National Marine Fisheries Service Endangered Species Act Section 7 Consultation

Biological Opinion

Agency: United States Department of Homeland Security, United States

Coast Guard

Activities Considered: An assessment of the United States Coast Guard's National

Ballast Water Management Program and Initial Numerical

Standard

Consultation Conducted by: Endangered Species Act Interagency Cooperation Division of

the Office of Protected Resources, National Marine Fisheries

Service

Approved by:

Date: 

### JUN 20 

## 2012

Section 7(a)(2) of the Endangered Species Act (ESA), as amended requires each Federal agency

to insure that any action they authorize, fund, or carry out is not likely to jeopardize the

continued existence of any endangered or threatened species or result in the destruction or adverse modification of critical habitat of such species (16 U.S.C. 1531 et seq.). When the

action of a Federal agency "may affect" a threatened or endangered species or critical habitat that

has been designated for them, that agency is required to consult with either the National Marine

Fisheries Service (NMFS) or the U.S. Fish and Wildlife Service (USFWS) (together, "the Services"), depending upon the species that may be affected by the action. In this case, NOAA

Fisheries conducted formal programmatic consultation with the U.S. Coast Guard on their ballast

water management program to determine whether the program complies with the requirements of

section 7(a)(2) of the Endangered Species Act (that is, the requirement to insure that their actions

are not likely to jeopardize the continued existence of endangered or threatened species or result

in the destruction or adverse modification of critical habitat that has been designated for those species).

This document represents NMFS' biological opinion (Opinion) on the United States Coast

Guard's (USCG) ballast water management program, authorized under the Nonindigenous Aquatic Nuisance Prevention and Control Act of 1990 (NANPCA) as reauthorized and amended

in the National Invasive Species Act of 1996 (NISA). Within this program, the USCG is proposing to establish, monitor, enforce, and if practicable and determined to be necessary, revise, a ballast water standard. Establishment of this standard is intended to ensure to the

maximum extent practicable that aquatic nuisance species are not discharged into waters of the

United States from vessels. The purpose ofthis consultation is to evaluate the ballast water management program and to analyze risks associated with introducing non-native species from discharged ballast water meeting the proposed numerical standard. This programmatic

12 consultation assesses the probable exposure and possi ble risks associated w ith the ballast water management standard and then reach conclusi ons on the USCG’s ballast water management program generally defined in the final rule (F R) for the allowable c oncentration of living organisms in ships’ ballast water discharged in waters of the United States. While this action addresses ballast water management , non-native species can also be introduced by water held in the sea chest or by reproduction of organisms attach ed to the hull of the ve ssel that are not within the scope of this consultation.

This Opinion has been prepared in accordance with section 7(a)(2) of the ESA of 1973, as amended (16 U.S.C. 1531 et seq.) and implem enting regulations at 50 CFR 402. However, consistent with a decision rendered by the Ninth Circuit Court of Appeals on August 6, 2004, we did not apply the regulatory definition of “destr uction or adverse modification of critical habitat” at 50 CFR 402.02. Instead, we relied on the statut ory provisions of the ESA to complete our analysis of the effects of the ac tion on designated critical habitat. Essential fish habitat (EFH) consultations, in accordance with Section 305(b)(2) of the Magnuson-Stevens Fishery Conservation and Management Act (MSA) (16 U.S.C. 1801, et seq. ) and implementing regulations at 50 CFR 600, are conduc ted at a regional level and are not within the scope of this

consultation.

This Opinion is based on our review of the October 21, 2011, letter re questing initiation of consultation, the final Programmatic Environmenta l Impact Statement (PEIS), draft PEISs, the FR, and supporting documentation; the draft U.S. recovery plans for Sacramento winter-run Chinook salmon, Central Valley spring-run Chinook salmon, Southern California steelhead, and California Central Valley steelhead trout; the fi nal U.S. recovery plans for smalltooth sawfish, Atlantic salmon, Upper Columbia River sp ring Chinook salmon, Puget Sound Chinook salmon, Hood Canal chum salmon, gulf sturgeon, shortn ose sturgeon, Ozette Lake sockeye salmon, Upper Columbia River steelhead, Middle Columbia River steelhead, blue whale, fin whale, humpback whale, southern reside nt killer whale, North Atlantic right whale, sperm whale, Hawaiian monk seal, eastern and western Steller sea lions, Nort hwest Atlantic population of loggerhead turtles, Pacific populations of the logg erhead turtle, Atlantic populations of green turtles, Pacific populations of the green turtle, ha wksbill turtle in the Atlantic, Pacific populations of the hawksbill turtle, leatherback turtle in the Atlantic, Pacific populations of the leatherback turtle, Pacific populations of the olive ridley turtle; the bi-national recovery plan for Kemp’s ridley turtle; white papers; primary literature ; past and current research, both published and unpublished; the documents that were used to list green sturgeon and smalltooth sawfish as threatened and endangered species (respectively); information pertinent to the performance of other USCG environmental enforcement programs; and monitoring report s from prior ballast water reporting and recordkeeping requirements.

Consultation History Prior to discussing the final PEIS and potenti al impacts to NMFS trust resources, NMFS met with the USCG multiple times between 2006 and 2011. These meetings and comments were part of the pre-consultation process. While formal programmatic consultation was not initiated until November 22, 2011, the agencies shared draft letters requesting initiation th at were important to the programmatic consultation process.3 On January 31, 2011, NMFS received a draft letter from the USCG requesting initiation of

consultation.

On March 9, 2011, NMFS provided a response to th e USCG letter, providing a complete species list and requesting a consistent use of th e environmental baseline for ESA purposes.

On March 29, 2011, NMFS and USCG had a teleconf erence to discuss ESA baseline and ballast water management standard impacts. NMFS stated that even if ballast water standards are an improvement over previous treatments, the draft PEIS did not support a co nclusion of not likely to adversely affect listed species or their critic al habitat. Previous determinations regarding ballast water exchange (BWE) had addressed upta ke and physical impacts (thermal, salinity), but not effects from non-native species. During the meeting, we also discussed whether to address the proposed International Maritime Organiza tion (IMO) standard or program including all

feasible alternatives and the USCG’s long term goal. NMFS recommended a programmatic approach because of the nature of the action.

On March 30, 2011, NMFS and USCG met to discu ss edits made to the final draft PEIS in response to NMFS’ comments and to establish the appropriate steps forward to initiate

consultation.

On May 12, 2011, the USCG provided NMFS with a dr aft request for initiation and asked NMFS to provide comments.

On May 23, 2011, NMFS responded that the informati on in the letter did not fully reflect the outcome of the March 30 meeting and the effects analysis did not seem to support the conclusion reached in the letter. Therefore, on May 24, 2011, NMFS and the USCG met to discuss the initiation request letter.

On June 22, 2011, USCG provided NMFS with anot her draft reflecting agreements reached on May 24, 2011.

On August 1, 2011, NMFS provided additional commen ts to the USCG reflecting agreements reached on May 24, 2012, and clarified on June 22, 2012.

On August 12, 2011, the USCG provided a new draft with changes made in response to NMFS’ comments. NMFS agreed with the draft.

On October 28, 2011. NMFS received the final, si gned request for initiation of programmatic consultation from the USCG, dated October 21, 2011. In the supporting documentation for that request, the USCG determined their action “may e ffect” listed species and their critical habitat, but left it up to NMFS to determine which speci es would be adversely affected and which were not likely to be adversely affected.

On November 9, 2011, NMFS hosted a meeting with USFWS and USCG to ensure all three agencies agreed on the proposed action and proposed determin ation of effects.4 A series of emails between November 14, 2011, and November 22, 2011, discussed NMFS’ intentions of initiating formal consultation to prod uce an Opinion rather than providing a letter of

concurrence.

On November 22, 2011, the USCG and NMFS agreed to formal programmatic consultation on the USCG ballast water management program as was requested in the October 28, 2012, letter (dated October 21, 2012).

On December 6, 2011, NMFS and USFWS attended a meeting at the USCG, where the USCG had several questions about the environmental base line and how that led to a likely to adversely

affect determination. There was also discussi on about the ESA timeline and when consultation could be completed. Additionally, all three agenci es agreed to meet on a biweekly basis so the USCG could be updated on the progress of the Opinions.

On December 20, 2011, NMFS, USFWS, and the USCG met. The main focus of the conversation was the scope of the proposed action. The USCG suggested the consultation should not be programmatic but rather only addre ss the interim final rule. The USCG requested case law on segmentation of the consultation process. Because no agreement was reached on the scope of the consultation, both agencies agreed to discuss this topic afte r the holidays on January 5, 2012.

On December 21, 2011, the USCG requested th at NMFS continue consultation under the programmatic scope as wa s initially requested.

On December 23, 2011, NMFS’ Office of General Council provided examples of general ESA

case law relevant to the ESA base line and the consultation process.

On January 5, 2012, NMFS, USFWS, and USCG met with the main topic again being the scope of the proposed action. NMFS provided a summa ry of the proposed action and the USCG agreed to provide any comments to both agencies by January 9, 2012.

On January 17, 2012, NMFS, USFWS, and USCG me t to discuss the continued disagreement on the scope of the proposed action. NMFS agreed to provide the Proposed Action section from the Opinion so both agencies could identify wher e the disagreement originated and whether the agencies could reach agreement. NMFS also reque sted information related to monitoring and the USCG provided information related to the oil wate r separator program which would be similar to the compliance monitoring plan in the FR.

On January 31, 2012, NMFS, USFWS, and USCG met to discuss the proposed action and effects analysis. NMFS also requested additional inform ation about the monitoring program and agreed to present those questions in an email following the meeting.

On February 1, 2012, NMFS received comments on the proposed action section. NMFS addressed the comments, accepting much of the language and explaining the reason for differences in other places. NMFS provided USCG with another draft of the proposed action on Febraury 2, 2012.5 On February 9, 2012, NMFS and USCG agreed on the proposed action se ction after a final delivery of comments and explanations by the USCG.

On March 6, 2012, NMFS, USFWS, and USCG met to update each other on the progress of the Opinion and FR. NMFS and USFWS both agreed th ey would have a draft of the Opinion ready to share on March 30, 2012.

On March 30, 2012, NMFS provided a draft Opinion to the USCG.

On April 18, 2012, the USCG provided NMFS with comments on the draft Opinion.

On May 8, 2012, after incorpora ting USCG’s comments, NMFS provided the USCG with a

second final draft.

On May 15, 2012, the USCG provided a second r ound of comments focusing on the difference between the practicability re view and periodic review.

On May 24, 2012, NMFS and the USCG agreed on the portion of the proposed action that identifies the differences between the pr acticability review a nd periodic review.

Legal and Policy Framework for Ballast Water Regulation by the USCG

Under the Nonindigenous Aquatic Nuisance Prev ention and Control Act of 1990 (NANPCA) as

reauthorized and amended in the National Invasi ve Species Act of 1996 (NISA), the USCG is the lead federal agency in implementing regulatio ns to reduce or preven t the introduction of nonindigenous species via shipping activities in United States (U S) waters. The purposes of NANPCA are to (1) prevent unint entional introduction and disper sal of non-indigenous species (NIS) into waters of the US through ballast water manageme nt and other requirements, (2) coordinate federally conducted, funded, or aut horized research; and to federally coordinate prevention control, information dissemination, a nd other activities rega rding the zebra mussel and other aquatic NIS, (3) develop and carry out environmentally sound control methods to prevent, monitor, and control unintentional in troductions of NIS from pathways other than ballast water exchange, (4) understand and mi nimize ecological and economic impacts of NIS that become established, including the zebra mussel, and (5) establish a pr ogram of research and technology development and assistance to states in the management and removal of zebra mussels. Under NANPCA, Congress directed th e USCG to develop a national program of specific regulations and guideline s intended to prevent or reduce the introduction and control the spread of NIS by ensuring to the maximum extent practicable that NIS are not discharged into waters of the US from vessels equi pped with ballast water tanks.

On April 8, 1993, the USCG published a final rule , “Ballast Water Management for Vessels Entering the Great Lakes” [58 FR 18330]. The final rule mandat ed the use of ballast water management practices by ships entering the Gr eat Lakes after operating outside the exclusive economic zone (EEZ). On December 30, 1994, the US CG published a final rule, “Ballast Water Management for Vessels Entering the Huds on River” [59 FR 67632], which expanded the6 mandatory requirements for the Great Lakes to portions of the Hudson River (north of the George Washington Bridge) that connect to th e Great Lakes. The management practices

mandated for use in the Great La kes and portions of the Hudson River are found in 33 CFR Part 151, Subpart C.

NISA reauthorized and amende d NANPCA and mandated the US CG to issue ballast water management guidelines and regulations applicab le to all US waters. Under NISA, Congress authorized the USCG to establish a mandatory ballast water management program [69 FR 44952], if shipping industry co mpliance under the initially es tablished national voluntary guidelines was insufficient. NISA also authorized the environmentally sound alternative methods, if the alternative methods are at least as effective as ballast water exchange. This includes any method, effort, or program to preven t or reduce invasive sp ecies introductions or control infestations, that minimizes adverse impacts to the structure and function of an ecosystem, and minimizes adverse effects on non target organisms and ecosystems (USCG 2003).

On May 17, 1999, the USCG published the interim ru le for the Implementation of the NISA [64 FR 26672] and finalized it on November 21, 2001 [66 FR 58381]. The rule set voluntary ballast water management guidelines applicable to all US waters, except for the Great Lakes and the Hudson River, where the mandatory requirement s remained in effect, and mandated reporting requirements. Both the voluntary manageme nt guidelines and the mandatory reporting requirements applied to vessels en tering US waters after operati ng in waters outside the EEZ. The rule also promoted ballast water management for all vessels operating within US waters (defined as the waters of all 50 states, the Dist rict of Columbia, Puerto Rico, Guam, American Samoa, the US Virgin Islands, and the Tr ust Territories of th e Pacific Islands).

As a result of several years of public comme nts and intra-governmental meetings, the USCG published a notice on September 26, 2003 [68 FR 55559], announcing USCG’s intent to prepare a PEIS to assess the impacts of developing a ballast water ma nagement standard. In the meantime, the USCG also published the final ru le on the mandatory ballast water management program for all US waters, under NISA, on July 28, 2004 [69 FR 44952]. This rule requires vessels that operate outside the US EEZ to use one of the foll owing management practices: (1) conduct mid-ocean ballast water exchange 200 nautic al miles from any shore, (2) retain ballast water onboard, or (3) use a USCG -approved alternative method.

Description of the Ballast Water Management Program

Since 2003, the USCG has been developing a ball ast water management program in cooperation with other federal agencies. The program is described by the USCG in the October 21, 2011, letter requesting initiation of consultation, the FR, and the final PEIS and framed by the NANPCA and the NISA. The goal of the USCG ’s ballast water management program is identified in NISA, and specifi ed in the final PEIS, as prev enting the unintentional introduction and dispersal of NIS into waters of the US through ballast water treatment and other requirements. This action will change language in 33 CFR 151 to reflect changes to the program. The USCG is also amending its regula tions to establish a type-approval requirement for ballast water management systems, which wi ll undergo, if necessary, separate consultation tiered from this programmatic consultation.7 Actions Authorized by the Program The USCG’s ballast water management program pr oposes to establish a nu merical standard and

allows for a stricter standard to be established via a series of pr acticability reviews. The USCG’s FR establishes the following ballast water management standard:

(1) For organisms greater than or equal to 50 micrometers in minimum dimension: discharge must include fewer than 10 living organisms per cubic meter (m 3 ) of ballast

water.

(2) For organisms less than 50 micrometers and greater than or equal to 10 micrometers: discharge must include fewer than 10 livi ng organisms per millil iter (mL) of ballast

water.

(3) Indicator microorgani sms must not exceed:

(i) For Toxicogenic Vibrio cholerae (serotypes O1 and O139): a concentration of less than 1 colony forming unit (cfu) per 100 mL.

(ii) For Escherichia coli : a concentration of fewer than 250 cfu per 100 mL.

(iii) For intestinal ente rococci: a concentration of fewer than 100 cfu per 100 mL.

Practicability Review and Periodic Review NISA requires the USCG to assess and, if neces sary, revise USCG’sballast water management regulations not less than every three years based onthe best scie ntific information available at the time of that review (16 U.S.C.4711(e)(1). The FR establishes apracticab ility review, completed no later than January 1, 2016, to de terminewhether technology and testing protocols are available to practicably implementa more stringent ballas t water discharge standa rd. The practicability review (33 CFR 151.2030(c)) will eval uate the capability of any id entified technology to achieve a more stringent ballast water management standard; the effectiveness of any identified technology in the shipboard environment; the comp atibility of any identified technology with vessel design and operation; the safety of any identified technology; whether the use of any identified technology would adversely affect the environment; the cost and cost-effectiveness of any identified technology; the economic impact of any identified tec hnology; the availability, accuracy, precision and cost effectiveness of methods to measure concentrations of organisms or chemicals per volume of treated ballast water; an d any other factors the USCG considers relevant

to the review.

Additionally, under NANPCA/NISA (16 U.S.C. 4711(e)(1)), the USCG will assess the

effectiveness of the USCG's ballast water mana gement regulations in reducing the introduction and spread of aquatic nuisance species by vesse ls and, as necessary and based on the best scientific information available at the time, revise the regulations and/or promulgate additional regulations pursuant to NANPCA/NI SA. Understanding the best scie ntific information available changes with the advent of time, the periodic reviews will likely include assessments of the practicability of technology or ba llast water testing capabilities that would allow for a more stringent ballast water discharge standard, including the capability of a new technology to meet a more stringent standard, the capab ility of testing methods to a llow for enforcement of a more stringent standard, the effectiven ess of approved technologies in the shipboard environment, and8 whether the use of any new technologies may have an adverse impact on the environment. To fulfill the NISA/NANPCA statutory mandate of ensu ring to the maximum extent practicable that aquatic nuisance species are not discharged into U.S. waters from vessels equipped with ballast tanks, the periodic reviews will likely entail a broader perspective than the review under 33 CFR 151.2030(c), focusing not solely on the practicability of implementing a more stringent discharge standard, but including assessments of practicable ballast water treatment as well as other ballast water, vessel management, and compliance assessm ent practices that could be implemented to achieve greater reductions in the risks of vesse l-mediated introductions of aquatic nuisance species.

Reporting and Record Keeping Vessels continue to be required to report and maintain records of their ballast water management and be subject to onboard testing during port and flag state control examinations as well as annual inspections (16 U.S.C. 4711(c)(2)(F). Th ese requirements are met by the reporting and recordkeeping regulations (33 CFR 151.2060 and 33 CFR 151.2070, respectively) and Shipboard Testing Requirements (46 CFR 162.060-28). Inform ation on ballast water management must be retained by the owner and/or operator for 24 mont hs, with the option of retaining information electronically for all but the most recent six m onths. The USCG has removed the ballast water reporting form from the CFR (previously an appendix to 33 CFR Subpart D) in order to streamline any future changes to the form. Th e form is still the proper form to satisfy the reporting requirements, but 33 CFR 151.2070 has been revised to reference the National Ballast Information Clearinghouse (NBIC) website as the form’s location. The Coast Guard has initiated a separate rulemaking to revise the form as needed.

Each vessel will be required to record vessel information, voyage information, ballast water information, ballast water management inform ation, information on ballast water to be discharged, information on sediment to be discha rged, and certification that the information is accurate. The required vessel information will be the name of the vessel, that vessel’s IMO number, vessel type, owner or operator, gross to nnage, call sign, and state of registry. The required voyage information will be date and port of arrival, the vessel agent, the last port of call, and next port of call. The required ballast wate r information will be total ballast water capacity,

total volume of ballast water onbo ard, total number of ballast ta nks, and total number of ballast tanks in ballast. The required ballast water mana gement information will be the total number of ballast tanks/holds that will be discharged into waters of the US, all alternative ballast water management methods used for each ballast ta nk, and identify whethe r there is a ballast management plan onboard and whether it was use d. The required information on ballast water to be discharged will be the origin of the ballas t water, including dates, locations, volumes, and temperatures; treatment information will incl ude dates, locations, volumes, method, percent exchanged (if using ballast water exchange), and s ea height at time of exchange or treatment; and the expected date, location, volume and salinity of any water to be discharged. The required information on ballast sediment to be discharged will be the name and lo cation of the facility where sediment disposal will take place.

Compliance Monitoring The USCG identifies the components of thei r compliance monitoring in 33 CFR 151.2075. The compliance monitoring allows the USCG to veri fy that vessels are equipped with approved9 technology. This section of the FR also authoriz es the USCG to board vessels and take samples

of ballast water as well as ballast sediments to verify treatment technologies are being used to

treat ballast water and to ensure vessels are mee ting the ballast water numerical standard. This monitoring is established to function in esse ntially the same way the oil-water separator monitoring program works.

The USCG has been monitoring vessels under th e ballast water management program since 1995. Compliance monitoring is an important component of that monitoring. To determine whether a vessel conducted BWE, which was the previous required balla st water treatment method, the USCG sampled ballast tanks to determ ine if the water in th e ballast tank had been exchanged or if it was still from the original port. As is proposed in the FR, the USCG will monitor self-reports, records, a nd continue to conduct on-board m onitoring to verify that ships are complying with the ballast water management program.

All vessels are required to in stall approved ballast water treatment technologies by the vessel’s first scheduled drydock after January 1, 2016. If a ship is unable to install the technology by that date, they may seek an extension. The USCG will assess the reason behind the need for an extension and if warranted may grant up to an additional 12 months (dependent on vessel type and treatment technology to be installed). The USCG will sample vessels based on a decision matrix to verify compliance with the ballast water treatment program. The primary goal is verifying that treatment technol ogies capable of meeting the pr oposed ballast water standard were installed. Two additional goa ls are verifying the treatment technologies are being used and they continue to meet the ballast water standard after their installation.

The USCG will continue to sample ballast tanks to ensure the ballast water treatment technology is effective (effectiveness monitoring) but the monitoring will depend on both the vessel type and the technology used. Because technologies will be approved on an individual basis, effectiveness monitoring for each particular technology will be evaluated in separate consultations. The tiered consultation is two part , an evaluation of how the technology performs prior to installation and taking wa ter and sediment samples from vessel ballast tanks to ensure the technology is being maintained and functioni ng properly after installa tion. The language for the effectiveness technology is included in the FR in 33 CFR 151.2075.

Mitigation Measures The FR include a section (33 CFR 151.2050) desc ribing appropriate mitigation measures that ships should take in addition to all required ballast wa ter discharge requirements. The mitigation measures are required and will have a positive effect on the environment. The mitigation measures direct vessel owners to avoid ballast water uptake and discharge in marine sanctuaries, marine preserves, marine parks, or coral reefs. Furthermore, the mitigation measures require vessel operators to clean their ballast tanks re gularly, discharge only the minimal amount of ballast water necessary, maintain a ballast water management plan, train vessel crew on ballast water treatment procedures, clean the anchor ch ains, and remove fouling organisms from vessel hulls, pipes, and tanks regularly.

Limitations of Coverage10 Certain vessels, including offshor e floating platforms, are exempted from this FR by NISA. The vessel exemptions are identifie d in 33 CFR 151.2015. Crude oil tankers engaged in coastwide trade and vessels operating exclusively within a single captain of the port zone (smaller than a bioregion) are exempted from the ballast wate r standard, reporting, or record keeping. Additionally, non sea going vessels ; sea going vessels of less th an 3,000 tons operating across captain of the port zones but not outside of the EEZ; and vessels that take on and discharge ballast water in the same captain of the port zone are exempted from meeting the ballast water standard, but must still report and maintain records as identified above. Exempted vessels are not required to submit a ballast water reporting fo rm to the USCG, but non-exempted vessels are required to report, regardless of whether a nd how they discharged ballast water.

Implementation Dates Non-exempt vessels are required to meet the ba llast water management standard as early as December 1, 2013 and all others by the first sc heduled drydock after January 1, 2016. Vessels

constructed on or after December 1, 2013, will be required to meet these standards on delivery, while vessels constructed before December 1, 2013, will need to meet the standard after their first drydocking after January 1, 2014, (vessels with ballast water cap acities of 1,500-5,000 m 3 ) or January 1, 2016 (vessels with ballas t water capacities of less than 1,500 m 3 and greater than 5,000 m 3 ). The USCG will phase out the ballast water exchange requirement and replace it with this numerical standard according to these implementation dates.

Enforcement Significant civil and criminal penalt ies are codified in the NISA and the FR that may be enforced against violators of the statut e and/or proposed regulations. The various monitoring programs described in this Opinion are part of the USCG enforcement program.

Approach to the Assessment

NMFS approaches its program specific section 7 an alyses through a series of steps. The first step identifies those aspects of proposed actions that are likely to have direct and indirect effects on the physical, chemical, and biotic environment. As part of this step, we identify the spatial extent of these direct and indirect effects, incl uding changes in the spatial extent over time. The results of this step represent the action area for th e consultation. The second step of our analyses identifies the listed and proposed species and designated or propos ed critical habitat that are likely to co-occur with these eff ects in space and time and the nature of that co-occurrence (these represent our exposure analyses). In this step of our analyses, we try to identify the primary constituent elements (PCEs), number, age, life stag e, and gender of these resources that are likely to be exposed to an action’s effects and the populations or subpopulati ons those individuals represent. Once we identify the li sted and proposed resources that are likely to be exposed to an

action’s effects and the nature of that exposure, we examine whet her and how those resources are likely to respond given their exposure (these represent our response analyses).

The final steps of our analyses—establishing th e risks those responses pose to listed and proposed resources—are different for listed spec ies and designated crit ical habitat (these represent our risk analyses). Our jeopardy dete rminations must be based on an action’s effects on the continued existence of threatened or endangered species as those “species” were listed,11 which may encompass the biological species, su bspecies, or distinct population segments of vertebrate species. Because the continued existe nce of listed species depends on the fate of the populations that comprise them, the viability (probability of extinc tion or probability of persistence) of listed and proposed species de pends on the viability of the populations that comprise the species. Similarly, the continued existence of populations are determined by the

fate of the individuals that co mprise them; populations grow or decline as the individuals that comprise the population live, die, grow, mature, migrate, and reproduce (or fail to do so). Our

destruction or adverse modifica tion determinations must be based on an action’s effects on the

conservation values of the essentia l features of critical habitat.

A programmatic review, however, typically analyzes the general environmen tal consequences of a broad scope of policy a lternatives under consid eration by an agency program, in this case, managing a ballast water program to aid in cont rolling the introduc tion and spread of non-native species from ship’s ballast water. Subsequent co nsultations that “tier” off of these programmatic consultations, when warranted, analyze the project and site specific effects typical of most consultations. As the proposed rule for the ball ast water management program states, there are multiple phases within the program, the first of which is establishing the ballast water standard identified in the FR. Additionally, many ballas t water management systems could be used to

achieve the ballast water standar d, so tiered consu ltations will evaluate any further impacts that may occur due to chemical, biological, or mechani cal treatment of ballast water. All subsequent

section 7 consultations conducte d by NMFS personnel will be designed to determine whether and to what degree the specific ac tion under review fits within th e general pattern identified in the “parent” or national programmatic consulta tion, and will determine whether the specific action is or is not likely to jeopardize the continued existence of endangered and threatened species or result in the destruction or adverse modification of designated critical habitat.

We begin our programmatic cons ultations by recognizing that an agency’s program normally represents the agency’s decision to authorize, fund, or carry out a su ite or class of activities that may require subsequent decision-making. When we conduct programmatic examinations of proposals such as the USCG’s na tional ballast water management program, we ask whether or to what degree the Federal action agency (in this case, USCG) has constructed a decision-making process that will consider the information, sta ndards and criteria that NMFS considers during consultations on specific actions. We also ask wh ether that decision-making process is likely to produce outcomes that will prevent endangered or threatened species and designated critical habitat from being exposed to physical, chemical, or biotic stressors that will directly or indirectly reduce the reproductive success of enda ngered or threatened individuals, increase the extinction risks of the population(s) those individuals represent, or increase the extinction risks of the species those populations comp rise. Specifically, we ask:

1. Has USCG structured the program so they wi ll know or be able to reliably estimate the probable individual and cumulative e ffects of the action? For exam ple, has USCG structured the program so it will know or be able to reliab ly estimate the probable number of non-native organisms that will be introduced under the aut hority of the program? Ha s USCG structured the program so it will know or be able to reliab ly estimate the probable location of those

introductions that will be authorized by the program?12 2. Has USCG structured the program so it will know or be able to reliably estimate the physical, chemical, or biotic stressors that are likely to be produced as a direct or indirect result of the ballast management that will be authori zed by the program (that is, the stressors produced by the introduction of non-native species to waters of the US)?

3. Has USCG structured the program so it will know or be able to reliably determine whether or to what degree operators have complied w ith the conditions, restri ctions, or mitigation measures the proposed action requires?

4. Has USCG structured the program so it will know or be able to reliably estimate whether or to what degree specific endangered or threatened species or designa ted critical habitat are likely to be exposed to (a) chemical, physical, or biol ogical stressors associat ed with ballast water management, or; (b) to the ecological consequences (habitat alterations or indirect effects to species) as a result of ballast wate r management in the United States?

5. Has USCG structured the program so it wi ll continuously identify, collect, and analyze information to detect impacts from ballast wa ter management to endangered or threatened species or designated critical ha bitat resulting from stressors at concentrations, intensities, durations, or frequencies that are known or suspected to produce physical, physiological, behavioral, or ecological responses that have potential individual or cumulative adverse consequences for individual or ganisms or primary constituent elements of critical habitat?

6. Has USCG structured the program to empl oy an analytical methodology that considers (a) the status and trends of endangere d or threatened species or desi gnated critical ha bitat; (b) the demographic and ecological status of populations and individuals of those species given their exposure to pre-existing stressor s in different drainages and wa tersheds; (c) th e direct and indirect pathways by which endangered or threaten ed species or designated critical habitat might be exposed to the consequences of organisms carri ed in ballast to waters of the United States; and (d) the physical, physiological, behavior, so ciobiological, and ecological consequences of exposing endangered or threatened species or designated critical habitat to biological stressors

from ballast water that are known or suspected to produce physical , physiological, behavioral, or ecological responses, given their pre-exis ting demographic and ecological condition?

7. Has USCG structured the program so it will be able to limit endangered or threatened species or designated critical habitat from being exposed to stressors resulting from a numerical ballast water standard (a) at concentra tions, durations, or frequencies that are potentially harmful to individual listed organisms, popula tions, or the species, or; (b) to ecological consequences that are potentially harmful to indi vidual listed organisms, popula tions, the species or primary constituent elements of designa ted critical habitat? How quick ly would USCG be able to implement preventive measures?

Because it is difficult to prevent free-ranging organisms from being affected by anthropogenic stressors once they have been exposed, the most effective management measures are designed to influence the likelihood, duration, and magnitude of the exposure itself. For that reason, our assessment focuses on whether and to what degree the program minimizes endangered and threatened species and designated critical ha bitat exposure to stress ors authorized by the13 proposed action.

Our assessment focused on whether and to what degree USCG structured the program to minimize endangered or threatened species or cri tical habitat that has been designated for those species from being exposed to st ressors resulting from ballast di scharges because such exposures commonly trigger a cascade of events with ulti mate consequences difficult to prevent. For example, once individual plants and animals are exposed to a stressor, their responses to the exposure is controlled by the direct and indirect consequences asso ciated with the exposure, their sensitivity to the stressor, other physical, chemical, or bi otic stressors that they are exposed to concurrently, their pre-existi ng physiological state, and their individu al fitness.

When individual listed plants or animals experi ence reductions in fitness, those reductions may reduce the abundance, reproduction rates, or growth rates (or increase varian ce in one or more of those rates) of the populations t hose individuals represent (see St earns 1992). Reductions in one or more of these variables (or one of the variables we derive from them) is a necessary condition for reductions in a population’s viability, which is itself a necessary condition for reductions in a species viability. On the other hand, when proposed or listed plants or animals exposed to an

action’s effects are not expected to experience reductions in fitness, we would not expect the action to have adverse consequences on the vi ability of the populations those individuals represent or the species t hose populations comprise (e.g., Brandon 1978, Mills and Beatty 1979, Stearns 1992, Anderson 2000). If we conclude that proposed or listed species are not likely to experience reductions in their fitness, we would conclude our assessment.

If, however, we conclude that proposed or listed species are likely to experience reductions in their fitness, we examine whether the program in cluded sufficient safeguards to ensure that the actions they authorize, fund, or otherwise car ry out would not reduce the viability of the populations those individuals repr esent (typically measured usi ng changes in the populations’ abundance, reproduction, spatial st ructure and connectivity, growth rates, or variance in these measures to make inferences about the population’s extinction risks). For those species likely to be adversely affected by the activities conducted under a program, we examine their status and the environment in which the species exists (in this Opinion, the Environmental Baseline and Status of the Species are examined in the section titled Action Area ), in detail, as a point of reference for determining if changes in popula tion viability are likely, and if, in turn, any changes in population viability is sufficient to reduce the viability of the species.

In addition to analyzing the ability of the USCG ’s ballast water management program to insure the action is not likely to jeopardize listed specie s or destroy or adversel y modify their critical habitat, we assess the probable consequences of the specific actions that will be authorized, funded or carried out by the progr am for endangered and threatened species and critical habitat that has been designated for those species. Specifically:

1. We examine activities that would be authorized by the proposed action.

These analyses summarize (a) establishing a numer ic al ballast water stan dard; (b) the geographic distribution of species introducti ons via ballast water; (c) the frequency of non-native species introductions originating from similar geographi c areas; (d) the stressors typically associated14 with non-native species, and; (e) any additional information on the actions to be authorized.

2. We determine the degree of spatial overlap between ballast water disc harges, listed species and designated critical habitat.

These analyses describe spatial overlap and a ny specific evidence (reports or studies) that particular endangered or threatened species or designated critical ha bitat have been or are likely to be exposed to the chemical, physical, or biolog ical stressors resulting from those activities. However, this does not represent a detailed e xposure analysis: we are merely establishing whether or to what degree endangered or threat ened species or designa ted critical habitat overlap, in space and time (e.g., some non-native populations may establish seasonally but perish during other seasons only to be re-established the following year and their residency may occur while migratory species are not in an area). Gi ven spatial and temporal overlap, we then have reason to ask whether or to what degree USCG's pr ogram can insure that these species or critical habitat are not likely to be exposed.

3. We conduct a detailed review of the liter ature available on the physical, physiological, behavioral, social, and ecologica l responses of endangered or th reatened species or primary

constituent elements of critical habitat given exposure to potential stressors, or to the ecological effects (that is, effects resulting from changes in populations of prey, predators, competitors, symbionts, etc.) of those stressors. Rather th an discuss the literature for each species, we organized the data using species gr oups (for example, Pacific salmon; sturgeon; sea turtles; etc.).

4. We summarize the probable consequences of the responses identified in the preceding section for populations of endangered and threaten ed species and designated critical habitat.

Rather than discuss the literature for each species, it is only n ecessary to discuss the risks of exposing species groups (for example, Pacifi c salmon; sturgeon; s ea turtles; etc.).

In this Opinion, we will present th e results of these analyses before we present the results of our review and evaluate the program using the sequence of seven questions we id entified previously. We use the results of these combined analyses to determine whether and to what degree USCG structured its program in ways that would protect endangered or threatened species or critical

habitat that has been de signated for those species.

Evidence Available for this Consultation

To conduct our analyses we considered the in formation contained in the USCG’s letter requesting initiation of consu ltation, the final PEIS, the FR , 33 CFR 151, NANPCA, NISA, and monitoring data from previous ballast water rules. We suppl emented this information using electronic searches of literature published in E nglish or with English abstracts using research platforms in the Online Computer Library Center’s First Search, CSA Illumina, and ISI Web of

Science. These platforms allow us to cross-search multiple databases for journals, open access resources, books, proceedings, Web s ites, doctoral dissertations and master’s theses for literature on the biological, ecological, and medical sciences. Particular databases we searched for this

consultation included BasicBiosis

, Disserta tions

, ArticleFirst

, Proceedings, Aquatic Sciences and Fisheries Abstracts, BioOne Abstract s & Indexes, and Water Resources Abstracts . Some of15 the databases provide access to documents pub lished from the 1960s through present, although references for many scientific journals containe d in these databases only date back to the 1970s or later. Through these databases we accessed the major journals dealing with the biology, ecology, distribution, status, and tr ends of the threatened and e ndangered species considered in this Opinion, and the impacts of invasive species on freshwater, es tuarine, and marine
ecosystems.

