# Endangered and Threatened Wildlife and Plants; Endangered Species Act Listing Determination for Alewife and Blueback Herring

> Briefs, arguments, decisions, and more.

URL: https://www.frixlaw.com/law-library/documents/fr%3A2013-19380

## Record

- **Collection:** Federal Register
- **Document type:** Notice
- **Published:** August 12, 2013
- **Citation:** 78 FR 48944

## Text

DEPARTMENT OF COMMERCE
National Oceanic and Atmospheric Administration
[Docket No. 111024651-3630-02]
RIN 0648-XA739
Endangered and Threatened Wildlife and Plants; Endangered Species Act Listing Determination for Alewife and Blueback Herring

AGENCY:

National Marine Fisheries Service (NMFS), National Oceanic and Atmospheric Administration (NOAA), Commerce.

ACTION:

Notice of a listing determination.

SUMMARY:

We, NMFS, have completed a comprehensive review of the status of river herring (alewife and blueback herring) in response to a petition submitted by the Natural Resources Defense Council (NRDC) requesting that we list alewife (
Alosa pseudoharengus
) and blueback herring (
Alosa aestivalis
) as threatened under the Endangered Species Act (ESA) throughout all or a significant portion of their range or as specific distinct population segments (DPS) identified in the petition. The Atlantic States Marine Fisheries Commission (ASMFC) completed a comprehensive stock assessment for river herring in May 2012 which covers over 50 river specific stocks throughout the range of the species in the United States. The ASMFC stock assessment contained much of the information necessary to make an ESA listing determination for both species; however, any deficiencies were addressed through focused workshops and working group meetings and review of additional sources of information. Based on the best scientific and commercial information available, we have determined that listing alewife as threatened or endangered under the ESA is not warranted at this time. Additionally, based on the best scientific and commercial information available, we have determined that listing blueback herring as threatened or endangered under the ESA is not warranted at this time.

DATES:

This finding is effective on August 12, 2013.

ADDRESSES:

The listing determination, list of references used in the listing determination, and other related materials regarding this determination can be obtained via the Internet at:
http://www.nero.noaa.gov/prot_res/CandidateSpeciesProgram/RiverHerringSOC.htm
or by submitting a request to the Assistant Regional Administrator, Protected Resources Division, Northeast Region, NMFS, 55 Great Republic Drive, Gloucester, MA 01930.

FOR FURTHER INFORMATION CONTACT:

Kim Damon-Randall, NMFS Northeast Regional Office, (978) 282-8485; or Marta Nammack, NMFS, Office of Protected Resources (301) 427-8469.

SUPPLEMENTARY INFORMATION:

Background

On August 5, 2011, we, the National Marine Fisheries Service (NMFS), received a petition from the Natural Resources Defense Council (NRDC), requesting that we list alewife (
Alosa pseudoharengus
) and blueback herring (
Alosa aestivalis
) under the ESA as threatened throughout all or a significant portion of their ranges. In the alternative, they requested that we designate DPSs of alewife and blueback herring as specified in the petition (Central New England, Long Island Sound, Chesapeake Bay, and Carolina for alewives, and Central New England, Long Island Sound, and Chesapeake Bay for blueback herring). The petition contained information on the two species, including the taxonomy, historical and current distribution, physical and biological characteristics of their habitat and ecosystem relationships, population status and trends, and factors contributing to the species' decline. The petition also included information regarding potential DPSs of alewife and blueback herring as described above. The following five factors identified in section 4(a)(1) of the ESA were addressed in the petition: (1) Present or threatened destruction, modification, or curtailment of habitat or range; (2) over-utilization for commercial, recreational, scientific, or educational purposes; (3) disease or predation; (4) inadequacy of existing regulatory mechanisms; and (5) other natural or man-made factors affecting the species' continued existence.

We reviewed the petition and determined that, based on the information in the petition and in our files at the time we received the petition, the petitioned action may be warranted. Therefore, we published a positive 90-day finding on November 2, 2011, and as a result, we were required to review the status of the species (e.g., anadromous alewife and blueback herring) to determine if listing under the ESA is warranted. We formed an internal status review team (SRT) comprised of nine NMFS staff members (Northeast Regional Office (NERO) Protected Resources Division and Northeast Fisheries Science Center staff) to compile the best commercial and scientific data available for alewife and blueback herring throughout their ranges.

In May 2012, the ASMFC completed a river herring stock assessment, which covers over 50 river-specific stocks throughout the ranges of the species in the United States (ASMFC, 2012; hereafter referred to in this determination as “the stock assessment”). In order to avoid duplicating this extensive effort, we worked cooperatively with ASMFC to use this information in the review of the status of these two species and identify information not in the stock assessment that was needed for our listing determination. We identified the missing required elements and held workshops/working group meetings focused on addressing information on stock structure, extinction risk analysis, and climate change.

Reports from each workshop/working group meeting were compiled and independently peer reviewed (the stock structure and extinction risk reports were peer reviewed by reviewers selected by the Center for Independent Experts, and the climate change report was peer reviewed by 4 experts identified during the workshops). These reports did not contain any listing advice or reach any ESA listing conclusions—such synthesis and analysis for river herring is solely within the agency's purview. We used this information to determine which extinction risk method and stock structure analysis would best inform the listing determination, as well as understand how climate change may impact river herring, and ultimately, we are using these reports along with the stock assessment and all other best available information in this listing determination.

Alewife and blueback herring are collectively referred to as “river herring.” Due to difficulties in distinguishing between the species, they are often harvested together in commercial and recreational fisheries, and managed together by the ASMFC. Throughout this finding, where there are similarities, they will be collectively referred to as river herring, and where there are distinctions, they will be identified by species.

Range

River herring can be found along the Atlantic coast of North America, from the Southern Gulf of St. Lawrence, Canada to the southeastern United States (Mullen
et al.,
1986; Schultz
et al.,
2009). The coastal ranges of the two

species overlap. Blueback herring range from Nova Scotia south to the St. John's River, Florida; and alewife range from Labrador and Newfoundland south to South Carolina, though their occurrence in the extreme southern range is less common (Collette and Klein-MacPhee, 2002; ASMFC, 2009a; Kocik
et al.,
2009).

In Canada, river herring (i.e., gaspereau) are most abundant in the Miramichi, Margaree, LaHave, Tusket, Shubenacadie and Saint John Rivers (Gaspereau Management Plan, 2001). They are proportionally less abundant in smaller coastal rivers and streams (Gaspereau Management Plan, 2001). Generally, blueback herring in Canada occur in fewer rivers than alewives and are less abundant in rivers where both species coexist (DFO 2001).

Habitat and Migration

River herring are anadromous, meaning that they mature in the marine environment and then migrate up coastal rivers to estuarine and freshwater rivers, ponds, and lake habitats to spawn (Collette and Klein-MacPhee, 2002; ASMFC, 2009a; Kocik
et al.,
2009). In general, adult river herring are most often found at depths less than 328 feet (ft) (100 meters (m)) in waters along the continental shelf (Neves, 1981; ASMFC, 2009a; Schultz
et al.,
2009). They are highly migratory, pelagic, schooling species, with seasonal spawning migrations that are cued by water temperature (Collette and Klein-MacPhee, 2002; Schultz
et al.,
2009). Depending upon temperature, blueback herring typically spawn from late March through mid-May. However, they spawn in the southern parts of their range as early as December or January, and as late as August in the northern portion of their range (ASMFC, 2009a). Alewives have been documented spawning as early as February in the southern portion of their range, and as late as August in the northern portion of the range (ASMFC, 2009a). The river herring migration in Canada extends from late April through early July, with the peak occurring in late May and early June. Blueback herring generally make their spawning runs about 2 weeks later than alewives do (DFO, 2001). River herring conform to a metapopulation paradigm (e.g., a group of spatially separated populations of the same species which interact at some level) with adults frequently returning to their natal rivers for spawning but with some limited straying occurring between rivers (Jones, 2006; ASMFC, 2009a).

Throughout their life cycle, river herring use many different habitats, including the ocean, estuaries, rivers, and freshwater lakes and ponds. The substrate preferred for spawning varies greatly and can include gravel, detritus, and submerged aquatic vegetation. Blueback herring prefer swifter moving waters than alewives do (ASMFC, 2009a). Nursery areas include freshwater and semi-brackish waters. Little is known about their habitat preference in the marine environment (Meadows, 2008; ASMFC, 2009a).

Landlocked Populations

Landlocked populations of alewives and blueback herring also exist. Landlocked alewife populations occur in many freshwater lakes and ponds from Canada to North Carolina as well as the Great Lakes (Rothschild, 1966; Boaze & Lackey, 1974). Many landlocked populations occur as a result of stocking to provide a forage base for game fish species (Palkovacs
et al.,
2007).

Landlocked blueback herring occur mostly in the southeastern United States and the Hudson River drainage. The occurrence of landlocked blueback herring is primarily believed to be the result of accidental stockings in reservoirs (Prince and Barwick, 1981), unsanctioned stocking by recreational anglers to provide forage for game fish, and also through the construction of locks, dams and canal systems that have subsequently allowed for blueback herring occupation of several lakes and ponds along the Hudson River drainage up to, and including Lake Ontario (Limburg
et al.,
2001).

Recent efforts to assess the evolutionary origins of landlocked alewives indicate that they rapidly diverged from their anadromous cousins between 300 and 5,000 years ago, and now represent a discrete life history variant of the species,
Alosa pseudoharengus
(Palkovacs
et al.,
2007). Though given their relatively recent divergence from anadromous populations, one plausible explanation for the existence of landlocked populations may be the construction of dams by either native Americans or early colonial settlers that precluded the downstream migration of juvenile herring (Palkovacs
et al.,
2007). Since their divergence, landlocked alewives have evolved to a point they now possess significantly different mouthparts than their anadromous cousins, including narrower gapes and smaller gill raker spacings to take advantage of year round availability of smaller prey in freshwater lakes and ponds (Palkovacs
et al.,
2007). Furthermore, the landlocked alewife, compared to its anadromous cousin, matures earlier, has a smaller adult body size, and reduced fecundity (Palkovacs
et al.,
2007). At this time, there is no substantive information that would suggest that landlocked populations can or would revert back to an anadromous life history if they had the opportunity to do so (Gephard, CT DEEP, Pers. comm. 2012; Jordaan, UMASS Amherst, Pers. comm. 2012).

The discrete life history and morphological differences between the two life history variants (anadromous and landlocked) provide substantial evidence that upon becoming landlocked, landlocked populations become largely independent and separate from anadromous populations and occupy largely separate ecological niches (Palkovacs and Post, 2008). There is the possibility that landlocked alewife and blueback herring may have the opportunity to mix with anadromous river herring during high discharge years and through dam removals which could provide passage over dams and access to historic spawning habitats restored for anadromous populations, where it did not previously exist. The implications of this are not known at this time.

In summary, genetics indicate that anadromous alewife populations are discrete from landlocked populations, and that this divergence can be estimated to have taken place from 300 to 5,000 years ago. Some landlocked populations of blueback herring do occur in the Mid-Atlantic and southeastern United States. Given the similarity in life histories between anadromous alewife and blueback herring, we assume that landlocked populations of blueback herring would exhibit a similar divergence from anadromous blueback herring, as has been documented with alewives.

A Memorandum of Understanding (MOU) between the U.S. Fish and Wildlife Service (USFWS) and NMFS (collectively, the Services) regarding jurisdictional responsibilities and listing procedures under the ESA was signed August 28, 1974. This MOU states that NMFS shall have jurisdiction over species “which either (1) reside the major portion of their lifetimes in marine waters; or (2) are species which spend part of their lifetimes in estuarine waters, if the major portion of the remaining time (the time which is not spent in estuarine waters) is spent in marine waters.”

Given that landlocked populations of river herring remain in freshwater throughout their life history and are genetically divergent from the anadromous species, pursuant to the aforementioned MOU, we did not

include the landlocked populations of alewife and blueback herring in our review of the status of the species and do not consider landlocked populations in this listing determination in response to the petition to list these anadromous species.

