females (Charleston Harbour, SC) and 3.1 years for males (Altamaha River, GA). Maximum age estimates were 22.6 for females (Rivière du Sud-Ouest, Quebec) and 15.4 for males (East River, Chester, N.S.). Commercially harvested yellow eels from the upper St. Lawrence River and Lake Ontario usually ranged in age from 17 to 26 years (Casselman unpub. data). Summary data on later-stage yellow eel ages from commercial fisheries (Casselman 2003) indicate that in northern locations, the range is 4-43 yr (St. Lawrence River-Lake Ontario and Atlantic coastal provinces) and in the Atlantic coastal states of the U.S., 2-20 yr. These data suggest that upper and lower estimates of three generations for this global assessment of A. rostrata should be in the region of 65–70 yr and 8–10 yr. The ASMFC Benchmark Stock Assessment (2012) indicated that a generation time for eels in the south is 3–5 yr and, in the north, 10–20 yr.
A more precise estimate of generation time could be made by combining individual estimates for various regions. Ten general regions were selected to estimate generation time across the range of the species. These regions and generation times are: 1) Caribbean islands – 4 yr; 2) Gulf of Mexico – 6 yr; 3) southern Atlantic states – 8 yr; 4) central Atlantic states – 10 yr; 5) northern Atlantic states – 12 yr; 6) Scotia-Fundy – 14 yr; 7) southern Gulf of St. Lawrence – 16 yr; 8) northern Gulf of St. Lawrence – 18 yr; 9) St. Lawrence River system and basin – 20 yr; 10) Mississippi River system – 12 yr. These estimates come from an extensive review of ages in the published literature. Comparing mean age of female silver eels and estimated generation time in seven regions ranging from the St. Lawrence River Basin to the southern Atlantic states provided in the Recovery Potential Assessment (DFO 2014), there was a highly significant correlation with a slope not significantly different than 1 with the estimated generation time provided here. From these data, it was concluded that a mean overall generation time for the American Eel could be estimated to be 12 yr.
General estimates of generation time used by others confirm the aforementioned estimates. In the DFO Recovery Potential Assessment (2014), it was considered that one generation time for American Eeels in Canada is approximately 16 yr. The ASMFC Benchmark Stock Assessment (2012) considered that 8–10 yr would be a good estimate for U.S. jurisdictions encompassed by the ASMFC.
|Continuing decline in area, extent and/or quality of habitat:||Yes|
|Generation Length (years):||12|
|Movement patterns:||Full Migrant|
|Congregatory:||Congregatory (and dispersive)|
|Use and Trade:||
The various life stages, ranging from glass eel to adult, of all Anguilla species are harvested and traded on a global scale for consumption, with current demand predominantly driven by East Asian markets, in particular Japan and mainland China. A concerning pattern of exploitation is already apparent – when one Anguilla species or population becomes overexploited, industry moves to the next in order to fulfil demand (Crook and Nakamura 2013).
Anguilla spp. are traded internationally as live eels for farming and consumption, as fresh, frozen and smoked/prepared eels for consumption and as skins and leather products for fashion accessories. Global trade data collated by FAO for live, fresh, frozen and smoked/prepared Anguilla species (non-species specific) is available for the period 1976–2009. According to FAO data, global annual Anguilla exports averaged around 20,000 tonnes in the late 1970s (valued annually at 55–95 million US Dollars), after which annual exports showed a steady increase to a maximum of over 130,000 tonnes in 2000 (valued at over 1,000 million US Dollars). Since then annual exports have been declining, to just over 80,000 tonnes in 2008 and 2009 (valued at over 800 million US Dollars). By weight, China and Taiwan are responsible for nearly 75% of these exports and Japan for over 75% of all imports (FAO 2013).