For our literature searches, we used paired combinations of the keywords: ballast water, discharge, invasive species, habitat, non-indigenous, and many others. We acquired references that, based on a reading of their ti tles and abstracts, appeared to comply with our keywords. If a reference’s title appeared releva nt, we acquired and reviewed the reference. We supplemented our electronic searches by searching the literature cited sections and biblio graphies of references we retrieved electronically to identify additiona l papers that had not been captured in our

electronic searches.

Collectively, this information provided the basis fo r our determination as to whether and to what degree listed resources under our jurisdiction are likely to be e ffected by the USCG’s ballast water management program, and whet her and to what degree the USCG can insure that its ballast water management program is not likely to je opardize the continued existence of proposed, endangered, or threatened species or result in the destruction or adverse modification of proposed or designated cri tical habitat.

USCG Decision Structure

Under the ballast water management program, one of the early and most important actions taken by the USCG is the review of available technolog ies to understand the be st available treatment technologies at the time of the proposal. Based on the feasibility of the technologies, the ballast

water standard is estab lished. The USCG also conducts annu al monitoring and reporting as well as a three-year periodic review. All of those reviews are intended to provide for flexibility and adaptive management within the USCG ballast water management program.

Figure 1 depicts a simplified model of the USCG ballast water management program, as NMFS understands it. Establishing a nati onal ballast water standard, an action that merits consultation, represents only one part of the overall program and decision making process in regulating ballast water discharges. During this consultation, we evaluated the proposed ballast water standard and the USCG’s decision-making process for when and how to establish a more stringent standard if appropriate. We reviewed the monitoring plan an d practicability review that the USCG analyzes

and considers to reach conclusions as to whethe r listed species would be affected by numerical

ballast water standards. We ev aluated this information to dete rmine whether and to what degree the USCG’s decision-making process insures that any activities it authorizes, funds, or carries out are not likely to, individua lly or cumulatively, jeopardi ze the continued existence of proposed or listed species or re sult in the destruction or adve rse modification of proposed or designated critical habitat.16 Figure 1. Diagram of USCG Ballast Water Management Program decision making process.**E. t~bl** **i. h** **B~ ~.t**  Wite r Pr o,rom l

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..... , .. ..17 Application of this Approach in this Consultation

In this consultation, we evaluated the USCG’s proposed ballast water management program, including the numerical standards, and whether the USCG has insu red that any action authorized, funded, or carried out under this ballast water management program is not likely to jeopardize the continued existence of any proposed, endangered, or threaten ed species, or result in the

destruction or adverse modificatio n of proposed or designated cri tical habitat. We began our analysis of the ballast water management program by exploring the USCG’s preferred alternative in relation to the ballast water management pr ogram. Specifically, we reviewed the rationale used to identify the preferred alternative, including all substantive decision criteria that

influenced the USCG’s decision to establish that specific ballast water standard. After evaluating the preferred altern ative, we assessed the ballas t water program’s monitoring programs: three-year periodic reviews, compliance monitoring, effectiveness monitoring, and environmental monitoring. The results of thos e monitoring actions will feed back into the decision making process for future pot ential numerical standards.

In addition, we asked: how does the USCG de termine a numerical standard, how frequently does the USCG conduct program reviews, what is the criteria that triggers an adaptive management adjustment to ballast water sta ndards, which ships are not bound by the USCG’s ballast water regulations, is the USCG able to influence improved ballast water standards through incentive programs or research funding, is the USCG able to determine when a ship has installed an acceptable technology, is the USCG able to determin e the technology still meets the

standard after it has been insta lled, is the USCG able to determ ine if the technology is able to meet the standard after having been used for y ears, is the USCG able to determine whether implementation of the numerical standard has e ffectively reduced or pr evented new invasive species, and what is the ability of the US CG to monitor newly introduced species?

Through the course of this consultation we lear ned that the decision to identify the proposed ballast water standard was dependent on the “max imum extent practicable ” mandate of NISA. This decision is informed by vessel and human safety, water quality considerations, NANCPA, NISA, economic impacts to commerce and societ y, and the proven ability of ballast water treatment techniques to achieve the proposed ballast water standards.

This consultation will evaluate the ballast water management program by describing the likelihood of listed species or their critical habitat being exposed to non-native species, introduced through the proposed ballast water sta ndard. To best analyze the likelihood of exposure, we will conduct a primary analysis of several large ports in the US with species compositions that encompass all listed functional groups. As a secondary analysis of exposure, we will conduct specific analyses of ports with listed species not identified in the primary analysis. To accomplish this and to understand th e responses of listed species, we will consider both listed species and introduced species as f unctional groups, such as salmonids, pinnipeds, and sea turtles or diseases, predators, and comp etitors, respectively. Once we have established the likely rate of exposure to listed species, us ing the response analysis, we will determine the proportion of invasive species th at would likely be harmful. Using the information from the exposure and response analyses, we can determin e the risks ballast wa ter discharge at the proposed standards may pose to listed species an d their critical habitat in order to reach a18 conclusion about the USCG’s program and its abili ty to insure the USCG is not likely to jeopardize any listed species or destroy/a dversely modify their critical habitat.

During this consultation, NMFS considered th e various ballast water discharge standards proposed by different states a nd authorized, funded, or carried out by other Federal agencies under different legal mandates. NMFS did not incl ude those other actions in the analysis, but instead evaluated: whether the approved state st andards are achievable, the outcomes of legal action against those enacted stat e standards, and the approval process for EPA vessel discharge permit program in order to understand how th e USCG program established the numerical

standards and whether there are other state or Federal laws that influence the ballast water
management program.

There are other consultations that may necessarily depend on the outcome of this consultation

and be tiered from this consultation in the future (see Figure 1). If th e USCG determines there may be an effect as a result of new technol ogies being developed to achieve the proposed numerical standard, NMFS will evaluate the eff ects of those technologies on proposed and listed resources. However, future consultations with EPA on its ve ssel discharge program or any

consultation on state standards with a federal nexus requiring section 7 consultation will be separate actions not considered under USCG’s ballast water mana gement program. Even if the ballast water program complies with the requirem ents of section 7(a)(2) of the Endangered Species Act (that is, the requirement to insure th at their actions are not likely to jeopardize the continued existence of endangered or threatened species or result in the destruction or adverse modification of critical habitat that has been designated for those species), the reinitiation criteria at the end of this Opinion could requ ire further analysis of this program.

We examined the USCG ballast water management program to see if it contains features that minimize the exposure of proposed, endangered, or threatened species, or their proposed or designated critical habitat to harmful non-native species. When exposure is not avoided, we broadly characterized the uptake and discharge of balla st water at US ports with specific emphasis on San Francisco Bay, Puget Sound, and Ches apeake Bay to describe the risk of listed resources being exposed to uptake or discharge of ballast water at all US ports. If, based on this information, we expect that listed resources are not likely to be exposed to non-native introduced species, then we conclude that the action would have “no effect” on those listed resources. If, based on this information, we determined that li sted individuals may be exposed to uptake of ballast water or non-native introduced species, but (a ) the probability of expos ure is so small that it would not be reasonable to expect exposure to occur, (b) there is no possibility or only a very small possibility that the indivi dual listed species would respo nd to exposure, (c) there is no possibility or only a small probabil ity of a negative response even if there is a response to uptake

in ballast water tanks or to intr oduced non-native species, or (d) th ere is no possibility or only a small probability that the individual would expe rience a reduction in i ndividual performance (or fitness), then we concluded that the USCG’s action is “not like ly to adversely affect” those proposed or listed resources.

If proposed or listed resources or th eir critical habitats may be leth ally or sub-lethally affected by actions the USCG’s ballast water management pr ogram authorizes, funds, or otherwise carries out, NMFS examines if the program includes suffici ent safeguards to insure that the incidental19 take of individuals does not o ccur in a manner that reduces the viability of the populations those individuals represent (typically measured using changes in the populations’ abundance, reproduction, spatial structure and connectivity, growth rates, or variance in these measures to make inferences about the populatio n’s extinction risks). Given th eir status and the environment in which the species exist, are those species lik ely to be adversely a ffected by the activities conducted under the proposed action also likely to suffer changes in population viability that would be sufficient to reduce the viability of the species those populations comprise?

ACTION AREA

The section 7 implementing regulations define th e “Action Area” of a Fede ral action as all areas to be affected, directly or i ndirectly, and not merely the imme diate area involved in the action (50 CFR 402.02). This Opinion assesses the cons equences of the USCG's establishment of

ballast water standards in all wa ters of the US. The USCG's pr oposed rule establishes the action area as all navigable waters within the US ex tending offshore to the outer edge of the EEZ, 200 miles offshore of the US coastline except wher e this region overlaps with the EEZ of another nation (Mexico, Canada, Russia, Cuba, the Federati on of Micronesian States, and The Bahamas). Furthermore, because vessels entering the Great Lakes are still required to conduct ballast water exchange beyond the EEZ, ballast water intake from areas beyond the EEZ along the East Coast of America are also part of this action area.

Status of Proposed and Listed Resources

The USCG ballast water management program may affect the threatened and endangered species

and critical habitat listed in Table 1. We excluded foreign species because the USCG has no jurisdiction to regulate ballast water in foreign countries.

Table 1. Proposed and listed resources in the action area. As terisks denote critical habitat in the action area. Double asterisks denote proposed critical habitat in the action area.

Common name (Distinct popula tion segment, evolutionarily significant unit, or subspecies) Scientific name Status

Cetaceans

Blue whale Balaenoptera musculus Endangered

Bowhead whale Balaena mysticetes Endangered

Fin whale Balaenoptera physalus Endangered Humpback whale Megaptera novaeangliae Endangered Killer whale (Southern Resident*) Orcinus orca Endangered North Atlantic right whale* Eubalaena glacialis Endangered North Pacific right whale* Eubalaena japonica Endangered

Sei whale Balaenoptera borealis Endangered Sperm whale Physeter macrocephalus Endangered Beluga whale (Cook Inlet)** Delphinapterus leucas Endangered

Pinnipeds

Guadalupe fur seal Arctocephalus townsendi Threatened Hawaiian monk seal*, ** Monachus schauinslandi Endangered Steller sea lion (Eastern*) Eumetopias jubatus Threatened Steller sea lion (Western*) Endangered

Marine Turtles20 Common name (Distinct popula tion segment, evolutionarily significant unit, or subspecies) Scientific name Status

Green sea turtle (Florida & Mexico’s Pacific coast colonies) Chelonia mydas Endangered Green sea turtle (All other areas*) Threatened

Hawksbill sea turtle* Eretmochelys imbricate Endangered Kemp’s ridley sea turtle Lepidochelys kempii Endangered Leatherback sea turtle*, ** Dermochelys coriacea Endangered Loggerhead sea turtle (northwestern Atlantic) Caretta caretta Threatened Loggerhead sea turtle (North Pacific) Endangered Loggerhead sea turtle (South Pacific) Threatened Olive ridley sea turtle (Mexico’s Pacific coast breeding colonies) Lepidochelys olivacea Endangered Olive ridley sea turtle (All other areas) Threatened

Anadromous Fishes

Atlantic salmon (Gulf of Maine*) Salmo salar Endangered Chinook salmon (California Coastal*) Oncorhynchus tschawytscha Threatened Chinook salmon (Central Valley Spring-run*) Threatened Chinook salmon (Lower Columbia River*) Threatened Chinook salmon (Upper Columbia River Spring-run*) Endangered Chinook salmon (Puget Sound*) Threatened Chinook salmon (Sacramento River Winter-run*) Endangered Chinook salmon (Snake River Fall-run*) Threatened Chinook salmon (Snake River Spring/Summer-run*) Threatened Chinook salmon (Upper Willamette River*) Threatened Chum salmon (Columbia River*) Oncorhynchus keta Threatened Chum salmon (Hood Canal Summer-run*) Threatened Coho salmon (Central California Coast*) Oncorhynchus kisutch Endangered Coho salmon (Lower Columbia River) Threatened Coho salmon (Southern Oregon & Northern California Coast*) Threatened Coho salmon (Oregon Coast*) Green sturgeon (Southern*) Acipenser medirostris Threatened Shortnose sturgeon Acipenser brevirostrum Endangered

Smalltooth sawfish* Pristis pectinata Endangered Bocaccio (Georgia Basin) Sebastes paucispinis Endangered Yelloweye rockfish (Georgia Basin) Sebastes pinniger Threatened Canary rockfish (Georgia Basin) Sebastes ruberrimus Threatened

Pacific eulachon** Tha;eichthys pacificus Threatened Sockeye salmon (Ozette Lake*) Oncorhynchus nerka Threatened Sockeye salmon (Snake River*) Endangered Steelhead (Central California Coast*) Oncorhynchus mykiss Threatened Steelhead (California Central Valley*) Threatened Steelhead (Lower Columbia River*) Threatened Steelhead (Middle Columbia River*) Threatened Steelhead (North er n California*) Threatened Steelhead (Puget Sound) Threatened Steelhead (Snake River*) Threatened Steelhead (South-Central California Coast*) Threatened Steelhead (Southern California*) Threatened Steelhead (Upper Columbia River*) Threatened Steelhead (Upper Willamette River*) Threatened Atlantic sturgeon (Gulf of Maine) Asipenser oxyrincus Threatened Atlantic sturgeon (New York Bight) Endangered Atlantic sturgeon (Chesapeake Bay) Endangered Atlantic sturgeon (Carolina) Endangered Atlantic sturgeon (South Atlantic) Endangered21 Common name (Distinct popula tion segment, evolutionarily significant unit, or subspecies) Scientific name Status

Marine Invertebrates

Elkhorn coral* Acropora palmata Threatened Staghorn coral* Acropora cervicornis Threatened

White abalone Haliotis sorenseni Endangered

Black abalone* Haliotis cracherodii Endangered

Marine Plants Johnson’s seagrass* Halophilia johnsonii Threatened Proposed for listing False killer whale (Hawaiian Insular)** Pseudorca crassidens Proposed Endangered Bearded seal (Beringia) Erignathus barbatus

nauticus Proposed Threatened

Ringed seal (Arctic) Phoca hispida hispida Proposed Threatened

Species not considered further We do not expect exposure of several listed or proposed species or critica l habitats that occur exclusively (or nearly so) in offshore or deepwa ter marine environments. Based upon review of

available literature de tailed further in our Response analysis , we discount the possibility that invasive species established from ballast water discharged in US waters (primarily US ports, which occupy nearshore or inland waters) would spr ead to offshore habitats. Inshore, estuarine, and freshwater habitats are fundamentally diffe rent in ecological terms and we discount the ability of an invasive species to establish in o ffshore regions as well as one of these areas where the large majority of ballast water discharge would occur. Based upon this, we discount the exposure of blue, fin, sei, sperm, North Pacific right, and Hawaiian insular false killer whales. For the same reasons, we also discount exposure to North Pacific right whal e critical habitat.

Several listed or proposed resource s also occur exclusively (excep t for extralimital occurrences) in areas where ballast water discharge is so lim ited that the probability of establishment of nonnative species and subsequent expos ure of listed resource is disc ountable. For some species, the isolated and/or offshore nature of the species was also a consideration. These include bowhead whales, Beringia DPS bearded seals, and Arctic D PS ringed seals.

The biology and ecology of species with anticipated exposure below informs the effects analysis for this Opinion. Summaries of the global status and trends of each species presented provide a foundation for the analysis of species as a whole.

Cetaceans

Humpback whale Description of the species

. Humpback whales are a cosmopolit an species that occur in the Atlantic, Indian, Pacific, and Southern oceans. Humpback whales migrate seasonally between warmer, tropical or sub-tropical waters in winter months (where they br eed and give birth to calves, although feeding occasionally occurs) and cooler, temperate or sub-Arctic waters in summer months (where they feed; (Gendron and Urban 1993). In both regions, humpback whales tend to occupy shallow, coastal waters. However, migrations are undertaken through deep, pelagic waters (Winn and Reichley 1985).22 Population designations. Populations have been relatively well defined for humpback whales.

North Atlantic. Humpback whales range from the mid- Atlantic bight and the Gulf of Maine

across the southern coast of Greenland and Icel and to Norway in the Barents Sea. Whales migrate to the western coast of Africa and th e Caribbean Sea during the winter. Humpback whales aggregate in four summer feeding area s: Gulf of Maine and eastern Canada, west Greenland, Iceland, and Norway (Katona and Beard 1990, Smith et al. 1999b, Boye et al. 2010). Increasing range and occurrence in the Mediterra nean Sea coincides with population growth and may represent reclaimed habitat from pre-commercial whaling (Frantzis et al. 2004, Genov et al. 2009). The principal breeding range for Atlantic humpback whales lies from the Antilles and northern Venezuela to Cuba (Winn et al. 1975, Balcomb III and Nichols 1982, Whitehead and Moore 1982). The largest breedi ng aggregations occur off the Gr eater Antilles where humpback whales from all North Atlantic feeding areas have been photo-identified (Katona and Beard 1990, Clapham et al. 1993, Mattila et al. 1994, Palsbøll et al. 1997, Smith et al. 1999b, Stevick

et al. 2003b). However, the possibility of historic and present breeding further north remains enigmatic but plausible (Smith and G.Pike 2009). Winter aggregations also occur at the Cape

Verde Islands in the eastern Nort h Atlantic and along Angola (Reiner et al. 1996, Reeves et al. 2002, Weir 2007, Cerchio et al. 2010). Accessory and historical aggregations also occur in the eastern Caribbean. To further highlight the “ope n” structure of humpba ck whales, a humpback whale migrated from the Indian Ocean to th e South Atlantic Ocea n, demonstrating that

interoceanic movements can occu r (Pomilla and Rosenbaum 2005). Genetic exchange at lowlatitude breeding groups between Northern and Southern Hemi sphere individuals and widerrange movements by males has been suggested to e xplain observed global gene flow (Rizzo and Schulte 2009). However, there is little genetic support for wide-scale interchange of individuals between ocean basins or across the equator.

North Pacific. Based on genetic and photo-id entification studies, the NMFS currently recognizes four stocks, likely corresponding to populati ons, of humpback whales in the North Pacific Ocean: two in the eastern North Pacific, one in the central North Pacific, and one in the western Pacific (Hill and DeMaster 1998). Gene flow between them may exist. Humpback

whales summer in coastal and inland waters from Point Conception, California, north to the Gulf of Alaska and the Bering Sea, and west along the Aleutian Islands to the Kamchatka Peninsula and into the Sea of Okhotsk (Nemoto 1957, To milin 1967, Johnson and Wolman 1984). These whales migrate to Hawaii, southern Japan, the Mariana Islands, and Mexico during winter. However, more northerly penetrations in Arctic waters occur on occasion (Hashagen et al. 2009). The central North Pacific populatio n winters in the waters around Hawaii while the eastern North Pacific population (also called th e California-Oregon-Washington- Mexico stock) winters along Central America and Mexico. However, Calambokidis et al. (1997) identified individuals from several populations wintering (and potentially breeding) in the area s of other populations, highlighting the potential fluidity of population structure. Humpb ack whales were recently found to migrate to the northwestern Hawaiian Island s, where singing has been recorded; this may represent an as yet undescribed breeding group, or expansion of breeding from the main Hawaiian Islands (Lammers et al. 2011). Herman (1979) presented extensive evidence that humpback whales associated with the main Hawa iian Islands immigrated there only in the past 200 years. Winn and Reichley ( 1985) identified genetic exchange between the humpback whales that winter off Hawaii and Mexico (with fu rther mixing on feeding areas in Alaska) and23 suggested that humpback whales that winter in Hawaii may have emigrated from Mexican wintering areas. A “population” of humpback whales winters in the South China Sea east through the Philippines, Ryukyu Retto, Ogasaw ara Gunto, Mariana Islands, and Marshall Islands, with occurrence in the Mariana Islands, Guam, Rota, and Saipan from January-March (Eldredge 1991, Darling and Mori 1993, Rice 19 98a, Eldredge 2003). During summer, whales from this population migrate to the Kuril Is lands, Bering Sea, Aleutian Islands, Kodiak, Southeast Alaska, and British Columbia to feed.

Humpback whales from both the eastern, west ern, and central North Pacific are known to congregate to feed in waters off Kodiak and the Shumagin Islands (Waite et al. 1999, Urban et
al. 2000, Calambokidis et al. 2001a, Witteveen et al. 2004, Calambokidis et al. 2009). The species is found in the Gulf of Alaska y ear-round, but are most abundant during summer (beginning in April and peaking in late-August to early September) when foraging opportunities draw many individuals into the region (Consiglieri et al. 1982, Baker et al. 1985, Straley 1990,

Waite et al. 1999, Stafford et al. 2007, Dahlheim et al. 2008). During this time, feeding aggregations are found throughout the Kodiak Archipelago, although pelagic areas may also be important foraging areas (MMC 2002, Baraff et al. 2005). Humpback whales have been found in particularly high numbers around Kodiak Isla nd in recent surveys, with a density of 54 individuals per 1,000 km 2 (Waite 2003, Zerbini et al. 2006). Sightings were most frequent during fall off Kodiak (Wynne and Witteveen 2005).

Arabian Sea. A separate population of humpback whales appears to reside in the Arabian Sea in the Indian Ocean off the coasts of Oman, Pa kistan, and India and movements of this group are poorly known (Mikhalev 1997, Rasmussen et al. 2007). Areas of the Mozambique Channel appear to be significant cal ving and wintering areas for humpback whales (Kiszka et al. 2010b).

Southern Hemisphere. Eight proposed stocks, or populations , of humpback whales occur in waters off Antarctica (Figure 2). Individuals fr om these stocks winter and breed in separate

areas and are known to return to the same areas. However, the degree (if a ny) of gene flow (i.e., adult individuals wintering in different breeding locations) is uncertain (Carvalho et al. 2011). Individuals from breeding grounds in Ecuador ar e somewhat heterogenous from individuals in other breeding areas, but appear to maintain a genetioc linkage (Felix et al. 2009). Based upon recent satellite telemetry, a revision of stocks A and G may be warranted to reflect stock movements within and between feeding areas separated east of 50o W (Dalla Rosa et al. 2008). In addition to being a breeding area, the west co ast of South Africa also appears to serve as a foraging ground due to upwelling of the Benguela Current (Barendse et al. 2010). Females appear in this area in large numbers well before their male counterparts, frequently accompanied by calves (Barendse et al. 2010). Female movement between breeding locations across years has been documented, bringing into question the ge netic discreteness of at least Southern Hemisphere populations (Stevick et al. 2011). However, mixing between some populations has not been found (such as between B2 and C1 gr oups). Sao Tome appears to be primarily a resting, nursing, and calving area with very little breeding occurring (Carvalho et al. 2011).24 Figure 2. Southern Hemisphere humpback stocks (populations)(IWC 2005).

Reproduction. Humpback whale calving and breeding gene rally occurs during winter at lower latitudes. Gestation takes about 11 months, follo wed by a nursing period of up to 1 year (Baraff and Weinrich 1993). Sexual maturity is reached at between 5-7 years of age in the western North Atlantic, but may take as long as 11 years in the North Pacific, and perhaps over 11 years (e.g., southeast Alaska, Gabriele et al. 2007). Females usually breed every 2-3 years, although consecutive calving is not unheard of (Glockne r-Ferrari and Ferrari 1985 as cited in NMFS 2005b, Clapham and Mayo 1987, 1990, Weinrich et al. 1993). Males appe ar to return to breeding grounds more frequently than do females (Herman et al. 2011). Larger females tend to produce larger calves that may have a greater chance of survival (Pack et al. 2009). In some Atlantic areas, females tend to prefer shallow nearshore waters for calving and rearing, even when these areas are extensively trafficked by humans (Picanco et al. 2009).

In calving areas, males sing long complex songs directed towards females, other males, or both. The breeding season can best be described as a floating lek or ma le dominance polygamy (Clapham 1996). Calving occurs in the shallo w coastal waters of c ontinental shelves and oceanic islands worldwide (Perry et al. 1999). Males “cort” fema les in escort groups and compete for proximity and presumably access to reproduce females (particularly larger females)(Pack et al. 2009). Although long-term relationships do not a ppear to exist between males and females, mature females do pair with ot her females; those indivi duals with the longest standing relationships also have the highest reproductive output, possibly as a result of improved feeding cooperation (Ramp et al. 2010).

Feeding. During the feeding season, humpback whales form small groups that occasionally aggregate on concentrations of food that may be stable for long-periods of times. Humpbacks use a wide variety of behaviors to feed on vari ous small, schooling prey including krill and fish (Jurasz and Jurasz 1979, Hain et al. 1982, Weinrich et al. 1992, Hain et al. 1995, Witteveen et al. 2011). The principal fish prey in the western North At lantic are sand lance, herring, and capelin (Kenney et al. 1985b). There is good evidence of some territoriality on feeding and calving areas (Tyack 1981, Clapham 1994, Cla pham 1996). Humpback whales are generally believed to fast while migrating and on breed ing grounds, but some individuals apparently feed while in low-latitude waters normally believed to be us ed exclusively for reproduction and calf-rearing25 (Danilewicz et al. 2009, Pinto De Sa Alves et al. 2009). Some individuals, such as juveniles, may not undertake migrations at all (Findlay and Best. 1995). Additional evidence, such as songs sung in northern latitudes during winter, provide additional suppor t to plastic seasonal distribution (Smith and G.Pike 2009). Relatively high rates of resighting in foraging sites in Greenland suggest whales return to the same areas year after year (Kragh Boye et al. 2010).

Average group size near Kodiak Isla nd is two to four individuals, although larger gr oups are seen near Shuyak and Sitkalidak islands and groups of 20 or more have been documented (Wynne et
al. 2005).

Status and trends. Humpback whales were originally li sted as endangered in 1970 (35 FR 18319), and this status remains under the ESA. (W inn and Reichley 1985) argued that the global humpback whale population consisted of at l east 150,000 whales in the early 1900s, mostly in the Southern Ocean. In 1987, the global population of humpback wh ales was estimated at about 10,000 (NMFS 1987). Although this estimate is out dated, it appears that humpback whale numbers are increasing. Table 2 provides estimates of historic and current abundance for ocean regions.

North Atlantic. The best available estimate of Nort h Atlantic abundance comes from 1992- 1993 mark-recapture data, which generated an es timate of 11,570 humpback whales (Stevick et
al. 2003a). Historical estimates have ra nged from 40,000-250,000 (Smith and G.Pike 2009). Smith and Reeves (2010) estimated that rough ly 31,000 individuals were removed from the North Atlantic due to whaling since the 1600s. Estimates of animals on Caribbean breeding grounds exceed 2,000 individuals (Balcomb III and Ni chols 1982). Several researchers report an increasing trend in abundance fo r the North Atlantic population, wh ich is supported by increased sightings within the Gulf of Maine feedi ng aggregation (Katona and Beard 1990, Barlow 1997,

Smith et al. 1999b, Waring et al. 2001). The rate of increase varies from 3.2-9.4%, with rates of increase slowing over the past two decades (Katona and Beard 1990, Barlow 1997, Stevick et al. 2003a). If the North Atlantic population has grown according to the estimated instantaneous rate of increase ( r = 0.0311), this would lead to an esti mated 18,400 individual whales in 2008 (Stevick et al. 2003a). Punt (2010) estimated the rate of increase for humpback whales in the Gulf of Maine to be 6.3% annually (1.2 SE). Pike et al. (2009) suggested th at the eastern and

northeastern waters off Iceland are areas of significant humpback utilization for feeding, estimating nearly 5,000 whales in 20 01 and proposing an annual growth rate of 12% for the area. The authors suggest that humpb ack whales in the area had proba bly recovered from whaling. However, recent data suggest that the upward gr owth may have slowed or ceased around Iceland according to analysis of survey data there (Pike et al. 2010).

North Pacific. The pre-exploitation population size may have been as many as 15,000 humpback whales, and current estimate s are 6,000-8,000 whales (R ice 1978, Calambokidis et al. 1997). It is estimated that 15,000 humpback whales resided in th e North Pacific in 1905 (Rice 1978). However, from 1905 to 1965, nearly 28,000 hu mpback whales were harvested in whaling operations, reducing the number of all Nort h Pacific humpback whale to roughly 1,000 (Perry et
al. 1999). Estimates have risen over time fr om 1,407-2,100 in the 1980s to 6,010 in 1997 (Baker 1985, Darling and Morowitz 1986, Bake r and Herman 1987, Calambokidis et al. 1997). Because estimates vary by methodology, they are not direc tly comparable and it is not clear which of26 these estimates is more accurate or if the change from 1,407 to 6,010 is the result of a real

increase or an artifact of m odel assumptions. Tentative estima tes of the eastern North Pacific stock suggest an increase of 6-7% annually, but fluctuations have included negative growth in the recent past (Angliss and Outlaw 2005). Ho wever, based upon surveys between 2004 and 2006, Calambokidis et al. (2008) estimated that the number of humpback whales in the North Pacific consisted of about 18,300 whal es, not counting calves (Barlow et al. (2009) provided a bias-corrected estimate of 20,800 individuals) a nd the population was growing at 4.9% annually. Almost half of these whales likely occur in wintering areas around th e Hawaiian Islands. Punt (2010) estimated the rate of increase for humpback whales in the eastern North Pacific to be 6.4% annually (0.9 SE) between 1992 and 2003 and 10.0% for Hawaii (3.32 SE).

Table 2. Summary of past and pres ent humpback wh ale abundance.

*Note: Confidence Intervals (C.I.) not provided by the authors were calculated from Coefficients of Variation (C.V.) where available, using the computa tion from Gotelli and Ellison (2004).

Arabian Sea. The population inhabiting the Arabian Sea likely numbers a few hundred individuals at most (Minton et al. 2008). This population likel y was much larger prior to exploitation in 1966 by Soviet whaling, with individuals found along not only Oman, but Yemen, Iran, Pakistan, and India (Slijper et al. 1964, Wray and Martin. 1983, Reeves et al. 1991, Mikhalev 2000, Minton et al. 2008).

Southern Hemisphere. The IWC recently compiled population data on humpback whales in the

Region Population, stock, o r study area Pre-exploitation

estimate 95% C.I. Current
estimate 95% C.I. Source

Global

-- 1,000,000

-- -- -- (Roman and Palumbi 2003)

North Atlantic

Basinwide 240,000 156,000-
401,000* 11,570 10,005-
13,135* (Roman and Palumbi 2003) (Stevick et al. 2001) in (Waring et al. 2004b)

Basinwide - Females -- -- 2,804 1,776-4,463 (Palsbøll et al. 1997)

Basinwide - Males -- -- 4,894 3,374-7,123 (Palsbøll et al. 1997)

Western North Atlantic from
Davis Strait, Iceland to the

West Indies >4,685*

-- -- -- *circa 1865; (Mitchell and Reeves 1983)

NMFS - Gulf of Maine stock

-- -- 845 CV=0.55 (NMFS 2008b)

NMFS - Gulf of Maine stock, including a portion o f

Scotian Shel f -- -- 902 177-1,627* (Clapham et al. 2003)

Northeast Atlantic - Barents and Norwegian Seas -- -- 889 331-1,447* (Øien 2001) in (Waring et
al. 2004b)

North Pacific Basinwide 15,000 -- 6,000-8,000 -- (Calambokidis et al. 1997)

NMFS - Western North

Pacific stock -- -- 394 329-459* (Angliss and Allen 2007)

NMFS - Central North

Pacific stock -- -- 4,005 3,259-4,751* (Angliss and Allen 2007)

NMFS - Eastern North

Pacific stock -- -- 1,391 1,331-1,451* (Carretta et al. 2008)

Indian

Ocean A rabian Sea

-- -- 56 35-255 Minton et al. (Minton et al. 2003) in (Bannister 2005)

Southern Hemisphere Basinwide 100,000 -- 19,851

-- (Gambell 1976, IWC 1996)

South of 60 o S

-- -- 4,660 2,897-6,423 (IWC 1996)27 Southern Hemisphere. Approximately 42,000 S outhern Hemisphere humpbacks can be found south of 60° S during the austral summer f eeding season (IWC 2007). However, humpback whales in this region experienced severe whaling pressure. Based upon whaling logs, particularly by Soviet vessels, at least 75,542 hum pback whales were harvested from Antarctic waters from 1946 through 1973, largely from ma nagement areas IV, V, and VI (Clapham et al. 2009). One-third of these catches occurred fr om 1959-1961 in Area V. These numbers support Southern Hemisphere humpbacks being well below their carrying capacities (Clapham et al. 2009). Recent surveys off the Brazilian br eeding grounds suggests a populations of 6,404 individuals in this area (Andriolo et al. 2010). Modelling efforts to bound the number of

individuals within Oceania have estimated 2,300-3,500 indivi duals divided amongst various populations/subpopulations (Constantine et al. 2010). A 2009 spike in calf mortality along western Australia brings into question whethe r carrying capacity has been reached by this population or other factors have increased mortality (Coughran and Gales 2010). Some vital rates of the humpback whale population summeri ng off eastern Australia (E1) were recently estimated, including adult annual survivial of 0.925, subadult survival of 0.70 (Hoffman et al. 2010). Growth rates for certain age classes in cluded 10.7% for adult females and 12.4% for juveniles (Hoffman et al. 2010). Punt (2010) estimated the rate of increase for humpback whales

off eastern and western Austra lia to be 10.9 and 10.1% annuall y, respectively (0.23 and 4.69 SE, respectively).

Natural threats. Natural sources and rates of mortalit y of humpback whales are not well known. Based upon prevalence of tooth marks, atta cks by killer whales appear to be highest among humpback whales migrating between Mexico and California, although populations throughout the Pacific Ocean appear to be targeted to some degree (Steigeret al. 2008). Juveniles appear to be the primary age group ta rgeted. Humpback whales engage in grouping behavior, flailing tails, and rolli ng extensively to fight off attacks. Calves remain protected near mothers or within a group and lone calves have been known to be protected by presumably

unrelated adults when confronted wi th attack (Ford and Reeves 2008).

Parasites and biotoxins from red-tide blooms ar e other potential causes of mortality (Perry et al. 1999). The occurrence of the nematode Crassicauda boopis appears to increase the potential for kidney failure in humpback whales and may be preventing some populations from recovering (Lambertsen 1992). Studies of 14 humpback wh ales that stranded al ong Cape Cod between November 1987 and January 1988 indicate they a pparently died from a toxin produced by dinoflagellates during this pe riod. One-quarter of humpback whales of the Arabian Sea population show signs of tattoo skin disease, which may reduce the fitness of afflicted individuals (Baldwin et al. 2010).

Anthropogenic threats. Three human activities are known to threaten humpback whales: whaling, commercial fishing, and shipping. Historic ally, whaling represente d the greatest threat to every population of whales and was ultimately responsible for listi ng several species as endangered.

Hum pback whales are also killed or injured duri ng interactions with commercial fishing gear. Like fin whales, humpback whales have been entangled by fishing gear off Newfoundland and Labrador, Canada. A total of 595 humpback whales were reported captured in coastal fisheries28 in those two provinces between 1969 and 1990, of which 94 died (Perkins and Beamish 1979, Lien 1994). Along the Atlantic coast of the US and the Maritime Provinces of Canada, there were 160 reports of humpback whales being en tangled in fishing gear between 1999 and 2005 (Cole et al. 2005c, Nelson et al. 2007c). Of these, 95 entangled humpback whales were confirmed, with 11 whales sustaining injuries an d nine dying of their wounds. Between 30 and 40% of humpback whales in the Arabian Sea show scarring from entanglements, with fishing effort on the rise (Baldwin et al. 2010).