Listing Species Under the Endangered Species Act

We are responsible for determining whether alewife and blueback herring are threatened or endangered under the ESA (16 U.S.C. 1531
et seq.
). Accordingly, based on the statutory, regulatory, and policy provisions described below, the steps we followed in making our listing determination for alewife and blueback herring were to: (1) Determine how alewife and blueback herring meet the definition of “species”; (2) determine the status of the species and the factors affecting them; and (3) identify and assess efforts being made to protect the species and determine if these efforts are adequate to mitigate existing threats.

To be considered for listing under the ESA, a group of organisms must constitute a “species.” Section 3 of the ESA defines a “species” as “any subspecies of fish or wildlife or plants, and any distinct population segment of any species of vertebrate fish or wildlife which interbreeds when mature.” Section 3 of the ESA further defines an endangered species as “any species which is in danger of extinction throughout all or a significant portion of its range” and a threatened species as one “which is likely to become an endangered species within the foreseeable future throughout all or a significant portion of its range.” Thus, we interpret an “endangered species” to be one that is presently in danger of extinction. A “threatened species,” on the other hand, is not presently in danger of extinction, but is likely to become so in the foreseeable future (that is, at a later time). In other words, the primary statutory difference between a threatened and endangered species is the timing of when a species may be in danger of extinction, either presently (endangered) or in the foreseeable future (threatened).

On February 7, 1996, the Services adopted a policy to clarify our interpretation of the phrase “distinct population segment of any species of vertebrate fish or wildlife” (61 FR 4722). The joint DPS policy describes two criteria that must be considered when identifying DPSs: (1) The discreteness of the population segment in relation to the remainder of the species (or subspecies) to which it belongs; and (2) the significance of the population segment to the remainder of the species (or subspecies) to which it belongs. As further stated in the joint policy, if a population segment is discrete and significant (i.e., it meets the DPS policy criteria), its evaluation for endangered or threatened status will be based on the ESA's definitions of those terms and a review of the five factors enumerated in section 4(a)(1) of the ESA.

As provided in section 4(a) of the ESA, the statute requires us to determine whether any species is endangered or threatened because of any of the following five factors: (1) The present or threatened destruction, modification, or curtailment of its habitat or range; (2) overutilization for commercial, recreational, scientific, or educational purposes; (3) disease or predation; (4) the inadequacy of existing regulatory mechanisms; or (5) other natural or manmade factors affecting its continued existence (section 4(a)(1)(A)(E)). Section 4(b)(1)(A) of the ESA further requires that listing determinations be based solely on the best scientific and commercial data available after taking into account efforts being made to protect the species.

Distribution and Abundance

United States

The stock assessment (described above) was prepared and compiled by the River Herring Stock Assessment Subcommittee, hereafter referred to as the `subcommittee,' of the ASMFC Shad and River Herring Technical Committee. Data and reports used for this assessment were obtained from Federal and state resource agencies, power generating companies, and universities.

The subcommittee conducted its assessment on the coastal stocks of alewife and blueback herring by individual rivers as well as coast-wide depending on available data. The subcommittee concluded that river herring should ideally be assessed and managed by individual river system, but that the marine portion of their life history likely influences survival through mixing in the marine portion of their range. However, coast-wide assessments are complicated by the complex life history of these species as well, given that factors influencing population dynamics for the freshwater portion of their life history can not readily be separated from marine factors. In addition, it was noted that data quality and availability varies by river and is mostly dependent upon the monitoring efforts that each state dedicates to these species, which further complicated the assessment.

The subcommittee also noted that most state landings records listed alewife and blueback herring together as `river herring' rather than identifying by species. These landings averaged 30.5 million pounds (lbs) (13,847 metric tons (mt)) per year from 1889 to 1938, and severe declines were noted coast-wide starting in the 1970s. Beginning in 2005, states began enacting moratoria on river herring fisheries, and as of January 2012, all directed harvest of river herring in state waters is prohibited unless states have submitted and obtained approved sustainable fisheries management plans (FMP) under ASMFC's Amendment 2 to the Shad and River Herring FMP.

The subcommittee summarized its findings for trends in commercial catch-per-unit-effort (CPUE); run counts; young-of-the-year (YOY) seine surveys; juvenile-adult fisheries independent seine, gillnet and electrofishing surveys; juvenile-adult trawl surveys; mean length; maximum age; mean length-at-age; repeat spawner frequency; total mortality (Z) estimates; and exploitation rates. Because the stock assessment contains the most recent and comprehensive description of this information and the subcommittee's conclusions, the following sections were taken from the stock assessment (ASMFC, 2012).

Commercial CPUE

Since the mid-1990s, CPUE indices for alewives showed declining trends in the Potomac River and James River (VA), no trend in the Rappahannock River (VA), and increasing trends in the York River (VA) and Chowan River (NC). CPUE indices available for blueback herring showed a declining trend in the Chowan River and no trend in the Santee River (SC). Combined species CPUE indices showed declining trends in Delaware Bay and the Nanticoke River, but CPUE has recently increased in the Hudson River (ASMFC, 2012).

Run Counts

Major declines in run sizes occurred in many rivers from 2001 to 2005. These declines were followed by increasing trends (2006 to 2010) in the Androscoggin River (ME), Damaraiscotta River (ME), Nemasket River (MA), Gilbert-Stuart River (RI), and Nonquit River (RI) for alewife and in the Sebasticook River (ME), Cocheco River (NH), Lamprey River (NH), and Winnicut River (NH) for both species combined. No trends in run sizes were evident following the recent major declines in the Union River (ME), Mattapoisett River (MA), and

Monument River (MA) for alewife and in the Exeter River (NH) for both species combined. Run sizes have declined or are still declining following recent and historical major declines in the Oyster River (NH) and Taylor River (NH) for both species, in the Parker River (MA) for alewife, and in the Monument River (MA) and Connecticut River for blueback herring (ASMFC, 2012).

Young-of-the-Year Seine Surveys

The young-of-the-year (YOY) seine surveys were quite variable and showed differing patterns of trends among rivers. Maine rivers showed similar trends in alewife and blueback herring YOY indices after 1991, with peaks occurring in 1995 and 2004. YOY indices from North Carolina and Connecticut showed declines from the 1980s to the present. New York's Hudson River showed peaks in YOY indices in 1999, 2001, 2005, and 2007. New Jersey and Maryland YOY indices showed peaks in 1994, 1996, and 2001. Virginia YOY surveys showed peaks in 1993, 1996, 2001, and 2003 (ASMFC, 2012).

Juvenile-Adult Fisheries-Independent Seine, Gillnet and Electrofishing Surveys

The juvenile-adult indices from fisheries-independent seine, gillnet and electrofishing surveys showed a variety of trends in the available datasets for the Rappahanock River (1991-2010), James River (2000-2010), St. John's River, FL (2001-2010), and Narragansett Bay (1988-2010). The gillnet indices from the Rappahannock River (alewife and blueback herring) showed a low and stable or decreasing trend after a major decline after 1995 and has remained low since 2000 (except for a rise in alewife CPUE during 2008). The gillnet and electrofishing indices in the James River (alewife and blueback herring) showed a stable or increasing trend. Blueback herring peak catch rates occurred in 2004, and alewife peak catch rates occurred in 2005. The blueback herring index from electrofishing in the St. John's River, FL, showed no trend after a major decline from 2001-2002. The seine indices in Narragansett Bay, RI (combined species) and coastal ponds (combined species) showed no trends over the time series. The CPUE for Narragansett Bay fluctuated without trend from 1988-1997, increased through 2000, declined and then remained stable from 2001-2004. The pond survey CPUE increased during 1993-1996, declined through 1998, increased in 1999, declined through 2002, peaked in 2003 and then declined and fluctuated without trend thereafter. The electrofishing indices showed opposing trends and then declining trends in the Rappahannock River (alewife and blueback herring) with catch rates of blueback herring peaking during 2001-2003, and catch rates of alewives lowest during the same time period (ASMFC, 2012).

Juvenile and Adult Trawl Surveys

Trends in trawl survey indices varied greatly with some surveys showing an increase in recent years, some showing a decrease, and some remaining stable. Trawl survey data were available from 1966-2010 (for a complete description of data see ASMFC (2012)). Trawl surveys in northern areas tended to show either an increasing or stable trend in alewife indices, whereas trawl surveys in southern areas tended to show stable or decreasing trends. Patterns in trends across surveys were less evident for blueback herring. The NMFS surveys showed a consistent increasing trend coast-wide and in the northern regions for alewife and the combined river herring species group (ASMFC, 2012).

Mean Length

Mean sizes for male and female alewife declined in 4 of 10 rivers, and mean sizes for female and male blueback herring declined in 5 of 8 rivers. Data were available from 1960-2010 (for a complete description of data see ASMFC (2012)). The common trait among most rivers in which significant declines in mean sizes were detected is that historical length data were available for years prior to 1990. Mean lengths started to decline in the mid to late 1980s; therefore, it is likely that declines in other rivers were not detected because of the shortness of their time series. Mean lengths for combined sexes in trawl surveys were quite variable through time for both alewives and blueback herring. Despite this variability, alewife mean length tended to be lowest in more recent surveys. This pattern was less apparent for blueback herring. Trend analysis of mean lengths indicated significant declines in mean lengths over time for alewives coast-wide and in the northern region in both seasons, and for blueback coast-wide and in the northern region in fall (ASMFC, 2012).

Maximum Age

Except for Maine and New Hampshire, maximum age of male and female alewife and blueback herring during 2005-2007 was 1 or 2 years lower than historical observations (ASMFC, 2012).

Mean Length-at-Age

Declines in mean length of at least one age were observed in most rivers examined. The lack of significance in some systems is likely due to the absence of data prior to 1990 when the decline in sizes began, similar to the pattern observed for mean length. Declines in mean lengths-at-age for most ages were observed in the north (NH) and the south (NC). There is little indication of a general pattern of size changes along the Atlantic coast (ASMFC, 2012).

Repeat Spawner Frequency

Examination of percentage of repeat spawners in available data revealed significant, declining trends in the Gilbert-Stuart River (RI—combined species), Nonquit River (RI—combined species), and the Nanticoke River (blueback herring). There were no trends in the remaining rivers for which data are available, although scant data suggest that current percentages of repeat spawners are lower than historical percentages in the Monument River (MA) and the Hudson River (NY) (ASMFC, 2012).

Total Mortality (Z) Estimates

With the exception of male blueback herring from the Nanticoke River, which showed a slight increase over time, there were no trends in the Z estimates produced using age data (ASMFC, 2012).

Exploitation Rates

Exploitation of river herring appears to be declining or remaining stable. In-river exploitation estimates have fluctuated, but are lower in recent years. A coast-wide index of relative exploitation showed a decline following a peak in the 1980s, and the index indicates that exploitation has remained fairly stable over the past decade. The majority of depletion-based stock reduction analysis (DB-SRA) model runs showed declining exploitation rates coast-wide. Exploitation rates estimated from the statistical catch-at-age model for blueback herring in the Chowan River also showed a slight declining trend from 1999 to 2007, at which time a moratorium was instituted. There appears to be a consensus among various assessment methodologies that exploitation has decreased in recent times. The decline in exploitation over the past decade is not surprising because river herring populations are at low levels and more restrictive regulations or moratoria have been enacted by states (ASMFC, 2012).

Summary of Stock Assessment Conclusions

Of the in-river stocks of alewife and blueback herring for which data were available and were considered in the stock assessment, 22 were depleted, 1 was increasing, and the status of 28 stocks could not be determined because the time-series of available data was too short. In most recent years, 2 in-river stocks were increasing, 4 were decreasing, and 9 were stable, with 38 rivers not having enough data to assess recent trends. The coast-wide meta-complex of river herring stocks in the United States is depleted to near historical lows. A depleted status indicates that there was evidence for declines in abundance due to a number of factors, but the relative importance of these factors in reducing river herring stocks could not be determined. Commercial landings of river herring peaked in the late 1960s, declined rapidly through the 1970s and 1980s and have remained at levels less than 3 percent of the peak over the past decade. Estimates of run sizes varied among rivers, but in general, declining trends in run size were evident in many rivers over the last decade. Fisheries-independent surveys did not show consistent trends and were quite variable both within and among surveys. Those surveys that showed declines tended to be from areas south of Long Island. A problem with the majority of fisheries-independent surveys was that the length of their time series did not overlap the period of peak commercial landings that occurred prior to 1970. There appears to be a consensus among various assessment methodologies that exploitation has decreased in recent times. The decline in exploitation over the past decade is not surprising because river herring populations are at low levels and more restrictive regulations or moratoria have been enacted by states (ASMFC, 2012).