Although American Eel has traditionally been consumed in small amounts domestically, in particular in the U.S. and Canada, harvesting is predominantly carried out with the purpose of export. Unlike in Europe or Asia, there is no tradition of farming eels in the Americas. It is well documented that American Eels were very important and widely used for sustenance, reverence, and practical purposes by Aboriginal peoples in prehistoric and historic times (Casselman 2003, MacGregor et al. 2009, Engler-Palma et al. 2013, Miller and Casselman 2014). Throughout the range, yellow and silver eels have also been fished commercially since European colonisation and more recently all life stages (excluding leptocephali) have been fished. The species was once extremely abundant in watersheds and tributaries in two of the largest reservoirs for the species — Lake Ontario and St. Lawrence River system and Mississippi River system — but has gradually declined since the turn of the 20th century, most rapidly from the 1970s to the present (Casselman 2003, MacGregor et al. 2013, Phelps et al. 2014).
Historically, glass eels and elvers have been commercially fished extensively along the Atlantic Coast of the U.S. for striped bass bait to supply farms in Asia (ASMFC 2012). Since 2001 glass eels are only fished in two states: Maine and South Carolina, with Maine greatly dominant in catch and activity (ASMFC 2012). Small yellow eels are still a popular bait fish for large striped bass fishery.
Global FAO catch data for American Eel is available for the period 1950–2011. In the 1950s and early 1960s, reported annual American Eel catch averaged 900 tonnes. This doubled to ~2,000 tonnes per year in the 1970s, after which there has been a gradual decline back to 1950 catch levels in the 2000s. Throughout this period, catch was split relatively evenly between the U.S. and Canada, with small intermittent annual catches of one to 50 tonnes reported by Mexico, the Dominican Republic and Cuba. In 2011, catches of young yellow eels from Mexico (140 t) and the Dominican Republic (72 t) were the highest ever recorded from these countries (see the Population section above). These catches include all life stages (FAO, 2013).
Historically, commercial harvest of yellow and silver eels in Canada and the U.S. has been approximately equal (MacGregor et al. 2009). From 1950 to 2010, U.S. Atlantic coast landings ranged from approximately 300 t in 1962 to 1,665 t in 1979 (ASMFC 2012). After an initial decline in the 1950s, landings increased to a peak in the 1970s and 1980s in response to higher demand from European food markets. In most regions, landings declined sharply in the 1990s and 2000s following a few years of peak landings. The value of U.S. commercial American Eel landings as estimated by NOAA Fisheries has varied from less than $100,000 (prior to the 1980s) to a peak of $6.4 million in 1997. Total landings value increased through the 1980s and 1990s, dropped in the late 1990s, and increased again in the 2000s. The average harvest along the Atlantic coast of the U.S. from 1980 to 2011 was 672 t. In recent years, over half of the harvest of yellow eels has occurred in the State of Maryland (ASMFC 2013). From 2004 to 2010, average annual harvest was 356 t but increased in 2011 to 529 t while harvest in Canada was 337 t and from Mexico and the Dominican Republic a greatly elevated harvest of 212 t. Total overall harvest of the American eel in 2011 was 1,078 t, with approximately 20% coming from the extreme southern part of the range.
American Eels, both live and frozen, have been exported to principal eel consuming markets in Europe and Asia for decades and consumed very little in North America (Miller and Casselman 2014). Between 1997 and 2007 the U.S. and Canada were globally the first and third most important exporters of frozen eel, together exporting 35% of the global total during this period, and on average 2,500 t per year. Over the last 15 years the U.S. and Canada have also been the first and third most important suppliers of live eels to Europe (all life stages), responsible for nearly 50% of all imports of live eels into the EU, averaging at 450 t a year (Crook 2010, Crook 2011).