More humpback whales are killed in collisions with ships than any other whale species except fin whales (Jensen and Silber 2003). Of 123 hum pback whales that stranded along the Atlantic coast of the US between 1975 and 1996, 10 (8.1%) s howed evidence of collisions with ships (Laist et al. 2001). Between 1999 and 2005, there were 18 reports of humpback whales being struck by vessels along the Atlantic coast of the US and the Maritime Provinces of Canada (Cole

et al. 2005c, Nelson et al. 2007c). Of these reports, 13 were confirmed as ship strikes and in seven cases, ship strike was determined to be the cause of death. In the Bay of Fundy, recommendations for slower vessel speeds to avoid right whale ship strike appear to be largely ignored (Vanderlaan 2008). However, new rule s for seasonal (June through December) slowing of vessel traffic to 10 knots and changing shipping lanes by less than one nautical mile to avoid the greatest concentrations of right whales are expected to reduce the chance of humpback whales being hit by ships by 9%. The first estimate of population- level effects of entanglement were recently produced, with over 12% of the Gulf of Maine population of humpbacks acquiring new scars from entanglement interactions annually (Mattila and Rowles 2010).

Organochlorines, including PCB and DDT, have b een identified from humpback whale blubber (Gauthier et al. 1997b). Higher PCB levels have been obser ved in Atlantic waters versus Pacific waters along the United States and levels tend to increase with individual age (Elfes et al. 2010). Although humpback whales in the Gu lf of Maine and off Southern California tend to have the highest PCB concentrations, overa ll levels are on par with ot her baleen whales, which are generally lower than odontocete cetaceans (Elfes et al. 2010). As with blue whales, these contaminants are transferred to young through the placenta, leaving newborns with contaminant loads equal to that of mothers before bioaccu mulating additional contaminants during life and passing the additional burden to the next generation (Metcalfe et al. 2004). Contaminant levels are relatively high in humpback whales as compar ed to blue whales. Humpback whales feed higher on the food chain, where prey carry highe r contaminant loads than the krill that blue

whales feed on.

Critical habitat. The NMFS has not designated critical habitat for humpback whales.

North Atlantic right whale Description of the species. All North Atlantic right wh ales compose a single population. Although not all individuals undergo the same mi gratory pattern, no subpo pulation structuring

has been identified.

Distribution. Right whales occur in sub-polar to tem per ate waters in all major ocean basins in the world, with a clear migratory pattern of hi gh latitudes in summer and lower latitudes in winter (Cummings 1985, Rice 1998b, Perry et al. 1999). The historical range of North Atlantic29 right whales extended as far s outh as Florida and northwestern Africa, and as far north as Labrador, southern Greenland, Iceland, and Norway (Reeves et al. 1978, Cummings 1985, Rice 1998b). Most sightings in the western North Atla ntic are concentrated within five primary habitats or high-use areas: coastal waters of the southeastern US, Cape Cod and Massachusetts Bays, the Great South Channel, the Bay of Fundy, and the Scotian Shelf (Winn et al. 1986). In 1994, the first three of these areas we re designated as critical hab itat for the North Atlantic right

whale.

North Atlantic right whales have been observe d from the mid-Atlantic Bight northward through the Gulf of Maine year-round, but are primar ily found along the north east US during summer and Florida during winter, with migratory routes in between. In New England, peak abundance of North Atlantic right whales in feeding areas occurs in Cape C od Bay beginning in late winter. In early spring (Late February to April), peak North Atlantic right whale abundance occurs in

Jordan and Wilkinson basins to the Great South Channel (Kenney et al. 1995, Nichols et al. 2008, Pace III and Merrick 2008). In late June and July, North Atlantic right whale distribution gradually shifts to the northern edge of Georges Bank. In late summer (August) and fall, much of the population is found in waters in the Ba y of Fundy, the western Gulf of Maine and around Roseway Basin (Winn et al. 1986, Kenney et al. 1995, Kenney et al. 2001, Pace III and Merrick 2008). However, year-to-year variation in space and time are known and likely result from patchy prey distribution (Nichols et al. 2008). Variation in the abundance and development of suitable food patches appears to modify the ge neral patterns of move ment by reducing peak numbers, stay durations and specific locales (Brown et al. 2001, Kenney 2001). In particular, large changes in the typical pa ttern of food abundance will dram atically change the general pattern of North Atlantic rightwhale habitat use (Kenney 2001).

Migration and movement. North Atlantic right whales e xhibit extensive migratory patterns, traveling along the eastern seaboard of the US and Canada between calving grounds off Georgia and Florida to northern feeding areas off of the northeast US and Canada in March/April and the

reverse direction in November/D ecember. The longest tracking of a North Atlantic right whale was a migration of 1,200 miles in 23 days the Bay of Fundy to Georgia (Mate and Baumgartner 2001). Migrations are typically w ithin 30 nautical miles of the co astline and in wa ters less than 160 feet deep. Although this patte rn is well-known, most of the population, particularly the males and non-pregnant females, is not found in the calving area and may not follow this pattern. It is unknown where the majority of the non-calving population spends the winter.

There have been a few recent sightings of Nort h Atlantic right whales far offshore, including those from Dutch ships indicati ng some individuals occur betw een 40° and 50°N, in waters influenced by the North Atlantic Current (the broad, eastward-flowing extension of the Gulf Stream). Right whales have been sighted offs hore (greater than 30 miles) during surveys flown

off the coast of northeastern Florida and south eastern Georgia from 1996 to 2001. These include three sightings in 1996, one in 1997, 13 in 1998, six in 1999, 11 in 2000, and six in 2001 (within each year, some were repeat sightings). Mate et al. (1997) recorded radio-tagged animals making extensive movements from the Gulf of Main e into deeper waters off the continental shelf (Mate et al. 1997). The frequency with which North At lantic right whales occur in offshore

waters in the southeastern US remains unclear. Occasionally, individuals are observed in distant locations, including the Gu lf of Mexico, Bermuda, the Gulf of St. Lawrence, Newfoundland,30 Greenland, Iceland, and northern No rway (an area known as a hist orical North Atlantic right whale feeding area Smith et al. 2006). The Norwegian sighting (September 1992) represents one of only two sightings this century of a right whale in Norwegian waters, and the first since 1926. Together, these long-range matches indicate an extended range for at least some individuals and perhaps the existence of important hab itat areas not presently well described.

Reproduction and demography. Data through the 1990s suggest s that mean calving interval increased since 1992 from 3.67 years to more than five years, a significan t trend that hampers North Atlantic right whale recovery (Best et al. 2001a, Kraus et al. 2007). This reproductive rate was approximately half that reported from studi ed populations of southe rn right whales (Best et
al. 2001b). This has been attributed to severa l possible causes, includ ing higher abortion or perinatal losses (Browning et al. 2009). An analysis of the age structure of Nort h Atlantic right whales suggests that the populat ion contains a smalle r proportion of juve nile whales than expected, which may reflect lowered recruitment and/or high juvenile mortality (Hamilton et al. 1998, Best et al. 2001a). In addition, it is possible that the apparently low reproductive rate is due in part to unstable age structure or to repr oductive senescence on the pa rt of some females. However, knowledge on either factor is poor. Even though investment in calves is high for North Atlantic right whales, an incident of calf exchange (probably acc identally and soon after birth) and subsequent adoption thro ugh weaning has been found (Frasier et al. 2010). Although North Atlantic right whales historically separa ted from their calves within one year, a shift appears to have taken place around 2001 where moth ers (particularly less experienced mothers) return to wintering grounds with their ye arling at a much greater frequency (71% overall)(Hamilton and Cooper. 2010). The significance of this change is unknown.

Habitat. Available evidence from North Atlantic right whale foraging and habitat studies shows that North Atlantic right whales focus foragi ng activities where physical oceanographic features such as water depth, current, and mixing fr onts combine to concen trate copepods (Wishner et al. 1988, Murison and Gaskin 1989, Mayo and Marx 1990, Baumgartner et al. 2003).

Feeding. North Atlantic right whales fast durin g the winter and feed during the summer, although some may opportuni stically feed during migration. North Atlantic right whales use their baleen to sieve copepods from dense pa tches, found in highly variable and spatially unpredictable locations in the Bay of Fundy, Ro seway Basin, Cape Cod Bay, the Great South Channel, and other areas off northern US and Canada (Pendleton et al. 2009). The primary prey of North Atlantic right whales is zoopla nkton, especially shrimp-like copepods such as Calanus (Kenney et al. 1985a, Beardsley et al. 1996). North Atlantic right whales feed largely by

skimm ing these prey from the ocean surface (Pivorunas 1979, Mayo and Marx 1990), but may feed anywhere in the water column (Wat kins and Schevill 1976, 1979, Goodyear 1993, Winn et
al. 1995). Feeding behavior has only been obs erved in northern areas and not on calving grounds or during migration (Kraus et al. 1993).

Status and trends. The Northern right whale was original ly listed as endangered in 1970 (35 FR 18319), and this status remained since the inception of the ESA in 1973. The early listing

included both the North Atlantic and the North Pacific populations, although subsequent genetic studies conducted by Rosenbaum (2000) resulted in strong eviden ce that North Atlantic and North Pacific right whales are separate specie s. Following a comprehensive status review,31 NMFS concluded that North Atlantic and North Paci fic right whales are separate species. In March 2008, NMFS published a final rule listing North Pacific and North Atlantic right whales as separate species (73 FR 12024).

North Atlantic right whales were formerly abunda nt, with an estimated 5,500 individuals present

in the 16 th century throughout the North Atlantic (Reeves 2001, Reeves et al. 2007). A review of the photo-id recapture database in June 2006, in dicated that only 313 individually recognized North Atlantic right whales were observed dur ing 2001. This represents a nearly complete census, and the estimated minimum population size. However, no estimate of abundance with an

associated coefficient of variation has been ca lculated for the populati on. The population growth rate reported for the period 1986 to 1992 by Knowlton et al. (1994) was 2.5%, suggesting the stock was showing signs of slow rec overy. However, work by Caswell et al. (1999) suggested that crude survival probability declined from about 0.99 in the ear ly 1980’s to about 0.94 in the

late 1990s. Additional work conducted in 1999 show ed that survival had i ndeed declined in the 1990s, particularly for adult females (Best et al. 2001a). Another workshop in September 2002

further confirmed the decline in this population (Clapham 2002).

Natural threats. Several researchers have s uggested that the recovery of North Atlantic right whales has been impeded by competition with other whales for food (Rice 1974, Scarff 1986). Mitchell (1975) analyzed trophic interactions among baleen whales in the western North Atlantic and noted that the foraging grounds of North Atla ntic right whales overl apped with the foraging grounds of sei whales. Both species f eed preferentially on copepods. Reeves et al. (1978) noted that several species of whales feed on copepods in the eastern North Pacifi c, so that the foraging pattern and success of right whales would be aff ected by other whales as well. Mitchell (1975) argued that the North Atlantic ri ght whale population had been depl eted by several centuries of whaling before steam-driven boats allowed whalers to hunt sei whales; from this, he hypothesized that the decline of the right whale population made more food available to sei whales and helped their population to grow. He then suggested that co mpetition with the sei whale population impedes or pr events the recovery of the right whale population.

Other natural factors influenci ng right whale recovery are possi ble, but unquantified. Right whales have been subjects of killer whale att acks and, because of their robust size and slow swimming speed, tend to fight ki ller whales when confronted (Ford and Reeves 2008). Similarly, mortality or debilitation from disease and red tide events are not known, but have the potential to be signifi cant problems in the recovery of ri ght whales because of their small population size.

Anthropogenic threats. Several human activities are known to threaten North Atlantic right whales: whaling, commercial fishing, shipping, a nd environmental contaminants. Historically, whaling represented the greatest threat to every population of right whales and was ultimately responsible for listing right whales as an endang ered species. As its legacy, whaling reduced North Atlantic right whales to about 300 indivi duals in the western North Atlantic Ocean; the number of North Atlantic right whales in the eastern North Atlantic Ocean is probably much smaller, although we cannot estimate the size of that population from the data available.

Of the current threats to North Atlantic right wh ales, entanglement in commercial fishing gear32 and ship strikes pose the greatest threats. Along the Atlantic coast of the US and the Maritime Provinces of Canada, there were 43 reports of No rth Atlantic right whales entangled in fishing gear between 1999 and 2005 (Cole et al. 2005a). Of the 39 reports that NMFS could confirm, North Atlantic right whales were injured in fi ve of the entanglements and killed in four entanglements. Recent efforts to disentangle right whales have met with success (Anonmyous. 2009b). In the same region, there were 18 reports of North Atlantic right whales being struck by vessels between 1999 and 2005 (Cole et al. 2005a, Nelson et al. 2007a). Of the 17 reports that NMFS could confirm, right whales were injured in two of the ship strikes and killed in nine.

Present recommendations for slow er vessel speeds in the Bay of Fundy appear to be largely ignored (Vanderlaan et al. 2008b). Proposed rules for seasonal (June through December) slowing of vessel traffic to 10 knots or changing shipping lanes by le ss than one nautical mile to avoid the greatest concentrations of right whales are predicted to be capable of reducing ship strike mortality by 62% in the Bay of Fundy region.

Concern also exists over climate change and it s effect on the ability of North Atlantic right whales to recover (Greene et al. 2003b). Specifically, the varia tions in oceanography resulting from current shifts and water temperatures can significantly affect the occurrence of the North Atlantic right whale’s primary food, copepod crust aceans. If climate changes such that current feeding areas cannot sustain North Atlantic right whales, the population may have to shift to reflect changes in prey distribut ion, pursue other prey types, or face prey shortage. Changes in calving intervals with sea surface temperature have already been documented for southern right whales (Leaper et al. 2006).

North Atlantic right whales, as with many mari ne mammals, are exposed to numerous toxins in their environment, many of wh ich are introduced by humans. Levels of chromium in North Atlantic right whale tissues are sufficient to be mutagenic and cau se cell death in lung, skin, or

testicular cells and are a concern for North Atlantic right whale recovery (Wise et al. 2008, Chen

et al. 2009). The organochlorines DDT, DDE, PCBs , dieldrin, chlordane, HCB, and heptachlor epoxide have been isolated from blubber sample s and reported concentrations may underestimate actual levels (Woodley et al. 1991). Mean PCB levels in North Atlantic right whales are greater than any other baleen whale species thus far measured, although less th an one-quarter of the levels measured in harbor porpoises (Van Scheppingen et al. 1996, Gauthier et al. 1997a). Organochlorines and pesticides, although variable in concentration by season, do not appear to currently threaten North Atlantic righ t whale health and recovery (Weisbrod et al. 2000). Flame retardants such as PBDEs (known to be carcinogenic) have also b een measured in North Atlantic right whales (Montie et al. 2010).

Critical habitat. Critical habitat is desi gnated for right whales in the North Atlantic. NMFS designated three areas in June 1994 as critical habitat for Eubalaena glacialis for feeding and calving (59 FR 28805). The critic al habitats for feeding cove r portions of the Great South Channel (east of Cape Cod), Massachusetts Bay and Cape Cod Bay, and Stellwagen Bank. Northern critical habitat was designated becau se of the concentration of right whales that feed in the area, apparently associated with complex o ceanographic features that drive prey density and

distribution. This area has come under considerable scrutiny with in the past few years because of the concern over ship strikes in this area. Boston serves as a major port facility and vessels transiting to and from the port cross critical ha bitat where North Atlantic right whale mortality33 occurs. Shipping traffic has generally increased in the recent past and co uld be considered to degrade the habitat due to the a dditional mortality and injury ri sk now present in the area. Although voluntary regulations are in place, th ese are frequently ignored and mandatory regulations are under considera tion. The southern critical habitats are along Georgia and northeastern Florida coasts (waters from the coas t out 15 nautical miles be tween the latitudes of

31°15’ N and 30°15’ N and from the coast out fi ve nautical miles between 30°15’ N and 28°00’ N). Southern critical habitat is designated to protected calving and breeding grounds for North Atlantic right whales, which gene rally calve and breed in shallow coastal waters. This critical habitat has generally fared better than northern critical habitat and significant degradation has not been clearly identified.

Southern resident killer whale Description of the species. Southern Resident killer whal es compose a single population that occurs primarily along Washington State and British Columbia. The listed entity consists of three family groups, identified as J, K, and L pods.

Distribution. They are found throughout the coastal waters off Washi ngton, Oregon, and

Vancouver Island and are known to travel as far s outh as central California and as far north as the Queen Charlotte Islands, British Columbia. However, there is limited information on the range of Southern Residents along the outer Pacific Coast, with only 25 confirmed sightings of J, K, and L pods between 1982 and 2006 (Krahn et al. 2004a).

Movement and habitat. Southern Residents are highly mob ile and can travel up to 100 miles per day (Erickson 1978, Baird 2000). Members of K and L pods once traveled a straightline distance of 584 miles from the northern Queen Ch arlotte Islands to Victoria, Vancouver Island, in seven days. Movements may be related to food availability.

Southern Resident killer whales spend a significant portion of the year in the inland waterways of the Strait of Georgia, Stra it of Juan de Fuca, and Puget Sound, particularly during the spring, summer, and fall, when all three pods are regularly present in the Georgia Basin (defined as the Georgia Strait, San Juan Isla nds, and Strait of Juan de Fuca)(Heimlich-Boran 1988, Felleman et
al. 1991, Olson 1998, Osborne 1999). Typically, K and L pods arrive in May or June and primarily occur in this core area until October or November. During this stay, both pods also make frequent trips lasting a few days to the outer coasts of Washington and southern Vancouver Island (Ford et al. 2000); however, J pod’s movements diffe r considerably and are present only intermittently in the Georgia Basin and Puget Sound. Late spring and early fall movements of Southern Residents in the Georgia Basin have remained fairly consistent since the early 1970s, with strong site fidelity shown to the region as a whole (NMFS 2005g). During late fall, winter, and early spring, the ranges and movements of the Southern Residents are less well known.

Offshore movements and distributi on are largely unknown for the Sout hern Resident population.

While the Southern Residents are in inland wate rs during the warmer mo nths, all of the pods concentrate their activities in Haro Strait, B oundary Passage, the southern Gulf Islands, the eastern end of the Strait of Juan de Fuca, and several localities in the southern Georgia Strait (Heimlich-Boran 1988, Felleman et al. 1991, Olson 1998, Ford et al. 2000). Individual pods are similar in their preferred areas of use, although there are some seasonal and temporal differences34 in certain areas visited (Olson 1998). For example, J pod is the only group to venture regularly

inside the San Juan Islands. The movements of Sout hern Resident killer whal es relate to those of their preferred prey, salmon. P ods commonly seek out and forage in areas where salmon occur, especially those associated with migrati ng salmon (Heimlich-Boran 1986, 1988, Nichol and Shackleton 1996).

Members of different pods do interact, but memb ers generally remain within their matrilinear group (Parsons et al. 2009). However, additional interac tion between pods has occurred over the past two decades, possibly in association with th e decline of the Southern Resident population as a whole (Parsons et al. 2009).

Feeding. Southern Resident killer whales are fi sh eaters, and predominantly prey upon salmonids, particularly Chinook salmon but are also known to consume more than 20 other species of fish and squid (Scheffer and Slipp 1948, Ford et al. 1998, Ford et al. 2000, Saulitis et
al. 2000, Ford and Ellis 2005, 2006). Killer whales s how a strong preference for Fraser River Chinook salmon (78% of identified prey ) during late spring to fall (Hanson et al. 2005, Ford and Ellis 2006, Hanson et al. 2010b). Chum salmon are also taken in significant amounts (11%), especially in autumn. Chinook are preferred despite much lower abundance in comparison to other salmonids (such as sockeye) presumably b ecause of the species’ large size, high fat and energy content, and year-round occurrence in the area. Killer whales also captured older (i.e., larger) than average Chinook (F ord and Ellis 2006). Throughout inland waters from May to September, Southern resident killer whal e diet is approximately 88% Chinook (Hanson et al. 2007b, Hanson et al. 2010a), with a shift to chum salmon in fall. Little is known about the winter and early spring diet of S outhern Residents. Early results from genetic analysis of fecal and prey samples indicate that Southern Residents consume Fraser River-origin Chinook, as well as salmon from Puget Sound, Washington and Ore gon coasts, the Columbia River, and Central Valley of California (Hanson et al. 2007a, Hanson et al. 2010a). However, recent studies suggest that members of L pod have undergone dietary shifts from Chinook salmon during fall months over the past decade (Krahn et al. 2009). Southern resident ki ller whales appear to be

more sensitive to vessel disturbance while feeding than during ot her activities (Ashe et al. 2010).

An area to the southwest of San Juan Isla nd appears to be a foraging “hotspot” (Ashe et al. 2010).

Growth and reproduction. Female Southern Resident killer whales give birth to their first surviving calf between the ages of 12 and 16 years (mean ~ 14.9 years) and produce an average of 5.4 surviving calves during a reproductive life span lasting about 25 years (Olesiuk et al. 1990a, Matkin et al. 2003). Females reach a peak of reproduction around ages 20-22 and decline

in calf production gradually over the next 25 years until reproductive senescence (Ward et al. 2009a). Older mothers tend to have greatercalving success than do their younger, lessexperienced counterparts (Ward et al. 2009b). Calving success also appears to be aided by the assistance of grandmothers (Ward et al. 2009b). The mean interval be tween viable calves is four years (Bain 1990). Males become sexually mature at body lengths ranging from 17 to 21 feet, which corresponds to between the ages of 10 to 17.5 years (mean ~ 15 years), and are presumed to remain sexually active throughout their adult lives (Christensen 1984, Perrin and Reilly 1984, Duffield and Miller 1988, Olesiuk et al. 1990a). Most mating is beli eved to occur from May to October (Nishiwaki 1972, Olesiuk et al. 1990a, Matkin et al. 1997). However, conception35 apparently occurs year-round because births of calves are reported in all months. Newborns measure seven to nine feet long and weigh about 440 lbs (Nishiwaki and Handa 1958, Olesiuk et
al. 1990a, Clark et al. 2000, Ford 2002). Mothers and offspr ing maintain highly-stable, life-long social bonds and this natal relationship is the basis for a matrilineal social structure (Bigg et al. 1990, Baird 2000, Ford et al. 2000). Some females may reach 90 years of age (Olesiuk et al. 1990a).

Status and trends. Southern Resident killer whales ha ve been listed as endangered since 2005 (70 FR 69903). In general, there is little in formation available regarding the historical

abundance of Southern Resident killer whales. Some evidence suggests that, unt il the mid- to late-1800s, the Southern Resident killer whal e population may have numbered more than 200 animals (Krahn et al. 2002a). This estimate was based, in part, on a recent genetic study that found that the genetic diversity of the Southe rn Resident population re sembles that of the Northern Residents (Barrett-L ennard 2000, Barrett-Lennard and Ellis 2001), and concluded that the two populations were possibly once similar in size. Unfortunately, lack of data prior to 1974 hinders long-term population analysis (NMFS 2005g) . The only pre- 1974 account of Southern Resident abundance is from Sheffer and Slipp (1948) and merely notes that the species was “frequently seen” during the 1940s in the Strait of Juan de Fuca, northern Puget Sound, and off the coast of the Olympic Peninsula, with sm aller numbers along Washington’s outer coast.

Olesiuk et al. (1990a) estimated the Southern Resident popul ation size in 1967 to be 96 animals.

Due to demand for marine mammals in zoos and marine parks, it is estimated that 47 killer whales, mostly immature, were taken from the Southern Resident population for public display between 1967 and 1973. By 1971, the level of removal decreased the population by about 30% to approximately 67 individuals (Olesiuk et al. 1990a). The population went then went through periods of decline and expansion for more than two decades. At the end of an 11-year growth cycle in 1995, the three Southern Resident pods – J, K, and L, reached a peak of 98 animals (NMFS 2008).

More recently, the Southern Resident population ha s continued to fluctuate in numbers. After growing to 98 whales in 1995, the population de clined by 17% to 81 whales in 2001 (-2.9% per year) before another slight in crease to 84 whales in 2003 (Ford et al. 2000, Carretta et al. 2005). The population grew to 90 whales in 2006, alt hough it declined to 87 in 2007 (NMFS 2008f). The most recent population abundance estimate of 87 Southern Residents consists of 25 whales in J pod, 19 whales in K pod, and 43 whales in L pod (NMFS 2008f).

Natural threats. The recent decline, unsta ble population status, a nd population structure (e.g., few reproductive age males and non-calving adult fema les) continue to be causes for concern. Moreover, it is unclear whethe r the recent increasing trend will continue. The relatively low number of individuals in this population makes it di fficult to resist/recover from natural spikes in mortality, including disease and fluctuations in prey availa bility (NMFS 2008f). Although

disease outbreaks have not been identified in this population, increased contam inant load (see below) may increase the susceptibilit y of individuals to disease.

Anthropogenic threats. Numerous threats to the continued survival of Southern Resident killer whales have been identified (NMFS 2008f). Many of these are human in origin. The primary prey of killer whales, salmon, has been severely reduced due to habitat loss and overfishing of36 salmon along the West Coast (NRC 1996c, Slaney et al. 1996, Gregory and Bisson 1997, Lichatowich 1999, Lackey 2003, Pess et al. 2003, Schoonmakeret al. 2003). Several salmon species are currently protected under the ESA, and are genera lly well below their former

numbers. A 50% reduction in ki ller whale calving has been correlated with years of low Chinook salmon abundance (Ward et al. 2009a).

Puget Sound also serves as a major port and drai nage for thousands of square miles of land. Contaminants entering Puget Sound and its surrounding waters accumulate in water, benthic sediments, and the organisms that live and eat here (Krahn et al. 2009). As the top marine predator, Southern Resident killer whales bioaccumu late these toxins in their tissues, potentially leading to numerous physiologi cal changes such as skeletal deformity, lowered disease resistance, and enzyme disruption (Krahn et al. 2009). Presently, the greatest contaminant threats are organochlorines, whic h include PCBs, pesticides, dioxi ns, furans, other industrial products, and the popularized chemical DDT (Ross et al. 2000a, CBD 2001b, Krahn et al. 2002a,

Cullon et al. 2009, Krahn et al. 2009). These chemicals tend to bioaccumulate in fatty tissues, such as whale blubber, persist over long periods in the environment, and can be transmitted from mother to offspring (Haraguchi et al. 2009, Krahn et al. 2009). Levels are much higher in fieldsampled individuals than those found in a captive killer whale (Bennett et al. 2009). A similar, but separate concern is the growth of the pe troleum industry in Puget Sound, which has the low potential to create a catastr ophic oil spill, or more likely, sm all but chronic releases of petrochemicals. As southern resi dent killer whales are normally exposed to high levels of whale watching, engine exhaust has been assessed as a possible threat and, in some cases, may contribute to health affects (Lachmuth et al. 2011).

Vessel activity also has been iden tified as a threat. This includes physical harm or behavioral modifications as well as habitat degradation/loss from US naval vessel sonar activities, ship strike, and heavy and continuous presence by wh ale-watching vessels. In 2005, a US vessel participating in sonar exercises apparently caus ed significant behavior ch anges in killer whale activity in the area, such that the whales vacated the area (NMFS 2005b). Although such activities are now receiving close scrutiny, the potent ial remains for these disruptions to occur, or as in other areas, the potential for auditory tr auma, stranding, and death. The increase in “background noise” resulting from vessel traffic a nd coastal development activities, although not directly traumatic, has the potenti al to influence or disrupt the acoustic system that Southern Resident killer whales use to navigate, co mmunicate, and forage (Bain and Dahlheim 1994, Gordon and Moscrop 1996, Erbe 2002a, Williams et al. 2002a, Williams et al. 2002c, NMFS 2008f, Holt et al. 2009). Commercial wh ale-watching in the regi on focuses primarily on

Southern Resident killer whales and has increased dram atically in the recent years (Osborne et
al. 1999, Baird 2001a, Erbe 2002a, MMMP 2002a, Koski 2004, 2006b, 2007a). Although mechanisms are in place to regulate the industry, concerns remain over persistent exposure to vessel noise, proximity to whales, which can cause behavioral changes, stre ss, or potentially the loss of habitat (Kruse 1991, Kriete 2002, Williams et al. 2002a, Williams et al. 2002c, Foote et
al. 2004a, Bain et al. 2006b, Bain et al. 2006c, NMFS 2008f, Wiley et al. 2008, Noren et al. 2009).

Critical habitat. Critical habitat for the DPS of Southern Resident killer wh ales was designated on November 29, 2006 (71 FR 69054). Three specifi c areas were designated; (1) the Summer37 Core Area in Haro Strait and waters around the San Juan Islands; (2) Puget Sound; and (3) the Strait of Juan de Fuca, which comprise approxi mately 2,560 square miles of marine habitat.

Three essential factors exist in these areas: water qua lity to support growth and development, prey species of sufficient quantity, quality and availability to support individual growth, reproduction and development, as well as overall population growth, and pa ssage conditions to allow for migration, resting, and foraging. Water quality has declined in recent years due to agricultural run-off, urban devel opment resulting in a dditional treated water discharge, industrial development, oil spills. The primary prey of sout hern residents, salmon, has also declined due to overfishing and reproductive impairment associated with loss of spawning habitat. The constant presence of whale-watching vessels and grow ing anthropogenic noise background has raised

concerns about the health of areas of growth and reproduction as well.

Cook Inlet beluga whale Description of the species. Beluga whales are wide ly distributed in Arctic and subarctic waters, and in Alaska five putative populations exist (B eaufort Sea, eastern Chukchi Sea, Bristol Bay, eastern Bering Sea, and Cook Inlet)(Angliss et al. 2001). Cook Inlet beluga whales are the only population that is listed under the ESA. Mito chondrial and nuclear DNA distinguish Alaskan beluga whales from those that occur in Hudson Strait, Baffin Bay and the St. Lawrence River, with the Cook Inlet population de monstrating the strong evidence of genetic isolation from the other Alaskan populations and ot her populations demonstrating w eak to moderate evidence of genetic isolation (O'Corry-Crowe et al. 2007, O'Corry-Crowe 2008, O'Corry-Crowe et al. 2010)

.

Distribution. Beluga whales are observed year-r ound in Cook Inlet although less is known

about their winter movements than summer movements (Hobbs et al. 2008). Data from satellite tagging studies suggest that movements of Cook Inlet beluga whales during summer months are short and largely focused around river estuarie s and inlets (e.g., Chickaloon Bay, Turnagain Arm, Susitna River, and Knik Arm in the upper inlet and in many cases the animals exhibited very little movement for weeks during the summer (Hobbs et al. 2005). Dense groupings in these areas during June and July are the fo cus of NMFS aerial surveys, but numbers drop substantially in the upper inlet by November (Hobbs et al. 2005). Outside of Cook Inlet in the Gulf of Alaska beluga whale sigh tings are extremely rare (Laidre et al. 2000). Hobbs et al. (2005) found that tagged beluga whales moved to farther offshore during winter months, but

remained within Cook Inlet. However travel distance appear ed to increase during winter months, and exhibited more widely dispersed patterns both within and among individuals (Hobbs

et al. 2005). Distribution during a ll months is likely influenced by prey distribution, where salmon and eulachon are concentrated in river mouths during summer m onths and other prey like

sand lance are found in mid and bottom waters of the inlet during winter months, albeit in more dispersed patterns leading to the wi der dispersal of the whales.

Based on past studies of the summ er distribution of beluga whales in Cook Inlet, it appears that the population has experienced a contraction in its overall distribution (Speckman and Piatt. 2000, Hobbs et al. 2008, Rugh et al. 2010). Aerial surveys in the 1970s indicated that at least 10% of the population used areas south of Kenai River and Kalgin Island (mid- to lower Cook Inlet) during summer months, whereas more r ecent surveys (1993-2007) observed more than 90% of the beluga whales in upper Cook Inle t in shallow waters. According to Hobbs et al. (2008) 90% of the whales in the 1970s were observed within 70 nmi of the western tip of38 Anchorage (Point Woronzof), whereas more rece ntly (1998-2007) 90% were detected within 20 nmi. Although the precise reason for the range contraction is not know n, the shrinking summer distribution likely reflects the re duction in the population size over the same intervals and the beluga whales’ preference for de nse aggregations of preferre d prey species. Through the distribution has shrunk since the 1970s, possi bly due to human hunting in excess of 50 individuals per year in the north ern portions of Cook Inlet (where human disturbance is highest; city and port of Anchorage), surviving individuals seemed to re-occupy this habitat and abandon southern habitat (Rugh et al. 2010). This suggests that the no rthern reaches of Cook Inlet is superior habitat to southern areas, even with the anthropogenic stresso rs present there (Rugh et
al. 2010). Overall, Cook Inlet beluga whale range has shrunk from greater than 7,000 km 2 to less than 3,000 km 2 (Rugh et al. 2010).

Habitat. During spring and summer months, beluga whales in Cook Inlet are typically concentrated near northern river mouths (Rugh et al. 2000). Cook Inlet expe riences some of the

most extreme tidal fluctuati ons in the world (see Hobbs et al. (2008) for a discussion), and beluga whales in the inlet have adapted to thes e tidal cycles and seemingly take advantage of them, although the precise causal reasons ar e not well known. Presumably, the feeding opportunities these tidal cycles pro ffer the beluga whale are a contri buting factor. Beluga whales

move further into inlets and ar ms during tidal floods and move back out during ebb tides in a relatively predictable manner, al though highly-localized differences do exist (Ezer 2010). Tidal

state is also believed to influence travel and m illing behavior (Garner and Mckee. 2010). Aerial surveys and predictive models of habitat us indi cate that beluga whale movement patterns are closely correlated to tidal patterns, flow accumulation and mudflats, with a preference for medium and high flow inlets of larger river basins (Goetz et al. 2007, Ezer et al. 2008). More information, however, is needed to link these ha bitat attributes to cau sative reasons for this preference. Besides feeding, studi es have suggested this preferen ce for tidal mudflats may also be attributed to calving and breeding, molting, or shelter from predator s like killer whales (Calkins 1989a, Huntington 2000b, Moore et al. 2000c, Shelden et al. 2003). Individual movements suggest utilization of a variety of areas, with little to no limitation to specific areas over time (Mcguire et al. 2010b).

Reproduction. Sexual maturity is believed to be attained at 4 to 10 years for females and at 8 to 15 for males (Nowak 1991, Suydam et al. 1999). Females typically pr oduce a single calf every 2 to 3 years following a 14-month gestation. Of te n adult females examined after stranding, seven showed evidence of lactation, post-partum condition, or pregnancy when they died (Burek et al. 2010). Calving in Cook Inlet is assumed to occur from mid- May to mid-July, with some occurrences in April and August (C alkins 1983, Huntington 2000a, Mcguire et al. 2010a). Areas specific to calving and rearing ha ve thus far not been identified, but movement to different areas

within Cook Inlet has been noted ; calves and their mothers appear to use the same habitats as other conspecifics (Mcguire et al. 2010a, Mcguire et al. 2010b). Alaska natives describe calving

areas within Cook Inlet as the nor thern side of Kachemak Bay in April and May, off the mouths of the Beluga and Susitna Rivers in May, and in Chickaloon Bay and Tu rnagain Arm during the summer. According to surveys by LGL (Funk et al. 2005) cow/calf pairs also make extensive

use of Knik Arm in the summer and fall. Neona tes are often not seen until June in Cook Inlet (Burns and Seaman 1986a). The warmer waters from rivers may help keep newborns warm relative to cooler inlet waters (Katona et al. 1983, Calkins 1989b). Mating follows the calving39 period.

Calculation of beluga whale age is based on growth layers in teet h. Some debate exists as to whether a beluga whale tooth cont ains two growth layer groups (GLG) per year or one growth layer per year (Hobbs et al. 2008). Due to this ambiguity, Hobbs et al. (2008) summarized life history parameters according to tooth growth la yers rather than years (Table 3 from (Hobbs et al. 2008)).

Table 3. Review of Female beluga life history paramete rs found in the published literature (from (Hobbs et al. 2008); GLG=growth layer groups).