Canada

The Department of Fisheries and Oceans (DFO) monitors and manages river herring runs in Canada. River herring runs in the Miramichi River in New Brunswick and the Maragree River in Cape Breton, Nova Scotia were monitored intensively from 1983 to 2000 (DFO, 2001). More recently (1997 to 2006) the Gaspereau River alewife run and harvest has been intensively monitored and managed partially in response to a 2002 fisheries management plan that had a goal of increasing spawning escapement to 400,000 adults (DFO, 2007). Elsewhere, river herring runs have been monitored less intensively, though harvest rates are monitored throughout Atlantic Canada through license sales, reporting requirements, and a logbook system that was enacted in 1992 (DFO, 2001).

At the time DFO conducted their last stock assessment in 2001, they identified river herring harvest levels as being low (relative to historical levels) and stable, to low and decreasing across most rivers where data were available (DFO, 2001). With respect to the commercial harvest of river herring, reported landings of river herring peaked in 1980 at slightly less than 25.5 million lbs (11,600 mt) and declined to less than 11 million lbs (5,000 mt) in 1996. Landings data reported through DFO indicate that river herring harvests have continued to decline through 2010.

Consideration as a Species Under the ESA

Distinct Population Segment Background

According to Section 3 of the ESA, the term “species” includes “any subspecies of fish or wildlife or plants, and any distinct population segment of any species of vertebrate fish or wildlife that interbreeds when mature.” Congress included the term “distinct population segment” in the 1978 amendments to the ESA. On February 7, 1996, the Services adopted a policy to clarify their interpretation of the phrase “distinct population segment” for the purpose of listing, delisting, and reclassifying species (61 FR 4721). The policy described two criteria a population segment must meet in order to be considered a DPS (61 FR 4721): (1) It must be discrete in relation to the remainder of the species to which it belongs; and (2) it must be significant to the species to which it belongs.

Determining if a population is discrete requires either one of the following conditions: (1) It is markedly separated from other populations of the same taxon as a consequence of physical, physiological, ecological, or behavioral factors. Quantitative measures of genetic or morphological discontinuity may provide evidence of this separation; or (2) it is delimited by international governmental boundaries within which differences in control of exploitation, management of habitat, conservation status, or regulatory mechanisms exist that are significant in light of section 4(a)(1)(D) of the ESA.

If a population is deemed discrete, then the population segment is evaluated in terms of significance. Factors to consider in determining whether a discrete population segment is significant to the species to which it belongs include, but are not limited to, the following: (1) Persistence of the discrete population segment in an ecological setting unusual or unique for the taxon; (2) evidence that loss of the discrete population segment would result in a significant gap in the range of the taxon; (3) evidence that the discrete population segment represents the only surviving natural occurrence of a taxon that may be more abundant elsewhere as an introduced population outside its historic range; or (4) evidence that the discrete population segment differs markedly from other populations of the species in its genetic characteristics.

If a population segment is deemed discrete and significant, then it qualifies as a DPS.

Information Related to Discreteness

To obtain expert opinion about anadromous alewife and blueback herring stock structure, we convened a working group in Gloucester, MA, on June 20-21, 2012. This working group meeting brought together river herring experts from state and Federal fisheries management agencies and academic institutions. Participants presented information to inform the presence or absence of stock structure such as genetics, life history, and morphometrics. A public workshop was held to present the expert working group's findings on June 22, 2012, and during this workshop, additional information on stock structure was sought from the public. Subsequently, a summary report was developed (NMFS, 2012a), and a peer review of the document was completed by three independent reviewers. The summary report and peer review reports are available on the NMFS Web site (see the
ADDRESSES
section above).

Steve Gephard of the Connecticut Department of Energy and Environmental Protection (CT DEP) presented a preliminary U.S. coast-wide genetic analysis of alewife and blueback herring data (Palkovacs
et al.,
2012, unpublished report). Palkovacs
et al.,
(2012, unpublished report) used 15 novel microsatellite markers on samples collected from Maine to Florida. For alewife, 778 samples were collected from spawning runs in 15 different rivers, and 1,201 blueback herring samples were collected from 20 rivers.

Bayesian analyses identified five genetically distinguishable stocks for alewife with similar results using both STRUCTURE and Bayesian Analysis of Population Structure (BAPS) software models. The alewife stock complexes identified were: (1) Northern New England; (2) Southern New England; (3)

Connecticut River; (4) Mid-Atlantic; and (5) North Carolina. For blueback herring, no optimum solution was reached using STRUCTURE, while BAPS suggested four genetically identifiable stock complexes. The stock complexes identified for blueback herring were: (1) Northern New England; (2) Southern New England; (3) Mid Atlantic; (4) and Southern. However, it should be noted that these Bayesian inferences of population structure provide a minimum number of genetically distinguishable groups. In the future, in order to better define potential stock complexes, further tests examining structure within designated stocks should be conducted using hierarchical clustering analysis and genetic tests.

The study also examined the effects of geography and found a strong effect of latitude on genetic divergence, suggesting a stepping stone model of population structure, and a strong pattern of isolation by distance, where gene flow is most likely among neighboring spawning populations. The preliminary results from the study found significant differentiation among spawning rivers for both alewife and blueback herring. Based on the results of their study, the authors' preliminary management recommendations suggest that river drainage is the appropriate level of management for both of the species. This inference was also supported by genetic tests which were conducted later. These tests suggest that there is substantial population structure at the drainage scale.

The authors noted a number of caveats for their study including: (1) Collection of specimens on their upstream spawning run may pool samples from what are truly distinct spawning populations within the major river drainages sampled, thereby, underestimating genetic structure within rivers (Hasselman, 2010); (2) a more detailed analysis of population structure within the major stocks identified (i.e., using hierarchical Bayesian clustering methods and genic test) would be useful for identifying any substructure within these major stocks; (3) neutral genetic markers used in this study represent the effects of gene flow and historical population isolation, but not the effects of adaptive processes, which are important to consider in the context of stock identification; (4) the analysis is preliminary, and there are a number of issues that need to be further investigated, including the effect of deviations in the Hardy-Weinberg Equilibrium model encountered in four alewife loci and the failure of STRUCTURE to perform well on the blueback herring dataset; and (5) hybridization may be occurring between alewife and blueback herring and may influence the results of the species-specific analyses.

Following the Stock Structure Workshop, additional analyses were run on the alewife dataset to examine the uniqueness of the (tentatively) designated Connecticut River alewife stock complex. Hybrids and misidentified samples were found and subsequently removed for this analysis, and the results were refined. By removing these samples from the Connecticut River alewife dataset, Palkovacs
et al.
(2012, unpublished report) found that, for alewife, the Connecticut and Hudson Rivers belong to the Southern New England stock. The analyses were further refined and Palkovacs
et al.
(2012, unpublished report) provided an updated map of the alewife genetic stock complexes, combining the tentative North Carolina stock with the Mid-Atlantic stock. This information and analysis is complete and is currently being prepared for publication. Thus, the refined genetic stock complexes for alewife in the coastal United States include Northern New England, Southern New England, and the Mid-Atlantic. For blueback herring, the identified genetic stocks include Northern New England, Southern New England, Mid-Atlantic and Southern (Palcovacs
et al.,
2012, unpublished report).

Bentzen
et al.
(2012) implemented a two-part genetic analysis of river herring to evaluate the genetic diversity of alewives in Maine and Maritime Canada, and to assess the regional effects of stocking on alewives and blueback herring in Maine. The genetic analysis of alewives and blueback herring along mid-coast Maine revealed significant genetic differentiation among populations. Despite significant differentiation, the patterns of correlation did not closely correspond with geography or drainage affiliation. The genetic analysis of alewives from rivers in Maine and Atlantic Canada detected isolation by distance, suggesting that homing behavior indicative of alewives' metapopulation conformance does produce genetically distinguishable populations. Further testing also suggested that there may be interbreeding between alewives and blueback herring (e.g., hybrids), especially at sample sites with impassible dams.

The unusual genetic groupings of river herring in Maine are likely a result of Maine's complex stocking history, as alewife populations in Maine have been subject to considerable within and out of basin stocking for the purpose of enhancement, recolonization of extirpated populations, and stock introduction. Alewife stocking in Maine dates back at least to 1803 when alewives were reportedly moved from the Pemaquid and St. George Rivers to create a run of alewives in the Damariscotta River (Atkins and Goode, 1887). These efforts were largely responsive to considerable declines in alewife populations following the construction of dams, over exploitation and pollution. Although there has been considerable alewife stocking and relocation throughout Maine, there are very few records documenting these efforts. In contrast, considerably less stocking of alewives has occurred in Maritime Canada. These genetic analyses suggest that river herring from Canadian waters are genetically distinct from Maine river herring.

All of the expert opinions we received during the Stock Structure Workshop suggested evidence of regional stock structure exists for both alewife and blueback herring as shown by the recent genetics data (Palkovacs
et al.,
2012, unpublished report; Bentzen
et al.,
unpublished data). However, the suggested boundaries of the regional stock complexes differed from expert to expert. Migration and mixing patterns of alewives and blueback herring in the ocean have not been determined, though regional stock mixing is suspected. Therefore, the experts suggested that the ocean phase of alewives and blueback herring should be considered a mixed stock until further tagging and genetic data become available. There is evidence to support regional differences in migration patterns, but not at a level of river-specific stocks.

In the mid-1980s, Rulifson
et al.
(1987) tagged and released approximately 19,000 river herring in the upper Bay of Fundy, Nova Scotia with an overall recapture rate of 0.39 percent. Alewife tag returns were from freshwater locations in Nova Scotia, and marine locations in Nova Scotia and Massachusetts. Blueback herring tag returns were from freshwater locations in Maryland and North Carolina and marine locations in Nova Scotia. Rulifson
et al.
(1987) suspected from recapture data that alewives and blueback herring tagged in the Bay of Fundy were of different origins, hypothesizing that alewives were likely regional fish from as far away as New England, while the blueback herring recaptures were likely not regional fish, but those of U.S. origin from the mid-Atlantic region. However, the low tag return numbers (n = 2) made it difficult to generalize about the natal rivers of

blueback herring caught in the Bay of Fundy. The results of this tagging study show that river herring present in Canadian waters may originate from U.S. waters and vice versa.

Metapopulations of river herring are believed to exist, with adults frequently returning to their natal rivers for spawning and some straying occurring between rivers—straying rates have been estimated up to 20 percent (Jones, 2006; ASMFC, 2009a; Gahagan
et al.,
2012). Given the available information on genetic differentiation coast-wide for alewife and blueback herring, it appears that stock complexes exist for both species.

River herring originating from Canadian rivers are delimited by international governmental boundaries. Differences in control of exploitation, management of habitat, conservation status, or regulatory mechanisms exist and, therefore, meet the discreteness criterion under the DPS policy; however, intermixing between both alewife and blueback herring from U.S. and Canadian coastal waters occurs, and the extent of this mixing is unknown.

Given the best available information, it is possible to determine that the various stocks of both alewife and blueback herring are discrete. The best available information suggests that the delineation of the stock complexes is as described above; however, future work will likely further refine these preliminary boundaries. Additionally, further information is needed on the oceanic migratory patterns of both species.

Information Related to Significance

If a population is deemed discrete, the population is evaluated in terms of significance. Significance can be determined using the four criteria noted above. Since the best available information indicates that the stock complexes identified for alewives and blueback herring are most likely discrete, the SRT reviewed the available information to determine if they are significant.

In evaluating the significance criterion, the SRT considered all of the above criteria. As indicated earlier, both alewives and blueback herring occupy a large range spanning almost the entire East Coast of the United States and into Canada. They appear to migrate freely throughout their oceanic range and return to freshwater habitats to spawn in streams, lakes and rivers. Therefore, they occupy many different ecological settings throughout their range.