Eel farming, which is responsible for over 90% of all Anguilla production worldwide (averaging at 280,000 tonnes per year since 2007, (FAO 2013)), is reliant on wild-caught juvenile eels or glass eels, as raising eel larvae to the glass eel stage in captivity has only had limited success as yet (Tanaka et al. 2003). The U.S. and Canada have exported A. rostrata glass eels to the principal East Asian eel farming countries/territories in small, irregular quantities for a number of decades. Between 1998 and 2010, live eel fry imports from A. rostrata range states into mainland China and Hong Kong ranged from 0.1 to 10 t per year. However, in 2011 combined imports increased suddenly - from just under 10 t in 2010 to nearly 54 t in 2011 and nearly 28 t in 2012; these high figures have been questioned, however, to date, no other data is available. This sudden change in export quantities coincided with the EU’s decision, at the end of 2010 to ban exports of European Eel, which had previously met a demand borne from a decline in A. japonica glass eel catches.
In 2012, for the first time in over ten years, Japan also imported live eel fry directly from the U.S. (Crook and Nakamura 2013, Crook in litt. 2013) - web-based advertisements in 2013 indicate that the price for A. rostrata glass eels has increased by orders of magnitudes in recent years. Maine and South Carolina are the only two U.S. States that allow commercial fishing of glass eels, however considerable levels of poaching and illegal trade, driven by the ever increasing prices offered for this commodity, have been reported from several States (Anon. 2012). Of the more southerly range of A. rostrata, Cuba, the Dominican Republic and Haiti have recently started exporting glass eels to Hong Kong and Korea (according to Asian Customs import data). A U.S.-based Asian-owned company advertises setting up camps along the U.S. coast and in the Caribbean to harvest and export glass eels to Asia (Glass Eel Farm 2012). However, at present only two companies (of 34 companies that applied for permits) are authorized to harvest glass eels in the Dominican Republic and illegal harvesting and trade via Haiti has become an issue of considerable concern to Dominican authorities (Anon. 2013).
In April 2012, the US Fish and Wildlife Service announced that it was considering submitting a CITES listing proposal for A. rostrata (USFWS 2012) due to concerns over the impact increased international trade in this species was having on the population, however no proposal was submitted.
Note: double-counting, under-reporting and misreporting must be taken into consideration when interpreting all available catch and trade data. See Crook (2010) for explanations of data issues.
Many have detailed the pressures that are believed to have caused, through synergistic interactions, the decline in recruitment, standing stock, and ultimately in the production of spawners of the American Eel. In an extensive review of status of the American Eel in Canada, COSEWIC (2012) identified the threats as: 1) barriers in freshwater resulting in the accumulative loss of formerly productive habitats; 2) turbine mortality of hydroelectric dams; 3) vulnerability of life stages to fisheries; 4) bioaccumulation of contaminants; 5) exotic swimbladder nematode parasite (Anguillicola crassus) potentially introduced from stocking; 6) climate change and shifting oceanic conditions. A recent report indicated that barriers to migration and fisheries were likely the greatest threats across the species’ range, while other threats would have more regional influences (Chaput et al. 2014). Many of the current major stressors affecting A. rostrata in North America are those facing other species of anguillid eels. Large barriers severely impede upstream migration of juvenile eels if no passage is provided, extirpating large areas of habitat suitable for growth and maturation of freshwater eels. Although eels are able to ascend many smaller barriers, recent studies have documented a tenfold reduction in eel density above each potentially passable barrier (Machut et al. 2007). Machut et al. also found significantly lower American Eel condition with increasing riparian urbanization in the Hudson River in the U.S.