Parameter Data Sources Age at sexual maturity 8-15 GLG (Brodie 1971, Sergeant 1973, Ognetov 1981, Seaman and Burns. 1981, Braham 1984, Burns and Seaman 1986b)

0% at 8-9 GLGs (Burns and Seaman 1986b) a

33% at 10-11 GLGs

94% at 12-13 GLGs 9.1 +/- 2.8 GLGs (Robeck et al. 2005) Age at color change (gray to white) 12 GLGs (Brodie 1971)

22 GLGs (Sergeant 1973) Age at 1 s t conception 54% at 8-9 GLGs (Burns and Seaman 1986b) b

41% at 10-11 GLGs

94% at 12-13 GLGs Age at senescence 42-43 GLGs (Brodie 1971) Pregnancy and birth rates Small fetuses: (Burns and Seaman 1986b)

0.055 at 0-11 GLGs

0.414 at 12-21 GLGs

0.363 at 22-45 GLGs

0.267 at 46-57 GLGs

0.190 at 58-77 GLGs

With full-term fetuses/neonates:

0.000 at 0-11 GLGs

0.326 at 12-21 GLGs

0.333 at 22-45 GLGs

0.278 at 46-51 GLGs

0.182 at 52-57 GLGs

0.125 at 58-77 GLGs Lifespan >60 GLGs (Oldest female estimated at 70+ GLGs) (Burns and Seaman 1986b)

64-65 GLGs (Khuzin 1961, Ohsumi 1979) 60-61 GLGs (Brodie 1971) 50-51 GLGs (Sergeant 1973)

Adult annual survival 0.96-0.97 (Beland et al. 1992) 0.955 (based on pilot whale data) (Brodie 1971) 0.935 (Lesage and Kingsley. 1998) 0.91-0.92 (Allen and Smith. 1978) 0.906 (includes natural & human-caused mortality) (Burns and Seaman 1986b)

0.84-0.905 (based on body length and lifespan (Ohsumi 1979)

Immature annual survival 0.905 (for neonates in first half year) (Sergeant 1973) Reproductive rate 0.010-012 (Perrin 1982) c40 Parameter Data Sources

0.11 d (Burns and Seaman 1986b)

0.13 d (Sergeant 1973)

0.09 d (Brodie 1971)

0.09-0.12 d (Braham 1984)

0.09-0.14 e (Braham 1984)

0.12 e (Sergeant 1973, Ray et al. 1984)

0.08-0.14 e (Davis and Evans 1982)

0.06-0.10 e (Davis and Finley. 1979)

0.08-0.10 e (Brodie et al. 1981) 0.08 (unknown) (Breton-Provencher 1980) Calving Interval <3 years (Burns and Seaman 1986b) f
2 yrs and 3 years (Sergeant 1973) g

a Alaska sample (52 whales). Sampling occring in June when most Alaskan beluga whales are born. Hobbs et al. (2008) note that it is possible that non-pregnant 8-9 GLGs beluga whales would have conceived before their 10-11

GLG birth date.

b Alaska sample of 22 whales.

c Based on literature review and adopted by the International Whaling Commission

d Based on annual calf production rates

e Based on calf counts

f For some female beluga whales. This was a tentative conclusion based on high conception rates noted in some females between the ages of 12-13 GLGs and 44-45 GLGs.

g Two-year intervals were for 25% of mature female belugas in eastern Canada (7 of 29 sampled); presumed after noting pregnancies occurred during lactation. Three-year intervals were for 75% of mature females in eastern Canada. Sergeant (1973) concluded that the “overlap of pregnancy and previous lactation is infrequent so that calving occurs about once in three years.”

Lifespan. Lifespan is known to exceed 30 y ears of age (Burns and Seaman 1986a).

Feeding. Analyses of beluga whale stomach cont ents indicate that beluga whales are opportunistic feeders (including incl uding octopi, squids, crabs, shrimps, clams, mussels, snails, sandworms, and fishes such as capelin, cod, he rring, smelt, flounder, sole, sculpin, lamprey, lingcod, salmon, trout, whitefish, northern pike, grayling, an d tomcod (Klinkhart 1966, Fay et al. 1984, Haley 1986, Perez 1990, Huntington 2000a)), but specific species form the bulk of the prey when they are seasonally abundant (Hobbs et al. 2008). For instance, eulachon ( Thaleichthys pacificus ) also known as smelt or candlefish, ar e a small anadromous fish return that their natal rivers in spring for spawning. In the Susitna River, the eulachon spawning migration has a bimodal peak, with fish enteri ng the estuary in May and again in June, and represents a significant biomass of prey, with estimates of se veral thousand fish entering the

river in the first wave and several million ente ring the river in June (Calkins 1989a). The

common name candlefish is derived from the fact the fish is so high in fat content during spawning, with up to 15% of total body weight as fat, that when caught and dried and strung on a

wick the fish could be bu rned like a candle (Payne et al. 1999). This high fat content confers a significant source of energy for beluga whales, in cluding calving whales that occur in the upper inlet during the same period (Calkins 1989a). The stomach contents of one beluga whale harvested in upper Cook Inlet in 1998 near the Susitna River cont ained only eulachon. Based on stomach sample analyses from 2002-2007 fish compose the majority of the prey species, with gadids (cod and walleye pollock) and salmonids co mposing the majority of the fish eaten (Hobbs

et al. 2008). Anadromous salmonids begin concentra ting at the river mouths and intertidal flats in upper Cook Inlet in late spring and early summe r as emigrating smolts and immigrating adult spawners (Fried et al. 1979, Hazard 1988). Like eulachon, salmon are another source of lipid-41 rich prey for the beluga whale and represent th e greatest percent frequency of occurrence of the prey species found in Cook Inle t beluga whale stomachs(Hobbs et al. 2008). As salmonid numbers dwindle in the fall and winter, beluga wh ales return to feed on nearshore or deeper water species including cod, sculpin, flounde r, sole, shrimp, crab and others (Hobbs et al. 2008). Feeding frequently occurs collaboratively. Si gnificantly, it appears that Cook Inlet beluga whales have switched diets during 1965 to 2007(Quakenbush and Nelson. 2010). Although the

exact nature of this switch is unknown, it appears that individual wh ales are currently feeding at a lower trophic level than they formerly did(Quakenbush and Nelson. 2010).

Status and trends. On October 22, 2008, NMFS listed the Cook Inlet beluga whale as endangered (73 FR 62919). The Cook Inlet population is relatively isolated compared to other beluga whale stocks and the Alas ka Peninsula is likely an effec tive physical barrier to genetic exchange, perhaps for thousands of years (Murray and Fay 1979, O’Corry-Crowe et al. 1997,

Laidre et al. 2000). As such, this population may be part icularly sensitive to decline. Because few robust data are available prior to the 1990s, hi storical population size is based upon surveys lacking strong statistical confid ence. These estimates range from 150 to 450 individuals between 1963 and 1988 (Klinkhart 1966, Sergeant and Br odie 1975, Murray and Fay 1979, Calkins 1984, Hazard 1988)(Table 4). However, the most r obust survey prior to 1994 occurred in 1979 and

estimated Cook Inlet to contain 1,293 individuals and may have been an underestimate of the population (Calkins 1989b). Based upon this surve y, the maximum stable size of the Cook Inlet population is believed to be 1,300 individuals and this value is used for management purposes.

Table 4. Estimated abundance of Cook Inlet beluga whales with coefficient of variation and 95% confidence

intervals.

Year Estimate 1 CV 95% CI 2

Lower Upper
1979 1,293

1994 653 0.43 291 1464

1995 491 0.44 215 1120

1996 594 0.28 347 1018

1997 440 0.14 335 578

1998 347 0.29 199 606

1999 367 0.14 279 482

2000 435 0.23 279 679

2001 386 0.087 326 458

2002 313 0.12 248 396

2003 357 0.107 290 440

2004 366 0.2 290 440

2005 278 0.18 196 394

2006 302 0.16 221 412

2007 375 0.14 285 492

2008 375 0.23 240 585

2009 2 321 0.18 226 456

1 All estimates, except 1979 estimate, reported in Hobbs & Shelden (2008). The 1979 estimate is from Calkins (1989a) as cited in NMFS 2008.

2 Data from R. Hobbs, pers. comm., to A. Garrett, Apr. 2010.

Systematic aerial surveys of beluga whales throughout Cook Inlet began in 1993. Surveys each

June from 1999 to 2005 have estimated abunda nce between 278 and 435 individuals, with a sharp 47% decline between 1994 and 1998 proceeding that (Hobbs et al. 2000, NMFS42 unpublished data). Data from June 2006 surv eys supported an abundance of 302 individuals (Angliss 2007). Trends support a 71 % probability that the populati on is declining and this trend is likely to continue (Hobbs et al. 2006, Lowry et al. 2006). Between 1979 and 1994, according to above noted population estimates, Cook Inlet beluga whales declined by 50%, with another 50% decline observed between 1994 and 1998. Using a growth fitted model Hobbs et al. (2008) observed an average annua l rate of decline of -2.91% (SE = 0.010) from 1994 to 2008, and a -15.1% (SE=0.047) between 1994 and 1998. A comparison with the 1999-2008 data suggests the ra te of decline at -1.45% (SE=0.014) per year (Hobbs et al. 2008). Reasons for population decline are somewhat nebulous, but mortality pressures include native harvests, strandings, an d killer whale predation (Mahoney and Shelden 2000, Angliss and Outlaw 2006, Angliss and Ou tlaw 2007b, Angliss 2007). However, Cook Inlet harvests have been severely restricted (1-2 whales annually) si nce 1999, which coincides with a significant trend in the stabilization of the beluga whale population. Despite this, the population has not yet shown signs of recovery to historical nu mbers. Although not currently factors, disease and habitat disturbance have th e potential to significantly impact the Cook Inlet population. According to the best available popu lation viability analys is, there is an 80% probability that the population is declining, a 26% probability that the popul ation will be extinct in 100 years (by 2108) and a 70% probability that the population will be extinct within 300 years (by 2308).

Natural threats. Cook Inlet beluga whales have significan t natural mortality from a variety of sources (NMFS 2006c). According to Hobbs et al. (2008), over 700 beluga whales have

stranded in Cook Inlet since 1988. Although many of these strandli ngs survive and refloat with the next tide, a portion do not (Mahoney et al. 2010). Killer whales are known to prey upon beluga whales in Cook Inlet and although exact removal numbers are unknown, small reductions in a small population can limit recovery (Burek et al. 2010, Mahoney et al. 2010). Herpes virus, meningoencephalitis, and parasites ha ve been identified as diseases in Cook Inlet beluga whales (Burek et al. 2010). However, the significance of disease and other causes of death is difficult to estimate due to the poor condition in which most dead-stranded beluga whales are found in Cook Inlet (Burek et al. 2010). Loss of genetic variability has b een raised as a possible reason for the failure of Cook Inlet beluga whales to recover; studies to evaluate this possibility are currently underway (Bechdel et al. 2010).

Anthropogenic threats. Cook Inlet beluga whales have been exposed to anthropogenic pressures in the past, including incidental take s in gillnet fisheries of roughly three to six individuals per year (Murray and Fay 1979, Burn s and Seaman 1986a). During the early 1900s

there was a short-lived commer cial whaling company, The Beluga Whaling Company, which operated at the Beluga River in upper Cook Inlet. The Company during its 5 years of operation harvest 151 belugas from 1917-1921 (Mahoney and Shelden. 2000). Although evidence does not presently exist for harassment of belugas, area s of high beluga use coincide with regions of

lowest ambient sound and have been known to evacuate high-use areas of hum ans (Huntington 2000a). Significant behavioral changes have been observed in Cook Inlet belugas in response to small boat proximity (NMFS 2006c). Propeller s cars are observed on belugas, although direct mortality has not been verified. Subsistence hunting has occurred since at least the 1700s and may occur in the future (Braund and Huntingt on. 2010). In 1999 and 2000 there was a voluntary moratorium on subsistence harvest (Carter and Ni elsen. 2010), and since substance harvest have43 been conducted under co-managemen t agreements. Since 2000, no more than 2 beluga whales have been taken in subsistence harvests in any one year (Hobbs et al. 2008). Take levels have been as high as 20% in the 1990s, but hunts ar e now tightly regulated and based upon population dynamics.

Although belugas in the St. Lawrence Seaway and Canadian Arctic have high contaminant burdens and subsequent biological impairments (such as cancer), belugas in Cook Inlet have

been found to have low concentrations of P CBs, organochlorines, and heavy metals (Becker et
al. 2000a, Kelly et al. 2009, Mcaloose and Newton. 2009, Tomy et al. 2009), although in situ concentrations of some organi c pollutants have incr eased over the past 20 years (Becker 2010). However, copper is two to three times higher in Cook Inlet beluga whales than beluga whales

from the eastern Beaufort Sea a nd the eastern Chukchi Sea, but is similar concentrations found in Hudson Bay beluga whales (Becker et al. 2000b). Studies indicate that PCBs and chlorinated pesticide concentrations are higher in male St. Lawrence Seaway beluga whales than females, reflecting the transference of body loads to the offspring that occurs during gestation and lactation (Becker et al. 2000b). However, discharges from indus trial activities that do not enter municipal treatment systems (petroleum, seafood processing, ship ballast, dredging), discharges from municipal wastewater treatment systems, runoff from urban areas, mining operations, military sites, airports and agricultural areas, and accidental spills or discharges of petroleum and other products remain concerns for stock recovery (Moore et al. 2000b, NMFS 2006c). Selenium from tooth samples has been found to be a good predictor of selenium levels in liver and muscle (selenium can be a pr edictor of mercury levels)(Kinghorn et al. 2008). Mercury levels also vary by age, as individuals grow older and switch dietary preferences (Loseto and Ferguson 2008).

Critical habitat. NMFS designated critical habitat for th e Cook Inlet beluga whale on April 11, 2011 (76 FR 20180). Two areas specific areas ar e proposed comprising 7,809 square kilometers of marine habitat. Area 1 encompasses 1,918 square kilometers (741 sq. mi.) of Cook Inlet

northeast of a line from the mouth of Threemile Creek (61o 08.5’N., 151 o 04.4’ W.) to Point Possession (61o 02.1’N., 150 o 24.3’ W.). This area is bounded by Anchorage, the MatansuskaSusitna Borough, and the Kenai Peninsula Borough. Th is area contains sha llow tidal flats, river mouths or estuarine areas and is important as fo raging and calving habitats. Area 1 also has the highest concentrations of beluga whales in the spring through fall as well as the greatest potential for adverse impact from anthropogenic threats. Area 1 contains many rivers with large eulachon and salmon runs, including 2 rivers in Turnagai n Arm (Twenty-mile Ri ver and Placer River) which are visited by beluga whales in the early spring. Use declines in the summer and increases again in August through the fall, coinciding with coho salmon returns. Also included in Area 1

is Knik Arm and the Susitna delta. Area 2 consis ts of 5,891 square kilometers (2,275 sq. mi.) of Cook Inlet, located south of Area 1, north of a line at 60o 25.0’N., and includes nearshore areas south of 60o 25.0’N. along the west side of the Inlet and Kachemak Bay on the east side of the lower inlet. Area 2 is used by C ook Inlet beluga whales in a disp ersed fashion for fall and winter feeding and as transit waters. Area 2 includes ne ar and offshore areas of the m id and upper Inlet,

and nearshore areas of the lower Inlet. Area 2 includes Tuxedni, Chinitna, and Kamishak Bays on the west coast and a portion of Kachemak Bay of the east coast. Dive studies indicate that beluga whales in this area dive to deeper depths and are at the surface less frequently than they are when they inhabit Area 1. The primary consti tuent elements essential to the conservation of44 Cook Inlet beluga whales are: (1 ) intertidal and subtidal waters of Cook Inlet with depths <30 ft. (MLLW) and within 5 miles of high and medium flow accumulation anadromous fish streams; (2) primary prey species consisting of four species of Pacific salmon (Chinook, coho, sockeye, and chum salmon), Pacific eulachon, Pacific c od, walleye pollock, saffron cod, and yellowfin sole; (3) the absence of toxins or other agents of a type or amount harmful to beluga whales; (4) Unrestricted passage within or between the crit ical habitat areas; and (5 ) absence of in-water

noise at levels result in the ab andonment of habitat by Cook Inle t beluga whales. The comment period on this proposed rule closed on February 1, 2010.

Pinnipeds

Steller sea lion Description of the species. Steller sea lions are distributed along the rim of the North Pacific Ocean from San Miguel Island (C hannel Islands) off Southern Ca lifornia to northern Hokkaido, Japan (Loughlin et al. 1984, Nowak 2003). Their centers of abundance and distribution are in Gulf of Alaska and the Aleutian Islands (N MFS 1992). In the Bering Sea, the northernmost major rookery is on Walrus Island in the Pribilo f Island group. The northernmost major haul-out is on Hall Island off the northwestern tip of St. Matthew Island. Their distribution also extends

northward from the western end of the Aleutian chain to sites along the eastern shore of the Kamchatka Peninsula. For management purposes, two stocks have been designated, but which represent a single population. These stocks likely have some ta xonomic basis at the sub-species level in both genetics and skull morphology (Phillips et al. 2009).

Distribution. The eastern DPS of Steller sea lions in cludes animals east of Cape Suckling, Alaska (144  W) south to California waters (55 FR 49204). The western DPS of Steller sea lions includes animals west of Cape Suckling, Alaska (144  W; 62 FR 24345). However, individuals move between rookeries and haul out sites regularly, even over long distances between eastern and western DPS locations (Calki ns and Pitcher 1982a, Raum-Suryan et al. 2002, Raum-Suryan

et al. 2004). Most adult Stelle r sea lions occupy rookeries during the summer pupping and breeding season and exhibit a high level of site fidelity. During the breeding season, some juveniles and non-breeding adults occur at or near the rookeries, but most are on haulouts (sites that provide regular retreat from the water on exposed rocky shoreline, gravel beaches, and wave-cut platforms or ice; (Rice 1998a, Ba n 2005, Call and Loughlin 2005). Adult males may disperse widely after the breedi ng season. Males that breed in California move north after the breeding season and are rarely seen in Califor nia or Oregon except from May through August (Mate 1973). During fall and winter many sea lions disperse from rookeries and increase use of haulouts, particularly on terrestrial sites but also on sea ice in the Bering Sea.

Reproduction. Female Steller sea lions reach sexual ma turity and first breed between three and eight years of age and the average age of reprodu cing females (generation time) is about 10 years (Pitcher and Calkins 1981, Calkins and Pitcher 1982b, York 1994). Th ey give birth to a single pup from May through July and then breed about 11 days after giving birth. Females normally ovulate and breed annually after ma turity although there is a high rate of reproductive failures. The gestation period is believed to be about 50 to 51 weeks (Pitcher and C alkins 1981). The

available literature indicates an overall reproductive (b irth) rate on the order of 55% to 70% or greater (Pike and Maxwell 1958, Gentry 1970, Pitcher and Calkins 1981). However, natality45 was reported to be low in the western DPS in recent years (2003-2009; 69%) versus earlier years (43%); (Maniscalco et al. 2010). Survival through the first th ree weeks can be less than 50% at some sites, while others can be over 90% (Kaplan et al. 2008). Twinning has been reported (Maniscalco and Parker. 2009).

Mothers with newborn pups will make their first foraging trip about a week after giving birth, but trips are short in duration a nd distance at first, then increas e as the pup gets older (Merrick and Loughlin 1997, Milette 1999, Pitcher et al. 2001, Milette and Tr ites 2003, Maniscalco et al. 2006). Females attending pups tend to stay within 37 km of the rookery (Calkins 1996, Merrick and Loughlin 1997). Newborn pups are wholly depe ndent upon their mother for milk during at

least their first three months of life, and observations suggest they continue to be highly dependent upon their mother through their first winter (Scheffer 1945, Porter 1997, Trites et al. 2006). Generally, female Steller sea lion will nur se their offspring until they are one to two years old (Gentry 1970, Sandegren 1970, Pitche r and Calkins 1981, Calkins and Pitcher 1982b,

Trites et al. 2006). Pups may enter the water after 2-4 weeks (Sandegren 1970).

Males reach sexual maturity at about the same tim e as females (three to seven years of age, reported in (Loughlin et al. 1987)), but generally do not reach physical maturity and participate in breeding until about eight to ten years of ag e (Pitcher and Calkins 1981). The sex ratio of pups at birth is assumed to be about 1:1 or bias ed toward slightly greater production of males, but non-pups are biased towards females (Pik e and Maxwell 1958, Ca lkins and Pitcher 1982b, NMFS 1992, Trites and Larkin 1992, York 1994).

Habitat. Steller sea lions are not known to make regular migrations but do move considerable distances. Adult males may disperse hundreds of miles after the breeding season (Calkins and Pitcher 1982b, Calkins 1986, Loughlin 1997). Adult fe males may travel far out to sea into water greater than 1,000 m deep (Merri ck and Loughlin 1997). Studies on immature Steller sea lions indicate three types of moveme nts: long-range trips (greater than 15 km and greater than 20 hours), short-range trips (less than 15 km and less than 20 hours), and transits to other sites (NMFS 2007a). Long-range trips started around 9 months of age and likely occur most frequently around the time of weaning, while short-range trips happe n almost daily. Young individuals generally rema in within 480 km of rookeries thei r first year before moving further away in subsequent years (Raum-Suryan et al. 2004). Many animals also use traditional rafting sites, which are places where they rest on th e ocean surface in a tightly packed group (Bigg 1985)NMFS unpublished data). Frontal features with small-scale temperature gradients appear to be attractive foraging sites for juvenile Steller sea lions (Lander et al. 2010). Large numbers

of Steller sea lions are found near the 200 m isobath ye ar round (Consiglieri et al. 1982). Foraging generally occurs within 8-24 km of shore (Fiscus and Braham 1976). However, foraging can occur hundreds of kilometers from shore over extended periods (Merrick et al. 1997).

Six m ajor rookeries and numerous haulouts occur in the action area, some of which are in decline (Sease and Gudmundson 2002, Wynne et al. 2005).

Feeding. Steller sea lions are generalist predator s that eat various fish (arrowtooth flounder, rockfish, hake, flatfish, Pacific salmon, Pacific herring, Pacific cod, sand lance, skates, cusk eel,46 lamprey, walleye, Atka mackerel), squids, a nd octopus and occasionally birds and marine mammals (Jones 1981, Pitcher and Fay 1982, Calkins and Goodwin 1988, Olesiuk et al. 1990b, Daniel and Schneeweis 1992, Brown et al. 2002, Sinclair and Zeppelin 2002, McKenzie and Wynne 2008, Womble and Conlon. 2010). Diet is likely strongly influenced by local and temporal changes in prey distribution and abundance (McKenzie and Wynne 2008, Sigler et al. 2009). Haulout selection appears to be driven at least in part by local prey density (Winter et al. 2009). Adult females embark on foraging trips of at night for 7-26 hours during the breeding season, while adult males rarely or ne ver eat while on breeding grounds (Andrews et al. 2001, Loughlin 2002).

Status and trends. Steller sea lions were originally li sted as threatened under the ESA on November 26, 1990 (55 FR 49204), following a dec line in the US of about 64% over the previous three decades. In 1997, the species wa s split into two separate populations based on demographic and genetic differences (Bickham et al. 1996, Loughlin 1997), and the western population was reclassified to endangered (62 FR 24345) while the eastern population remained threatened (62 FR 30772). The Steller sea lion is also listed as endangered on the 2007 IUCN Red List (Group 1996).

Loughlin et al. (1984) estimated the worldwide populati on of Steller sea lions was between 245,000 and 290,000 animals (including pups) in the la te 1970s. Though the genetic differences

between the eastern and western DPSs were not known at the time, Loughlin et al. (1984) noted that 90% of the worldwide populat ion of Steller sea lions was in the western DPS in the early 1980s (75% in the US and 15% in Russia) and 10% in the eastern DPS. Loughlin et al. (1984)

concluded that the total worldw ide population size (both DPSs) was not significantly different from that estimated by Kenyon and Rice (1961) for the years 1959 and 1960, though the distribution of animals had change d. Steller sea lions collected in the Gulf of Alaska during the early 1980s showed evidence of reproductive failure and reduced rates of body growth that were

consistent with nutritional limitati on (Calkins and Goodwin 1988, Calkins et al. 1998, Pitcher et
al. 1998). After conducting a range-wide survey in 1989, Loughlin et al. (1992) noted that the worldwide Steller sea lion population had declin ed by over 50% in the 1980s, to approximately 116,000 animals, with the entire decline occu rring in the range of the western DPS.

The eastern stock seems to be performing better than the western stock. Trend counts in Oregon were relatively stable in th e 1980s, showing a gradual increase in numbers since 1976 (NMFS 2005d). Numbers in California, however, have declined to fewer th an 2,000 non-pups, from

counts between 1927 and 1947 that were as hi gh as 7,000 non-pups (NMFS 2005d). The count

from Central California in 2000 reached the se cond lowest recorded count of 349 non-pups (in 1992 the count was as low as 276 non-pups). In S outheast Alaska, counts of non-pups at trend sites increased by 56% from 1979 to 2002 from 6,376 animals to 9,951 (Sease et al. 2001, NMFS 2005d). Counts of non-pups at British Columbia trend s ites increased nearly 260% between 1982 and 2002 (NMFS 2005d).

The NMFS considers th is population stable or in creasing, and multiplies pup counts by a factor of 4.5 (based on (Calkins and Pitcher 1982b) or 5.1 (T rites and Larkin 1996) to estimate the total population size (Angliss and Outlaw 2008). Pup count data from 2002 through 2005 from across the range of the eastern population, multiplied by a factor of 4.5 or 5.1 results in a population47 estimate of 48,519 or 54,989 animals. In 2005, 5,510 pups were counted in Alaska, 3,318 pups were counted in British Columbia in 2002, 1,136 pups were counted in Oregon in 2002, and 818

counted in California in 2004. The curre nt minimum population ranges from 58,334-72,223 non-pup individuals (Allen and Angliss 2010). The NMFS calculates this estimate by adding non-pup counts taken in 2002 in Southeast Alaska , to counts of animals in Washington in 2002 as well as counts of pups and non-pups in Ca nada in 1998, Oregon in 2002, California in 2004, and southeastern Alaska in 2005 (Angliss and Outlaw 2008).

Estimated annual mortality is 0.22 for ages 0-2, dropping to 0.07 at age 3, then increasing gradually to 0.15 by age 10 and 0.20 by age 20 (York 1994). Population modeling suggests decreased juvenile survival likely played a major role in the decline of sea lions in the central Gulf of Alaska during 1975-1985 (Pascual a nd Adkison 1994, York 1994, Holmes and York 2003).

Natural threats. Killer whale predation, particular ly on the western DPS under reduced population size, may cause signifi cant reductions in the stock (N MFS 2008g). Sleeper sharks are also significant pr edators of Steller sea lions. Frid et al. (2009) suggested that risk of predation in nearshore waters by killer wh ales and offshore predation risk by sleeper sharks limited the use of Pacific herring in deep water and walleye Pollock in shallow water.

Steller sea lions have tested positive for severa l pathogens, but disease levels are unknown (FOC 2008). Similarly, parasites in this species are co mmon, but mortality resulti ng from infestation is unknown. However, significant negative effects of these factors may occur in combination with stress, which reduces immune capability to resist infections and infestations. If other factors, such as disturbance, injury, or difficulty feeding occur, it is more likely that disease and parasitism can play a greater role in population reduction.

Anthropogenic threats. Steller sea lions were historically and recently subjected to substantial mortality by humans, primarily due to co mmercial exploitation a nd both sanctioned and unsanctioned predator control, (Bonnot 1928, Rowley 1929, Scheffer 1945, Bonnot and Ripley 1948, Scheffer 1950, Pearson and Verts 1970, Bigg 1988, Atkinson et al. 2008, NMFS 2008g). Several dozen individuals may become entangl ed and drown in commercial fishing gear (Atkinson et al. 2008, NMFS 2008g). Several hundred individuals are removed by subsistence hunters annually in controlled and authorized harvests. Occasional ha rvest occur in Canada (FOC 2008). Additional mortality (362 from 1990 to 2003) has occurred from shooting of sea lions interfering in aquaculture opera tions along British Columbia (FOC 2008).

Significant concern also exists regarding competition between co mmercial fisheries and Steller sea lions for the same resource: stocks of pollo ck, Pacific cod, and Atka mackerel. Significant evidence exists that supports the western DPS declin ing as a result of change in diet and resulting declines in growth, birth rates, and su rvival (Calkins and Goodwin 1988, Calkins et al. 1998,

Pitcher et al. 1998, Trites and Donnelly 2003, Atkinson et al. 2008). As a result, limitations on fishing grounds, duration of fishing season, and monitoring have been established to prevent

Steller sea lion nutritional deficiencies as a result of inadequate prey availability.

Contaminants are a considerable issue for Steller sea lions. Roughly 30 individuals died as a48 result of the Exxon Valdez oil spill and contained particularly high levels of PAH contaminants, presumably as a result of the spill. Bl ood testing confirmed h ydrocarbon exposure. Subsequently, premature birth rates incr eased and pup survival decreased (Calkins et al. 1994, Loughlin et al. 1996). Organochlorines, including PCBs and DDT (and their metabolites), have

been identified in Steller sea li ons in greater concentrations th an any other pinniped during the 1980s, although levels appear to be declining (Barron et al. 2003, Hoshino et al. 2006). The

levels of PCBs have been found to have twice the burden in individuals from Russia than from western Alaska (4.3 ng/g wet weight versus 2.1 ng/g wet weight; (Myers et al. 2008). Levels of DDT in Russian pups were also on average twic e that in western Alaska pups (3.3 ng/g wet weight blood versus 1.6 ng/g wet we ight). PCB levels in the ki dneys of some adult males are high enough that reproductive and immune function may have been compromised (Wang et al. 2011). The source of contamination is likely fr om pollack, which have been found to contain organochlorines throughout the Gu lf of Alaska, but higher in re gions occupied by the eastern DPS of Steller sea lions (NM FS 2008g). Heavy metals, including mercury, zinc, copper, metallothionien, and butyltin have been identif ied in Steller sea lion tissues, but are in concentrations lower than other pinnipeds (Noda et al. 1995, Kim et al. 1996, Castellini 1999,

Beckmen et al. 2002, NMFS 2008g). Mercury may be of highe r significance, w ith liver levels being measured at levels above thos e necessary to impact fish (Holmes et al. 2008). However, contaminants leading to mortal ity in Steller sea lions have not been identified (NMFS 2008g).

Contaminant burdens are lower in females than ma les, because contaminants are transferred to

the fetus in utero as well as through lactation (Lee et al. 1996, Myers et al. 2008). However, this means that new generations tend to start with high er levels of contaminants than their parents originally had. Steller sea lion contaminants ar e of additional concern because contaminants in the body tend to be mobilized as fat reserves are us ed, such as when prey availability is low; a situation that is likely occurri ng for Steller sea lions today.

Critical habitat. Critical habitat was designated on August 27, 1993 for both eastern and western DPS Steller sea lions in California, Or egon, and Alaska (58 FR 45269). Steller sea lion critical habitat includes all major rookeries in California, Oregon, and Alaska as well as major

haulouts in Alaska and includes a 37 km buffer around these locations. Essential features of

Steller sea lion critical habitat include the physical and biological habitat features that support reproduction, foraging, rest, and refuge, and include terrestrial, air and a quatic areas. Specific

terrestrial areas include rookeries and haul -outs where breading, pupping, refuge and resting occurs. More than 100 major ha ulouts are docum ented. The prin cipal, essential aquatic areas

are the nearshore waters around rookeries and haul outs, their forage reso urces and habitats, and traditional rafting sites. Air z ones around terrestrial and aquatic ha bitats are also designated as

critical habitat to reduce disturban ce in these essential areas. Speci fic activities that occur within the habitat that may disrupt the essential life f unctions that occur there include: (1) wildlife viewing, (2) boat and airplane traffic, (3) research activities, (4) timber ha rvest, (5) hard mineral extraction, (6) oil and gas explor ation, (7) coastal development and pollutant discharge, and

others.

In addition, British Columbia has established prot ective areas in which St eller sea lion rookeries occur at Triangle Island and Cape St. James (C anada 2008). Several othe r haul-out sites occur within Canadian national and provincial parks. Further, the Canadian government is moving to

establish a marine wildlife area for the Scott Isla nds, where Steller sea lions haul-out and breed.49 Hawaiian monk seal Description of the species. Hawaiian monk seals are found pr imarily in the NWHI, which extend more than 2,000 km m iles northwest of the MHI

. Major breeding subpopulations occur at FFS, Laysan Island, Lisianski Island, Pearl and Hermes Reef, Midway Atoll, and Kure Atoll (Carretta et al. 2001)

. Smaller groups are found at Nihoa an d Necker Islands, seals have been observed at Gardner Pinnacles, Maro Reef, and Johnston Atoll, and several dozen seals are distributed throughout MHI (Carretta et al. 2001, NMFS 2007b)

. Midway was an important breeding rookery at one time, but is no longer used (Reeves 1992)

. However, all Hawaiian monk seals represent a single popul ation, with genetic connectiv ity high enough to maintain population-level genetic diffe rentiation {Schultz, 2011 #79451

. Reported sightings on each of the eight MHI have become increasingly common, and births have been reported on all of the MHI except Lanai and Hawaii

.

Habitat and feeding. Virtually all terrestrial substrat es, including emergent reefs and shipwrecks, are used by monk seals. Sandy beaches with shallow protected water near shore are the primary haul-out areas, for pupping, nursing, and resting, although pups are born on a variety of substrates (Gilmartin 1983). Seals use vegeta tion behind beaches as shelter from wind and

rain.

Pinniped movements are generally based on fora ging. Oceanographic features, such as thermal changes that might concentrate pr ey densities, can affect indivi dual seal foraging behavior (Field

et al. 2001)

. Hawaiian monk seal distribution, destinations , routes, food sources, and causes of movements when not traveling betw een islands are not well known

. Approximately 10-15% of Hawaiian monk seals migrate among the bree ding populations (Johnson and Kridler 1983)

. Interisland movement appears to be more likely when the islands are close together

. For example, movement between Kure Atoll, Midway Atoll, Pearl and Hermes Reef appear to be fairly common, while movement between FFS and Kure At oll (a distance of 2,000 km) is not known to

occur

. The western subpopulations (Pearl and Hermes Reef, Midway Islands, and Kure Atoll) exhibit a higher degree of migration compared to the more isolated s ubpopulations at Laysan, Lisianski, and FFS (NMFS 2007b; Table 3).

Table 5

. Migration rates per Hawaiian monk seal among subpopulations per year

. Data are from 1995- 2008, excluding translocations.

From/To Nihoa Necker FFS Laysan Lisianski PHR Midway Kure

Nihoa 0.7727 0 0.2273 0 0 0 0 0

Necker 0 0.4259 0.5741 0 0 0 0 0

FFS 0.0027 0.0065 0.9888 0.0019 0 0 0 0 Laysan 0 0 0.0043 0.9610 0.0334 0.0013 0 0

Lisianski 0 0 0.0011 0.0484 0.9434 0.0057 0.0007 0.0007

PHR 0 0 0.0004 0.003 0.0082 0.9450 0.0305 0.0130 Midway 0 0 0 0.0013 0 0.0707 0.7639 0.1641

Kure 0 0 0 0 0.0011 0.0168 0.0786 0.9035

At the breeding islands, monk seals feed on octo pus, spiny lobster, eels, and bottom-dwelling and reef fish (Gilmartin 1983; Goodman-Lowe 1998; Rice 1960). Considered foraging generalists, monk seals e xhibit significant differences in di et between islands, age, and sex groups (NMFS 2007b). No research or monitoring e ffort has been identified that will effectively50 measure or index monk seal prey abundance at the major breeding atolls (NMFS 2007b).