As described earlier, the Palkovacs
et al.
(2012, unpublished report) study assessed the genetic composition of alewife and blueback herring stocks within U.S. rivers using 15 neutral loci and documented that there are at least three stock complexes of alewife in the United States and four stock complexes of blueback herring in the United States. Palkovac
et al.
(2012, unpublished report) showed a strong effect of latitude on genetic divergence, suggesting that although most populations are genetically differentiated, gene flow is greater among neighboring runs than among distant runs. The genetic data are consistent with the recent results of the ASMFC stock assessment (2012), which noted that even among rivers within the same state, there are differences in trends in abundance indices, size-at-age, age structure and other metrics, indicating there are localized factors affecting the population dynamics of both species.

Neutral genetic markers such as microsatellites have a longstanding history of utilization in stock designation for many anadromous fish species (Waples, 1998). However, these markers represent the effects of gene flow and historical population isolation and not the effects of adaptive processes. The effects of adaptive genetic and phenotypic diversity are also extremely important to consider in the context of stock designation, but are not captured by the use of neutral genetic markers. Therefore, the available genetic data are most appropriately used in support of the discreteness criterion, rather than to determine significance.

Determining whether a gap in the range of the taxon would be significant if a stock were extirpated is difficult to determine with anadromous fish such as river herring. River herring are suspected to migrate great distances between their natal rivers and overwintering areas, and therefore, estuarine and marine populations are comprised of mixed stocks. Consequently, the loss of a stock complex would mean the loss of riverine spawning subpopulations, while the marine and estuarine habitat would most likely still be occupied by migratory river herring from other stock complexes. As it has been shown that gene flow is greater among neighboring runs than among distant runs, we might expect that river herring would re-colonize neighboring systems over a relatively short time frame. Thus, the loss of one stock complex in itself may not be significant; the loss of contiguous stock complexes may be. The goal then for river herring stock complexes is to maintain connectivity between genetic groups to support proper metapopulation function (spatially separated populations of the same species that interact, recolonize vacant habitats, and occupy new habitats through dispersal mechanisms (Hanski and Gilpin, 1991)).

DPS Determination

Evidence for genetic differentiation exists for both alewife and blueback herring, allowing for preliminary identification of stock complexes; however, available data are lacking on the significance of each of these individual stock complexes. Therefore, we have determined that there is not enough evidence to suggest that the stock complexes identified through genetics should be treated under the DPS policy as separate DPSs. The stock complexes may be discrete, but under the DPS policy, they are not significant to the species as a whole. Furthermore, given the unknown level of intermixing between Canadian and U.S. river herring in coastal waters, the Canadian stock complex should also not be considered separately under the DPS policy.

Throughout the rest of this determination, the species will be referred to by species (alewife or blueback herring), as river herring where information overlaps, and by the identified stock complexes (Palkovacs
et al.,
2012, unpublished report) for each species as necessary. While the individual stock complexes do not constitute separate DPSs, they are important components of the overall species and relevant to the evaluation of whether either species may be threatened or endangered in a significant portion of their overall range. Therefore, we have evaluated the threats to, and extinction risk of the overall species and each of the individual stock complexes as presented below. For this analysis, the identified stock complexes for alewife (Figure 1) in the coastal United States for the purposes of this finding will include Northern New England, Southern New England, the Mid-Atlantic, and Canada; and stock complexes for blueback herring (Figure 2) will include Northern New England, Southern New England, Mid-Atlantic, Southern Atlantic, and Canada. While the SRT concluded that there was not sufficient information at this time to determine with any certainty whether alewife or blueback herring stock complexes constitute separate DPSs, they recognized that future information on behavior, ecology and genetic population structure may reveal significant differences, showing fish to be uniquely adapted to each stock complex. We agree with this conclusion. Thus, we are not identifying DPSs for either species.

EN12AU13.004

EN12AU13.005

Foreseeable Future and Significant Portion of Its Range

The ESA defines an “endangered species” as “any species which is in danger of extinction throughout all or a significant portion of its range,” while a “threatened species” is defined as “any species which is likely to become an endangered species within the foreseeable future throughout all or a significant portion of its range.” NMFS and the U.S. Fish and Wildlife Servce (USFWS) recently published a draft policy to clarify the interpretation of the phrase “significant portion of the range” in the ESA definitions of “threatened” and “endangered” (76 FR 76987; December 9, 2011). The draft policy provides that: (1) If a species is found to be endangered or threatened in only a significant portion of its range, the entire species is listed as endangered or threatened, respectively, and the ESA's protections apply across the species' entire range; (2) a portion of the range of a species is “significant” if its

contribution to the viability of the species is so important that, without that portion, the species would be in danger of extinction; (3) the range of a species is considered to be the general geographical area within which that species can be found at the time USFWS or NMFS makes any particular status determination; and (4) if the species is not endangered or threatened throughout all of its range, but it is endangered or threatened within a significant portion of its range, and the population in that significant portion is a valid DPS, we will list the DPS rather than the entire taxonomic species or subspecies.

The Services are currently reviewing public comment received on the draft policy. While the Services' intent is to establish a legally binding interpretation of the term “significant portion of the range,” the draft policy does not have legal effect until such time as it may be adopted as final policy. Here, we apply the principles of this draft policy as non-binding guidance in evaluating whether to list alewife or blueback herring under the ESA. If the policy changes in a material way, we will revisit the determination and assess whether the final policy would result in a different outcome.

While we have determined that DPSs cannot be defined for either of these species based on the available information, the stock complexes do represent important groupings within the range of both species. Thus, in our analysis of extinction risk and threats assessment below, we have evaluated whether either species is at risk rangewide and within any of the individual stock complexes so that we can evaluate whether either species is threatened or endangered in a significant portion of its range.

We established that the appropriate period of time corresponding to the foreseeable future is a function of the particular type of threats, the life-history characteristics, and the specific habitat requirements for river herring. The timeframe established for the foreseeable future takes into account the time necessary to provide for the conservation and recovery of each species and the ecosystems upon which they depend, but is also a function of the reliability of available data regarding the identified threats and extends only as far as the data allow for making reasonable predictions about the species' response to those threats. As described below, the SRT determined that dams and other impediments to migration have already created a clear and present threat to river herring that will continue into the future. The SRT also evaluated the threat from climate change from 2060 to 2100 and climate variability in the near term (as described in detail below).

Highly productive species with short generation times are more resilient than less productive, long lived species, as they are quickly able to take advantage of available habitats for reproduction (Mace
et al.,
2002). Species with shorter generation times, such as river herring (4 to 6 years), experience greater population variability than species with long generation times, because they maintain the capacity to replenish themselves more quickly following a period of low survival (Mace
et al.,
2002). Given the high population variability among clupeids, projecting out further than three generations could lead to considerable uncertainty in the probability that the model will provide an accurate representation of the population trajectory for each species. Thus, a 12 to 18 year timeframe (e.g., 2024-2030), or a three-generation time period, for each species was determined by the Team to be appropriate for use as the foreseeable future for both alewife and blueback herring. We agree with the Team that a three-generation time period (12-18 years) is a reasonable foreseeable future for both alewife and blueback herring.

Connectivity, population resilience and diversity are important when determining what constitutes a significant portion of the species' range (Waples
et al.,
2007). Maintaining connectivity between genetic groups supports proper metapopulation function, in this case, anadromy. Ensuring that river herring populations are well represented across diverse habitats helps to maintain and enhance genetic variability and population resilience (McElhany
et al.,
2000). Additionally, ensuring wide geographic distribution across diverse climate and geographic regions helps to minimize risk from catastrophes (e.g., droughts, floods, hurricanes, etc.; McElhany
et al.,
2000). Furthermore, preventing isolation of genetic groups protects against population divergence (Allendorf and Luikart, 2007).

Threats Evaluation

As described above, Section 4(a)(1) of the ESA and NMFS implementing regulations (50 CFR 424) states that we must determine whether a species is endangered or threatened because of any one or a combination of the following factors: (A) Current or threatened habitat destruction or modification or curtailment of habitat or range; (B) overutilization for commercial, recreational, scientific, or educational purposes; (C) disease or predation; (D) inadequacy of existing regulatory mechanisms; and (E) other natural or man-made factors affecting the species' continued existence. This section briefly summarizes the findings regarding these factors.

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

Past, present, and reasonably foreseeable future factors that have the potential to affect river herring habitat include, but are not limited to, dams and hydropower facilities, dredging, water quality (including land use change, water withdrawals, discharge and contaminants), climate change and climate variability. As noted above, river herring occupy a variety of different habitats including freshwater, estuarine and marine environments throughout their lives, and thus, they are subjected to habitat impacts occurring in all of these different habitats.

Dams and Other Barriers

Dams and other barriers to upstream and downstream passage (e.g., culverts) can block or impede access to habitats necessary for spawning and rearing; can cause direct and indirect mortality from injuries incurred while passing over dams, through downstream passage facilities, or through hydropower turbines; and can degrade habitat features necessary to support essential river herring life history functions. Man-made barriers that block or impede access to rivers throughout the entire historical range of river herring have resulted in significant losses of historical spawning habitat for river herring. Dams and other man-made barriers have contributed to the historical and current declines in abundance of both blueback and alewife populations. While estimates of habitat loss over the entire range of river herring are not available, estimates from studies in Maine show that less than 5 percent of lake spawning habitat and 20 percent of river habitat remains accessible for river herring (Hall
et al.,
2010). As described in more detail below, dams are also known to impact river herring through various mechanisms, such as habitat alteration, fish passage delays, and entrainment and impingement (Ruggles 1980; NRC 2004). River herring can undergo indirect mortality from injuries such as scale loss, lacerations, bruising, eye or fin damage, or internal hemorrhaging when passing through turbines, over spillways, and through bypasses (Amaral
et al.,
2012).

The following summary of the effects of dams and other barriers on river herring is taken from Amendment 2 to the Interstate Fishery Management Plan for Shad and River Herring (hereafter, referred to as “Amendment 2” and cited as “ASMFC, 2009”). Because it includes a detailed description of barriers to upstream and downstream passage, it is the best source of comprehensive information on this topic. Please refer to Amendment 2 for more information.

Dams and spillways impeding rivers along the East Coast of the United States have resulted in a considerable loss of historical spawning habitat for shad and river herring. Permanent man-made structures pose an ongoing barrier to fish passage unless fishways are installed or structures are removed. Low-head dams can also pose a problem, as fish are unable to pass over them except when tides or river discharges are exceptionally high (Loesch and Atran, 1994). Historically, major dams were often constructed at the site of natural formations conducive to waterpower, such as natural falls. Diversion of water away from rapids at the base of falls can reduce fish habitat, and in some cases cause rivers to run dry at the base for much of the summer (MEOEA, 2005; ASMFC, 2009).

Prior to the early 1990s, it was thought that migrating shad and river herring suffered significant mortality going through turbines during downstream passage (Mathur and Heisey, 1992). Juvenile shad emigrating from rivers have been found to accumulate in larger numbers near the forebay of hydroelectric facilities, where they become entrained in intake flow areas (Martin
et al.,
1994). Relatively high mortality rates were reported (62 percent to 82 percent) at a hydroelectric dam for juvenile American shad and blueback herring, depending on the power generation levels tested (Taylor and Kynard, 1984). In contrast, Mathur and Heisey (1992) reported a mortality rate of 0 percent to 3 percent for juvenile American shad (2 to 6 in fork length (55 to 140 mm)), and 4 percent for juvenile blueback herring (3 to 4 in fork length (77 to 105 mm)) through Kaplan turbines. Mortality rate increased to 11 percent in passage through a low-head Francis turbine (Mathur and Heisey, 1992). Other studies reported less than 5 percent mortality when large Kaplan and fixed-blade, mixed-flow turbines were used at a facility along the Susquehanna River (RMC, 1990; RMC, 1994). At the same site, using small Kaplan and Francis runners, the mortality rate was as high as 22 percent (NA, 2001). At another site, mortality rate was about 15 percent where higher revolution, Francis-type runners were used (RMC, 1992; ASMFC, 2009).

Additional studies reported that changes in pressure had a more pronounced effect on juveniles with thinner and weaker tissues as they moved through turbines (Taylor and Kynard, 1984). Furthermore, some fish may die later from stress, or become weakened and more susceptible to predation, and as such, losses may not be immediately apparent to researchers (Gloss, 1982) (ASMFC, 2009).