Habitat loss resulting from impoundment may contribute to reduced eel abundance in eastern Canada (Jessop 2000). In the St Lawrence watershed alone, 13% (14,000 km²) of yellow eel rearing habitat is no longer accessible (DFO 2014). However, in the late 1990s and early 2000s a number of eel ladders were constructed on barriers which attempted to reduce this loss. Throughout the United States, it is reported that as much as 84% of historic stream length is now inaccessible to eels (DFO 2014). The Susquehanna River, the largest river system on the East Coast of the U.S., once supported a large numbers of eels; however the construction of Conowingo Dam, a large hydroelectric dam located near the mouth of the river as well as additional upstream dams severely limited the access to this habitat (SRAFRC 2010). Major habitat perturbations in the St. Lawrence River took place in the 1950s (e.g., construction of the St. Lawrence Seaway and hydroelectric dams), about 30 years before recruitment started declining; this long delay argues against these perturbations being primary causes of the decline (Castonguay et al. 1998). However, an unpublished, long-term CPUE series (1930–1965) provides evidence that the decline occurred simultaneously with the habitat alterations (see Castonguay et al. 1998). Irrespective of the causes of declines, it has been proposed that changes in flow regulation, along with changes in temperature, have affected migration behaviour (e.g., Verreault et al. 2012).
Passage through turbines at hydropower dams during downstream migration represents a major source of eel mortality (Ritter et al. 1997). Turbine-induced mortality ranges from 5 to 97%, depending on turbine type, flow rate, and length of the fish (Hadderingh 1990, McCleave pers. comm., Bussye et al. 2014). Cumulative direct turbine mortality from the two large hydroelectric dams on the St. Lawrence River system (Moses-Saunders and Beauharnois) accounts for a mortality of 40% of the emigrating eels (MacGregor et al. 2009). This had a substantial impact on spawner escapement from this once large, highly fecund northern population. Without development of suitable eel passage at hydroelectric dams, such mortality continually affects escapement at all hydroelectric facilities across the entire range.
The commercial harvest of all continental life stages of eels has been, and remains, a cause of significant mortality (see the Use and Trade section above for details). How accurately, effort-based commercial harvest data represent true population trends however, remains a strong discussion point due to the potential carrying capacity of unfished habitat. However, for the purposes of the current assessment, these areas, while being recognised as having potential importance, were not considered unless fisheries-independent monitoring data was available. Worldwide demand for eels is greater than can be supplied by harvest of wild populations, and eel farming – widespread in Japan, Taiwan, Korea and China and to a lesser extent in Europe – has become a major source of marketable eels (Jessop 2000). In Asia, wild-caught glass eels and elvers are farmed to marketable size. Since captive reproduction of American eels is not yet feasible (although see Oliveira and Hable 2010), the intensive farming industry in East Asia is dependent upon an annual supply of wild-caught glass eels and elvers (Moriarty and Dekker 1997). When Asian domestic stocks are inadequate, a strong market develops for American Eels. The Asian market for American glass eels and elvers was strong from 1972 to 1977, declined dramatically in 1978, began to strengthen in the 1990s, and is now at record-high levels (Crook and Nakamura 2013). Increased worldwide demand for glass eels and elvers has led to increased poaching of the American Eel in the United States (ASMFC 2013) though the extent of this is difficult to quantify.
In addition to threats from harvest, they also face other challenges. American Eels from the St. Lawrence River have historically been identified as being heavily contaminated with chemicals such as PCBs, Mirex, and various pesticides (Hodson et al. 1994, Couillard et al. 1997, Castonguay et al. 1998). Lethal toxicity from chemical contamination has been known to occur in St. Lawrence eels for more than 30 years (Castonguay et al. 1998). The highest concentrations of chemicals in migrating silver eels in the St. Lawrence River are in the gonads; chemical levels in eggs could exceed the thresholds of toxicity for larvae (Hodson et al. 1994). Due to PCB body burdens in eels, commercial harvest of eels for human consumption has been banned throughout the Hudson River south of Glen’s Falls, NY; however, PCBs are currently being removed from the Hudson River. More recent studies indicate that levels of toxins are lower (Byer, Alaee et al. 2013), and while in some cases the negative effects are considered to be minimal (e.g. dioxins; Byer, Lebeuf et al. 2013), there is still regional concern in relation to other xenobiotics.