Seals forage within the barrier reefs of the atolls and on the leeward slopes of reefs and islands, as well as nearby seamounts and submerged r eefs and banks (Pullin and Stewart 2006). Several recent studies of the foraging patterns of Ha waiian monk seals near rookeries in the NWHI provide insight into their divi ng behavior. Dive de pths appear to differ slightly between rookeries as well as between ag e and sex classes. Stewart et al. (2006) found that throughout the six NWHI breeding colonies, most dives were less than 150 meters, but found some dives exceeding 550 meters. At Pearl and Hermes Reef, mo st dives reach 8-40 meters, with some dives

to three- to four-fold greater depths (Harington et al. 2004). However, at Laysan Island, this

trend was reversed and dives were mu ch deeper (800 to 1,150 feet; Harington et al. 2004). Most dives at FFS were to depths of 4-40 meters, but some dives exceeded 500 meters (Abernathy 1999). Parrish et al. (2002) noted a tendency to wards night diving at FFS.

Hawaiian monk seals tend to dive within the wa ter column, rather than to the sea floor, regardless of site (Pullin and Stewart 2006). Some work using Crittercams on seals at FFS indicates that most time spent underwater was fo r resting and socializi ng, not feeding. Despite

the reef fishes of the coral sh allows, adult seals forage on the slopes of the atoll and neighboring banks (Parrish et al. 2000). This is corroborated by the comparison between the diet composition of tagged seals and the composition of fish in each of four ecological zones (defined by depth)(Parrish and Abernathy 2006). Foraging has been shown to vary by age, with older juveniles (years 2 and 3) focusing on shallow at oll depths (10-30 meters) and yearlings feeding

in sand fields at 50-100 meters. It is possible that the shift in foraging behavior with age is dependent on the physical strength to flip small ro cks to find prey, rather than increasing dive duration or depth (Parrish et al. 2005).

Reproduction. Hawaiian monk seals do not form breed ing colonies or harems (Johanos et al. 1994; Kenyon and Rice 1959). Mating, which occurs in water and is rarely observed, is inferred from male-female association pattern s and from mounting injuries (Johanos et al. 1994). Breeding is asynchronous, lasting from February through September (Johanos et al. 1994). In recent years, fewer than 200 individuals ar e born annually (NMFS 2007b). Some islands, particularly in the southeastern Hawaiian Islands, have strong ma le-biased sex ratios, leading to female deaths from male aggression; a situat ion which has been sign ificantly corrected by selective male removal (5:1, up to 25:1)(Johanos et al. 2010).

Females typically give birth for the firs t time between ages of 5 and 10 (Antonelis et al. 2006). Pupping patterns vary greatly a nd not all females give birth in consecutive years (Johanos et al. 1994; Kenyon and Rice 1959). Females that do give birth in cons ecutive years pup later each season, while fe males that skip a year or more give birth earlier the next season. The mean

interval for births in consecutive years was found to be 381 days (Johanos et al. 1994). Birth rates vary depending on breeding location and year, with approximately 30-70% of all adult females giving birth in any given year (J ohanos and Ragen 1994, Rage n and Lavigne 1999). Females give birth from February to August, peaking in late Marc h/early April (Johanos et al. 1994), although pupping has been recorded year round. They prefer to give birth on beaches near

shallow water and coral reefs su rrounding the area, apparently to afford protection to the pup (Westlake and Gilmartin 1990).51 Newborn pups weigh 16-17 kg and measure 95-100 cm long (Kenyon and Rice 1959). Pups are black at birth and undergo a post-natal molt late in the nursing period. Nursing lasts, on average, 39 days (Johanos et al. 1994), during which time the mother remains constantly near her pup in and out of water (Kenyon and Rice 1959). The mo ther does not eat during nursing and rapidly loses weight (Gerrodette and DeMaster 1990). After weaning, the pup is abandoned to live off

of fat stores until it learns to feed on its own, while the mother swims offshore to feed (Gerrodette and Gilmartin 1990; Johnson and Johnson 1984; Kenyon and Rice 1959; Wirtz 1968). At weaning, pups normally weigh be tween 43-73 kg (Kenyon and Rice 1959). Rice (1964) suggested that adult females weigh a pproximately 205 kg and are about 2.3 m long, and the average adult male is smaller, at about 170 kg and 2.1 m.

Although nursing monk seal mothers generally avoi d other adult seals, occasional pup switches do occur (Johnson and Johnson 1978, Boness 1990), and mothers sometimes foster a pup if her own is lost (Alcorn and Henderson 1984, Gerrodette et al. 1992). If switched pups are of similar size, survival for the first year is minimally a ffected; however if a larger pup switches with a small one, the larger pup will have a longer nurs ing period and the smaller pup’s probability of survival will be reduced (NMFS 2007b).

Status and trends. The Hawaiian monk seal was listed as endangered under the ESA on November 23, 1976 (41 FR 51611). Hawaiian monk seals are considered one of the most endangered groups of pinnipeds on the planet becau se all of their populations are either extinct (Caribbean monk seal) or close to extinction (Medite rranean and Hawaiian monk seals). Two periods of anthropogenic decline have been reported; the first declin e occurred in the 1800s when sealers, crews of wrecked vessels, and guano and feather hunters nearly hunted monk seals to extinction (Dill and Bryan 1912, Kenyon and Rice 1959).

Following the initial collapse, expeditions to the NWHI reported increas ing seal numbers and partial recovery to slightly more than 1,000 individuals (Bailey 1952, Rice 1960). However, a

second decline occurred from the late 1950s to the mid-1970s; the population declined by roughly 50% by the 1980s (NMFS 1991). The total population in the FFS, Laysan Island, Lisianski Island, Pearl and Herm es Reef, Kure Atoll, and Midway, Necker, and Nihoa was estimated to be 1,501 in 1984, 1,976 seals in 1986, and 1,580 in 1992 (Ragen 1993). For the years 1985 to 1993 the mean beach counts declin ed by approximately 5% per year. This downward trend is expected to continue, mainly because of poor pup and juvenile survival in
recent years.

The best estimate of the total population of the species is 1,202 seals and the minimum population size estimate for the Hawaiian m onk seal is 1,176 seals (NMFS 2007b). Data collected in 2008 suggest that the species popu lation is now 1,146 (NMFS 2009a). A log-linear regression of estimated abundance from 1998 to 2006 suggests the population has declined on average -3.9% per year, and models predict that the total population of the species will fall below 1,000 monk seals within 5 years (NMFS 2007b). Tr ends in abundance vary considerably among the six main subpopulations.

A recent five-year status review conducte d by NMFS recommends that the Hawaiian monk52 seals’ endangered status should remain th e same (72 FR 46966, August 2007). The population dynamics at the different subpopulations have va ried considerably, and current demographic variability among the island populations probably refl ects a combination of di fferent histories of human disturbance and management (Gerrode tte and DeMaster 1990, Craig and Ragen 1999), and varying environmental conditions (Baker et al. 2007; Baker and Thompson 2007; Craig and Ragen 1999; Polovina et al. 1994). The current status of the Hawaiian monk seal is dire, due to low juvenile survival and the number of aging breeding females in the population. Consequently, NMFS is currently exploring deve lopment of a captive care program for juvenile

Hawaiian monk seals to enhan ce their potential for survival and recovery (NMFS 2007b).

The total of mean, non-pup, beach counts at th e main reproductive subpopulations in 2005 was approximately 67% lower than in 1958 (Benson et al. 2007b). A log-linear regression of estimated abundance from 1998 (the first year for which a reliable total abundance estimate was obtained) to 2006 estimates that abundance de clined by 3.9% annually (Figure 3)(NMFS 2007b).

Figure 3. Trends in abundance of Hawaiian monk seals at the six main NWHI subpopulations combined, 1998- 2006. This graph does not include abundance estimates for Necker, Nihoa, or the MHI. Error bars indicate ±2

standard errors or kn own minimum abundance. The fitted trend line re veals an estimated decline of 3.9% (NMFS 2007b).

Trends vary among the six main subpopulations. Non-pup beach counts at FFS have decreased by 73% from 1989 to 2005 (Benson et al. 2007b). At one time, FFS accounted for over 50% of the total non-pup beach counts among the NWHI subpopulations; however, that proportion has dropped to 25%, although FFS still maintains the single largest subpopulation (NMFS 2007b). The annual number of births has dropped from a high of 127 in 1988 to 39 in 2006, and survival from weaning to age two has declined from a hi gh near 90% in the mid-1980s to a low of 8% in 1997 (NMFS 2007b). Shark predation and prey av ailability are two potentially responsible

factors.53 Populations at Laysan and Lisianski Islands de clined sharply after th e late 1950s. In 1994, 21

adult male Hawaiian monk seals were relocated from Laysan Island to the MHI in an attempt to equalize the sex ratio at Laysan Island, and b each counts increased from 1995 to 2000, but have declined in the following years, while the Lisi anski subpopulation has remain ed relatively stable, yet low, since the 1970s. Marine debris and low fecundity are factors th at might contribute to the lack of subpopulation growth at Lisianski Island. And while the decline in abundance in Laysan may be related to female mortality cau sed by male aggression, j uvenile survival is relatively good for most cohorts, and the lack of recovery on Laysan is not understood (NMFS 2007b).

Until recently, the three westernmost subpopulations , Kure, Midway and Pearl and Hermes Reef exhibited substantial growth. Beach counts on Kure increased 5% per year from 1983 to 2000, declined in 2000-2001, and are now slowly increasi ng. At Midway, beach counts increased from 1995 to 2000, and have since declined. The subpopul ation at Pearl and Hermes Reef increased after the mid-1970s. Prior to 1999, beach count incr eases of up to 7% per year were observed. This is the highest estimate of the maximum ne t productivity rate obser ved for this species. Since 2000, low juvenile survival, thought to be due largely to f ood limitation, has been widespread with rare exceptions in the NWHI , resulting in the popul ation decline (Benson et al. 2007b), and several recent cohorts at the three westernmost sites indicate a drop in juvenile survival (NMFS 2007b).

Sightings of Hawaiian monk seals ha ve occurred on at least three oc casions at the remote Pacific location of Johnston Atoll (excluding nine adul t males relocated there from Laysan Island in 1984).

The decrease in survival rates of immature animal s, including a decline in survival from birth to weaning, and survival from weaning to age 2 year s has contributed to a dramatically imbalanced age structure for all six of the main NWHI subpopulations (Figure 4)(NMFS 2007b, 2009a). Although studies show that the relationship between size of pups and first year survival vary between subpopulations and over time, site-speci fic analyses do suppor t girth and year as predictors of first-year survival at each locat ion. When conditions for survival are worse, the relationship between size and surv ival strengthens. The simplest explanation fo r this is food limitation (Baker 2008).54 Figure 4. Age distribution for the Hawaiian monk seal population in the NWHI (MMRP unpublished data in NMFS 2009a).

Sightings and births are increas ing in the MHI, although systema tic surveys were not conducted before 2000, and counts do not represent total abunda nce, as they do not account for seals in the water, and not every seal on land is detected. In 2000, the count in the MH I was 45 seals, and in 2001, 52 were counted. In 2005, the total number of unique seals identified was 77, based on non-systematic sightings. Annual births have increased since the mid-1990s. Although this could be a positive indication for the survival of the species, the increased chance of contraction of diseases such as leptospirosis and toxopl asmosis from wild and domestic animals, and increased interactions with humans, including fi shermen, boaters, and divers raise conservation concerns which do not apply to the NWHI (NMFS 2007b). Th e only available estimate of abundance in the MHI is 152 individuals, with an annual population growth rate of 7% {Baker, 2011 #79839}. Survival to one year of age is 77% in the NHI, versus 42-57% in the NWHI (Baker et al. 2011).

Hawaiian monk seal pups weaned in the MHI exhib it higher girths and lengths compared to pups from the NWHI, as a result of pr e- and post-partum mate rnal investment – a partial reflection of prey availability – cont radicting the studies that indicate bett er foraging conditions in the NWHI. Suggested explanations for this include a higher per cap ita availability of prey in the MHI, similar absolute preferred prey densities when apex predators are not in cluded in the biomass, and increased prey availability due to redu ced competition from apex predators (Baker and Johanos 2004).55 Critical habitat. Critical habitat was or iginally designated on April 30, 1986 (51 FR 16047), and was extended on May 26, 1988 (53 FR 18988; CF R 226.201). The critical habitat includes all beach areas, sand spits and islets, including all beach crest vegetation to its deepest extent inland, lagoon waters, inner reef waters, and ocea n waters out to a depth of 20 fathoms (37 m) around the following: Kure Atoll (28°24 ′ N, 178°20 ′ W), Midway Islands, except Sand Island and its harbor (28°14 ′ N, 177°22 ′ W), Pearl and Hermes Reef (27°55 ′ N, 175° W), Lisianski Island (26°46 ′ N, 173°58 ′ W), Laysan Island (25°46 ′ N, 171°44 ′ W), Maro Reef (25°25 ′ N,

170°35 ′ W), Gardner Pinnacles (25°00 ′ N, 168°00 ′ W), French Frigate Shoals (23°45 ′ N, 166°00 ′ W), Necker Island (23°34 ′ N, 164°42 ′ W), Nihoa Island (23°03.5 ′ N, 161°55.5 ′ W). The marine component of this habitat was designated prim arily as feeding areas for Hawaiian monk seals, while terrestrial habitat serves as pupping a nd nursing habitat for mothers and pups. Both components are currently under significant degradation pressure. Because the marine critical habitat is in the Papah ā naumoku ā kea Marine National Monument, fishing is forbidden in the

critical habitat. A result of this is the estab lishment of large predatory fishes (sharks and jacks) that compete with Hawaiian monk seals for prey resources. This may be a cause of seal malnourishment seen throughout many islands in the region.

On June 2, 2011, NMFS proposed to extend critic al habitat in the NWHI to Sand Island (Midway) and ocean waters out to a depth of 500 (76 FR 27988). The area around the MHI (Kaula Island, Niihau, Kauai, Oahu, Maui Nui, and Hawaii) to a deth of 500 m and inland to a distance of 5 m from shore was also proposed.

Sea turtles

Leatherback sea turtle

Distribution. Leatherbacks range farther than any other sea turtle species, having evolved physiological and anatomical adaptations that allow them to exploit cold waters (Frair et al. 1972, Greer et al. 1973, USFWS 1995). High-lat itude leatherback range in cludes in the Atlantic includes the North and Barents Seas, Newfoundland and Labrador , Argentina, and South Africa (Threlfall 1978, Goff and Lien 1988, Márquez 1990, Hughes et al. 1998, Luschi et al. 2003,

Luschi et al. 2006). Pacific ranges extend to Alaska, Chile, and New Zealand (Gill 1997, Brito 1998, Hodge and W ing 2000).

Leatherbacks also occur in Mediterranean a nd Indian Ocean waters (Casale 2003, Casale et al. 2003, Hamann et al. 2006a). Associations ex ist with continental shelf and pelagic environments and sightings occur in offshore waters of 7-27  ̊ C (CETAP 1982). Juvenile leatherbacks usually stay in warmer, tropical waters >21  ̊ C (Eckert 2002). Males and females show some degree of natal homing to annual breeding sites (James et al. 2005).

Population designations. Leatherbacks break into four nesti ng aggregations: Pacific, Atlantic, and Indian oceans, and the Caribbean Sea. Detailed population structure is unknown, but is likely dependent upon nesting beach location.

Atlantic Ocean. Previous genetic analyses of leatherbacks using only mitochondrial DNA (mtDNA) resulted in an earlier determination that within the Atlantic ba sin there are at least three genetically different nesting populations: the St. Croix nesting population (US Virgin56 Islands), the mainland nesting Caribbean popula tion (Florida, Costa Rica, Suriname/French Guiana), and the Trinidad nesting population (Dutton et al. 1999). Further genetic analyses using microsatellite markers in nuclear DNA al ong with the mtDNA data and tagging data has

resulted in Atlantic Ocean leatherbacks now being divided into seven groups or breeding populations: Florida, Northern Caribbean, We stern Caribbean, Southern Caribbean/Guianas, West Africa, South Africa, and Brazil (TEWG 2007a).

Caribbean Sea. Nesting occurs in Puerto Rico, St . Croix, Costa Rica, Panama, Colombia, Trinidad and Tobago, Guyana, Suriname, and French Guiana (Márquez 1990, Spotila et al. 1996, Bräutigam and Eckert 2006b).

Indian Ocean. Nesting is reported in South Africa, India, Sri Lanka, and the Andaman and Nicobar islands(Hamann et al. 2006a).

Pacific Ocean. Leatherbacks are found from tropical wate rs north to Alaska within the North

Pacific and is the most common sea turtle in th e eastern Pacific north of Mexico (Stinson 1984a, Eckert 1993a, Wing and Hodge 2002). The west coast of Central America and Mexico hosts nesting from September-March, although Costa Rican nesting peaks during April-May (ChacónChaverri and Eckert 2007, LGL Ltd. 2007). Leat herback sea turtles disappeared from India before 1930, have been virtually extinct in Sri Lanka since 1994, and appear to be approaching extinction in Malaysia (Spotila et al. 2000). In Fiji, Thailand, a nd Australia, leatherback sea turtles have only been known to nest in low dens ities and scattered sites. Leatherback nesting aggregations occur widely in the Pacific, including China, Malaysia, Papua New Guinea, Indonesia, Thailand, Australia, Fiji, the Solo mon Islands, and Central America (Limpus 2002,

Dutton et al. 2007). Significant nesting also occurs along the Central American coast (Márquez 1990). Although not generally known to nest on Ja panese shores, two nest s were identified in the central Ryukyu Islands in 2002 (Kamezaki et al. 2002).

Nesting beaches also occur in Mexico and Cost a Rica (nesting occurs Oc tober through March) are a separate population from the western Pacific beaches. In Costa Rica, leatherbacks nest at Playa Naranjo in Santa Rosa National Park, th e second-most important nesting beach on the Pacific coast (Yañez et al. 2010), Rio Oro on the Osa Peninsula, and at various beaches in Las Baulas National Park, which includes Playa Langos ta and Playa Grande and contains the largest colony of leatherbacks in the Pacific (Spotila 20 04a). Females typically lay six clutches per season (average nine days between ne sts), which incubate for 58–65 days (Lux et al. 2003). Limited nesting also occurs along Nicaragua, Pa nama, El Salvador, Vanuatu, and Guatemala.

Growth and reproduction. It has been thought th at leatherbacks reach sexual maturity somewhat faster than other sea turtles (except Ke mp’s ridley), with an estimated range of 3-6 years (Rhodin 1985) to 13-14 years (Zug and Parh am 1996). However, recent research suggests otherwise, with western North Atlantic leathe rbacks possibly not maturing until as late as 29 years of age (Avens and Goshe 2007). Female l eatherbacks nest frequently (up to 10 nests per year and about every 2-3 years). During each nesting, females produce 100 eggs or more per clutch and 700 eggs or more pe r nesting season (Schultz 1975). However, up to ~30% of the eggs can be infertile. Thus, the actual proportion of eggs that can result in hatchli ngs is less than this seasonal estimate. The eggs incu bate for 55-75 days before hatching.57 Habitat. Leatherbacks occur throughout marine waters, from nearshore habitats to oceanic environments (Schroeder and Thompson 1987, Shoop and Kenney 1992a, Grant and Ferrell 1993, Starbird et al. 1993). Movements are largely de pendent upon reproductive and feeding cycles and the oceanographic features that conc entrate prey, such as frontal systems, eddy features, current boundaries, and coastal retention areas (Co llard 1990, Davenport and Balazs 1991, Frazier 2001, HDLNR 2002, Benson et al. 2011). Aerial surveys off the western US support continental slope waters as having greater leatherback occurrence than shelf waters (Green et al. 1992, Carretta and Forney 1993, Green et al. 1993, Bowlby et al. 1994). Nesting sites appear to be related to beaches with relatively high expo sure to wind or wind-generated waves (Santana Garcon et al. 2010).

Areas above 30o N in the Atlantic appear to be popular foraging locations (Fossette et al. 2009b). Northern foraging areas were proposed for wate rs between 35o and 50o N along North American, Nova Scotia, the Gulf of Saint-Laurent, in the western and northern Gulf Stream, the Northeast Atlantic, the Azores front and northeast of the Azores Islands, north of the Canary Islands. Southern foraging was proposed to occur betw een 5o and 15o N in the Mauritania upwelling, south of the Cape Verde islands , over the Guinea Dome area, a nd off Venezuela, Guyana and

Suriname.

Migration and movement. Leatherback sea turtles migrate throughout open ocean convergence zones and upwelling areas, along continental marg ins, and in archipelag ic waters (Morreale et al. 1994, Eckert 1998, 1999). In a single year, a l eatherback may swim more than 9,600 km to nesting and foraging areas throughout ocean basins (Eckert 1998, Ferraroli et al. 2004, Hays et
al. 2004, Eckert 2006, Eckert et al. 2006, Sale et al. 2006, Benson et al. 2007a, Benson et al. 2007c). Much of this travel may be due to m ovements within current and eddy features, moving individuals along (Sale and Luschi 2009). Return to nesting beaches may be accomplished by a form of geomagnetic navigation and use of local cues (Sale and Luschi 2009). Leatherback

females will either remain in nearshore wate rs between nesting events, or range widely, presumably to feed on available prey (Byrne et al. 2009, Fossette et al. 2009a).

Fossette et al. (2009b) identified three main migratory strategies in leatherbacks in the North Atlantic (almost all of studied individuals were female). One involved 12 individuals traveling to northern latitudes during summer/fall and retu rning to waters during winter and spring. Another strategy used by six individuals was similar to this, but instead of a southward movement in fall, individuals overwintered in northern latitudes (30-40o N, 25-30o W) and moved into the Irish Sea or Bay of Biscay dur ing spring before moving south to between 5 and 10o in winter, where they remained or returned to the northwest Atlantic. A third strategy, which was followed by three fem ales remaining in tropical waters for the first year subsequent to nesting and moving to northern latitudes during summer/fall and sp ending winter and spring in

latitudes of 40-50o N.

Satellite tracking data reveal that leatherback females leavi ng Mexican and Central American nesting beaches migrate towards the equator and into Southern Hemisphere waters, some passing the Galápagos Islands, and disperse south of 10oS (Dutton et al. 2006, Shillinger et al. 2010). However, observations of leatherbacks in the Galápagos Islands are rare (Zárate et al. 2010). Nesting site selection in the s outhwest Pacific appears to favor sites with higher wind and wave58 exposure, possibly as a means to aid hatchling dispersal (Garcon et al. 2010). Individuals nesting in Malayasia undergo migrations to tropic al feeding areas, taking 5-7 months to arrive there from nesting locations (Benson et al. 2011). Additional foraging occurs in temperate locations, including across the Pacific basin aloi ng the US west coast; individuals take 10-12 months to migrate here (Benson et al. 2011). Individuals nesti ng during the boreal summer move to feeding areas in the North China Sea, while boreal winter nesters moved across the Equator to forage in the Southern Hemisphere (Benson et al. 2011).

Sex ratio. A significant female bias exists in all l eatherback populations t hus far studied. An examination of strandings and in-water sighting da ta from the US Atlantic and Gulf of Mexico

coasts indicates that 60% of individuals were female. Studies of Suriname nesting beach temperatures suggest a female bias in hatchlings , with estimated percentages of females hatched over the course of each season at 75.4, 65.8, and 92.2% in 1985, 1986, and 1987, respectively (Plotkin 1995). Binckley et al. (1998) found a heavy female bias upon examining hatchling gonad histology on the Pacific coast of Costa Rica, and estimated male to female ratios over three seasons of 0:100, 6.5:93.5, and 25.7:74.3. James et al. (2007) also found a heavy female bias (1.86:1) as well as a primarily large sub-a dult and adult size distribution. Leatherback sex determination is affected by nest temperature, with higher temperatur es producing a greater proportion of females (Mrosovks y 1994, Witzell et al. 2005b).

Feeding. Leatherbacks may forage in high-invertebra te prey density areas formed by favorable features (Ferraroli et al. 2004, Eckert 2006). Although leatherb acks forage in coastal waters, they appear to remain primarily pe lagic through all lif e stages (Heppell et al. 2003b). The location and abundance of prey, including medus ae, siphonophores, and salpae, in temperate and boreal latitudes likely ha s a strong influence on leatherback di stribution in these areas (Plotkin 1995). Leatherback prey are frequently found in th e deep-scattering layer in the Gulf of Alaska (Hodge and Wing 2000). North Pacific foraging gr ounds contain individual s from both eastern and western Pacific rookeries, al though leatherbacks from the easte rn Pacific generally forage in the Southern Hemisphere al ong Peru and Chile (Dutton et al. 1998, Dutton et al. 2000, Dutton 2005-2006). Mean primary productivity in all fora ging areas of western Atlantic females is 150% greater than in eastern Paci fic waters, likely resulting in twice the reproductive output of eastern Pacific females (Saba et al. 2007). Leatherbacks have been observed feeding on jellyfish in waters off Washington State and Ore gon (Eisenberg and Frazier 1983, Stinson 1984a).

Status and trends. Leatherback sea turtles received protection on June 2, 1970 (35 FR 8491) under the Endangered Species Cons ervation Act and, since 1973, have been listed as endangered under the ESA, but declines in nesting have continued worldwide. Breeding females were initially estimated at 29,000-40,000, but were late r refined to ~115,000 (Pritchard 1971, 1982). Spotila et al. (1996) estim ated 34,500 females, but la ter issued an update of 35,860 (Spotila 2004b). The species as a whole is declining and local populations are in danger of extinction (NMFS 2001b, a).

Nesting aggregations occur along Gabon, Sao Tome and Principe, French Guiana, Suriname, and Florida (Márquez 1990, Spotila et al. 1996, Bräutigam and Eckert 2006b) . Widely dispersed but fairly regular African nesting also occurs between Mauritania and Angola (Fretey et al. 2007). Many sizeable populations (perha ps up to 20,000 females annually) of leatherbacks are known to59 nest in West Africa (Fretey 2001b). The popula tion of leatherbacks nesting on Gabon beaches has been suggested as being the world’s la rgest, with 36,185-126,480 clutches being laid by 5,865-20,499 females annually from 2002-2007 (Witt et al. 2009). The total number of females utilizing Gabon nesting beaches is estimated to be 15,730- 41,373 (Witt et al. 2009). North

Atlantic leatherback s likely number 34,000-94,000 individual s, with females numbering 18,800 and the eastern Atlantic segm ent numbering 4,700 (TEWG 2007a). Trends and numbers include only nesting females and are not a complete de mographic or geographic cross-section. In 1996, the entire Western Atlantic population was characterized as stab le at best (Spotila et al. 1996), with roughly 18,800 nesting females. A subsequent analysis indicated that by 2000, the western Atlantic nesting population had decreased to about 15,000 nesting females (NMFS 2011l). Spotila et al. (1996) estimated that the entire Atlantic basin, including all ne sting beaches in the Americas, the Caribbean, and West Africa, tota led approximately 27,600 nesting females, with an estimated range of 20,082-35,133. This is consistent with other estimates of 34,000-95,000 total adults (20,000-56,000 adult females; 10,000-21,000 nesting females)(TEWG 2007b).

The largest nesting aggregation in the western North Atlantic occurs in French Guiana and Suriname, likely belongs to a metapopulat ion whose limits remain unknown (Rivalan et al. 2006). Heppell et al. (2003a) concluded that leatherbac ks generally show less genetic structuring than green and hawksbill sea turtles. The French Guiana nesting aggregation has declined ~15% annually since 1987 (NMFS 2001a). However, from 1979-1986, the number of nests increased ~15% annually, possibly indicatin g the current decline may be linked with the erosion cycle of Guiana beaches (NMFS 2006e). Guiana nesting may have increased again in the early 2000s (NMFS 2006e). Suriname nestin g numbers have recently increased from more than 10,000 nests annually since 199 9 and a peak of 30,000 nests in 2001. Overall, Suriname and French Guiana nesting trends towards an increase (Hilterman and Goverse 2003, Girondot et
al. 2007). Florida (March-July) and US Caribbean nesting sin ce the early 1980s has increased ~0.3% and 7.5% per year, respectively, but lags be hind the French Guiana coast and elsewhere in magnitude (NMFS/SEFSC 2001). This positive grow th was seen within major nesting areas for the stock, including Trinidad, Guyana, and the combined beaches of Suriname and French Guiana (TEWG 2007b). Using both Bayesian m odeling and regression analyses, the TEWG (2007b) determined that the Southern Caribbean/G uianas stock had demonstrated a long-term, positive population growth rate (using nes ting females as a proxy for population).

The Caribbean coast of Costa Rica and extending through Chiriquí Beach, Panama, represents the fourth largest known leatherback rookery in the world (Troeng et al. 2004). Examination of data from three index nesting beaches in the region (Tortuguero, Gandoca, and Pacuare in Costa Rica) using various Bayesian a nd regression analyses indicated that the nesting population likely was not growing during 1995-2005 (T EWG 2007b). Other modeling of the nesting data for Tortuguero indicates a 67.8% decl ine between 1995 and 2006 (Troëng et al. 2007).

In Puerto Rico, the primary nest ing beaches are at Fajardo and on the is land of Culebra. Nesting between 1978 and 2005 ranged between 469-882 nest s, and the population has been growing since 1978, with an overall annual growth rate of 1.1% (TEWG 2007b). At the primary nesting beach on St. Croix, the Sandy Point National Wildlif e Refuge, nesting has fluctuated from a few hundred nests to a high of 1,008 in 2001, and the average annual growth rate has been approximately 1.1% from 1986-2004 (TEWG 2007b).60 The Florida nesting stock comes ashore primarily al ong the east coast of Florid a. This stock is of growing importance, with total nests between 800-900 per year in the 2000s following nesting totals fewer than 100 nests per year in the 1980s (NMFS 2011l). Using data from the index nesting beach surveys, the TEWG (2007b) estimate d a significant annual nesting growth rate of

1% between 1989 and 2005. Stewart et al. (2011) evaluated nest counts from 68 Florida beaches over 30 years (1979-2008) and found that nesting in creased at all beaches with trends ranging from 3.1%-16.3% per year, with an overall incr ease of 10.2% per year. In 2007, a record 517

leatherback nests were observed on the index beaches in Florida, with 265 in 2008, and then an increase to a new record of 615 nests in 2009, and a slight decline in 2010 back to 552 nests (FWC Index Nesting Beach database). This up-a nd-down pattern is thought to be a result of the cyclical nature of leatherback nesting, similar to the biennial cycle of green turtle nesting.

The most recent population estimate for leatherbac k sea turtles from the North Atlantic as a whole is between 34,000-90,000 adult individuals (20,000-56,000 adult females)(TEWG 2007b).

Heavy declines have occurred at all major Pacific basin rookeries, as well as Mexico, Costa Rica, Malaysia, India, Sri Lanka, Thailand, Trinidad, Tobago, and Papua New Guinea. This includes a nesting decline of 23% between 1984-1996 at Mexiquillo, Michoacán, Mexico (Sarti et al. 1996). According to reports from the late 1970 s and early 1980s, three beaches on the Pacific coast of Mexico supported as many as half of all leatherback turtle nests for the eastern Pacific. Since the early 1980s, the eastern Pacific Mexican population of adu lt female leatherback turtles has declined to slightly more than 2 00 individuals during 1998-1999 and 1999-2000 (Sarti et al. 2000). Spotila et al. (2000) reported the dec line of the leatherback tu rtle population at Playa Grande, Costa Rica, which had been the fourth largest nesting colony in the world. Between 1988 and 1999, the nesting colony declined from 1,367 to 117 female leatherback turtles. Based on their models, Spotila et al. (2000)estimated that the colony c ould fall to less than 50 females by 2003-2004. Fewer than 1,000 females nested on the Pacific coast of Mexico from 1995-1996

and fewer than 700 females are estim ated for Central America (Spotila et al. 2000). The number of leatherback turtles nesting in Las Baulas National Park de clined rapidly during the 1990s, from about 1,500 females during the 1988–89 nes ting season, to about 800 in 1990–91 and 1991–92 to 193 in 1993–94 (Williams et al. 1996) and 117 in 1998–99 (Spotila et al. 2000). Spotila (2004a) reported that between 59 and 435 leatherbacks nest at Las Baulas each year depending on the El Niño–La Niña cycle. Only an Indonesian nesting assemblage has remained relatively abundant in the Pacific basin. The largest exta nt leatherback nesti ng assemblage in the Indo-Pacific lies on the northern Vogelkop coast of Irian Jaya (West Papua), Indonesia, with roughly 3,000 nests recorded annually (Putrawidjaja 2000, Suárez et al. 2000) (Dutton et al. 2007). The Western Pacific leatherback metapopul ation harbors the last remaining nesting aggregation of significant size in the Pacific with approximately 2700–4500 breeding females (Dutton et al. 2007, Hitip euw et al. 2007). The total number of ne sts per year for the JamursbaMedi leatherback nesting population ranged between a high of 6,373 nests in 1996 and a low of 1,537 nests in 2010 (Hitipeuw et al. 2007).

Declines in the western Pacific is equally severe . Nesting at Terengganu, Ma laysia is 1% of that in 1950s (Chan and Liew 1996). The South China Sea and East Pacific nesting colonies have undergone catastrophic collapse. Overall, Pacifi c populations have declined from an estimated61 81,000 individuals to <3,000 total adults and subadults (Spotila et al. 2000). The number of nesting leatherbacks has declined by an estima ted 95% over the past 20 years in the Pacific (Gilman 2009). Drastic overharvesting of eggs a nd mortality from fishing activities is likely responsible for this tremendous decline (Sarti et al. 1996, Eckert 1997).

Based on the survey and tagging work, it was esti mated that 400-500 female leatherbacks nest annually on Great Nicobar Island (Andrews et al. 2002). The number of nesting females using the Andaman and Nicobar Islands combined was estimated around 1,000 (Andrews and Shanker 2002).

Natural threats. Sea turtles face predation primarily by sh arks and to a lesser extent by killer whales (Pitman and Dutton 2004). Hatchlings are preyed upon by herons, gulls, dogfish, and sharks. Leatherback hatching success is particularly sensitive to nesting site selection, as nests that are overwashed have significantly lower hatchi ng success and leatherbacks nest closer to the high-tide line than other sea turtle species (Caut et al. 2009b).

Anthropogenic threats. Leatherback nesting and marine environm ents are facing increasing impacts through widespread development a nd tourism along nesting beaches (Hamann et al. 2006a, Maison 2006, Hernandez et al. 2007, Santidrián Tomillo et al. 2007). Structural impacts to beaches include building and piling construc tion, beach armoring and renourishment, and sand extraction (Lutcavage et al. 1997b, Bouchard et al. 1998). In some areas, timber and marine

debris accumulation as well as sand mining reduce avai lable nesting habitat (Chacón Chaverri 1999, Formia et al. 2003, Laurance et al. 2008, Bourgeois et al. 2009). Lights on or adjacent to nesting beaches alter nes ting adult behavior and is often fatal to emerging hatchlings as they are drawn to light sources and away from the s ea (Witherington and Bjorndal 1991, Witherington 1992, Cowan et al. 2002, Deem et al. 2007, Bourgeois et al. 2009). Plastic ingestion is very

common in leatherbacks and can block gastrointestin al tracts leading to death (Mrosovsky et al. 2009). Along the coast of Peru, 13% of 140 leathe rback carcasses were found to contain plastic bags and film (Fritts 1982). Although global warm ing may expand foraging habitats into higher latitude waters, increasing temperatures may increase feminization of nests (Mrosovsky et al. 1984, James et al. 2006, McMahon and Hays 2006, Hawkes et al. 2007b). Rising sea levels may also inundate nests on some beaches. Egg collecti on is widespread and attributed to catastrophic declines, such as in Malaysia. Harvest of females along nesting beaches is of concern

worldwide.

Bycatch, par ticularly by longline fisheries, is a major source of mortality for leatherback sea turtles (Crognale et al. 2008, Gless et al. 2008, Fossette et al. 2009a, Petersen et al. 2009).