Changes to the river system, resulting in delayed migration among other things, were also identified in Amendment 2 as impacting river herring. Amendment 2 notes that when juvenile alosines delay out-migration, they may concentrate behind dams and become more susceptible to actively feeding predators. They may also be more vulnerable to anglers that target alosines as a source of bait. Delayed out-migration can also make juvenile alosines more susceptible to marine predators that they may have avoided if they had followed their natural migration patterns (McCord, 2005a). In open rivers, juvenile alosines gradually move seaward in groups that are likely spaced according to the spatial separation of spawning and nursery grounds (Limburg, 1996; J. McCord, South Carolina Department of Natural Resources, personal observation). Releasing water from dams and impoundments (or reservoirs) may lead to flow alterations, altered sediment transport, disruption of nutrient availability, changes in downstream water quality (including both reduced and increased temperatures), streambank erosion, concentration of sediment and pollutants, changes in species composition, solubilization of iron and manganese and their absorbed or chelated ions, and hydrogen sulfide in hypolimnetic (water at low level outlets) releases (Yeager, 1995; Erkan, 2002; ASMFC, 2009).

Many dams spill water over the top of the structure where water temperatures are the warmest, essentially creating a series of warm water ponds in place of the natural stream channel (Erkan, 2002). Conversely, water released from deep reservoirs may be poorly oxygenated, at below-normal seasonal water temperature, or both, thereby causing loss of suitable spawning or nursery habitat in otherwise habitable areas (ASMFC, 2009).

Reducing minimum flows can reduce the amount of water available and cause increased water temperature or reduced dissolved oxygen levels (ASMFC, 1985; ASMFC, 1999; USFWS
et al.,
2001). Such conditions have occurred along the Susquehanna River at the Conowingo Dam, Maryland, from late spring through early fall, and have historically caused large fish kills below the dam (Krauthamer and Richkus, 1987; ASMFC, 2009).

Disruption of seasonal flow rates in rivers can impact upstream and downstream migration patterns for adult and juvenile alosines (ASMFC, 1985; Limburg, 1996; ASMFC, 1999; USFWS
et al.,
2001). Changes to natural flows can also disrupt natural productivity and availability of zooplankton that larval and early juvenile alosines feed on (Crecco and Savoy, 1987; Limburg, 1996; ASMFC, 2009).

Although most dams that impact diadromous fish are located along the lengths of rivers, fish can also be affected by hydroelectric projects at the mouths of rivers, such as the large tidal hydroelectric project at the Annapolis River in the Bay of Fundy, Canada. This particular basin and other surrounding waters are used as foraging areas during summer months by American shad from all runs along the East Coast of the United States (Dadswell
et al.,
1983). Because the facilities are tidal hydroelectric projects, fish may move in and out of the impacted areas with each tidal cycle. While turbine mortality is relatively low with each passage, the repeated passage in and out of these facilities may cumulatively result in substantial overall mortalities (Scarratt and Dadswell, 1983; ASMFC, 2009).

Additional man-made structures that may obstruct upstream passage include: tidal and amenity barrages (barriers constructed to alter tidal flow for aesthetic purposes or to harness energy); tidal flaps (used to control tidal flow); mill, gauging, amenity, navigation, diversion, and water intake weirs; fish counting structures; and earthen berms (Durkas, 1992; Solomon and Beach, 2004). The impact of these structures is site-specific and will vary with a number of conditions including head drop, form of the structure, hydrodynamic conditions upstream and downstream, condition of the structure, and presence of edge effects (Solomon and Beach, 2004). Road culverts are also a significant source of blockage. Culverts are popular, low-cost alternatives to bridges when roads must cross small streams and creeks. Although the amount of habitat affected by an individual culvert may be small, the cumulative impact of multiple culverts within a watershed can be substantial (Collier and Odom, 1989; ASMFC, 2009).

Roads and culverts can also impose significant changes in water quality.

Winter runoff in some states may include high concentrations of road salt, while stormwater flows in the summer may cause thermal stress and bring high concentrations of other pollutants (MEOEA, 2005; ASMFC, 2009).

Sampled sites in North Carolina revealed river herring upstream and downstream of bridge crossings, but no herring were found in upstream sections of streams with culverts. Additional study is underway to determine if river herring are absent from these areas because of the culverts (NCDENR, 2000). Even structures only 8 to 12 in (20 to 30 cm) above the water can block shad and river herring migration (ASMFC, 1999; ASMFC, 2009).

Rivers can also be blocked by non-anthropogenic barriers, such as beaver dams, waterfalls, log piles, and vegetative debris. These blockages may hinder migration, but they can also benefit by providing adhesion sites for eggs, protective cover, and feeding sites (Klauda
et al.,
1991b). Successful passage at these natural barriers often depends on individual stream flow characteristics during the fish migration season (ASMFC, 2009).

Dredging

Wetlands provide migratory corridors and spawning habitat for river herring. The combination of incremental losses of wetland habitat, changes in hydrology, and nutrient and chemical inputs over time, can be extremely harmful, resulting in diseases and declines in the abundance and quality. Wetland loss is a cumulative impact that results from activities related to dredging/dredge spoil placement, port development, marinas, solid waste disposal, ocean disposal, and marine mining. In the late 1970s and early 1980s, the United States was losing wetlands at an estimated rate of 300,000 acres (1,214 sq km) per year. The Clean Water Act and state wetland protection programs helped decrease wetland losses to 117,000 acres (473 sq km) per year, between 1985 and 1995. Estimates of wetlands loss vary according to the different agencies. The U.S. Department of Agriculture (USDA) attributes 57 percent of wetland loss to development, 20 percent to agriculture, 13 percent to the creation of deepwater habitat, and 10 percent to forest land, rangeland, and other uses. Of the wetlands lost between 1985 and 1995, the USFWS estimates that 79 percent of wetlands were lost to upland agriculture. Urban development and other types of land use activities were responsible for 6 percent and 15 percent of wetland loss, respectively.

Amendment 2 identifies channelization and dredging as a threat to river herring habitat. The following section, taken from Amendment 2, describes these threats.

Channelization can cause significant environmental impacts (Simpson
et al.,
1982; Brookes, 1988), including bank erosion, elevated water velocity, reduced habitat diversity, increased drainage, and poor water quality (Hubbard, 1993). Dredging and disposal of spoils along the shoreline can also create spoil banks, which block access to sloughs, pools, adjacent vegetated areas, and backwater swamps (Frankensteen, 1976). Dredging may also release contaminants, resulting in bioaccumulation, direct toxicity to aquatic organisms, or reduced dissolved oxygen levels (Morton, 1977). Furthermore, careless land use practices may lead to erosion, which can lead to high concentrations of suspended solids (turbidity) and substrate (siltation) in the water following normal and intense rainfall events. This can displace larvae and juveniles to less desirable areas downstream and cause osmotic stress (Klauda
et al.,
1991b; ASMFC, 2009).

Spoil banks are often unsuitable habitat for fishes. Suitable habitat is often lost when dredge disposal material is placed on natural sand bars and/or point bars. The spoil is too unstable to provide good habitat for the food chain. Draining and filling, or both, of wetlands adjacent to rivers and creeks in which alosines spawn has eliminated spawning areas in North Carolina (NCDENR, 2000; ASMFC, 2009).

Secondary impacts from channel formation include loss of vegetation and debris, which can reduce habitat for invertebrates and result in reduced quantity and diversity of prey for juveniles (Frankensteen, 1976). Additionally, stream channelization often leads to altered substrate in the riverbed and increased sedimentation (Hubbard, 1993), which in turn can reduce the diversity, density, and species richness of aquatic insects (Chutter, 1969; Gammon, 1970; Taylor, 1977). Suspended sediments can reduce feeding success in larval or juvenile fishes that rely on visual cues for plankton feeding (Kortschal
et al.,
1991). Sediment re-suspension from dredging can also deplete dissolved oxygen, and increase bioavailability of any contaminants that may be bound to the sediments (Clark and Wilber, 2000; ASMFC, 2009).

Migrating adult river herring avoid channelized areas with increased water velocities. Several channelized creeks in the Neuse River basin in North Carolina have reduced river herring distribution and spawning areas (Hawkins, 1979). Frankensteen (1976) found that the channelization of Grindle Creek, North Carolina removed in-creek vegetation and woody debris, which had served as substrate for fertilized eggs (ASMFC, 2009).

Channelization can also reduce the amount of pool and riffle habitat (Hubbard, 1993), which is an important food-producing area for larvae (Keller, 1978; Wesche, 1985; ASMFC, 2009).

Dredging can negatively affect alosine populations by producing suspended sediments (Reine
et al.,
1998), and migrating alosines are known to avoid waters of high sediment load (ASMFC, 1985; Reine
et al.,
1998). Fish may also avoid areas that are being dredged because of suspended sediment in the water column. Filter-feeding fishes, such as alosines, can be negatively impacted by suspended sediments on gill tissues (Cronin
et al.,
1970). Suspended sediments can clog gills that provide oxygen, resulting in lethal and sub-lethal effects to fish (Sherk
et al.,
1974 and 1975; ASMFC, 2009).

Nursery areas along the shorelines of the rivers in North Carolina have been affected by dredging and filling, as well as by erection of bulkheads; however, the degree of impact has not been measured. In some areas, juvenile alosines were unable to enter channelized sections of a stream due to high water velocities caused by dredging (ASMFC, 2000 and 2009).

Water Quality

Nutrient enrichment has become a major cumulative problem for many coastal waters. Nutrient loading results from the individual activities of coastal development, marinas and recreational boating, sewage treatment and disposal, industrial wastewater and solid waste disposal, ocean disposal, agriculture, and aquaculture. Excess nutrients from land based activities accumulate in the soil, pollute the atmosphere, pollute ground water, or move into streams and coastal waters. Nutrient inputs are known to have a direct effect on water quality. For example, nutrient enrichment can stimulate growth of phytoplankton that consumes oxygen when they decay, which can lead to low dissolved oxygen that may result in fish kills (Correll, 1987; Tuttle
et al.,
1987; Klauda
et al.,
1991b); this condition is known as eutrophication.

In addition to the direct cumulative effects incurred by development activities, inshore and coastal habitats are also threatened by persistent increases in certain chemical discharges. The combination of incremental losses of wetland habitat, changes in hydrology, and nutrient and chemical inputs produced over time can

be extremely harmful to marine and estuarine biota, including river herring, resulting in diseases and declines in the abundance and quality of the affected resources.

Amendment 2 identified land use changes including agriculture, logging/forestry, urbanization and non-point source pollution as threats to river herring habitat. The following section, taken from Amendment 2, describes these threats.

The effects of land use and land cover on water quality, stream morphology, and flow regimes are numerous, and may be the most important factors determining quantity and quality of aquatic habitats (Boger, 2002). Studies have shown that land use influences dissolved oxygen (Limburg and Schmidt, 1990), sediments and turbidity (Comeleo
et al.,
1996; Basnyat
et al.,
1999), water temperature (Hartman
et al.,
1996; Mitchell, 1999), pH (Osborne and Wiley, 1988; Schofield, 1992), nutrients (Peterjohn and Correll, 1984; Osborne and Wiley, 1988; Basnyat
et al.,
1999), and flow regime (Johnston
et al.,
1990; Webster
et al.,
1992; ASMFC, 2009).

Siltation, caused by erosion due to land use practices, can kill submerged aquatic vegetation (SAV). SAV can be adversely affected by suspended sediment concentrations of less than 15 ppm (15 mg/L) (Funderburk
et al.,
1991) and by deposition of excessive sediments (Valdes-Murtha and Price, 1998). SAV is important because it improves water quality (Carter
et al.,
1991). SAV consumes nutrients in the water and as the plants die and decay, they slowly release the nutrients back into the water column. Additionally, through primary production and respiration, SAV affects the dissolved oxygen and carbon dioxide concentrations, alkalinity, and pH of the waterbody. SAV beds also bind sediments to the bottom resulting in increased water clarity, and they provide refuge habitat for migratory fish and planktonic prey items (Maldeis, 1978; Monk, 1988; Killgore
et al.,
1989; ASMFC, 2009).