Oceanic effects on long-term patterns of American Eel recruitment are poorly understood, but they may play a role in the changing abundance of eels along the Atlantic coast of North America (Castonguay, Hodson, Moriarty et al. 1994). At the larval stage, changes in marine primary production and thus food availability associated with climate change have been suggested as a cause of population decline in the American as well as the European and Japanese Eels (Bonhommeau et al. 2008). Changes in frontal locations or oceanic currents may also play a role in determining recruitment to the coast (Miller et al. 2009). It has recently been discovered that predation of adult eels may be a further impact in the oceanic environment (Béguer-Pon et al. 2012, Wahlberg et al. 2014).
Given the relative lack of understanding of the threats we have attempted to quantify this using the IUCN ‘Threat Classification Scheme’ coding, however, this is by no means definitive.
In the late 1990s and early 2000s there were several conservation initiatives in existence relating to fisheries and barriers to migration. However, the primary initiative for conservation concern and action for the American eel was initiated with the 2003 Quebec Declaration of Concern (Dekker et al. 2003) to restore the world’s anguillid eels, produced at the AFS 2003 International Eel Symposium (Casselman and Cairns 2009). But by comparison with the European Eel, integrated management has not been initiated. In December 2010, European Union member states suspended all exports and imports of European Eel commodities, placing further pressure on the American Eel to be used as seedstock for farms.
In Canada, the federal Department of Fisheries and Oceans manages American Eels in the Maritime provinces, with the powers of management transferred to the province for both Ontario and Quebec. As a result of significant declines in the upper St. Lawrence River and Lake Ontario, the Ontario government established a harvest quota. However, because of the precipitous decline in abundance, commercial quotas were never met, although they were reduced annually. In 2004, the Ontario Ministry of Natural Resources officially closed all commercial fisheries of eels as a result of the dramatic decline of eels in Lake Ontario and the St. Lawrence River (Casselman 2003). In 2007 eels in Ontario were officially classified as “Endangered” under the Ontario Endangered Species Act and a Recovery Strategy was prepared to provide advice to government (MacGregor et al. 2013). The Recovery Strategy in Ontario is currently being assessed by the Ontario government and a formal ‘Response Statement’ is being prepared. In Ontario, there is emphasis on restoring, protecting, and diversifying suitable habitat with proposed 10% increases in availability of this habitat every year for five years (MacGregor et al. 2013). A reduction in cumulative mortality (50% at the watershed level by 2050) is also designed to enhance escapement across the species range in Canada in another target set out in the DFO American Eel management plan (MacGregor et al. 2009). COSEWIC, in 2006, reviewed the status of the American Eel in Canada and recommended “Special Concern” but with the caveat that the species should be reassessed in five years. In May 2012, COSEWIC upgraded the Canadian status to “Threatened”.
In 2008, Aboriginals assembling in Ottawa, Canada, concerned about the decline of eels, prepared an Aboriginal People’s American Eel Resolution, suggesting that eels in Canada be considered “Threatened” and a willingness to give up long-standing rights with eels to see them restored. The resolution expresses concern about the harvest of glass eels and encourages co-operative action and the application of precautionary principles. This early resolution was signed by three large Aboriginal First Nations: the Algonquins of Ontario; the St. Lawrence Iroquois, represented by the Mohawks of Akwesasne; and the Mi'kmaq of the Maritimes. The Algonquins of Ontario have provided documents on the importance of the species and the need to see it returned to its historic range and their ancestral lands to guarantee their association in perpetuity (Algonquins of Ontario 2012).
In the United States, the U.S. Fish and Wildlife Service and the U.S. National Marine Fisheries Service reviewed the status of the American Eel in 2007 in response to a petition it received in 2004 to list it under the U.S. Endangered Species Act (ESA). A 90-day review determined that protection under the ESA may be warranted. After examining all available information about the eel population from Greenland south along the North American coast to Brazil in South America and as far inland as the Great Lakes and the Mississippi River drainage, the Service found declines of eel populations in some areas. After a more extensive review the Service proposed that declines of eel populations in some areas had not put the overall population in danger of extinction. In 2010 the Service received another petition seeking to extend federal protection to the American Eel. The Service found that this petition, from the Council for Endangered Species Act Reliability, presented substantial new information that warranted the initiation of another more extensive status review of the species (FR 2011). This review will be competed in 2015 and will determine whether listing the species under the ESA as Threatened is warranted.