Wallace et al. (2010) estimated that between 1990 a nd 2008, at least 85,000 sea turtles were captured as bycatch in fisheries worldwide. Th is estimate is likely at least two orders of magnitude low, resulting in a likely bycatch of nearly half a million sea turtles annually (Wallace

et al. 2010); many of these turtles are expected to be leatherb acks. Donoso and Dutton (2010) found that 284 leatherbacks were bycaught between 2001 and 2005 as part of the Chilean longline fishery, with two indivi duals observed dead; leatherbacks were the most frequently bycaught sea turtle species. Between 8-17 leatherback turtles likely died annually between 1990 and 2000 in interactions with the California/Oreg on drift gillnet fishery; 500 leatherback turtles are estimated to die annually in Chilean and Peruvian fisheries; 200 leatherback turtles are62 estimated to die in direct harvests in Indone sia; and, before 1992, the North Pacific driftnet fisheries for squid, tuna, and bill fish captured an estimated 1,000 leatherback turtles each year, killing about 111 of them each year . An estimated 6,363 leatherback sea turtles were documented as caught by the US Atlantic tuna and swordfish longlin e fisheries between 1992- 1999 (NMFS SEFSC 2001). Currently, the US tuna and swordfish longline fisheries managed under the HMS FMP are estimated to capture 1,764 leatherbacks (no more than 252 mortalities) for each 3-year period starting in 2007 (NMFS 2004). In 2010, there were 26 observed

interactions between leat herback sea turtles and l ongline gear used in th e HMS fishery (Garrison and Stokes 2011a, 2011b). All leatherbacks were rele ased alive, with all gear removed for the majority of captures. While 2010 total estimat es are not yet available, in 2009, 285.8 (95% CI: 209.6-389.7) leatherback sea turtles are estimated to have been taken in the longline fisheries managed under the HMS FMP based on the obs erved takes (Garrison and Stokes 2010).

Lewison et al. (2004) estimated that 30,000-60,000 leatherb acks were taken in all Atlantic longline fisheries in 2000 (including the US Atlantic tuna an d swordfish longline fisheries, as well as others). Finkbeiner et al. (2011) estimated that annual bycatch intera ctions total 1,400 individuals annually for US Atlantic fisheries (r esulting in roughly fourty mortalities) and one

hundred interactions in US Pacifi c fisheries (resulting in about ten mortalities). Mortality of leatherbacks in the US shrimp fishery is now estimated at 54 turtles per year. Data collected by the NEFSC Fisheries Observer Program from 1994 through 1998 (excluding 1997) indicate that a total of 37 leatherbacks were incidentally capture d (16 lethally) in drift gillnets set in offshore

waters from Maine to Florida dur ing this period. Observer covera ge for this period ranged from 54 to 92%. Trinidad and Tobago's Institute for Marine Affairs estimated that more than 3,000 leatherbacks were captured incidental to gillnet fishing in the coas tal waters of Trinidad in 2000. Half or more of the gravid tu rtles in Trinidad and Tobago wate rs may be killed (Lee Lum 2003), though many of the turtles do not die as a result of drowning, but rather because the fishermen butcher them in order to get th em out of their nets (NMFS 2001b).

We know little about the effects of contaminants on leatherback sea turtles. The metals arsenic, cadmium, copper, mercury, selenium, and zinc bioaccumulate, with cadmium in highest

concentration in leatherbacks versus any other marine vertebrate (Gordon et al. 1998, Caurant et
al. 1999). A diet of primarily jellyfish, which ha ve high cadmium concentr ations, is likely the cause (Caurant et al. 1999). Organochlorine pesticides have also been found (Mckenzie et al. 1999). PCB concentrations are reportedly equiva lent to those in some marine mammals, with liver and adipose levels of at least one congener being exceptiona lly high (PCB 209: 500-530 ng/g wet weight Davenport et al. 1990, Oros et al. 2009).

Critical habitat. On March 23, 1979, leatherback critical habitat was identified adjacent to Sandy Point, St. Croix, US Virgin Islands fr om the 183 m isobath to mean high tide level between 17° 42’12” N and 65°50’00” W (44 FR 17710) . This habitat is essential for nesting, which has been increasingly threatened sin ce 1979, when tourism increased significantly, bringing nesting habitat and peopl e into close and frequent proxi mity. However, studies do not currently support signif icant critical hab itat deterioration.

On Janauary 5, 2010, the NMFS proposed to designate critical habitat for leatherback sea turtles in waters along Washington State (Cape Flattery to the Umpqua River; 63,455 km 2 ) and California (Point Arena to point Vincente; 119,400 km 2 ). The primary constituent elements of63 these areas include (1.) the occurrence of pr ey species, primarily scyphomedusae of the order Semaeostomeae ( Chrysaora, Aurelia, Phacellophora, and Cyanea ) of sufficient condition, distribution, diversity, and a bundance to support individual as well as population growth, reproduction, and development and (2) migratory pathway conditions to allow for safe and timely passage and access to/from/within high use foraging areas.

Hawksbill sea turtle

Distribution. The hawksbill has a circumglobal distribution throughout tropical and, to a lesser extent, subtropical waters of th e Atlantic, Indian, and Pacific o ceans. Satellite tagged turtles have shown significant variation in movement and migration patterns. In the Caribbean, distance traveled between nesting and foraging locations ranges from a few kilometers to a few hundred kilometers (Byles and Swimmer 1994, Miller et al. 1998, Hillis-Starr et al. 2000, Horrocks et al. 2001, Prieto et al. 2001, Lagueux et al. 2003).

Population designation. Populations are distinguished generally by ocean basin and more specifically by nesting location. Our understanding of population structure is relatively poor. For example, genetic analysis of hawksbill sea turtles foraging off the Cape Verde Islands identified three closely-related ha plotypes in a large majority of individuals sampled that did not match those of any known nesting population in the Western Atlantic, where the vast majority of nesting has been documented (McClellan et al. 2010, Monzon-Arguello et al. 2010).

Migration and movement. Upon first entering the sea, neonatal hawksbills in the Caribbean are believed to enter an oceanic phase that may invol ve long distance travel and eventual recruitment to nearshore foraging habitat (Boulon 1994). In the marine environment, the oceanic phase of juveniles (i.e., the "lost years") remains one of the most poorly understood aspects of hawksbill life history, both in terms of wher e turtles occur and how long they remain oceanic. Nesting site

selection in the southwest Paci fic appears to favor sites with higher wind and wave exposure, possibly as a means to aid hatchling dispersal(Garcon et al. 2010).

Habitat. Hawksbill sea turtles are highly migratory and use a wide range of broadly separated

localities and habitats during their lifetimes (Musick and Limpus 1997, Plotkin 2003). Small juvenile hawksbills (5-21 cm st raight carapace length) have b een found in association with Sargassum spp. in both the Atlantic and Pacific oceans (Musick and Limpus 1997) and observations of newly hatched hawksbills attracte d to floating weed have been made (Hornell 1927, Mellgren et al. 1994, Mellgren and Mann 1996). Post-oceanic hawksbills may occupy a range of habitats that include coral reefs or ot her hard-bottom habitats, sea grass, algal beds, mangrove bays and creeks (Musick and Limpus 1997, Bjorndal and Bolten 2010), and mud flats (R. von Brandis, unpublished data in NMFS and US FWS 2007e). Eastern Pacific adult females have recently been tracked in saltwater mangr ove forests along El Salvador and Honduras, a habitat that this species was not previously known to occupy (Gaos et al. 2011). Individuals of multiple b reeding locations can occupy the same foraging habitat (Bowen et al. 1996, Bass 1999,

Diaz-Fernandez et al. 1999, Bowen et al. 2007, Velez-Zuazo et al. 2008). As larger juveniles, some individuals may associate with the same feeding locality for more than a decade, while others apparently migrate from one site to another (Musick and Limpus 1997, Mortimer et al. 2003, Blumenthal et al. 2009). Larger individuals may prefer deeper habitats than their smaller counterparts (Blumenthal et al. 2009). Nesting sites appear to be related to beaches with64 relatively high exposure to wind or wi nd-generated waves (Santana Garcon et al. 2010).

Hawksbill sea turtles appear to be rare visitors to the Gulf of Mexico, with Florida being the only Gulf state with regular sightings (Rabalai s and Rabalais 1980, H ildebrand 1983, Witzell 1983, NMFS and USFWS 1993, Rester and Condrey 1996) . Individuals stranded in Texas are generally young (hatchlings or y earlings) originating from Mexi can nesting beaches (Hildebrand 1983, Amos 1989, Collard and Ogren 1990, Landry and Costa 1999).

Growth and reproduction. The best estimate of age at sexual maturity for hawksbill sea turtles is 20-40 years (Chaloupka and Limpus 1997, Crouse 1999). Reproductive females undertake periodic (usually non-annual) migrations to their natal beaches to nest. Movements of reproductive males are less well know n, but are presumed to involve migrations to their nesting beach or to courtship stations along the migrat ory corridor (Meylan 1999). Females nest an average of 3-5 times per season (Meylan and Donnelly 1999, Richardson et al. 1999). Clutch size up to 250 eggs; larger than that of other sea turtles (Hirth 1980). Reproductive females may exhibit a high degree of fidel ity to their nest sites.

The life history of hawksbills cons ists of a pelagic stage that la sts from hatching until they are approximately 22-25 cm in straight carapace length (Meylan 1988, Meylan and Donnelly 1999), followed by residency in coastal developmental habitats.

Feeding. Dietary data from oceanic stage hawksbills are limited, but indicate a combination of plant and animal mate rial (Bjorndal 1997b).

Status and trends. Hawksbill sea turtles received protection on June 2, 1970 (35 FR 8495) under the Endangered Species Cons ervation Act and since 1973 have been listed as endangered under the ESA. Although no historical record s of abundance are known, hawksbill sea turtles are considered to be severely depleted due to the fragmentation and low use of current nesting beaches (NMFS and USFWS 2007e). Worldwid e, an estimated 21,212-28,138 hawksbills nest each year among 83 sites. Among the 58 sites for w ith historic trends, all show a decline during the past 20 to 100 years. Among 42 sites for whic h recent trend data are available, 10 (24%) are increasing, three (7%) are stable and 29 (69%) are decreasing. Encouragingly, nesting range along Mexico and Central America appears not to have contracted and estimates continue to increase as additional dedicated study is conducted in the eastern Pacific (Gaos et al. 2010a).

Atlantic Ocean. Atlantic nesting sites include: Antigua (Jumby Bay), the Turks and Caicos, Barbados, the Bahamas, Puerto Rico (Mona Is land), the US Virgin Islands, the Dominican Republic, Sao Tome, Guadaloupe, Trinidad and Tobago, Jamaica, Martinique, Cuba (Doce Leguas Cays), Mexico (Yucatan Peninsula), Cost a Rica (Tortuguero National Park), Guatemala, Venezuela, Bijagos Archipelago, Guinea-Bissau, and Brazil.

Population increase has been grea ter in the Insular Caribbean th an along the Western Caribbean

Mainland or the eastern Atlant ic (including Sao Tomé and E quatorial Guinea). Nesting populations of Puerto Rico appeared to be in decline until the early 1990s , but have universally increased during the survey pe riods. Mona Island now hosts 199-332 nesting females annually, and the other sites combined host 51-85 nesting females annually (R.P. van Dam and C.E. Diez,65 unpublished data in NMFS and USFWS 2007e) C.E. Diez, Chelonia, Inc., in litt. to J. Mortimer 2006). The US Virgin Islands have a long histor y of tortoiseshell trade (Schmidt 1916). At Buck Island Reef National Monument, protection has been in force since 1988, and during that time, hawksbill nesting has increased by 143% to 56 nesting females annually, with apparent spill over to beaches on adjacent St. Croix (Z. Hillis -Starr, National Park Service, in litt. to J. Mortimer 2006). However, St. John populations did not increase, perhaps due to the proximity of the legal turtle harvest in the British Virgin Islands (Z. Hillis-Starr, National Park Service, in litt. to J. Mortimer 2006). Populations have also been identified in Belize and Brazil as genetically unique (Hutchinson and Dutton 2007). An estimated 50-200 nests are laid per year in the Guinea-Bissau (Catry et al. 2009).

Pacific Ocean. American Samoa and Western Samoa host fewer than 30 females annually (Tuato'o-Bartley et al. 1993, Grant et al. 1997). In Guam, only 5-10 females are estimated to nest annually (G. Balazs, NMFS, in litt. to J. Mo rtimer 2007; G. Davis, NMFS, in litt. to J. Mortimer 2007) and the same is true for Hawaii, but there are indications that this population is increasing (G. Balazs, pers. comm. in NMFS a nd USFWS 2007e). Addi tional populations are

known from the eastern Pacifi c (potentially extending from Mexico through Panama), northeastern Australia, and Mala ysia (Hutchinson and Dutton 2007) . El Salvador is now known to host the majority of hawksbill turtle nesting activity in the eastern Pacific, with 79.6% (n= 5430) of all nesting observation records, and Mexico hosting the majority of records of hawksbill turtles at sea, with 60.3% (n= 544) of all in-water observation records (Gaos et al. 2010b). Total number of nesting females for the Central P acific hawksbill population was estimated at 940 – 1,200 females annually for the last few years, with an overall downward trend (NMFS and USFWS 2007b).

Indian Ocean. The Indian Ocean hosts several populati ons of hawksbill sea turtles (Spotila 2004a, Hutchinson and Dutton 2007). These include western Australian, Andaman and Nicobar islands, Maldives, Seychelles, Burma, East Af rica, Egypt, Oman, Saudi Arabia, Sudan, and

Yemen.

Natural threats. Sea turtles face predation primarily by sh arks and to a lesser extent by killer whales. All sea turtles except leatherbacks can undergo “cold stunning” if water temperatures drop below a threshold level, which can be lethal . The only other significa nt natural threat to hawksbill sea turtles is from hybridization of hawksbill s with other species of sea turtles. This is especially problematic at certain sites where hawksbill numbers are particularly low (Mortimer and Donnelly in review). Predators (primarily of eggs and hatchlings) include dogs, pigs, rats, crabs, sea birds, reef fishes, gr oupers, feral cats, and foxes (Bell et al. 1994, Ficetola 2008). In some areas, nesting beaches can be almost comple tely destroyed and all nests can sustain some level of depredati on (Ficetola 2008).

Anthropogenic threats. Threats to hawksbill sea turtles are largely anthropogenic, both historically and currently. Imp acts to nesting beaches include th e construction of buildings and pilings, beach armoring and renourishm ent, and sand extraction (Lutcavage et al. 1997b,

Bouchard et al. 1998). Because hawksbills prefer to nest under vegetation (Mortimer 1982, Horrocks and Scott 1991), they are particul arly impacted by beachfront d evelopment and clearing of dune vegetation (Mortimer and Donnelly in review). The presence of lights on or66 adjacent to nesting beaches alters the behavior of nesting adults (Witherington 1992) and is often fatal to emerging hatchlings as they are attracted to light sources and drawn away from the water (Witherington and Bjorndal 1991). One of the most detrimental human threats to

hawksbill sea turtles is the intensive harvest of eggs from nesting beaches. Between 1950 and 1992, approximately 1.3 million hawksbill shells were collected to supply tortoiseshell to the Japanese market, the world’s largest. Japan st opped importing tortoisesh ell in 1993 in order to comply with CITES (Limpus and Miller 2008).

In addition to impacting the terrestrial zone, an thropogenic disturbances also threaten coastal marine habitats. These impacts include contamina tion from herbicides, pesticides, oil spills, and other chemicals, as well as structural degrad ation from excessive boa t anchoring and dredging (Francour et al. 1999, Lee Long et al. 2000, Waycott et al. 2005). Hawksbills are typically associated with coral reefs, which are among th e world’s most endangered marine ecosystems (Wilkinson 2000). Although primarily spongivorous, bycatch of hawksbill sea turtles in the swordfish fishery off Sout h Africa occurs (Petersen et al. 2009). Finkbeiner et al. (2011) estimated that annual bycat ch interactions to tal at least 20 individuals annually for US Atlantic fisheries (resulting in less than ten mortalities) and no or very few interactions in US Pacific

fisheries.

Future impacts from climate change and global warming may result in significant changes in hatchling sex ratios. The fact that hawksb ill turtles exhibit temperature-dependent sex determination (Wibbels 2003) suggests that there may be a skewing of future hawksbill cohorts toward strong female bias (since warmer te mperatures produce more female embryos).

Critical habitat. On September 2, 1998, the NMFS established critical habitat for hawksbill sea turtles around Mona and Monito Islands, Puerto Ri co (63 FR 46693). Aspects of these areas that are important for hawksbill sea turtle survival and recovery include important natal development habitat, refuge from predation, shelter betw een foraging periods, and food for hawksbill sea turtle prey.

Kemp’s ridley sea turtle

Distribution. The Kemp's ridley was formerly known only from the Gulf of Mexico and along

the Atlantic coast of th e US (TEWG 2000a). However, recent records support Kemp’s ridley sea turtles distribution extending into the Medite rranean Sea on occasion (Tomas and Raga 2008). The vast majority of individuals stem from breeding beaches at Rancho Nuevo on the Gulf of

Mexico coast of Mexico.

Movement and migration. Tracking of post-nesting females from Rancho Nuevo and Texas

beaches indicates that turtles m ove along coastal migratory corridors either to the north or south from the nesting beach (Byles 1989b, By les and Plotkin 1994, Renaud 1995b, Renaud et al. 1996, Shaver 1999, 2002). These migratory corridor s appear to extend throughout the coastal

areas of the Gulf of Mexico and mo st turtles appear to travel in waters less than roughly 164 feet in depth. Turtles that headed north and east trav eled as far as southwest Florida, whereas those

that headed south and east traveled as far as the Yucatan Peninsula, Mexico (Morreale et al. 2007).67 Following migration, Kemp’s ridley sea turtles se ttle into resident f eeding areas for several months (Byles and Plotkin 1994, Morreale et al. 2007). Females may begin returning along relatively shallow migratory corridors toward the nes ting beach in the winter in order to arrive at the nesting beach by early spring.

Reproduction. Mating is believed to occur about three to four weeks prior to the first nesting (Rostal 2007), or late March through early to mid April. It is presumed that most mating takes place near the nesting beach (Morreale et al. 2007, Rostal 2007). Fe males initially ovulate within a few days after successful mating and la y the first clutch approximately two to four weeks later; if a turtle nests more than on ce per season, subsequent ovulations occur within approximately 48 hours after each nesting (Rostal 2007).

Approximately 60% of Kemp's ridley nesting oc curs along an approximate 25-mile stretch of beach near Rancho Nuevo, Tamaulipas, Mexico from April to July, with limited nesting to the north (100 nests along Texas in 2006) and south (s everal hundred nests near Tampico, Mexico in 2006 USFWS 2006). Nesting at this location may be particularly important because hatchlings can more easily migrate to foraging grounds (Putman et al. 2010). The Kemp's ridley sea turtle tends to nest in large aggregations or arri badas (Bernardo and Plot kin 2007). The period between Kemp's ridley arribadas averages appr oximately 25 days, but the precise timing of the arribadas is unpredictable (Rostal et al. 1997, Bernardo and Plotkin 2007). Like all sea turtles, Kemp's ridley sea turtles nest multiple times in a single nesting season. The most recent analysis suggests approximately 3.075 nests per nesting season per female (Rostal 2007). The annual average number of eggs per nest (clutch size) is 94 to 100 and e ggs typically take 45 to 58 days to hatch, depending on temperatures (Marquez-M. 1994, USFWS 2000, 2001, 2002, 2003, 2004, 2005a, 2006, Rostal 2007). The period between nesting seasons for each female is approximately 1.8 to 2.0 years. The nesting beach at Rancho Nuevo may produce a "natural" hatchling sex ratio that is female-biased, which can potentially increase egg production as those turtles reach sexual maturity (Coyne and Landry Jr. 2007, Wibbels 2007).

Growth. Kemp's ridleys require approximately 1.5 to two years to grow from a hatchling to a size of approximately 7.9 inches long, at which si ze they are capable of making a transition to a benthic coastal immature stage, but can range from one to four years or more. Based on the size of nesting females, it is assumed that turtles mu st attain a size of appr oximately 23.6 inches long prior to maturing (Marquez-M. 1994). Growth models based on mark-recapture data suggest that a time period of seven to nine years woul d be required for this growth from benthic immature to mature size (S chmid and Witzell 1997b, Snover et al. 2007b). Currently, age to sexual maturity is believed to range from a pproximately 10 to 17 years for Kemp's ridleys (Caillouet Jr. et al. 1995, Schmid and Witzell 1997a, Snover et al. 2007a, Snover et al. 2007b). However, estimates of 10 to 13 years predominate in previous studies (Caillouet et al. 1995, Schmid and Witzell 1997b, TEWG 2000a).

Habitat. Stranding data indicate that immature turtles in this benthic stage are found in coastal

habitats of the entire Gulf of Mexico and US Atlantic coast (TEWG 2000a, Morreale et al. 2007). Developmental habitats for juveniles occur throughout the entire co astal Gulf of Mexico

and US Atlantic coast northward to N ew England (Schmid 1998, Wibbels et al. 2005, Morreale

et al. 2007). Key foraging areas in the Gulf of Me xico include Sabine Pass, Texas; Caillou Bay68 and Calcasieu Pass, Louisiana; Big Gulley, Alabama; Cedar Keys, Florida; and Ten Thousand Islands, Florida. Foraging areas studied al ong the Atlantic coast include Pamlico Sound, Chesapeake Bay, Long Island Sound, Charleston Har bor, and Delaware Bay. Near-shore waters of 120 feet or less provide the primary marine ha bitat for adults, although it is not uncommon for adults to venture into deeper waters.

Benthic coastal waters of Louisiana and Texas s eem to be preferred foraging areas for Kemp’s ridley sea turtles (particularly passes and beach fronts), although individu als may travel along the entire coastal margin of the Gu lf of Mexico. Sightings are less frequent dur ing winter and spring, but this is likely due to lesser sighting effort duri ng these times (Shoop and Kenney 1992b, Keinath et al. 1996).

Feeding. Kemp’s ridley diet consists mainly of swimming crabs, but may also include fish, jellyfish, and an array of mollusks. Immature Kemp’s ridleys off southwest Florida documented predation on benthic tunicates, a previously undocumented food s ource for this species (Witzell and Schmid 2005).

Status and trends. The Kemp’s ridley sea turtle was listed as endangered on December 2, 1970 (35 FR 18319). Internationally, the Kemp’s ridley is considered the most endangered sea turtle (NRC 1990c, USFWS 1999).

During the mid 20th century, the Kemp's ridley wa s abundant in the Gulf of Mexico. Historic

information indicates that tens of thousands of Kemp’s ridleys nested near Rancho Nuevo, Mexico, during the late 1940s (H ildebrand 1963). From 1978 through the 1980s, arribadas were 200 turtles or less, and by 1985, the total number of nests at Rancho Nuevo had dropped to approximately 740 for the entire nesting season, or a projection of r oughly 234 turtles (USFWS and NMFS 1992, TEWG 2000a). Beginning in the 1990s , an increasing number of beaches in Mexico were being monitored for nesting, and the total number of nests on all beaches in Tamaulipas and Veracruz in 2002 was over 6,000; the rate of increase from 1985 ranged from 14-16% (TEWG 2000a, USFWS 2002, Heppell et al. 2005). In 2006, approximately 7,866 nests

were laid at Rancho Nuevo with the total num ber of nests for all the beaches in Mexico estimated at about 12,000 nests, which amount ed to about 4,000 nesting females based upon three nests per female per season (Rostal et al. 1997, USFWS 2006, Rostal 2007). Considering remigration rates, the population included appr oximately 7,000 to 8,000 adult female turtles at that time (Marquez et al. 1989, TEWG 2000a, Rostal 2007). Most recently, the 2007 nesting season included an arribada of over 4,000 turtle s over a three-day period at Rancho Nuevo (P. Burchfield, pers. comm. in NMFS and USFW S 2007d). The increased recruitment of new

adults is illustrated in the proportion of first time nesters, which has increased from 6% in 1981 to 41% in 1994. Average population growth wa s estimated at 13% pe r year between 1991 and 1995 (TEWG 1998a). In 2008, there were 17,882 ne sts in Mexico (Gladys Porter Zoo 2008), and nesting in 2009 reached 21,144 (Gladys Port er Zoo 2010). In 2010, nesting declined significantly, to 13,302 (Gladys Porter Zoo 2010) but it is too early to determine if this is a onetime decline or if is indicative of a change in the trend. Population m odelling used by the TEWG (2000b) projected that Kemp’s ridleys could reac h the recovery plan’s intermediate recovery goal of 10,000 nesters by the year 2015. Recent calcu lations of nesting females determined from nest counts show that the population trend is in creasing towards that recovery goal, with an69 estimate of 4,047 nesters in 2006 and 5,500 in 2007 (NMFS and USFWS 2007c).

Nesting has also expanded geographically, with a headstart program reestablishing nesting on South Padre Island starting in 1978. Growth re mained slow until 1988, when rates of return started to grow slowly (Shaver and Wibbels 2007b). Nesting rose from 6 in 1996 to 128 in 2007, 195 in 2008, and 197 in 2009. Texas nesting then expe rienced a decline similar to that seen in Mexico for 2010, with 140 nests (N ational Park Service data, http://www.nps.gov/pais/naturescience/strp.htm), but nesting rebounded in 2011 with a record 199 nests (National Park Service data, h ttp://www.nps.gov/pais/naturescience/currentseason.htm).

Natural threats. Sea turtles face predation primarily by sharks and to a lesser extent by killer whales. All sea turtles except leatherbacks can undergo “cold stunning” if water temperatures drop below a threshold level, which can pose leth al effects. Kemp’s ridley sea turtles are particularly prone to this phenomenon along Cape Cod (Innis et al. 2009). In the last five years (2006-2010), the number of cold-stunned turtles on Cape Cod beaches averaged 115 Kemp’s ridleys.

Anthropogenic threats. Population decline has been curtaile d due to the virtual elimination of sea turtle and egg harvesting, as we ll as assistance in hatching and ra ising hatchlings (head-start). However, habitat destruction remains a concern in the form of bottom trawling and shoreline development. Trawling destroys habitat utilized by Kemp’s ridl ey sea turtles for feeding and construction activities can produce hazardous runoff. Bycatch is also a source of mortality for Kemp’s ridley sea turtles (McClellan et al. 2009). Finkbeiner et al. (2011) estimated that annual bycatch interactions tota l at least 98,300 individual s annually for US Atlan tic fisheries (resulting in 2,700 mortalities or more). The vast majority of fisheries interactions with sea turtles in the US are either Kemp’s ridley’s or loggerhead sea turtles (Finkbeiner et al. 2011).

Toxin burdens in Kemp’s ridley sea turtles include DDT, DDE, PCBs, PFOA, PFOS, chlordane, and other organochlorines (Lake et al. 1994, Rybitski et al. 1995, Keller et al. 2004a, Keller et
al. 2005). These contaminants have the potenti al to cause deficien cies in endocrine, developmental and reproductive health, and are known to depress immune function in loggerhead sea turtles (Keller et al. 2006, Storelli et al. 2007b). Along with loggerheads, Kemp’s ridley sea turtles have higher levels of PCB and DDT than leatherback and green sea turtles (Pugh and Becker 2001a). Organochlorin es, including DDT, DDE, DDD, and PCBs have

been identified as bioaccumulativ e agents and in greatest concen tration in subcutaneous lipid tissue (Rybitski et al. 1995). Concentrations ranged from 7.46 mu g/kg to 607 mu g/kg, with a mean of 252 mu g/kg in lipid tissue. Five P CB congeners composed most of the contaminants: 153/132, 138/158, 180, 118, and 187 in order of concentra tion. PCBs have also been identified in the liver, ranging in concentra tion from 272 ng/g to 655 ng/g of wet weight, values that are several fold higher than in ot her sea turtle species (Lake et al. 1994). However, concentrations are reportedly 5% of that which causes reprodu ctive failure in snapping turtles. DDE was identified to range from 137 ng/ g to 386 ng/g wet weight. Tran s-nonachlor was found at levels between 129 ng/g and 275 ng/g wet weight. Bl ood samples may be appropriate proxies for organochlorines in other body tissues (Keller et al. 2004a).70 Perfluorinated compounds in the forms of PFOA and PFOS have been identified in the blood of Kemp’s ridley turtles at con centrations of 39.4 ng/mL and 3. 57 ng/mL, respectively (Keller et al. 2005). PFCAs have also been detected. It is likely that age and ha bitat are linked to PFC

bioaccumulation.

Oil can also be hazardous to Kemp’s ridley tur tles, with fresh oil causing significant mortality and morphological changes in hatchlings, but aged oil having no detectable effects (Fritts and McGehee 1981). Blood levels of metals are lower in Kemp’s ridley sea turtles than in other sea turtles species or similar to them, with coppe r (215 ng/g to 1,300 ng/g), lead (0 to 34.3 ng/g), mercury (0.5 ng/g to 67.3 ng/g), silver (0.042 ng /g to 2.74 ng/g), and zinc (3,280 ng/g to 18,900 ng/g) having been iden tified (Orvik 1997, Innis et al. 2008). It is likely that blood samples can

be used as an indicator of meta l concentration. Mercury has been identified in all turtle species studied, but are generally an order of magnitude lo wer than toothed whales. The higher level of contaminants found in Kemp’s ridley sea turtles ar e likely due to this species tendency to feed higher on the food chain than ot her sea turtles. Females from sexual maturity through reproductive life should have lower levels of cont aminants than males because contaminants are shared with progeny through egg formation.

Critical habitat. NMFS has not designated critical ha bitat for Kemp’s ridley sea turtle.

Green sea turtle

Distribution. Green sea turtles have a circumglobal di stribution, occurring throughout tropical, subtropical waters, and, to a le sser extent, temperate waters.

Population designation. Populations are distinguished generally by ocean basin and more specifically by nesting location (Table 6). Based upon genetic differences, two or three distin ct regional clades may exist in the Pacific: western Pacific and South Pacific islands, eastern Pacific, and central Pacific, including the rookery at French Frigate Shoals, Hawaii (Dutton et al. 1996, Dutton and Balazs In review). In the eastern Pacific, green sea turtles forage from San Diego Bay, California to Mejillones, Chile. Individuals along the southern fo raging area originate from Gala pagos Islands nesting beaches,

while those in the Gulf of Calif ornia originate primarily from Mi choacán. Green turtles foraging in San Diego Bay and along the Pacific coast of Baja California or iginate primarily from

rookeries of the Islas Revill agigedos (Dutton 2003a).

Table 6. Locations and most recent abundance estimates of thr eatened green sea turtles as annual nesting females (AF), annual nests (AN), annual egg production (EP), and annual egg harvest (EH).

Location Most recent abundance Reference

Western Atlantic Ocean Tortuguero, Costa Rica 17,402-37,290 AF (Troëng and Rankin 2005) Aves Island, Venezuela 335-443 AF (Vera 2007) Galibi Reserve, Suriname 1,803 AF (Weijerman et al. 1998) Isla Trindade, Brazil 1,500-2,000 AF (Moreira and Bjorndal 2006)

Central Atlantic Ocean Ascension Island, UK 3,500 AF (Broderick et al. 2006)

Eastern Atlantic Ocean Poilao Island, Guinea-Bissau 7,000-29,000 AN (Catry et al. 2009) Bioko Island, Equatorial Guinea 1,255-1,681 AN (Tomas et al. 1999)

Mediterranean Sea71 Location Most recent abundance Reference Turkey 214-231 AF (Broderick et al. 2002) Cyprus 121-127 AF (Broderick et al. 2002)

Israel / Palestine 1-3 AF (Kuller 1999) Syria 100 AN (Rees et al. 2005)

Western Indian Ocean Eparces Islands 2,000-11,000 AF (Le Gall et al. 1986)

Comoros Islands 5,000 AF S. Ahamada, pers. comm. 2001 Seychelles Islands 3,535-4,755 AF J. Mortimer, pers. comm. 2002 Kenya 200-300 AF (Okemwa and Wamukota 2006)

Northern Indian Ocean Ras al Hadd, Oman 44,000 AN S. Al-Saady, pers. comm. 2007 Sharma, Yemen 15 AF (Saad 1999) Karan Island, Saudi Arabia 408-559 AF (Pilcher 2000) Jana and Juraid Islands, Saudi Arabia 643 AN (Pilcher 2000) Hawkes Bay and Sandspit, Pakistan 600 AN (Asrar 1999) Gujarat, India 461 AN (Sunderraj et al. 2006)

Sri Lanka 184 AF (Kapurisinghe 2006)

Eastern Indian Ocean Thamihla Kyun, Myanmar <250,000 EH (Thorbjarnarson et al. 2000) Pangumbahan, Indonesia 400,000 EH (Schulz 1987) Suka Made, Indonesia 395 AN C. Limpus, pers. comm. 2002

Western Australia 3,000-30,000 AN R. Prince, pers. comm. 2001

Southeast Asia

Gulf of Thailand 250 AN Charuchinda pers. comm. 2001

Vietnam 239 AF (Hamann et al. 2006b) Berau Islands, Indonesia 4,000-5,000 AF (Schulz 1984) Turtle Islands, Philippines 1.4 million EP (Cruz 2002) Sabah Turtle Islands, Malaysia 8,000 AN (Chan 2006) Sipadan, Malaysia 800 AN (Chan 2006) Sarawak, Malaysia 2,000 AN (Liew 2002) Enu Island (Aru Islands) 540 AF Dethmers, in preparation Terengganu, Malaysia 2,200 AN (Chan 2006)

Western Pacific Ocean

Heron Island and southern Great Barrier Reef areas, Australia 5,000-10,000 AF (Maison et al. 2010)

Raine Island and northern Great Barrier Reef areas, Australia Coringa-Herald National Nature Reserve,

Australia 10,000-25,000 AF 1,445 AF (Limpus et al. 2003) (Maison et
al. 2010) (Maison et al. 2010)

Guam Phoenix Islands, Kiribati 45 AF

100-300 AF (Cummings 2002) (Maison et al. 2010)

Ogasawara Islands, Japan

Micronesia

Marshall Islands

New Caledonia 500 AF 500-1,000 AF

100-500 AF 1,000-2,000 AF (Chaloupka et al. 2007) (Maison et al. 2010) (Maison et al. 2010) (Maison et al. 2010)

Central and Eastern Pacific Ocean French Frigate Shoals, Hawaii 400 AF (Balazs and Chaloupka 2006) Michoacán, Mexico 1,395 AF C. Delgado, pers. comm. 2006

Central American Coast 184- 344 AN (López and Arauz 2003) Galapagos Islands, Ecuador 1,650 AF (Zárate et al. 20 06)

Growth and reproduction. Most green sea turtles exhibit particularly slow growth rates, which

have been attributed to their la rgely plant-eating diet (Bjorndal 1982). Growth rates of juveniles72 vary substantially among populations, ranging from <1 cm/year (Green 1993) to >5 cm/year (McDonald Dutton and Dutton 1998), likely due to differences in diet quality, duration of foraging season (Chaloupka et al. 2004), and density of turtles in foraging areas (Bjorndal et al. 2000, Seminoff et al. 2002b, Balazs and Chaloupka 2004). If individuals do not feed sufficiently, growth is stunted and apparently does not compensate even when greater-thanneeded resources are available (Roark et al. 2009). In general, there is a tendency for green sea turtles to exhibit monotonic grow th (declining growth rate with size) in the Atlantic and nonmonotonic growth (growth spurt in mid size classes) in the Pacific, althou gh this is not always the case (Chaloupka and Musick 1997, Seminoff et al. 2002b, Balazs and Chaloupka 2004). It is estimated that green sea turtles reach a maxi mum size just under 100 cm in carapace length (Tanaka 2009). A female-bias has been identifie d from studies of green sea turtles (Wibbels 2003).