Decreased water quality from sedimentation became a problem with the advent of land-clearing agriculture in the late 18th century (McBride, 2006). Agricultural practices can lead to sedimentation in streams, riparian vegetation loss, influx of nutrients (e.g., inorganic fertilizers and animal wastes), and flow modification (Fajen and Layzer, 1993). Agriculture, silviculture, and other land use practices can lead to sedimentation, which reduces the ability of semi-buoyant eggs and adhesive eggs to adhere to substrates (Mansueti, 1962; ASMFC, 2009).

From the 1950s to the present, increased nutrient loading has made hypoxic conditions more prevalent (Officer
et al.,
1984; Mackiernan, 1987; Jordan
et al.,
1992; Kemp
et al.,
1992; Cooper and Brush, 1993; Secor and Gunderson, 1998). Hypoxia is most likely caused by eutrophication, due mostly to non-point source pollution (e.g., industrial fertilizers used in agriculture) and point source pollution (e.g., urban sewage).

Logging activities can modify hydrologic balances and in-stream flow patterns, create obstructions, modify temperature regimes, and add nutrients, sediments, and toxic substances into river systems. Loss of riparian vegetation can result in fewer refuge areas for fish from fallen trees, fewer insects for fish to feed on, and reduced shade along the river, which can lead to increased water temperatures and reduced dissolved oxygen (EDF, 2003). Threats from deforestation of swamp forests include: siltation from increased erosion and runoff; decreased dissolved oxygen (Lockaby
et al.,
1997); and disturbance of food-web relationships in adjacent and downstream waterways (Batzer
et al.,
2005; ASMFC, 2009).

Urbanization can cause elevated concentrations of nutrients, organics, or sediment metals in streams (Wilber and Hunter, 1977; Kelly and Hite, 1984; Lenat and Crawford, 1994). More research is needed on how urbanization affects diadromous fish populations; however, Limburg and Schmidt (1990) found that when the percent of urbanized land increased to about 10 percent of the watershed, the number of alewife eggs and larvae decreased significantly in tributaries of the Hudson River, New York (ASMFC, 2009).

Water Withdrawal/Outfall

Water withdrawal facilities and toxic and thermal discharges have also been identified as impacting river herring, and the following section is summarized from Amendment 2.

Large volume water withdrawals (e.g., drinking water, pumped-storage hydroelectric projects, irrigation, and snow-making) can alter local current characteristics (e.g., reverse river flow), which can result in delayed movement past a facility or entrainment in water intakes (Layzer and O'Leary, 1978). Planktonic eggs and larvae entrained at water withdrawal projects experience high mortality rates due to pressure changes, shear and mechanical stresses, and heat shock (Carlson and McCann, 1969; Marcy, 1973; Morgan
et al.,
1976). While juvenile mortality rates are generally low at well-screened facilities, large numbers of juveniles can be entrained (Hauck and Edson, 1976; Robbins and Mathur, 1976; ASMFC, 2009).

Fish impinged against water filtration screens can die from asphyxiation, exhaustion, removal from the water for prolonged periods of time, removal of protective mucous, and descaling (DBC, 1980). Studies conducted along the Connecticut River found that larvae and early juveniles of alewife, blueback herring, and American shad suffered 100-percent mortality when temperatures in the cooling system of a power plant were elevated above 82 °F (28°C); 80 percent of the total mortality was caused by mechanical damage, 20 percent by heat shock (Marcy, 1976). Ninety-five percent of the fish near the intake were not captured by the screen, and Marcy (1976) concluded that it did not seem possible to screen fish larvae effectively (ASMFC, 2009).

The physical characteristics of streams (e.g., stream width, depth, and current velocity; substrate; and temperature) can be altered by water withdrawals (Zale
et al.,
1993). River herring can experience thermal stress, direct mortality, or indirect mortality when water is not released during times of low river flows and water temperatures are higher than normal. Water flow disruption can also result in less freshwater input to estuaries (Rulifson, 1994), which are important nursery areas for river herring and other anadromous species (ASMFC, 2009).

Industrial discharges may contain toxic chemicals, such as heavy metals and various organic chemicals (e.g., insecticides, solvents, herbicides) that are harmful to aquatic life (ASMFC, 1999). Many contaminants can have harmful effects on fish, including reproductive impairment (Safe, 1990; Mac and Edsall, 1991; Longwell
et al.,
1992). Chemicals and heavy metals can move through the food chain, producing sub-lethal effects such as behavioral and reproductive abnormalities (Matthews
et al.,
1980). In fish, exposure to polychlorinated biphenyls (PCBs) can cause fin erosion, epidermal lesions, blood anemia, altered immune response, and egg mortality (Post, 1987; Kennish
et al.,
1992). Steam power plants that use chlorine to prevent bacterial, fungal, and algal growth present a hazard to all aquatic life in the receiving stream, even at low concentrations (Miller
et al.,
1982; ASMFC, 2009).

Pulp mill effluent and other oxygen-consuming wastes discharged into rivers and streams can reduce dissolved oxygen concentrations below what is

required for river herring survival. Low dissolved oxygen resulting from industrial pollution and sewage discharge can also delay or prevent upstream and downstream migrations. Everett (1983) found that during times of low water flow when pulp mill effluent comprised a large percentage of the flow, river herring avoided the effluent. Pollution may be diluted in the fall when water flows increase, but fish that reach the polluted waters downriver before the water has flushed the area will typically succumb to suffocation (Miller
et al.,
1982; ASMFC, 2009).

Effluent may also pose a greater threat during times of drought. Such conditions were suspected of interfering with the herring migration along the Chowan River, North Carolina, in 1981. In the years before 1981, the effluent from the pulp mill had passed prior to the river herring run, but drought conditions caused the effluent to remain in the system longer that year. Toxic effects were indicated, and researchers suggested that growth and reproduction might have been disrupted as a result of eutrophication and other factors (Winslow
et al.,
1983; ASMFC, 2009).

Klauda
et al.
(1991a) provides an extensive review of temperature thresholds for alewife and bluback herring. In summary, the spawning migration for alewives most often occurs when water temperatures range from 50-64 °F (10-18 °C), and for bluebacks when temperatures range from 57-77 °F (14-25 °C). Alewife egg deposition most often occurs when temperatures range between 50-72 °F (10 and 22 °C), and for bluebacks when temperatures range between 70-77 °F (21 and 25 °C). Alewife egg and larval development is optimal when temperatures range from 63—70 °F (17-21 °C), and for bluebacks when temperatures range from 68-75 °F (20-24 °C) (temperature ranges were also presented and discussed at the Climate Workshop (NMFS, 2012b)). Thermal effluent from power plants outside these temperature ranges when river herring are present can disrupt schooling behavior, cause disorientation, and may result in death. Sewage can directly and indirectly affect anadromous fish. Major phytoplankton and algal blooms that reduced light penetration (Dixon, 1996) and ultimately reduced SAV abundance (Orth
et al.,
1991) in tidal freshwater areas of the Chesapeake Bay in the 1960s and early 1970s may have been caused by ineffective sewage treatment (ASMFC, 2009).

Water withdrawal for irrigation can cause dewatering or reduced streamflow of freshwater streams, which can decrease the quantity of both spawning and nursery habitat for anadromous fish. Reduced streamflow can reduce water quality by concentrating pollutants and/or increasing water temperature (ASMFC, 1985). O'Connell and Angermeier (1999) found that in some Virginia streams, there was an inverse relationship between the proportion of a stream's watershed that was agriculturally developed and the overall tendency of the stream to support river herring runs. In North Carolina, cropland alteration along several creeks and rivers significantly reduced river herring distribution and spawning areas in the Neuse River basin (Hawkins, 1979; ASMFC, 2009).

Atmospheric deposition occurs when pollutants (e.g. nitrates, sulfates, ammonium, and mercury) are transferred from the air to the earth's surface. Pollutants can get from the air into the water through rain and snow, falling particles, and absorption of the gas form of the pollutants into the water. Atmospheric pollutants can result in increased eutrophication (Paerl
et al.,
1999) and acidification of surface waters (Haines, 1981). Atmospheric nitrogen deposition in coastal estuaries can lead to accelerated algal production (or eutrophication) and water quality declines (e.g., hypoxia, toxicity, and fish kills) (Paerl
et al.,
1999). Nitrate and sulfate deposition is acidic and can reduce stream pH (measure of the hydronium ion concentration) and elevate toxic forms of aluminum (Haines, 1981). When pH declines, the normal ionic salt balance of the fish is compromised and fish lose body salts to the surrounding water (Southerland
et al.,
1997). Sensitive fish species can experience acute mortality, reduced growth, skeletal deformities, and reproductive failure (Haines, 1981).

Climate Change and Climate Variability

Possible climate change impacts to river herring were noted in the stock assessment (ASMFC, 2012) based on regional patterns in trends (e.g., trawl surveys in southern regions showed declining trends more frequently compared to those in northern regions). However, additional information was needed on this topic to inform our listing decision, and as noted above, we held a workshop to obtain expert opinion on the potential impacts of climate change on river herring (NMFS, 2012b).

As discussed at the workshop, both natural climate variability and anthropogenic-forced climate change will affect river herring (NMFS, 2012b). Natural climate variability includes the Atlantic Multidecadal Oscillation, the North Atlantic Oscillation, and the El Niño Southern Oscillation. During the workshop, it was noted that impacts from global climate change induced by human activities are likely to become more apparent in future years (Intergovernmental Panel on Climate Change (IPCC), 2007). Results presented from the North American Regional Climate Change Assessment Program (NARCCAP—a group that uses fields from the global climate models to provide boundary conditions for regional atmospheric models covering most of North America and extending over the adjacent oceans) suggest that temperature will warm throughout the years over the northeast, mid-Atlantic and Southeast United States (comparing 1968-1999 to 2038-2069; NMFS, 2012b). Additionally, it was noted that there is an expected but less certain increase in precipitation over the northeast United States during fall and winter during the same years (NMFS, 2012b). In conjunction with increased evaporation from warmer temperatures, the Northeast and mid-Atlantic may experience decrease in runoff and decreased stream flow in late winter and early spring (NMFS, 2012b). Additionally, enhanced ocean stratification could be caused by greater warming at the ocean surface than at depth (NMFS, 2012b).

Many observed changes in river herring biology related to environmental conditions were noted at the workshop, but few detailed analyses were available to distinguish climate change from climate variability. One analysis by Massachusetts Division of Marine Fisheries showed precipitation effects on spawning run recruitment at Monument River, MA (1980-2012; NMFS, 2012b). Jordaan and Kritzer (unpublished data) showed normalized run counts of alewife and blueback herring have a stronger correlation with fisheries and predators than various climate variables at broad scales (NMFS, 2012b). Once fine-scale (flow related to fishways and dams) data were used, results indicate that summer and fall conditions were more important. Nye
et al.
(2012) investigated climate-related mechanisms in the marine habitat of the United States that may impact river herring. Their preliminary results indicate the following: (1) A shift in northern ocean distribution for both blueback herring and alewife depending on the season; (2) decrease in ocean habitat within the preferred temperature for alewife and blueback herring in the spring; and (3) effects of climate change on river herring populations may depend on the current condition (e.g.,

abundance and health) of the population, assumptions, and temperature tolerances (e.g., blueback herring have a higher temperature tolerance than alewife).