American Eels along the Atlantic Coast of the U.S. are managed by the ASMFC . Increasing demand in the 1990s produced what were considered to be unsustainable harvest levels of eels. In 1999, the ASMFC American Eel Fisheries Management Plan (FMP) was put in place to ensure ecological stability and promote sustainable fisheries. This included yearly monitoring of glass eel recruitment followed in 2006 by a mandatory catch and effort monitoring program. Since the late 1990s, American Eel fisheries are managed by the ASMFC in territorial seas and inland waters along the U.S. Atlantic coast from Maine to Florida. Increasing demand for eel by Asian markets and domestic bait fisheries, coupled with concern about declining eel abundance and limited assessment data, spurred development of the first Interstate Fishery Management Plan (FMP) for American Eel in the mid-1990s. In 2006 the ASMFC initiated the first stock assessment for American Eels. After extensive peer review the stock assessment was published in 2012. The 2012 benchmark stock assessment found that the American Eel population is depleted in U.S. waters. The stock is at or near historically low levels due to a combination of historical overfishing, habitat loss, food web alterations, predation, turbine mortality, environmental changes, toxins and contaminants, and disease (ASMFC 2012). Through multiple addendums to the American Eel Fishery Management Plan some small reductions in mortality across some life stages has been achieved, however, the need for, and extent of, additional quotas and/or reductions on the glass and yellow eels life stages have not been resolved (ASMFC 2013).
Dam removal and installation of eel ladders along the U.S. portion of the range is now a common restoration strategy, however, strategies to protect downstream migration though hydroelectric facilities are rare. Eel passage is currently being addressed on Conowingo Dam and some eel passage has been provided. Fish passage on other large hydro dams is addressed during the U.S. Federal Energy Regulatory Commission’s licensing or relicensing process. Most states are working to improve upstream eel passage on unregulated dams. A recent study shows that dam removal can increase American Eel abundances in the upstream areas (Hitt et al. 2012).
Increasing spawner escapement is critically important in recovery strategies, particularly involving fishing and turbine mortality, which need to be reduced and mitigated (MacGregor et al. 2013), however, governance remains a major problem for the American Eel and integrated management for this panmictic species is lacking. Another possible mitigation measure involves supplementary stocking. Such stocking programs were undertaken from 2005 to 2010 in the St. Lawrence River system, on the Richelieu River, and in the upper St. Lawrence River and Lake Ontario. The contribution of these stocked eels to the production of silver eels and subsequently to recruitment in the next generation is unknown; however, there were unexpected effects. For example, a study by Pratt and Threader (2011), of the stocking programme in the upper St.Lawrence River system showed strong evidence that stocked eels survived and grew quickly - female stocked eels started to emigrate within four years after stocking at a small size (compared to wild females), similar to the size reported for migrating wild females in their site of original capture in Nova Scotia. This said, a considerable number of the stocked eels became male in an environment where only females had previously been observed (Pratt and Mathers 2011, Pratt and Threader 2011) and the effects of this shift in sex bias and reduced size are unknown. The authors themselves highlight the limitations of this study in the context of the species' range and more long-term data are required to fully determine the effectiveness of stocking programmes. Restocking programmes, while having been shown to increase escapement in some instances, must be carefully planned and executed in river systems that will benefit from an increase in freshwater population. It is obvious that conservation concerns about the status of the American Eel are now considerable and that international cooperation and coordinated management are needed to protect a species that now has substantially reduced habitat and increasing anthropogenic stressors, with greatly reduced abundance and distribution.
|Amended reason:||This amended assessor removes one of the assessor's names and moves it to the contributors field at his request.|
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