Consistent with slow growth, ag e-to-maturity for green sea turtle s appears to be the longest of any sea turtle species and ranges from ~20-40 years or more (Balazs 1982, Frazer and Ehrhart 1985b, Chaloupka and Musick 1997, Hirth 1997b, Li mpus and Chaloupka 1997, Zug and Glor 1998, Seminoff et al. 2002b, Zug et al. 2002, Chaloupka et al. 2004). Estimates of reproductive longevity range from 17 to 23 years (Carr et al. 1978, Fitzsimmons et al. 1995, Chaloupka et al. 2004). Considering that mean duration between fe males returning to nest ranges from 2 to 5 years (Hirth 1997b), these reproduct ive longevity estimates suggest that a female may nest 3 to

11 seasons over the course of her life. Each fema le deposits 1-7 clutches (usually 2-3) during the breeding season at 12-14 day intervals. Mean clutch size is highly variable among populations, but averages 110-115 eggs/nest. Females usuall y have 2-4 or more years between breeding seasons, whereas males may mate every year (B alazs 1983). Based on reasonable means of three nests per season and 100 eggs per nest (Hirth 199 7b), a female may deposit 9 to 33 clutches, or about 900 to 3,300 eggs, during her lifetime. Nesting sites appear to be re lated to beaches with relatively high exposure to wind or wi nd-generated waves (Santana Garcon et al. 2010).

Once hatched, sea turtles emerge and orient toward s a light source, such as light shining off the ocean. They enter the sea in a “frenzy” of swimming activity, which decreases rapidly in the first few hours and gradually over the first several weeks (Ischer et al. 2009, Okuyama et al. 2009). Factors in the ocean environment have a major influence on reproduction (Limpus and Nicholls 1988, Chaloupka 2001, Solow et al. 2002). It is also appa rent that during years of heavy nesting activity, density dependent fact ors (beach crowding and digging up of eggs by nesting females) may impact hatchling production (Tiwari et al. 2005, 2006). Precipitation, proximity to the high tide line, and nest depth can also significantly affect nesting success (Cheng et al. 2009). Precipitation can also be significan t in sex determ ination, with greater nest moisture resulting in a higher proportion of ma les (Leblanc and Wibbels 2009). Green sea turtles often return to the same foragi ng areas following nesting migrations (Godley et al. 2002,

Broderick et al. 2006). Once there, they move within spec ific areas, or home ranges, where they routinely visit specific localitie s to forage and rest (Seminoff et al. 2002a, Godley et al. 2003,

Makowski et al. 2006, Seminoff and Jones 2006, Taquet et al. 2006). It is also apparent that some green sea turtles remain in pelagic habita ts for extended periods, perhaps never recruiting to coastal foraging sites (Pelletier et al. 2003).73 In general, survivorship tends to be lower for juveniles and subadults th an for adults. Adult survivorship has been calculated to range from 0.82-0.97 versus 0.58-0.89 for juveniles (Seminoff et al. 2003, Chaloupka and Limpus 2005, Troëng and Chaloupka 2007), with lower values coinciding with areas of human impact on green sea turtles and th eir habitats (Bjorndal et
al. 2003, Campbell and Lagueux 2005).

Migration and movement. Green sea turtles are highly mobile and undertake complex movements through geographically disparate habita ts during their lifetimes (Musick and Limpus 1997, Plotkin 2003). The periodic migration between nesting sites and foraging areas by adults is a prominent feature of their life history. After departing as hatchlings and residing in a variety of marine habitats for 40 or more years (L impus and Chaloupka 1997), green sea turtles make their way back to the same beach from which they hatched (Carr et al. 1978, Meylan et al. 1990). At approximately 20-25 cm carapace length, juven iles leave pelagic habitats and enter benthic foraging areas (Bjorndal 1997a). Green sea turtles spend the majority of their lives in coastal foraging grounds. These areas include both open coastline and protecte d bays and lagoons. While in these areas, green sea turtles rely on ma rine algae and seagrass as their primary dietary constituents, although some populations also forage heavily on invertebrates. There is some

evidence that individuals move from shallow se agrass beds during the day to deeper areas at night (Hazel 2009). However, avoidance of areas of greater than 10 m when moderate depths of 5-10 m with sea grass beds has been found, with speed and displ acement from capture locations being similar at night as during the daytime (Senko et al. 2010a).

Habitat. Green turtles appear to prefer waters that usually remain around 20o C in the coldest month, but may occur considerably north of these regions during wa rm-water events, such as El Niño. Stinson (1984b) found green turtles to appear most frequently in US coastal waters with temperatures exceeding 18o C. Further, green sea turtles seem to occur pr eferentially in drift lines or surface current convergences, probably b ecause of the prevalence of cover and higher prey densities that associate with flotsam. For example, in the western Atlantic Ocean, drift lines commonly containing floating Sargassum spp. are capable of providing juveniles with shelter (NMFS and USFWS 1998a). Underwater resting s ites include coral recesse s, the underside of ledges, and sand bottom areas that are relativel y free of strong currents and disturbance. Available information indicates that green turtle resting areas are near feeding areas (Bjorndal and Bolten 2000). Strong site fidelity appears to be a characteristic of ju veniles green sea turtles along the Pacific Baja coast (Senko et al. 2010b).

Green sea turtles in the Gulf of Mexico tend to remain along the coast (lagoons, channels, inlets, and bays), with nesting primarily occurring in Fl orida and Mexico and infrequent nesting in all other areas (NMFS and USFWS 1991, Meylan et al. 1995a, USAF 1996, Landry and Costa 1999). Foraging areas seem to be based upon seag rass and macroalgae abundanc e, such as in the Laguna Madre of Texas. However, green sea turtles may also occur in offshore regions, particularly during migra tion and developm ent.

Feeding. While offshore and sometimes in coastal habitats, green sea turtles are not obligate plant-eaters as widely believed, a nd instead consume invertebrates such as jellyfish, sponges, sea pens, and pelagic prey (Godley et al. 1998, Heithaus et al. 2002, Seminoff et al. 2002a, Hatase et
al. 2006, Parker and Balazs in press). A shift to a more herbivorous diet occurs when individuals74 move into neritic habitats, as vegetable mater replaces an omnivorous diet at around 59 cm in carapace length off Mauritania (Cardona et al. 2009). This transition may occur rapidly starting at 30 cm carapace length, but animal prey continue to constitute an important nutritional component until individuals reach about 62 cm (Cardona et al. 2010). Foraging within seagrass ecosystems by green sea turtles can be signi ficant enough to alter habitat and ecological parameters, such as species composition (Lal et al. 2010).

Status and trends. Federal listing of the green sea tur tle occurred on July 28, 1978, with all populations listed as threatened except for the Florida and Pacific coast of Mexico breeding populations, which are endangered (43 FR 32800). The International Union for Conservation of Nature (IUCN) has classified th e green turtle as “endangered.”

No trend data are available for almost half of the important nesting s ites, where numbers are based on recent trends and do not span a full gr een sea turtle generation, and impacts occurring over four decades ago that caused a change in ju venile recruitment rates may have yet to be manifested as a change in nesting abundance. The numbers also only reflect one segment of the population (nesting females), who are the only segment of the population for which reasonably good data are available and are cautiously used as one measur e of the possible trend of populations.

Table 12 summarizes nesting abundance for 46 nes ting sites worldwide. These include both large and small rookeries believed to be representative of the overa ll trends for their respective regions. Based on the mean annual reproducti ve effort, 108,761-150,521 females nest each year among the 46 sites. Overall, of the 26 sites for which data enable an assessment of current trends, 12 nesting populations are increasing, 10 are stab le, and four are d ecreasing. Long-term continuous datasets of 20 years are available for 11 sites, all of which ar e either increasing or stable. Despite the apparent global increase in numbers, the positive overall trend should be viewed cautiously because trend data are available for just over half of all sites examined and very few data sets span a full green sea turtle generation (Seminoff 2004b).

Atlantic Ocean. Primary sites for green sea turtle nesti ng in the Atlantic/Caribbean include: (1) Yucatán Peninsula, Mexico; (2) Tortuguero, Cost a Rica; (3) Aves Island, Venezuela; (4) Galibi Reserve, Suriname; (5) Isla Trindade, Brazi l; (6) Ascension Island, United Kingdom; (7) Bioko Island, Equatorial Guinea; and (8) Bijagos Achipelago, Guinea-Bissau (NMFS and USFWS 2007a). Nesting at all of these sites was consider ed to be stable or increasing with the exception of Bioko Island and the Bijagos Archipelago wher e the lack of sufficie nt data precludes a meaningful trend assessment for either si te (NMFS and USFWS 2007a). Seminoff (2004a) reviewed green sea turtle nesting data for eight sites in the wester n, eastern, and central Atlantic. Seminoff (2004a) concluded that al l sites in the central and west ern Atlantic showed increased nesting, with the exception of nesting at Aves Is land, Venezuela, while both sites in the eastern

Atlantic demonstrated decreased nesting. These sites are not incl usive of all green sea turtle nesting in the Atlantic. Howeve r, other sites are not believed to support nesting levels high enough that would change the overa ll status of the species in the Atlantic (NMFS and USFWS 2007a).75 By far, the most important nesting concentration fo r green sea turtles in the western Atlantic is in Tortuguero, Costa Rica (NMFS a nd USFWS 2007a). Nesting in the area has increased considerably since the 1970s and nest count data from 1999-2003 suggest nesting by 17,402- 37,290 females per year (NMFS and USFWS 2007a). The number of females nesting per year on beaches in the Yucatán, at Aves Island, Galib i Reserve, and Isla Trindade number in the hundreds to low thousands, depending on th e site (NMFS and USFWS 2007a).

The vast majority of green sea turtle nesting within the southeastern US occurs in Florida (Johnson and Ehrhart 1994, Meylan et al. 1995b). Green sea turtle ne sting in Florida has been increasing since 1989 (Florida Fish and Wildlif e Conservation Commi ssion, Florida Marine Research Institute Index Nesting Beach Survey Da tabase). Since establishment of index beaches in 1989, the pattern of green tur tle nesting shows biennial peak s in abundance with a generally positive trend during the ten years of regular mo nitoring. This is perhaps due to increased protective legislation thr oughout the Caribbean (Meylan et al. 1995b). A total statewide average (all beaches, including index beaches) of 5,039 green turtle nests were laid annually in Florida between 2001 and 2006, with a low of 581 in 2001 and a high of 9,644 in 2005 (NMFS and USFWS 2007a). Data from index nesting beaches s ubstantiate the dramatic increase in nesting. In 2007, there were 9,455 green turtle nests foun d just on index nesting beaches, the highest since index beach monitoring began in 1989. Th e number fell back to 6,385 in 2008, further dropping under 3,000 in 2009, but that consecutiv e drop was a temporary deviation from the normal biennial nesting cycle for green turtles, as 2010 saw an increase back to 8,426 nests on the index nesting beaches (FWC Index Nesting B each Survey Database). Occasional nesting has been documented along the Gulf coast of Florida (Meylan et al. 1995b). More recently, green turtle nesting occurred on Bald Head Island, Nort h Carolina; just east of the mouth of the Cape Fear River; on Onslow Island; and on Cape Hatteras National Seashore. In 2010, a total of 18 nests were found in North Carolina, 6 nests in South Carolina, and 6 ne sts in Georgia (nesting databases maintained on www.seaturtle.org). In creased nesting has also been observed along the Atlantic coast of Florida, on beaches where onl y loggerhead nesting was observed in the past (Pritchard 1997). Recent modeling by Chaloupka et al. (2008a)using data se ts of 25 years or more has resulted in an estimate of the Florida nesting stock at the Archie Carr National Wildlife Refuge growing at an annual rate of 13.9%, a nd the Tortuguero, Costa Rica, population growing

at 4.9%.

There are no reliable estimates of the number of immature green sea turtles that inhabit coastal areas of the southeastern US. However, inform ation on incidental captu res of immature green

sea turtles at the St. Lucie Power Plant in St . Lucie County, Florida, shows that the annual number of immature green sea turtles captured by their offshore cooling wa ter intake structures has increased significantly. Green sea turtle annual captures averag ed 19 for 1977-1986, 178 for 1987-1996, and 262 for 1997-2001 (Florida Power and Light Company St. Lucie Plant 2002). More recent unpublished data shows 101 cap tures in 2007, 299 in 2008, 38 in 2009 (power output was cut—and cooling water intake concomita ntly reduced—for part of that year) and 413

in 2010. Ehrhart et al. (2007) documented a significant increas e in in-water abundance of green turtles in the Indian River Lagoon area.

Pacific Ocean. Green turtles are thought to be declini ng throughout the Pacifi c Ocean, with the exception of Hawaii, from a com bination of overexploitation and habitat loss (Eckert 1993b,76 Seminoff et al. 2002a). In the western Pacific, the only major (>2,000 nesting females) populations of green turtles occu r in Australia and Malaysia, w ith smaller colonies throughout the area. Indonesian nesting is widely distributed, but has expe rienced large declines over the past 50 years. Hawaii green turtles are genetical ly distinct and geographi cally isolated, and the population appears to be increasi ng in size despite the prevalen ce of fibropapillomatosis and spirochidiasis (Aguirre et al. 1998).

The East Island nesting beach in Hawaii is show ing a 5.7% annual growth rate over >25 years (Chaloupka et al. 2008a). In the Eastern Pacific, m itochondrial DNA analysis has indicated three key nesting populations: Michoacán, Mexi co; Galapagos Islands, Ecuador; and Islas Revillagigedos, Mexico (Dutton 2003b). The num ber of nesting females per year exceeds 1,000 females at each site (NMFS and USFWS 2007a) . However, historically, >20,000 females per year are believed to have nest ed in Michoacán alone (Clifton et al. 1982, NMFS and USFWS 2007a). Thus, the current number of nesting females is still far below histor ical levels. Datasets over 25 years in Chichi-jima, Japan; Heron Isla nd, Australia; and Raine Island, Australia, show increases in abundance (Chaloupka et al. 2008a).

Indian Ocean. One of the largest nesting sites for green sea turtles worldwide occurs on the beaches of Oman where an estimated 20,000 green sea turtles nest annually (Hirth 1997a). Only

the Comoros Island index site in the western Indi an Ocean showed evidence of increased nesting (Seminoff 2004a).

Natural threats. Herons, gulls, dogfish, and sharks prey upo n hatchlings. Adults face predation primarily by sharks and to a le sser extent by killer whales. All sea turtles except leatherbacks can undergo “cold stunning” if water temperatures drop below a threshold level, which can be lethal. For unknown reasons, the frequency of a disease called fibropapi llomatosis is much higher in green sea turtles than in other spec ies and threatens a large number of existing subpopulations. Extremely high incidence has been reported in Hawaii, where affliction rates peaked at 47-69% in some foraging areas (Murakawa et al. 2000). A to-date unidentified virus may aid in the development of fibropapillomatosis (Work et al. 2009). Predators (primarily of eggs and hatchlings) also include dogs, pigs, rats, crabs, sea bi rds, reef fishes, and groupers (Witzell 1981, Bell et al. 1994). Green sea turtles with an abundance of barnacles have been found to have a much greater probabili ty of having health issues (Flint et al. 2009).

Anthropogenic threats. Major anthropogenic impacts to the nesting and marine environment affect green sea turtle survival and recovery. At nesting beaches, green sea turtles rely on intact dune structures, native vegetation, and normal b each temperatures for nesting (Ackerman 1997). Structural impacts to nesting habitat include th e construction of buildings and pilings, beach armoring and renourishment, and sand extraction (Lutcavage et al. 1997b, Bouchard et al. 1998). These factors may directly, through loss of beach habitat, or indirectl y, through changing thermal profiles and increasing erosion, serve to decrease the am ount of nesting ar ea available to nesting females, and may evoke a change in the natural behaviors of adults a nd hatchlings (Ackerman 1997, Witherington et al. 2003, 2007). On the Pacific coast of Mexico in the mid-1970s, >70,000 green turtle eggs were harvested every ni ght. The presence of lights on or adjacent to nesting beaches alters th e behavior of nesting adults (With erington 1992) and is often fatal to emerging hatchlings as they ar e attracted to light sources a nd drawn away from the water77 (Witherington and Bjorndal 1991). In addition to impacting the terrestrial zone, anthropogenic

disturbances also threaten coasta l marine habitats, part icularly areas rich in seagrass and marine algae. These impacts include contamination from herbicides, pesticides, oil spills, and other chemicals, as well as structural degradati on from excessive boat anchoring and dredging (Francour et al. 1999, Lee Long et al. 2000, Waycott et al. 2005). Ingestion of plastic and other

marine debris is another source of morbidity and mortality (Stamper et al. 2009). Green sea

turtles stranded in Brazil were all found to ha ve ingested plastics or fishing debris (n=34), although mortality appears to have results in three cases (Tourinho et al. 2009). Low-level bycatch has also been documented in longline fisheries (Petersen et al. 2009). Further, the introduction of alien algae species threatens the stability of some coastal ecosystems and may lead to the elimination of preferred dietary sp ecies of green sea turtle s (De Weede 1996). Very few green sea turtles are bycaught in US fisheries (Finkbeiner et al. 2011). However, a legal fishery operates in Madagascarthat harvested about 10,000 green turtles annually in the mid1990s.

Sea level rise may have signifi cant impacts upon green tu rtle nesting on Pacific atolls. These low-lying, isolated locations c ould be inundated by rising water levels associated with global warming, eliminating nesting habitat (Baker et al. 2006, Fuentes et al. 2010). Fuentes et al. (2010) predicted that rising temper atures would be a much greater threat in the long term to the hatching success of sea turtle tu rtles in general and green sea tu rtles along northeastern Australia particularly. Green sea turtles emerging from nest s at cooler temperatures likely absorb more yolk that is converted to body tissue than do hatchlings from warmer nests (Ischer et al. 2009). Predicted temperature rises may approach or exceed the upper thermal tolerance limit of sea turtle incubation, causi ng widespread failure of nests (Fuentes et al. 2010). Although the timing of loggerhead nesting depends upon sea-surface temp erature, green sea turtles do not appear to be affected (Pike 2009).

Green sea turtles have been f ound to contain the organochlorines chlordane, lindane, endrin, endosulfan, dieldrin, DDT and PCB (Miao et al. 2001, Gardner et al. 2003). Levels of PCBs found in eggs are considered far higher than wh at is fit for human cons umption (van de Merwe et
al. 2009). The heavy metals copper, lead, manga nese, cadmium, and nickel have also been

found in various tissues and life st ages (Barbieri 2009). Arsenic also occurs in very high levels in green sea turtle eggs (van de Merwe et al. 2009). These contaminants have the potential to cause deficiencies in endocrine, developm ental, and reproductive health, and depress immune function in loggerhead sea turtles (Keller et al. 2006; Storelli et al. 2007). Exposure to sewage effluent may also result in green sea turtle eggs harboring antibiotic-resista nt strains of bacteria (Al-Bahry et al. 2009). DDE has not been found to influe nce sex determination at levels below cytotoxicity (Podreka et al. 1998, Keller and McClellan-Green 2004). To date, no tie has been found between pesticide concentr ation and susceptibility to fibropapillomatosis, although degraded habitat and pollution have been tied to the incidence of the disease (Aguirre et al. 1994, Foley et al. 2005). Flame retardants have been measured from healthy individuals (Hermanussen et al. 2008). It has been theorized that exposure to tumor-promoting compounds produced by the cyanobacteria Lyngbya majuscule could promote the development of fibropapillomatosis (Arthur et al. 2008). It has also been theori zed that dinoflagellates of the

genus Prorocentrum that produce the tumorogenic com pound okadoic acid may influence the development of fibropapillomatosis (Landsberg et al. 1999).78 Critical habitat. On September 2, 1998, critical habitat fo r green sea turtles was designated in coastal waters surrounding Culebra Island, Puerto Rico (63 FR 46693). Aspects of these areas that are important for green sea turtle survival and recovery include important natal development habitat, refuge from predation, shelter between foraging periods, and food for green sea turtle

prey.

Loggerhead sea turtle

Distribution. Loggerheads are circumglobal occurring throughout the temperate and tropical regions of the Atlantic, Pacific, and Indian o ceans. Loggerheads are the most abundant species

of sea turtle found in US coastal waters.

Population designations. Five groupings represent loggerhead sea turtles by major sea or ocean basin: Atlantic, Pacific, and Indian oceans, as well as Caribbean and Mediterranean seas. As with other sea turtles, populations are frequently divided by nesting aggregation (Hutchinson and Dutton 2007). On March 16, 2010, the NMFS proposed to designate nine distinct population segments (DPSs) of loggerhead sea turtles: South Atlantic Ocean and southwest Indian Ocean as threatened as well as Mediterranean Sea, Nort h Indian Ocean, North Pacific Ocean, northeast Atlantic Ocean, northwest Atlantic Ocean, Sout h Pacific Ocean, and s outheast Indo-Pacific Ocean as endangered (75 FR 12598).

Atlantic Ocean. Western Atlantic nesting locations incl ude The Bahamas, Brazil, and numerous

locations from the Yucatán Peninsula to No rth Carolina (Addison and Morford 1996, Addison 1997, Marcovaldi and Chaloupka 2007). This gr oup comprises five nesting subpopulations: Northern, Southern, Dry Tortugas, Florida Panhandle, and Yucatán. Additional nesting occurs on Cay Sal Bank (Bahamas), Cuba, the Bahami an Archipelago, Quintana Roo (Yucatan Peninsula), Colombia, Brazil, Ca ribbean Central America, Venezu ela, and the eastern Caribbean

Islands. Genetic studies indicate that, although females routinely re turn to natal beaches, males may breed with females from multiple populat ions and facilitate gene flow Bowen et al. (2005). In the eastern Atlantic, we know of five rookeri es from Cape Verde, Greece, Libya, Turkey, and

the western Africa coast.

Indian Ocean. Loggerhead sea turtles nest along the In dian Ocean in Oman, Yemen, Sri Lanka, Madagascar, South Africa, and possibly Mozambique.

Pacific Ocean. Pacific Ocean rookeries are limited to the western portion of the basin. These sites include Australia, Ne w Caledonia, New Zealand, Indon esia, Japan, and the Solomon

Islands.

Population structure in the Pacifi c is comprised of a northwestern Pacific nesting aggregation in Japan and a smaller southwestern nesting aggr egation in Australia and New Caledonia (NMFS 2006e). Genetics of Japanese nesters suggest that this subpopulation is co mprised of genetically distinct nesting colonies (Hatase et al. 2002a). Almost all loggerheads in the North Pacific seem to stem from Japanese nesting beaches (Bowen et al. 1995; Resendiz et al. 1998). The fidelity of nesting females to their nesting beach allowed differentiation of thes e subpopulations and the loss of nesting at a beach means a significant loss of diversity and the beach is unlikely to be79 recolonized (NMFS 2006e).

Reproduction and growth. Loggerhead nesting is confined to lower latitudes temperate and subtropic zones but absent from tr opical areas (NRC 1990b, NMFS and USFWS 1991b, Witherington et al. 2006b). The life cycle of loggerhead sea turtles can be divided into seven stages: eggs and hatchlings, small juveniles, large juveni les, subadults, novice breeders, first year emigrants, and matu re breeders (Crouse et al. 1987). Hatchling loggerheads migrate to the ocean (to which they are drawn by n ear ultraviolet light Kawamura et al. 2009), where they are generally believed to lead a pelagic existe nce for as long as 7-12 years (NMFS 2005c). Loggerheads in the Mediterranean, similar to thos e in the Atlantic, grow at roughly 11.8 cm/yr for the first six months and slow to roughly 3.6 cm/yr at age 2.5-3.5. As adults, individuals may experience a secondary growth pulse associated with shifting into ner itic habitats, although growth is generally monotypi c (declines with age Casale et al. 2009a, Casale et al. 2009b). Individually-based variables likely have a hi gh impact on individual-to-i ndividual growth rates (Casale et al. 2009b). At 15-38 years, loggerhead sea turtles become sexually mature, although the age at which they reach maturity varies widely among populations (Frazer and Ehrhart 1985a, Frazer et al. 1994, NMFS 2001b, Witherington et al. 2006, Casale et al. 2009b). However, based on new data from tag returns, strandings, and nesting surveys, NMFS (2001b) estimated ages of maturity ranging from 20-38 year s and benthic immature stage lasting from 14- 32 years.

Loggerhead mating likely occurs along migration routes to nesting beaches, as well as in offshore from nesting beaches several weeks pr ior to the onset of nesting (Dodd 1988a, NMFS and USFWS 1998d). Females usually breed every 2-3 years, but can vary from 1-7 years (Richardson et al. 1978, Dodd 1988a). Females lay an av erage of 4.1 nests per season (Murphy and Hopkins 1984) , although recent satellite tele metry from nesting females along southwest Florida support 5.4 nests per female per season, with increasing numbers of eggs per nest during the course of the season (Tucker 2009). The au thors suggest that this finding warrants revision of the number of females nesting in the region. The western Atlantic breeding season is MarchAugust. Nesting sites appear to be related to beaches with relatively high exposure to wind or wind-generated waves (Santana Garcon et al. 2010).

The Japanese rookeries are the most significant ne sting sites for loggerheads in the North Pacific, with nesting occurring on the Japanese ma inland, except for Hokkaido, as well as the Ryukyu Islands to the south (Kamezaki 1989, Uchida and Nishiwaki 1995, Kamezaki et al. 2003, Sea Turtle Association of Japan 2010). Nesting generally occurs through summer and fall (AprilAugust, peaking in July), with females re turning every two to three years (Iwamoto et al. 1985). Nesting females lay at least three nests of 60- 115 eggs per nest each seaso n, with roughly two weeks between nests (Iwamoto et al. 1985, Eckert 1993b, Nishimura 1994). Between nests, females appear to swim offshore into the Kuro shio Current, possibly to speed egg development (NMFS and USFWS 1998b, Sato et al. 1998).

Nesting in the Gulf of Mexico does occur, although primarily in Fl orida, with rare nests do occur along North and South Padre Island in Texas (Hildebrand 1983, Dodd 1988b).

Migration and movement. Loggerhead hatchlings migrate offshore and become associated80 with Sargassum spp. habitats, driftlines, and other convergence zones (Carr 1986). After 14-32 years of age, they shift to a be nthic habitat, where im mature individuals forage in the open ocean and coastal areas along continental shelves, bays, lagoons, and estuaries (NMFS 2001b, Bowen

et al. 2004). Adult loggerheads make lengthy migr ations from nesting beaches to foraging grounds (TEWG 1998b). In the Gulf of Mexico, la rger females tend to disperse more broadly after nesting than smaller individuals, which tend to stay closer the nesting location (Girard et al. 2009). In the North Atlantic, loggerheads travel north during spring and summer as water temperatures warm and return south in fall and winter, but occur offshore year-round assuming adequate temperature. For imma ture individuals, this movement occurs in two patterns: a northsouth movement over the continental shelf with mi gration south of Cape Hatteras in winter and movement north along Virginia for summer fora ging, and a not-so-season al oceanic dispersal

into the Gulf Stream as far north as the 10-15  ̊ C isotherm (Mansfield et al. 2009). Wallace et al. (2009) suggested differences in growth rate based upon these foraging strategies. There is conflicting evidence that immature loggerheads roam the oceans in currents and eddies and mix from different natal origins or distribute on a la titudinal basis that corre sponds with their natal beaches (Monzon-Arguello et al. 2009, Wallace et al. 2009). McCarthy et al. (2010) found that movement patterns of loggerhead sea turtles were more convoluted when sea surface temperatures were higher, ocean depths shallower, ocean currents stronger, and chlorophyll a

levels lower.

Individuals in the western Paci fic also show wide-ranging moveme nts. Loggerheads hatched on

beaches in the southwest Pacific travel have be en found to range widely in the southern portion of the basin, with individuals fr om populations nesting in Austra lia found as far east as Peruvian coast foraging areas still in the juvenile stage (Boyle et al. 2009). Individuals hatched along Japanese coasts have been found to migrate to waters off Baja Californi a via the North Pacific Subtropical Gyre (and the Kuroshio Extension) to feed for several years before migrating back to western Pacific waters to breed (Bowen et al. 1995, Resendiz et al. 1998, Polovina et al. 2000, Nichols 2005, Polovina et al. 2006). Adult loggerheads also resi de in oceanic waters off Japan (Hatase et al. 2002b). Habitat use off Japan may further be partitioned by sex and size (Hatase et
al. 2002b, Hatase et al. 2002c, Hatase and Sakamoto 2004). Loggerheads returning to Japanese waters seem to migrate along nutri ent-rich oceanic fronts (Nichols et al. 2000, Polovina et al. 2000, Kobayashi et al. 2008). Individuals bycaught and sa tellite tracked in Hawaii longline

fisheries show individual moveme nt north and south within a th ermal range of 15-25o C, or 28- 40o N, with juveniles following the 17-20o C isotherm (Nichols et al. 2000, Polovina et al. 2004, Kobayashi et al. 2008). The Transition Zone Chlorophyll Front and Kuroshio Extension Current are likely important foraging area s for juvenile loggerheads (Polovina et al. 2004). The Kuroshio Current off Japan may be significant for juvenile and adult logg erheads as a wintering

areas for those individuals not migrating south (Hatase et al. 2002c).

Sighting and stranding records support logger head sea turtles to be common, year-round

residents of the Gulf of Mexi co, although their abundance is much greater in the northeastern region versus the northwestern (Fritts et al. 1983, Landry and Costa 1999, Davis et al. 2000). Loggerheads may occur in both offshore habitats (particularly around oil platforms and reefs, where prey and shelter are available; (Fritts et al. 1983, Rosman et al. 1987, Lohoefener et al. 1990, Gitschlag and Herczeg 1994, Davis et al. 2000), as well as shallow bays and sounds (which may be important developm ental habitat for late juveniles in the eastern Gulf of Mexico;81 (Lohoefener et al. 1990, USAF 1996, Davis et al. 2000). Offshore abundance in continental slope waters increases during the wi nter in the eastern Gulf of Me xico, as cooler inshore waters force individuals into warmer offshore areas (Davis et al. 2000).

Gender, age, and survivorship. Although information on males is limited, several studies identified a female bias, although a single study has found a strong male bias to be possible (Dodd 1988a, NMFS 2001b, Rees and Margaritoulis 2004).

Additionally, little is known about longevity, although D odd (1988a) estimated the maximum female life span at 47-62 years. Heppell et al. (2003a) estimated annual su rvivorship to be 0.81 (southeast US adult females), 0.78-0.91 (Austr alia adult females), 0.68-0.89 (southeast US benthic juveniles, and 0.92 (Austr alia benthic juveniles). Surviv al rates for hatchlings during their first year are likely very low (Heppell et al. 2003a, Heppell et al. 2003b).

Feeding. Loggerhead sea turtles are omnivorous and opportunistic feeders through their lifetimes (Parker et al. 2005). Hatchling loggerheads feed on macroplankton associated with Sargassum spp. communities (NMFS and USFWS 1991b). Pelagic and benthic juveniles forage on crabs, mollusks, jellyfish, and vegetati on at or near the surface (Dodd 1988a, Wallace et al. 2009). Loggerheads in the deep, offshore waters of the western North Pacific feed on jellyfish, salps, and other gelatinous animals (Dodd Jr. 1988b, Hatase et al. 2002b). Sub-adult and adult loggerheads prey on benthic inve rtebrates such as gastropods, mo llusks, and decapod crustaceans in hard-bottom habitats, although fish and plants are also occasionally eaten (NMFS and USFWS 1998d). Stable isotope analysis and study of orga nisms on turtle shells has recently shown that although a loggerhead population may feed on a va riety of prey, individuals composing the population have specialized diets (Reich et al. 2010, Vander Zanden et al. 2010).

Status and trends. Loggerhead sea turtles were listed as threatened under the ESA of 1973 on July 28, 1978 (43 FR 32800). The NMFS recently determined that a petition to reclassify loggerhead turtles in the western North Atlantic Ocean as endangered may be warranted due to

the substantial scientific and commercial in formation presented. Consequently, NMFS has initiated a review of the status of the species and is currently soliciting additional information on the species status and ecology, as well as areas that ma y qualify as critical habitat (73 FR 11849; March 5, 2008).

There is general agreement that the number of nesting females provides a useful index of the species’ population size and stab ility at this life stage, ev en though there are doubts about the ability to estimate the overa ll population size (Bjorndal et al. 2005). An important caveat for population trends analysis based on nesting beach data is that this may reflect trends in adult nesting females, but it may not reflect overall population growth rates well. Adult nesting

females often account for less than 1% of total population numbers. Th e global abundance of nesting female loggerhead turtle s is estimated at 43,320–44,560 (Spotila 2004b).

Atlantic Oc ean. In the eastern Atlantic, the Cape Verde Islands support the only known loggerhead nesting assemblage, which is of at least intermediate size (Fretey 2001a); 1,071 nests were observed in 2009 (Lino et al. 2010). In 2000, researcher s tagged over 1,000 nesting females (Erhart et al. 2003). Annual data from monitoring projects in Cyprus, Greece, Israel,82 Tunisia, and Turkey reveal total annual nesting in the Medite rranean ranging of 3,375-7,085 nests per season (Margaritoulis et al. 2003). Libya and the West African coast host geneticallyunique breeding populations of loggerhead sea turt les as well (Hutchinson and Dutton 2007). A recently discovered nesting site along the sout hern Italian shores of the Ionian Sea found particularly high genetic diversity amongst nesting females (Garofalo et al. 2009). Nesting at Dalyan Beach, Turkey does not have an appa rent trend, with between 50 and 286 nests laid annually for the past 19 years (Turkozan and Yilmaz 2008).

The greatest concentration of loggerheads occurs in the Atlantic Ocean and the adjacent Caribbean Sea, primarily on the Atlantic coast of Florida, with other major nesting areas located on the Yucatán Peninsula of Mexico, Columbia, Cuba, South Africa (Márquez 1990, EuroTurtle 2006 as cited in LGL Ltd. 2007).

Among the five subpopulations, loggerhead fe males lay 53,000-92,000 nests per year in the southeastern US and the Gulf of Mexico, a nd the total number of nesting females is 32,000- 56,000. All of these are currently in decline or data are insuff icient to access trends (TEWG 1998a, NMFS 2001b). Loggerheads from western Nort h Atlantic nesting a ggregations may or may not feed in the same regions from which they hatch. Loggerhead sea turtles from the northern nesting aggregation, which represents ab out 9% of the loggerhea d nests in the western North Atlantic, comprise 25-59% of individuals foraging from Georgia up to the northeast US (Sears 1994, Norrgard 1995, Sears et al. 1995, Rankin-Baransky 1997, Bass et al. 1998). Loggerheads associated with the South Florida nesting aggregati on occur in higher frequencies in the Gulf of Mexico (where they represen t ~10% of the loggerhead captures) and the Mediterranean Sea (where they represent ~45% of loggerhead sea turtles captured). About 4,000 nests per year are laid along the Brazilian coast (Ehrhart et al. 2003).

The northern recovery unit along Georgia, South Ca rolina, and North Carolina has a forty-year time-series trend showing an ove rall decline in nesti ng, but the shorter comprehensive survey data (20 years) indicate a st able population (GDNR, NCWRC, a nd SCDNR nesting data located at www.seaturtle.org). NMFS scientists have es timated that the northern subpopulation produces 65% males (NMFS 2001b).

The peninsular Florida recovery unit is the largest loggerhead nesti ng assemblage in the northwest Atlantic. A near-complete nest census (all beaches including index nesting beaches)

undertaken from 1989 to 2007 showed a mean of 64,513 loggerhead nests per year, representing approximately 15,735 nesting females annually (NMFS and USFWS 2008). The statewide estimated total for 2010 was 73,702 (FWRI nesting da tabase). An analysis of index nesting beach data shows a 26% nesting decline be tween 1989 and 2008, and a mean annual rate of decline of 1.6% despite a larg e increase in nesting for 2008, to 38,643 nests (FWRI nesting database)(NMFS and USFWS 2008, Witherington et al. 2009). In 2009, nesting levels, while still higher than the lows of 2004, 2006, and 2007, dropped below 2008 levels to approximately 32,717 nests, but in 2010 a large increase was seen, with 47,880 nests on the index nesting beaches (FWRI nesting database). The 2010 inde x nesting number is the largest since 2000. With the addition of data through 2010, the nesti ng trend for the nowrthwestern Atlantic DPS is slightly negative and not statis tically different from zero ( no trend)(NMFS and USFWS 2010).83 Preliminary, unofficial reports i ndicate that 2011 nesting may be a high nesting year on par with

2010.