Although preliminary, Nye
et al.
(2012) indicate that climate change will impact river herring. The results (also supported by Nye
et al.,
2009) indicate that both blueback herring and alewife have and will continue to shift their distribution to more northerly waters in the spring, and blueback herring has also shifted its distribution to more northerly waters in the fall (1975-2010) (Nye
et al.,
2012). Additionally, Nye
et al.
(2012) found a decrease in habitat (bottom waters) within the preferred temperature for alewife and blueback herring in the spring under future climate predictions (2020-2060 and 2060-2100). They concluded that an expected decrease in optimal marine habitat and natal spawning habitat will negatively affect river herring populations at the southern extent of their range. Additionally, Nye
et al.
(2012) infer that this will have negative population level effects and cause population declines in southern rivers, resulting in an observed shift in distribution which has already been observed. Nye
et al.
(2012) also found that the effects of climate change on river herring populations may depend on the current condition (e.g., abundance and health) of the population, assumptions, and temperature tolerances. Using the model, projections of alewife distribution and abundance can be predicted for each year, but for ease of interpretation, 2 years of low and high relative abundance were chosen to illustrate the effects of population abundance and temperature on alewife distribution. The low and high abundance years were objectively chosen as the years closest to −1 and +1 standard deviation from overall mean abundance. Two years closest to the −1 and +1 standard deviation from mean population abundance were selected to reflect the combined effect of warming with low and high abundance of blueback herring. The difference in species response (as noted below) may reflect the different temperature tolerances (9-11 °C for blueback herring and 4-11 °C for alewife) as indicated by the southern limit of their ranges. Blueback herring may be able to tolerate higher temperature as their range extends as far south as Florida, but the southern extent of the alewife's range is limited to North Carolina. For both species, the Nye
et al.
(2012) analysis indicates that, if robust populations of these species are maintained, declines due to the effects of climate change will be reduced. Their specific results include the following:

• Alewife: At low population size, coast-wide abundance is projected to decrease with less suitable habitat and patchy areas of high density in the Gulf of Maine and Georges Bank in 2060-2100. At high population size, abundance is projected to increase slightly from 2020-2060 (+4.64 percent) but is projected to decrease (−39.14 percent) and become more patchy in 2060-2100.

• Blueback herring: Abundance is projected to increase at both high and low population size throughout the Northeast United States, especially in the mid-Atlantic and Georges Bank. However, at low abundance the increase is minimal and remains at a level below the 40-year mean. The percentage change due to climate change (factoring only temperature) is +29.93 percent for the time period 2020-2060 and +55.81 percent from 2060-2100.

We hoped to obtain information during the workshop on potential impacts of climate change by region, including information on species, life stage, indicators, potential impacts, and available data/relevant references (NMFS, 2012b). Although we did obtain information on each of these categories, substantial data gaps in the species information were apparent (NMFS, 2012b). For example, although no specific information on impacts of ocean acidification on river herring was presented, possible effects on larval development, chemical signaling (olfaction), and de-calcification of prey were noted (NMFS, 2012b). Additional research is needed to identify the limiting factor(s) for river herring populations. As Nye
et al.
(2012) noted, the links between climate and river herring biology during freshwater stages are unclear and will require additional time to research and thoroughly analyze. This conclusion is supported by the results of the workshop, which noted numerous potential climate effects on the freshwater stages, but little synthesis has been accomplished to date. The preliminary analysis of Nye
et al.
(2012) indicates that water temperatures in the rivers will be warmer, and there will be a decrease in the river flow in the northeast and Mid-Atlantic in late winter/early spring.

Although current information indicates climate change is and will continue to impact river herring (e.g., Nye
et al.,
2012), climate variability rather than climate change is expected to have more of an impact on river herring from 2024-2030. Several studies have shown that the climate change signal is readily apparent by the end of the 21st century (Hare
et al.,
2010; Hare
et al.,
2012). At intermediate time periods (e.g., 2024-2030), the signal of natural climate variability is likely similar to the signal of climate change. Thus, a large component of the climate effect on river herring in 2024-2030 will be composed of natural climate variability, which could be either warming or cooling.

Summary and Evaluation of Factor A

Dams and hydropower facilities, water quality and water withdrawals from urbanization and agricultural runoff, dredging and other wetland alterations are likely the causes of historical and recent declines in abundance of alewife and blueback herring populations. Climate variability rather than climate change is expected to have more of an impact on river herring from 2024-2030 (NMFS' foreseeable future for river herring). Nye
et al.,
(2012) conducted a preliminary analysis investigating climate-related mechanisms in the marine habitat of the United States that may impact river herring, and found that changes in the amount of preferred habitat and a potential northward shift in distribution as a result of climate change may affect river herring in the future (e.g., 2020-2100). Thus, the level of threat posed by these potential stressors is evaluated further in the qualitative threats assessment as described below.

B. Overutilization for Commercial, Recreational, Scientific, or Educational Purposes

Directed Commercial Harvest

This following section on river herring fisheries in the United States is from the stock assessment (ASMFC, 2012).

Fisheries for anadromous species have existed in the United States for a very long time. They not only provided sustenance for early settlers but a source of income as the fisheries were commercialized. It is difficult to fully describe the characteristics of these early fisheries because of the lack of quantifiable data.

The earliest commercial river herring data were generally reported in state and town reports or local newspapers. In 1871, the U.S. Fish Commission was founded (later became known as the U.S. Fish and Fisheries Commission in 1881). This organization collected fisheries statistics to characterize the biological and economic aspects of commercial fisheries. Data describing historical river herring fisheries were

available from two of this organization's publications—the Bulletin of the U.S. Fish Commission (renamed Fishery Bulletin in 1971; Collins and Smith, 1890; Smith, 1891) and the U.S. Fish Commission Annual Report (USFC, 1888-1940). In the stock assessment, the river herring data were transcribed and when available, dollar values were converted to 2010 dollar values using conversion factors based on the annual average consumer price index (CPI) values, which were obtained from the U.S. Bureau of Labor Statistics. Note that CPI values are not available for years prior to 1913 so conversion factors could not be calculated for years earlier than 1913 (ASMFC, 2012).

There are several caveats to using the historical fisheries data. There is an apparent bias in the area sampled. In most cases, there was no systematic sampling of all fisheries; instead, sampling appeared to be opportunistic, concentrating on the mid-Atlantic States. It is also difficult to assess the accuracy and precision of these data. In some instances, the pounds were reported at a fine level of detail (e.g., at the state/county/gear level), but details regarding the specific source of the data were often not described. The level of detail provided in the reports varied among states and years. Additionally, not all states and fisheries were canvassed in all years, so absence of landings data does not necessarily indicate the fishery was not active as it is possible that the data just were not collected. For these reasons, these historical river herring landings should not be considered even minimum values because of the variation in detail and coverage over the time series. No attempt was made to estimate missing river herring data since no benchmark or data characteristics could be found, and the stock assessment subcommittee also did not attempt to estimate missing data in a time series at a particular location because of the bias associated with these estimates (ASMFC, 2012).

During 1880 to 1938, reported commercial landings of river herring along the Atlantic Coast averaged approximately 30.5 million lbs (13,835 mt) per year. The majority of river herring landed by commercial fisheries in these early years are attributed to the mid-Atlantic region (NY-VA). The dominance of the mid-Atlantic region is, in part, due to the apparent bias in the spatial coverage of the canvass (see above). From 1920 to 1938, the average annual weight of reported commercial river herring landings was about 22.8 million lbs (10,351 mt). The value of the commercial river herring landings during this same time period was approximately 2.87 million dollars (2010 USD) (ASMFC, 2012).

Domestic commercial landings of river herring were presented in the stock assessment by state and by gear from 1887 to 2010 where available. Landings of alewife and blueback herring were collectively classified as “river herring” by most states. Only a few states had species-specific information recorded for a limited range of years. Commercial landings records were available for each state since 1887 except for Florida and the Potomac River Fisheries Commission (PRFC), which began recording landings in 1929 and 1960, respectively. It is important to note that historical landings presented in the stock assessment do not include all landings for all states over the entire time period and are likely underestimated, particularly for the first third of the time series, since not all river landings were reported (ASMFC, 2012).

Total domestic coast-wide landings averaged 18.5 million lb (8,399 mt) from 1887 to 1928 (
See table 2.2
in ASMFC (2012)). During this early time period, landings were predominately from Maryland, North Carolina, Virginia, and Massachusetts (overall harvest is likely underestimated because landings were not recorded consistently during this time). Virginia made up approximately half of the commercial landings from 1929 until the 1970s, and the majority of Virginia's landings came from the Chesapeake Bay, Potomac River, York River, and offshore harvest. Coast-wide landings started increasing sharply in the early 1940s and peaked at over 68.7 million lb (31,160 mt) in 1958 (
See Table 2.2,
ASMFC, 2012). In the 1950s and 1960s, a large proportion of the harvest came from Massachusetts purse seine fisheries that operated offshore on Georges Bank targeting Atlantic herring (G. Nelson, Massachusetts Division of Marine Fisheries, Pers. comm., 2012). Landings from North Carolina were also at their highest during this time and originated primarily from the Chowan River pound net fishery. Severe declines in landings began coast-wide in the early 1970s and domestic landings are now a fraction of what they were at their peak, having remained at persistently low levels since the mid-1990s. Moratoria were enacted in Massachusetts (commercial and recreational in 2005), Rhode Island (commercial and recreational in 2006), Connecticut (commercial and recreational in 2002), Virginia (for waters flowing into North Carolina in 2007), and North Carolina (commercial and recreational in 2007). As of January 1, 2012, river herring fisheries in states or jurisdictions without an approved sustainable fisheries management plan, as required under ASMFC Amendment 2 to the Shad and River Herring FMP, were closed. As a result, prohibitions on harvest (commercial or recreational) were extended to the following states: New Jersey, Delaware, Pennsylvania, Maryland, DC, Virginia (for all waters), Georgia and Florida (ASMFC, 2012).

Pound nets were identified as the dominant gear type used to harvest river herring from 1887 through 2010. Seines were more prevalent prior to the 1960s, but by the 1980s, they were rarely used. Purse seines were used only for herring landed in Massachusetts, but made up a large proportion of the landings in the 1950s and 1960s. Historically, gill nets made up a small percentage of the overall harvest. However, even though the actual pounds landed continued to decline, the proportion of gill nets that contributed to the overall harvest has increased in recent years (ASMFC, 2012).

Foreign fleet landings of river herring (reported as alewife and blueback shad) are available through the Northwest Atlantic Fisheries Organization (NAFO). Offshore exploitation of river herring and shad (generally <7.5 in (190 mm) in length) by foreign fleets began in the late 1960s and landings peaked at about 80 million lbs (36,320 mt) in 1969 (ASMFC, 2012).

Total U.S. and foreign fleet harvest of river herring from the waters off the coast of the United States (NAFO areas 5 and 6) peaked at about 140 million lb (63,560 mt) in 1969, after which landings declined dramatically. After 1977 and the formation of the Fishery Conservation Zone, foreign allocation of river herring (to both foreign vessels and joint venture vessels) between 1977 and 1980 was 1.1 million lb (499 mt). The foreign allocation was reduced to 220,000 lb (100 mt) in 1981 because of the condition of the river herring resource. In 1985, a bycatch cap of no more than 0.25 percent of total catch was enacted for the foreign fishery. The cap was exceeded once in 1987, and this shut down the foreign mackerel fishery. In 1991, area restrictions were passed to exclude foreign vessels from within 20 miles (32.2 km) of shore for two reasons: 1) In response to the increased occurrence of river herring bycatch closer to shore and 2) to promote increased fishing opportunities for the domestic mackerel fleet (ASMFC, 2012).

In-river Exploitation

The stock assessment subcommittee calculated in-river exploitation rates of the spawning runs for five rivers (Damariscotta River (ME—alewife),

Union River (ME—alewife), Monument River (MA—both species combined), Mattapoisett River (MA—alewife), and Nemasket River (MA—alewife)) by dividing in-river harvest by total run size (escapement plus harvest) for a given year. Exploitation rates were highest (range: 0.53 to 0.98) in the Damariscotta River and Union River prior to 1985, while exploitation was lowest (range: 0.26 to 0.68) in the Monument River. Exploitation declined in all rivers through 1991 to 1992. Exploitation rates of both species in the Monument River and of alewives in the Mattapoisett River and Nemasket River were variable (average = 0.16) and, except for the Nemasket River, declined generally through 2005 until the Massachusetts moratorium was imposed. Exploitation rates of alewives in the Damariscotta River were low (<0.05) during 1993 to 2000, but they increased steadily through 2004 and remained greater than 0.34 through 2008. Exploitation in the Damariscotta dropped to 0.15 in 2009 to 2010. Exploitation rates of alewives in the Union River declined through 2005 but have remained above 0.50 since 2007 (ASMFC, 2012).

According to the stock assessment, exploitation of river herring appears to be declining or remaining stable. In-river exploitation was highest in Maine rivers (Damariscotta and Union) and has fluctuated, but it is currently lower than levels seen in the 1980s. Also, in-river exploitation in Massachusetts rivers (Monument and Mattapoisett) was declining at the time a moratorium was imposed in 2005. The coast-wide index of relative exploitation also declined following a peak in the late 1980s and has remained fairly stable over the past decade. Exploitation rates declined in the DB-SRA model runs except when the input biomass-to-
K
ratio in 2010 was 0.01. Exploitation rates estimated from the statistical catch-at-age model for blueback herring in the Chowan River (see the NC state report in the stock assessment) also showed a slight declining trend from 1999 to 2007, at which time a moratorium was instituted. There appears to be a consensus among various assessment methodologies that exploitation has decreased in recent times. The stock assessment indicates that the decline in exploitation over the past decade is not surprising because river herring populations are at low levels and more restrictive regulations or moratoria have been enacted by states (ASMFC, 2012).

Past high exploitation may also be a reason for the high amount of variation and inconsistent patterns observed in fisheries-independent indices of abundance. Fishing effort has been shown to increase variation in fish abundance through truncation of the age structure, and recruitment becomes primarily governed by environmental variation (Hsieh
et al.,
2006; Anderson
et al.,
2008). When fish species are at very low abundances, as is believed for river herring, it is possible that the only population regulatory processes operating are stochastic fluctuations in the environment (Shepherd and Cushing, 1990) (ASMFC, 2012).

Canadian Harvest

Fisheries in Canada for river herring are regulated through limited seasons, gears, and licenses. Licenses may cover different gear types; however, few new licenses have been issued since 1993 (DFO, 2001). River-specific management plans include closures and restrictions. River herring used locally for bait in other fisheries are not accounted for in river-specific management plans (DFO, 2001). DFO estimated river herring landings at just under 25.5 million lb (11,577 mt) in 1980, 23.1 million lb (10,487 mt) in 1988, and 11 million lb (4,994 mt) in 1996 (DFO, 2001). The largest river herring fisheries in Canadian waters occur in the Bay of Fundy, southern Gulf of Maine, New Brunswick, and in the Saint John and Miramichi Rivers where annual harvest estimates often exceed 2.2 million lb (1,000 mt) (DFO, 2001). Recreational fisheries in Canada for river herring are limited by regulations including area, gear and season closures with limits on the number of fish that can be harvested per day; however, information on recreational catch is limited. Licenses and reporting are not required by Canadian regulations for recreational fisheries, and harvest is not well documented.

Incidental Catch

The following section on river herring incidental catch in the United States is from the stock assessment (ASMFC, 2012).

Three recent studies estimated river herring discards and incidental catch (Cieri
et al.,
2008; Wigley
et al.,
2009; Lessard and Bryan, 2011). The discard and incidental catch estimates from these studies cannot be directly compared as they used different ratio estimators based on data from the Northeast Fishery Observer Program (NEFOP), as well as different raising factors to obtain total estimates. Cieri
et al.
(2008) estimated the kept (i.e., landed) portion of river herring incidental catch in the Atlantic herring fishery. Cieri
et al.
(2008) estimated an average annual landed river herring catch of approximately 71,290 lb (32.4 mt) in the Atlantic herring fishery for 2005-2007, and the corresponding coefficient of variation (CV) was 0.56. Cournane
et al.
(2010) extended this analysis with additional years of data. Further work is needed to elucidate how the landed catch of river herring in the directed Atlantic herring fishery compares to total incidental catch across all fisheries. Since this analysis only quantified kept river herring in the Atlantic herring fishery, it underestimates the total catch (kept plus discarded) of river herring across all fishing fleets. Wigley
et al.
(2009) quantified river herring discards across fishing fleets that had sufficient observer coverage from July 2007-August 2008. Wigley
et al.
(2009) estimated that approximately 105,820 lb (48 mt) were discarded during the 12 months (July 2007 to August 2008), and the estimated precision was low (149 percent CV). This analysis estimated only river herring discards (in contrast to total incidental catch), and noted that midwater trawl fleets generally retained river herring while otter trawls typically discarded river herring.

Lessard and Bryan (2011) estimated an average incidental catch of river herring and American shad of 3.3 million lb (1,498 mt)/yr from 2000-2008. The methodology used in this study differed from the Standardized Bycatch Reporting Methodology (SBRM) (the method used by NOAA's Northeast Fisheries Science Center (NEFSC) to quantify bycatch in stock assessments) (Wigley
et al.,
2007; Wigley
et al.,
2012). Data from NEFOP were analyzed at the haul level; however, the sampling unit for the NEFOP database is at the trip level. Within each gear and region, all data, including those from high volume fisheries, appeared to be aggregated across years from 2000 through 2008. However, substantial changes in NEFOP sampling methodology for high volume fisheries were implemented in 2005, limiting the interpretability of estimates from these fleets in prior years. Total number of tows from the fishing vessel trip report (VTR) database was used as the raising factor to estimate total incidental catch. The use of effort without standardization makes the implicit assumption that effort is constant across all tows within a gear type, potentially resulting in a biased effort metric. In contrast, the total kept weight of all species is used as the raising factor in SBRM. When quantifying incidental catch across multiple fleets, total kept weight of all species is an appropriate surrogate for effective fishing power because it is

likely that all trips will not exhibit the same attributes. Lessard and Bryan (2011) also did not provide precision estimates, which are imperative for estimation of incidental catch.

The total incidental catch of river herring was estimated as part of the work for Amendment 14 to the Atlantic Mackerel, Squid and Butterfish (MSB) Fishery Management Plan, that includes measures to address incidental catch of river herring and shads. From 2005-2010, the total annual incidental catch of alewife ranged from 41,887 lb (19.0 mt) to 1.04 million lb (472 mt) in New England and 19,620 lb (8.9 mt) to 564,818 lb (256.4 mt) in the Mid-Atlantic. The dominant gear varied across years between paired midwater trawls and bottom trawls. Corresponding estimates of precision (COV) exhibited substantial interannual variation and ranged from 0.28 to 3.12 across gears and regions. Total annual blueback herring incidental catch from 2005 to 2010 ranged from 30,643 lb (13.9 mt) to 389,111 lb (176.6 mt) in New England and 2,645 lb (1.2 mt) to 843,479 lb (382.9 mt) in the Mid-Atlantic. Across years, paired and single midwater trawls exhibited the greatest blueback herring catches, with the exception of 2010 in the mid-Atlantic where bottom trawl was the most dominant gear. Corresponding estimates of precision ranged from 0.27 to 3.65. The temporal distribution of incidental catches was summarized by quarter and fishing region for the most recent 6-year period (2005 to 2010). River herring catches occurred primarily in midwater trawls (76 percent, of which 56 percent were from paired midwater trawls and the rest from single midwater trawls), followed by small mesh bottom trawls (24 percent). Catches of river herring in gillnets were negligible. Across gear types, catches of river herring were greater in New England (56 percent) than in the Mid-Atlantic (44 percent). The percentages of midwater trawl catches of river herring were similar between New England (37 percent) and the Mid-Atlantic (38 percent). However, catches in New England small mesh bottom trawls were three times higher (18 percent) than those from the Mid-Atlantic (6 percent). Overall, the highest quarterly catches of river herring occurred in midwater trawls during Quarter 1 in the Mid-Atlantic (35 percent), followed by catches in New England during Quarter 4 (16 percent) and Quarter 3 (11 percent). Quarterly catches in small mesh bottom trawls were highest in New England during Quarter 1 (7 percent) and totaled 3 to 4 percent during each of the other three quarters.

Recreational Harvest

The Marine Recreational Fishery Statistics Survey (MRFSS) provided estimates of numbers of fish harvested and released by recreational fisheries along the Atlantic coast. The stock assessment subcommittee extracted state harvest and release estimates for alewives and blueback herring from the MRFSS catch and effort estimates files available on the web (
http://www.sefsc.noaa.gov/about/mrfss.htm
). Historically, there were few reports of river herring taken by recreational anglers for food. Most often, river herring were taken for bait. MRFSS estimates of the numbers of river herring harvested and released by anglers are very imprecise and show little trend. Thus, the stock assessment concluded that these data are not useful for management purposes. MRFSS concentrates their sampling strata in coastal water areas and does not capture any data on recreational fisheries that occur in inland waters. Few states conduct creel surveys or other consistent survey instruments (diary or log books) in their inland waters to collect data on recreational catch of river herring. Some data are reported in the state chapters in the stock assessment; but the stock assessment committee concluded that data are too sparse to conduct any systematic comparison of trends (ASMFC, 2012).

Scientific Monitoring and Educational Harvest

Maine, New Hampshire, Massachusetts and Rhode Island estimate run sizes using electronic counters or visual methods. Various counting methods are used at the Holyoke Dam fish lift and fishways on the Connecticut River. Young of year (YOY) surveys are conducted through fixed seine surveys capturing YOY alewife and blueback herring generally during the summer and fall in Maine, Rhode Island, Connecticut, New York, New Jersey, Maryland, District of Columbia, Virginia and North Carolina. Rhode Island conducts surveys for juvenile and adult river herring at large fixed seine stations. Virginia samples river herring using a multi-panel gill net survey and electroshocking surveys. Florida conducts electroshocking surveys to sample river herring. Maine, New Hampshire, Massachusetts, Rhode Island, Maryland, and North Carolina collect age data from commercial and fisheries independent sampling programs, and length-at-age data. All of these scientific monitoring efforts are believed to have minimal impacts on river herring populations.

Summary and Evaluation of Factor B

Historical commercial and recreational fisheries for river herring likely contributed to the decline in abundance of both alewife and blueback herring populations. Current directed commercial and recreational alewife and blueback herring fisheries, as well as commercial fishery incidental catch may continue to pose a threat to these species. Since the 1970s, regulations have been enacted in the United States on the directed harvest of river herring in an attempt to halt or reverse their decline with the most recent regulations being imposed in January 2012. Additionally, there are regulations in Canada on river herring harvest. Historical landings data and current fishery effort is the best available information to describe the impact that the commercial fishery may be having on river herring.

Moratoria are in place on directed catch of these species throughout most of the United States; however, they are taken as incidental catch in several fisheries. The extent to which incidental catch is affecting river herring has not been quantified and is not fully understood. Thus, the level of threat posed by directed and indirect catch is evaluated further in the qualitative threats assessment as described below. Scientific collections or collections for educational purposes do not appear to be significantly affecting the status of river herring, as they result in low mortality.

C. Disease and Predation

Disease

Little information exists on diseases that may affect river herring; however, there are reports of a variety of parasites that have been found in both alewife and blueback herring. The most comprehensive report is that of Landry
et al.
(1992) in which 13 species of parasites were identified in blueback herring and 12 species in alewives from the Miramichi River, New Brunswick, Canada. The parasites found included one monogenetic trematode, four digenetic trematodes, one cestode, three nematodes, one acanthocephalan, one annelid, one copepod and one mollusk. The same species were found in both alewife and blueback herring with the exception of the acanthocephalan, which was absent from alewives.

In other studies, Sherburne (1977) reported piscine erythrocytic necrosis (PEN) in the blood of 56 percent of prespawning and 10 percent of

postspawning alewives in Maine coastal streams. PEN was not found in juvenile alewives from the same locations. Coccidian parasites were found in the livers of alewives and other finfish off the coast of Nova Scotia (Morrison and Marryatt, 1990). Marcogliese and Compagna (1999) reported that most fish species, including alewife, in the St. Lawrence River become infected with trematode metacercariae during the first years of life. Examination of Great Lakes fishes in Canadian waters showed larval
Diplostomum
(trematode) commonly in the eyes of alewife in Lake Superior (Dechtiar and Lawrie, 1988) and Lake Ontario (Dechtiar and Christie, 1988), though intensity of infections was low (<9/host). Heavy infections of
Saprolegnia,
a fresh and brackish water fungus, were found in 25 percent of Lake Superior alewife examined, and light infections were found in 33 percent of Lake Ontario alewife (Dechtiar and Lawrie, 1988). Larval acanthocephala were also found in the guts of alewife from both lakes.
Saprolegnia
typically is a secondary infection, invading open sores and wounds, and eggs in poor environmental conditions, bu

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Source: Frix Law Library, https://www.frixlaw.com/law-library/documents/fr%3A2013-19380. Public record. Not legal advice.