Because of its size, the south Fl orida subpopulation of loggerheads may be critical to the survival of the species in the Atlantic, and in the past it was considered second in size only to the Oman nesting aggregation (NMFS and USFWS 1991b, NMFS 2006e). The South Florida population increased at ~5.3% per year fr om 1978-1990, and was initially in creasing at 3.9-4.2 % after 1990. An analysis of nesting data from 1989-2005, a period of more c onsistent and accurate surveys than in previous years, showed a detectable trend and, more recently (1998-2005), has shown evidence of a declining trend of appr oximately 22.3% (FFWCC 2007b, a, Witherington et al. 2009). This is likely due to a decline in the number of nesting females within the population (Witherington et al. 2009). Nesting data from the Archie Ca rr Refuge (one of the most important nesting locations in southeast Florida) over the last 6 years shows nests declined from approximately 17,629 in 1998 to 7,599 in 2004, also suggesting a decrease in population size 1

. Loggerhead nesting is thought to consist of just 60 nesting females in the Caribbean and Gulf of Mexico (NMFS 2006f). Based upon the small sizes of almost all nesting aggregations in the Atlantic, the large numbers of individuals killed in fisheries, and the decline of the only large nesting aggregation, we suspect that the exti nction probabilities of loggerhead sea turtle populations in the Atlantic are onl y slightly lower than those of populations in the Pacific.

Zurita et al. (2003) found a statistically si gnificant increase in the num ber of nests on seven of the beaches on Quintana Roo, Mexico, from 198 7-2001, where survey effort was consistent during the period. However, nesting has declin ed since 2001, and the previously reported increasing trend appears to have b een temporary (NMFS and USFWS 2008).

Mediterranean Sea. Nesting in the Mediterranean is confin ed almost exclusively to the eastern basin. The highest level of nes ting in the Mediterranean occurs in Greece, with an average of 3,050 nests per year.

Pacific Ocean. Abundance has declined dramatically ov er the past 10-20 years, although loggerheads range widely from Alaska to Chile (NMFS and USFWS 1998d). Pacific nesting is limited to two major locations, Australia a nd Japan, although low level nesting may occur outside of Japan in areas surr ounding the South China Sea (Kamezaki et al. 2003, Chan et al. 2007). Eastern Australia supported one of the ma jor global loggerhead nesting assemblages until recently (Limpus 1985). Now, less than 500 female s nest annually, an 86% reduction in the size of the annual nesting population in 23 years (Limpus and Limpus 2003). The status of loggerhead nesting colonies in southern Japa n and the surrounding regi on is uncertain, but approximately 1,000 female loggerhead turtles ma y nest there; a 50-90% decline compared to historical estimates (Dodd Jr. 1988b, Bolten et al. 1996, STAJ 2002, Kamezaki et al. 2003). Nine major nesting beaches (greater than 100 ne sts per season) and six “submajor” beaches (10– 100 nests per season) exists, including Yakushima island where 40% of nesting occurs (Kamezaki et al. 2003). Nesting declined from an init ial peak of approximately 6,638 nests in

1 While this is a long period of decline relative to the pa st observed nesting pattern at this location, aberrant ocean surface temperatures complicate the analysis and interpreta tion of these data. Althoug h caution is warranted in interpreting the decreasing nesting trend given inherent annu al fluctuations in nesting and the short time period over which the decline has been noted, the recent nesting d ecline at this nesting beach is reason for concern.84 1990–1991, followed by a steep decline to a low of 2,064 nests in 1997 (Conant et al. 2009). During the past decade, nesting increased gradually to 5,167 nests in 2005 (Conant et al. 2009), declined and then rose agai n to a record high of 11,082 nests in 2008, and then 7,495 and 10,121 nests in 2009 and 2010, respectively (STAJ 2008, 2009, 2010).

In addition, loggerheads uncommonly occur in US Pacific waters, and there have been no documented strandings of loggerheads on the Ha waiian Islands in nearly 20 years (1982-1999 stranding data). There are very few records of loggerheads nesting on any of the many islands of the central Pacific, and the speci es is considered rare or vagr ant in this region (USFWS 1998). Overall, Gilman (2009) estimated that the number of loggerheads nesting the Pacific has declined by 80% in the past 20 years.

Indian Ocean. The largest known nesting aggregati on occurs on Masirah and Kuria Muria Islands in Oman (Ross and Barwani 1982). Extra polations resulting from partial surveys and tagging in 1977-1978 provided broad estimates of 19,000-60,000 females nesting annually at Masirah Island, while a more recent partial survey in 1991 provided an estimate of 23,000 nesting females (Ross 1979, Ross and Barwani 1982, Baldwin 1992, Ross 1998). Over 3,000 nests per year have been record ed on the Al-Halaniyat Islands, while along the Oman mainland of the Arabian Sea, about 2,000 nests are deposited per year (Salm 1991, Salm et al. 1993). Based upon genetic analyses, additional populations nest in Yemen, Sri Lanka, and Madagascar (Hutchinson and Dutton 2007). In the southwestern Indian Ocean, the highest concentration of nesting occurs on the coast of Tongaland, South Africa (Baldwin et al. 2003). The total number of females nesting annually in South Africa is estimated to be between 500-2,000 (Baldwin et al. 2003). In the Eastern Indian Ocean, all known nesting sites are found in Western Australia (Dodd Jr. 1988a). An estimated 800-1,500 logge rheads nest annually on Dirk Hartog Island beaches along Western Australia (Baldwin et al. 2003).

Natural threats. Sea turtles face predation primarily by sh arks and to a lesser extent by killer whales. All sea turtles except leatherbacks can undergo “cold stunning” if water temperatures drop below a threshold level, which can pose lethal effects. In January 2010, an unusually large cold-stunning event occurred throughout the southeast US, with well over 3,000 sea turtles (mostly greens but also hundreds of loggerh eads) found cold-stunned. Most survived, but several hundred were found dead or died after being discovered in a cold-st unned state. Eggs are commonly eaten by raccoons and ghos t crabs along the eastern US (Barton and Roth 2008). In the water, hatchlings are hunted by herons, gulls , dogfish, and sharks. Heavy loads of barnacles are associated with unhealthy or dead stranded loggerheads (Deem et al. 2009).

Anthropogenic threats. Anthropogenic threats impacting lo ggerhead nesting habitat are numerous: coastal development and constructio n, placement of erosion control structures, beachfront lighting, vehicular and pedestrian traffic, sand extraction, beach erosion, beach nourishment, beach pollution, removal of na tive vegetation, and planting of non-native vegetation (Baldwin 1992, USFWS 1998, Margaritoulis et al. 2003, Mazaris et al. 2009b). Surprisingly, beach nourishment also hampers nes ting success, but only in the first year postnourishm ent before hatching success increases (Brock et al. 2009). Loggerhead sea turtles face

numerous threats in the marine environment as well, including o il and gas exploration, marine pollution, trawl, purse seine, hook and line, gill net, pound net, longline, and trap fisheries,85 underwater explosions, dredgi ng, offshore artificial lighti ng, power plant entrapment, entanglement in debris, ingesti on of marine debris, marina a nd dock construction and operation, boat collisions, and poaching. At least in the Mediterannean Sea, Anthorpogenic threats appear to disproportionally impact larger (more fecund) loggerheads (Bellido et al. 2010).

The major factors inhibiting their recovery includ e mortalities caused by fishery interactions and degradation of the beaches on which they nest. Shrimp trawl fisheries account for the highest number of captured and killed l oggerhead sea turtles. Along the Atlantic coast of the US, the

NMFS estimated that shrimp trawls capture almost 163,000 loggerhead sea turtles each year in the Gulf of Mexico, of which 3,948 die. Howe ver, more recent estimates from suggest interactions and mortality ha s decreased from pre-regulator y periods, with a conservative estimate of 26,500 loggerheads captured annually in US Atlantic fisheries causing mortality to 1,400 individuals per year (Finkbeiner et al. 2011). Pacific bycatch is much less, with about 400 individuals bycaught annually in US fisheries re sulting in at least 20 mortalities (Finkbeiner et
al. 2011). Each year, various fisheries capture about 2,000 loggerhead sea turtles in Pamlico Sound, of which almost 700 die. As a result of the 2006 and 2007 tri-national fishermen’s exchanges in 2007 a prominent Baja California Sur fleet retired its bottom-set longlines (Peckham et al. 2008) (Peckham and Maldonado-Diaz 2011). Pr ior to this closure, the longline fleet interacted with an estimated 1,160-2,174 logger heads annually, with near ly all (89 percent) of the takes resulting in mortalities(Peckham et al. 2008). Offshore longline tuna and swordfish longline fisheries are also a se rious concern for the survival and recovery of loggerhead sea turtles and appear to affect the largest individuals more th an younger age classes (Bolten et al. 1994, Aguilar et al. 1995, Howell et al. 2008, Tomás et al. 2008, Carruthers et al. 2009, Marshall

et al. 2009, Petersen et al. 2009). Deliberate hunting of loggerh eads for their meat, shells, and eggs has declined from previous exploitation leve ls, but still exists and hampers recovery efforts (Lino et al. 2010). In the Pacific, loggerh ead turtles are captured, injure d, or killed in numerous Pacific fisheries including:

 Japanese longline fisheries in the west ern Pacific Ocean and South China Seas

 direct harvest and commercial fisher ies off Baja Calif ornia, Mexico

 commercial and artisanal swordfish fisherie s off Chile, Columbia, Ecuador, and Peru

 purse seine fisheries for tuna in the eastern tropical Pacific Ocean

 California/Oregon drift gillnet fisheries (NMFS 2006e)

Wallace et al. (2010) estimated that between 1990 a nd 2008, at least 85,000 sea turtles were captured as bycatch in fisheries worldwide. Th is estimate is likely at least two orders of magnitude low, resulting in a likely bycatch of nearly half a million sea turtles annually (Wallace

et al. 2010); many of these are expected to be loggerhead sea turtles.

Marine debris ingestion can be a widespread issue for loggerhead sea turtles. More than onethird of loggerheads found strande d or bycaught had injected m ari ne debris in a Mediterranean study, with possible mortality result ing in some cases (Lazar and Gra č an 2010).

Climate change may also have significant implic ations on loggerhead populat ions worldwide. In addition to potential loss of nesting habitat due to sea level rise, loggerhead sea turtles are very sensitive to temperature as a determinant of se x while incubating. Ambient temperature increase86 by just 1o-2o C can potentially change hatchling sex ratios to all or nearly all female in tropical and subtropical areas (Hawkes et al. 2007a). Over time, this can reduce genetic diversity, or even population viability, if males become a small proportion of populations (Hulin et al. 2009). Sea surface temperatures on loggerhead foraging gr ounds correlate to the timing of nesting, with higher temperatures leading to earlier nesting (Mazaris et al. 2009a, Schofield et al. 2009). Increasing ocean temperatures may also lead to reduced primary productivity and eventual food availability. This has been proposed as pa rtial support for reduced nesting abundance for loggerhead sea turtles in Japan; a finding that could have broader implications for other populations in the future if individuals do not shift feedi ng habitat (Chaloupka et al. 2008b). Warmer temperatures may also decrease the energy needs of a developing embryo (Reid et al. 2009).

Tissues taken from loggerheads sometimes cont ain very high levels of organochlorines chlorobiphenyl, chlordanes, lindane, endrin, en dosulfan, dieldrin, PFO S, PFOA, DDT, and PCB (Rybitski et al. 1995, Mckenzie et al. 1999, Corsolini et al. 2000, Gardner et al. 2003, Keller et
al. 2004a, Keller et al. 2004b, Keller et al. 2005, Alava et al. 2006, Perugini et al. 2006, Storelli

et al. 2007a, Monagas et al. 2008, Oros et al. 2009). It appears that levels of organochlorines have the potential to suppress the immune syst em of loggerhead sea turtles and may affect metabolic regulation (Keller et al. 2004c, Keller et al. 2006, Oros et al. 2009). These

contaminants could cause deficiencies in e ndocrine, developmental, and reproductive health (Storelli et al. 2007a). It is likely that the omnivorous nature of loggerheads makes them more prone to bioaccumulating toxins than other sea turtle species (Godley et al. 1999, Mckenzie et al. 1999).

Heavy metals, including arseni c, barium, cadmium, chromium, iron, lead, nickel, selenium, silver, copper, zinc, and manganese , have also been found in a vari ety of tissues in levels that increase with turtle size (Godley et al. 1999, Saeki et al. 2000, Anan et al. 2001, Fujihara et al. 2003, Gardner et al. 2006, Storelli et al. 2008, Garcia-Fernandez et al. 2009). These m etals likely originate from plants and seem to have high transfer coefficients (Anan et al. 2001, Celik

et al. 2006, Talavera-Saenz et al. 2007).

Loggerhead sea turtles have higher mercury leve ls than any other se a turtle studied, but concentrations are an order of magnit ude less than many toothed whales (Godley et al. 1999, Pugh and Becker 2001b). Arsenic occurs at levels several fold more concentrated in loggerhead

sea turtles than marine mammals or seabirds.

Also of concern is the spread of antimicrobial agents from human society into the marine environment. Loggerhead sea turtles may harbor antibiotic-resistant bacteria, which may have developed and thrived as a result of high use and discharge of antimicrobial agents into freshwater and marine ecosystems (Foti et al. 2009).

Critical habitat. The NMFS has not designated critical habitat for loggerhead sea turtles.

Anadromous fish

Gulf of Maine Atlantic salmon87 Description of the species. Gulf of Maine (GOM) DPS Atlantic salmon occur along the Atlantic coast from the Androscoggin River (Mai ne) in the south to the St. Croix River on the

US-Canadian border. The lower Penobscot River has three primary trib utaries that contain Atlantic salmon: Cove Brook, Kenduskeag Str eam, Kennebec and Ducktrap rivers. The estimated population of Atlantic salmon in the lower Penobscot Ri ver and its tributaries is less

than 20 adult Atlantic salmon. A tlantic salmon are also listed in the Dennys River, East Machias River, Machias River, Pleasant River, Narraguagus River, and Sheepscot River.

Distribution. The Atlantic salmon is an an adromous fish species that is native to the basin of

the North Atlantic Ocean from the Arctic Circle to Portugal in the eastern Atlantic Ocean, from Iceland and southern Greenland, and from the U ngava region of northern Quebec south to the Connecticut River (Scott and Crossman 1973a). In the US, Atlantic salmon historically ranged from Maine south to Long Island Sound. However, the central New England and Long Island Sound DPSs have been extirpated (65 FR 69459).

Habitat. The salmon’s preferred spawning habitat is co arse gravel or rubbl e substrate (up to 3.5 inches in diameter) with adequate water circ ulation to keep the bur ied eggs well oxygenated (Peterson 1978). Water depth at spawning sites is typically between one and 2 feet deep, and water velocity averages 2 feet per second (Beland 1984). Spawning sites, or redds, average 8 feet long and 4.5 feet wide and are often located at the downstream end of riffles where water percolates through the gravel or wher e upwellings of groundwater occur (Moir et al. 1998). The annual egg production is approxi mately 240 eggs per 1,075 feet 2 of fluvial habitat (Chaput et al. 1998).

Movement, growth, and reproduction. Adult Atlantic salmon as cend the rivers of New England beginning in the spring a nd continuing into the fall, w ith peak numbers occurring in June. Although spawning does not o ccur until late fall, the majority of Atlantic salmon in Maine enter freshwater between May and mid-July (M eister 1958, Baum 1997). Salmon that return in early spring spend nearly 5 months in the river before spawning, of ten seeking cool water refuge (e.g., deep pools, springs, and mouths of smaller tributaries) during the summer months. Once an adult salmon enters a river, rising river temperatures and water flows stimulate upstream migration. Approximately 80% of salmon return to their home river after two years at sea, measuring approximately 2.5 feet long and weighing approximatel y 10 pounds (USFWS 2005b). A minority (10 to 20%) of Maine salmon return as sm aller fish, or grilse, after only one winter at sea and still fewer return after three years at s ea. A spawning run of sa lmon with representation of several age groups ensures some level of genetic exchange among generations. Once in freshwater, adult salmon cease feeding during thei r up-river migration. Spawning occurs in late October through November. Spawning sites are pos itioned within flowing water, particularly where upwelling of groundwater occurs, allowing for percolation of water through the gravel (Danie et al. 1984). These sites are most often positio ned at the head of a riffle (Beland et al. 1982), the tail of a pool; or the ups tream edge of a gravel bar where water depth is decreasing, water velocity is increasing (White 1942, McLaughlin and Knight 1987), and hydraulic head allows for permeation of water through the redd (a gravel depression where eggs are deposited).

A single female may create several redds before de positing all of her eggs. Female anadromous Atlantic salmon produce a total of 1,500-1,800 eggs per kilogram of body weight, yielding an88 average of 7,500 eggs per two sea-winter (SW) female (an adult female that has spent two winters at sea before returning to spawn) (Baum and Meister 1971).

After spawning, most Atlantic salm on move immediately downstream to backwater habitats near the head of tide (Cunjak et al. 1998, Fay et al. 2006). Upon returning to salt water, the spawned salmon or kelt resume feeding. If the salmon surv ives another one or two years at sea, it will

return to its home river as a repeat spawner. From 1967 to 2003, approximately 3% of the wild and naturally reared adults that returned to ri vers where adult returns are monitored--mainly the Penobscot River--were repeat spawne rs (USASAC 2004). Hatchery fish also return to the rivers into which they are stocked (Gorsky et al. 2009).

In late March or April, the eggs hatch into al evins. Alevins remain in the redd for about six weeks and are nourished by their yolk sac. Alev ins emerge from the gravel about mid-May, generally at night, and begin actively feed ing (Gustafson-Greenwood and Moring 1991). Survival from the egg to fry stage in Maine is estimated to range from 15-35% (Jordan and Beland 1981). Survival rates of eggs and larvae are a function of stream gradient, overwinter temperatures, interstitial flow, predation, di sease, and competition (Bley and Moring 1988b). Once larval fry emerge from the gravel and begin active feeding they are referred to as fry. The majority of fry (>95%) emerge from redds at night (Gustafson-Marjenan and Dowse 1983). The survival rate of fry is affected by stream grad ient, overwintering temperatures and water flows, and the level of predation and competition (Bley and Moring 1988a).

Within days, the free-swimming fry enter the pa rr stage, moving downstream to areas with adequate cover (rocks, vegeta tion, overhanging banks, and woody de bris), water depths ranging from approximately four to 24 inches, veloci ties between 1foot and 3 feet per second, and temperatures near 61oF (Beland 1984). When they finally reach their desired habitats, parr will actively defend territories that vary in size depending on the amount of food available and the density of other parr in the area (Symons 1971, McCormick et al. 1998, Armstrong et al. 1999). Some male parr become sexually mature and can su ccessfully spawn with s ea-run adult females. Water temperature, appetite, parr density, photo period, the level of competition and predation, and food supply may all influence the growth rate of parr (Lundqvist 1980, Randall 1982, Hearn 1987, Fausch 1988, Metcalfe et al. 1988, Elliot 1991, Nicieza and Metcalfe 1997). Maine Atlantic salmon parr densities are typically between thr ee and nine parr per 1,075 feet 2 , with years up to 16 parr per 1,075 feet 2 not uncommon (Beland 1996). There is no evidence of density-dependent limitations at densities of 13 parr per 1,075 feet 2 (Whalen et al. 2000). Parr feed on larvae of mayflie s, stoneflies, caddisflies, chironom ids, blackflies, annelids, and mollusks, as well as numerous terrestrial insect s that fall into the river (Scott and Crossman 1973a).

In a parr’s second or third spring, when it has grown 5 to 6 inches long, physiological, morphological, and behavioral changes occur (S chaffer and Elson 1975). This process, called smoltification, prepares parr for the dramatic ch ange in osmoregulatory needs that comes with

the transition from a freshwater to a saltwater habitat (Hoar 1976, McLeese et al. 1994,

McCormick et al. 1998). In southern latitudes, incl uding New England, most parr smolt after one year, but in cooler areas, they may take tw o to four years in fres hwater before smolting (McCormick et al. 1998). Most smolts in New England ri vers enter the sea during May and June89 to begin their ocean migration. Maine rivers produce approximately three smolts per 1,075 feet 2

of habitat.

Atlantic salmon of US origin are highly migratory, undertaki ng long marine migrations from the

mouths of US rivers into the northwest Atlantic Ocean, where they are distributed seasonally over much of the region (Reddin 1985). The ma rine phase starts with smoltification and subsequent migration through the natal river and estuary. Upon completion of the physiological transition to saltwater, the post- smolt stage grows rapidly and ha s been documented moving in small, loosely aggregated schools near the surf ace (Dutil and Coutu 1988). After entering the nearshore waters of Canada, the post-smolts beco me part of a mixture of stocks of Atlantic

salmon from various North American streams. Post-smolts appear to feed opportunistically on macroinvertebrates, amphipods, euphausiids, and fish (Hansen and Pethon 1985, Hansen and Quinn 1998, Andreassen et al. 2001). Once they mature to adu lt salmon, they trav el individually and primarily eat capelin, herring, and sand lance (Hansen and Pethon 1985, Reddin 1985, Hansen and Quinn 1998).

Status and trends. The GOM DPS of anadromous Atlant ic salmon was listed by the USFWS and NMFS as an endangered species on N ovember 17, 2000 (65 FR 69495). The GOM DPS encompasses all naturally reproducing remnant popul ations of Atlantic salmon downstream of

the former Edwards Dam site on the Kennebec Rive r northward to the mouth of the St. Croix River. To date, Atlantic salmon are listed in the Dennys, East Maccias, Machias, Pleasant, Narraguagus, Ducktrap, and Sheepscot Rivers, Ke nduskeag Stream, and Cove Brook. Naturally reproducing Atlantic salm on in the Penobscot River and its tributaries downstream of the former Bangor Dam are listed as endangered. The USFW S’s GOM DPS river-speci fic hatchery-reared fish are also included as part of the listed entity (73 FR 51415).

Anadromous Atlantic salmon were native to nearly every major coastal river north of the Hudson River, New York (USFWS 2005b). The annual historic Atlantic salmon adult population returning to US rivers has been esti mated to be between 300,000 and 500,000 (Stolte 1981, Beland 1984). The largest historical salmon runs in New England were likely in the Connecticut, Merrimack, Androscoggin, Kennebec, and Penobscot Rivers.

By the early 1800s, Atlantic salmon runs in New England had been severely depleted, reducing the distribution in the southern half of its range . Restoration efforts we re initiated in the mid1800s, but there was little success (Stolte 1981). Th ere was a brief period of success in the late

19 th century when limited runs were reestablishe d in the Merrimack and Connecticut Rivers by artificial propagation, but these r uns were extirpated by the end of the century. By the end of the

19 th century, three of the five largest salm on populations in New England (Connecticut, Merrimack, and Androscoggin Rivers) had been eliminated.

Abundance of adult Atlantic salmon is estimated using traps at a fishway, or through redd (nest) counts. Total trap counts, which include wild and hatchery fish, and total number of redds

counted in GOM DPS between 1997 and 2004 are de picted in Figure 5. Such counts typically

underestimate the actual retu rns of Atlantic salmon, but can give an idea of trends over time for index reaches and watershed. Juvenile smolt pr oduction is another measure of population trends, growth rate, and densities.90 0 1,000 2,000 3,000 4,000 5,000 6,000 7,000

1967 1970 1973 1976 1979 1982 1985 1988 1991 1994 1997 2000 2003 2006

yearadult returns Penobscot River All Other GOM DPS Rivers

Figure 5. Adult returns to the GOM DPS 1967-2007.

Recently, Fay et al. (2006) used Population Viability An alysis (PVA) techniques to determine

the conservation status of Atlantic salmon in th e GOM DPS. Composite spawner data used to populate the model included adult return and rod kill estimates from the Penobscot River, adult spawner and rod kill estimates for the Narraguagu s River, and adult spawner estimates for the GOM DPS. Using two time seri es, 1984 to 2004 and 1991 to 2004, Fay et al. (2006) calculated the negative population growth rates (for 1980-2004, lambda = 0.9690, variance = 0.0261; for 1991-2004, lambda = 0.9471, variance = 0.0142). From this, the estimated risk of extinction (defined herein as the number of spawners that falls below 100 individual s) within 100 years is 61% and 75% (or 28% and 45% in 40 y ears), for each respective data set.

Natural threats. Geographic features, such as waterfal ls, pose natural barriers to salmon migration to spawning habitat. A variety of di seases affect Atlantic salmon, but are exacerbated by the presence of farming pens near river mouths. Atlantic salmon are prey for a variety of predators, including seals, porpoise s, dolphins, otters, minks, birds, sharks, and a variety of other fishes at various salmon life stages.

Anthropogenic threats. Humans pose numerous threats to Atlantic salmon survival and recovery (see USFWS 2005b for a review). Water quality in both marine, estuarine, and aquatic habitats suffers from both point and non-point so urce pollution, both biolog ical (bacteria) and

chemical. Riverine environments are becomi ng acidified, which can cause physiological stress in adults and altered developmental biology in eggs or hatchlings. In association with acidification, aluminum toxicity can lead to os moregulation failure. This is because Atlantic salmon are highly sensitive to pH changes a nd many runs of Atlantic salmon in Sweden, Norway, and Canada have been severely depleted or extirpated due to acidity changes in river systems resulting from industrial activity (Watt 1981, Watt et al. 1983, Watt et al. 2000, Sandøy and Langåker 2001). Pesticide use a nd its immigration into Maine wa terways is also of concern.91 For example, atrazine can significantly impa ir water balance in salmon even at low concentrations, resulting in a reduced ability for salmon to move betw een fresh and salt water (Jagoe and Haines 1990, Staurnes et al. 1993, WWF 2001). At levels that presently occur in stream environments, male salmon also experience impaired olfactory reception in being able to detect female pheromones (Waring and Moore 1 998). Thus, male reproduction activity is not cued to that of females and has the potential to severely reduce recru itment. Nonylphenols are also severely detrimental to juvenile salmon. Th ese chemicals also reduce the ability of smolts to transition between fresh and salt water, l eading to mortality, as well as imitate female hormones leading to eggs that do not hatch (Fairchild et al. 1999, WWF 2001). Sedimentation due to erosion and development in and around a quatic waterways can de grade salmon habitat

and the habitat of their invertebra te prey. Excessive nutrient load, as in marine systems, can lead to a bloom of plant growth and subsequent death, which reduces oxygen levels to anoxic

conditions. This can lead to extensiv e habitat loss and salmon mortality.

Although changes overtly seem minor, increases in Maine’s river temperatures can have broad impacts on salmon recovery, including changes in fish physiology, prey abundance and distribution, loss of spawning activity, and other effects (USFWS 2005b, Holbrook et al. 2009). As in Pacific salmon species, Atlantic salmon dec line originated largely from manmade barriers across rivers preventing moveme nt to and from spawning and marine habitats. Although many

of these barriers have since been modified or removed, modern construc tion (bridges, culverts, etc.) that do not consider A tlantic salmon needs can hinde r recovery efforts (Holbrook et al. 2009). When water temperatures exceed 22o C during spawning runs, Atlantic salmon tend to have poorer success in passi ng obstacles than (Holbrook et al. 2009).

Atlantic salmon fisheries have been discontinued in the US , Canada, and Greenland. A high threat is posed by farm-raised salmon due to the potential for these fish to escape (instances of thousands of fish escaping are known) and interbreed with wild salmon, thereby affecting the genetics of Atlantic salmon as a species. Recent evidence shows that supportive breeding programs, where wild Atlantic salmon are captured and bred in captivity and young are released early in life, produce fish that are genetically, morphologically, a nd behaviorally different from truly wild progeny (Blanchet et al. 2008). The presence of disease and parasites in farm-raised salmon pens can also have a deleteri ous effect on wild Atlantic salmon.

Climate change has the potential to be a strong negative influence on Altantic salmon. Remaining occupied habitat is at the southern e dge of the ESU’s range. To survive, populations have adapted to distinct physic al and environmental conditions here (Saunders 1981). Climate models predict significant, extended warming (IPCC 2001b). Although periods of North Atlantic warming and cooling have occurred, ch anges have not been uniform as global warming is, changing sea temperatures, wind currents, fr esh water input, and mixing of the ocean’s surface layer. Small thermal changes can critical ly affect biological functions, such as protein metabolism, response to aquatic contaminants, reproductive performance, smolt development, and species distribution limits (K eleher and Rahel 1996, McCormick et al. 1997, Reid et al. 1997, Som ero and Hofmann 1997, Van de r Kraak and Pankhurst 1997, Welch et al. 1998). Atlantic salmon smolt growth is known to change with temperature, with a temperature increase from 57o to 64oF resulting in a greater than 10% decrease in growth rate (Handeland et al. 2008).92 It should be noted that positive effects may also be realized by climate change and specifically warmer water temperature. Increased opportun ities for growth in spring and summer could increase the percentage of fish that enter the upper size distribu tion of a population and smolt the following spring (Thorpe 1977, Thorpe et al. 1980, Thorpe 1994). In addition, warmer rearing temperatures during the late wi nter and spring have been show n to advance the timing of the parr-smolt transformation in Atlantic salmon (Solbakken et al. 1994). There is, however, an optimal temperature range and a limit for growth after which salmon parr will stop feeding due to thermal stress. During this time, protein de gradation and weight loss will increase with rising water temperature (McCarthy and Houlihan 199 7). The NRC (2004) concluded that some degree of climate warming or change in hydrologic regime could be tolerated if other problems affecting Atlantic salmon are reduced.

Critical habitat. On June 19, 2009, 45 specific areas occupied by Atlantic salmon at the time of listing (approximately 19,571 km of perennial river, stream, and estuary habitat and 799 square kilometers of lake habitat within the range of the GOM DPS) were es tablished for Atlantic salmon critical habitat (74 FR 29300). The PCEs for this critical habitat include:

 Deep, oxygenated pools and cover (e.g., boulders , woody debris, vegetation, etc.), near freshwater spawning sites, necessary to s upport adult migrants during the summer while they await spawning in the fall.

 Freshwater spawning sites that contain clean, permeable gravel and cobble substrate with oxygenated water and cool water temperat ures to support spawning activity, egg incubation, and larval development as well as support emergence, territorial development and feeding activities of Atlantic salmon fry.

 Freshwater rearing sites with space to accomm odate growth and survival of Atlantic salmon parr.

 Freshwater rearing sites with a combination of river, stream, and lake habitats that accommodate parr’s ability to occupy many niches and maximize parr production.

 Freshwater rearing sites with cool, oxygenated (6 mg/L) water and diverse food resources (mayflies, stoneflies, chironomid s, caddisflies, blackflies, aq uatic annelids, and mollusks,

as well as numerous terrestrial invertebrate s, alewives, dace, or minnows) to support growth and survival of Atlantic salmon parr.

 Freshwater and estuary migrat ory sites free from physical and biological barriers that delay or prevent access of adult salmon seeking spawning grounds needed to support recovered populations or prevent emigration of smolts to the marine environment.

 Freshwater and estuary migrati on sites with pool, lake, and in stream habitat that provide cool, oxygenated water and cover items (e.g., b oulders, woody debris, and vegetation) to serve as temporary holding and resting areas during upstream migratio n of adult salmon.

 Freshwater and estuary migration sites with abundant, diverse native fish communities to serve as a protective buffer against predation.

 Freshwater and estuary m igration sites with sufficiently cool water temperatures and

water flows that coincide with diurna l cues to stimulate smolt migration.

 Freshwater migration sites with water chemis try (particularly pH) needed to support sea water adaptation of smolts.

These PCEs have undergone significant degradat ion over in the recent past. Dams, along with93 degraded substrate and cover, water quality, water temperature, and biological communities, have reduced the quality and quantity of habitat available to Atlantic salmon populations within the DPS. A combined total of twenty FERC- licensed hydropower dams on the Penobscot River significantly impede the migration of Atlantic salmon and other diadromous fish to historically accessible spawning and rearing hab itat. Agriculture and urban development largely affect the

lower third of the Penobscot River below the Pi scataquis River sub-basi n by reducing substrate and cover, reducing water quality, and eleva ting water temperatures . Introductions of smallmouth bass and other non-indigenous species significantly degrade habitat quality throughout the mainstem Penobscot and portions of the Mattawamkeag, Piscataquis, and lower Penobscot sub-basins by altering pr edator/prey relationships. Sim ilar to smallmouth bass, recent Northern pike introductions threaten habitat in the lower Penobscot River below the Great Works

Dam.

Today, dams are the greatest impediment, outside of marine survival, to the recovery of salmon in the Penobscot, Kennebec and Androscoggin river basins (Fay et al. 2006). Hydropower dams significantly impede the migrati on of Atlantic salmon and other diadromous fish and either reduce or eliminate access to historically accessibl e spawning and rearing habitat. In addition to hydropower dams, agriculture and urban developm ent largely affect the lower third of the Merrymeeting Bay recovery unit by reducing substrate and cover, reducing water quality, and elevating water temperatures. Additionally, smallmouth bass and brown trout introductions, along with other non-indigenous species, signifi cantly degrade habitat quality throughout the Merrymeeting Bay recovery unit by altering natu ral predator/prey relationships.

Impacts to substrate and cover, water quality, water temperature, biological communities, and migratory corridors, among a host of other factors, have impacted the qu ality and quantity of habitat available to Atlantic salmon populations within the Downeast Coastal recovery unit. Two hydropower dams on the Union river, and to a lesser extent the sma ll ice dam on the lower Narraguagus River, limit access to spawning and rear ing habitat within these two watersheds. In the Union River, physical and biological features have been most notably limited by high water temperatures and abundant smallmouth bass populat ions associated with impoundments. In the Pleasant River and Tunk Stream, which collectiv ely contain over 4,300 units of spawning and rearing habitat, pH has been identified as possibly being the predominate limiting factor. The Machias, Narraguagus, and East Machias rive rs contain the highest quality habitat and collectively account for approxima tely 40 percent of the spawni ng and rearing habitat in the Downeast Coastal recovery unit.

Chinook salmon Description of the species. Chinook salmon are the largest of the Pacific salmon and historically ranged from the Ve ntura River in California to Point Hope, Alaska in North America, and in northeastern Asia from Hokkai do, Japan to the Anadyr River in Russia (Healey 1991a). In addition, Chinook salmon have been re ported in the Canadian Beaufort Sea (McPhail and Lindsey 1970). We discuss the distribution, status, and critical habitats of the species 2 of

2 We use the word “species” as it has been defined in sec tion 3 of the ESA, which incl ude “species, subspecies, and any distinct population segment of any species of vertebra te fish or wildlife which interbreeds when mature (16 U.S.C 1533).” Pacific salmon that have been listed as en dangered or threatened were listed as “evolutionarily94 endangered and threatened Chinook salmon separately, and summarize their common dependence on waters of the United States. Ho wever, because Chinook salmon in the wild are virtually indistinguishable between listed species, and are the same biol ogical species we begin this section describing thos e characteristics common acro ss the listed species.

Of the Pacific salmon species considered here in, Chinook salmon exhibit arguably one of the most diverse and complex life history strategies with multiple races wi thin which there is substantial variation. One form, the “stream-type,” resides in freshwater for a year or more following emergence and the “ocean-type” migrates to the ocean within their first year. The ocean-type typifies populations north of 56o N (H ealey 1991a). Within each race, there is often variation in age at seaward migration, age of maturity, timing of spawning migrations, male precocity, and female fecundity.

Reproduction. The general Chinook salmon life cycle sp ans fresh and marine waters, with one reproductive event per adult (Chinook salmon are semelparous and die after spawning). Spawning migrations generally o ccur in the spring and fall, although the precise timing of spawning migrations and spawning varies across populations and can vary within populations. Temperature and stream flow can significantly in fluence the timing of upstream migrations and spawning, and the selection of spawning habitat (Geist et al. 2009, Hatten and Tiffan 2009). For Klamath River Chinook, temperatures above 21. 9o C (mean average body temperature), 20.6o C (mean minimum daily body temperature), or 23. 1o C (mean maximum daily body temperature) completely inhibited upstream spawning migration in rivers; these values are close to the upper lethal limits for this and other salmonid species (Strange 2010). .A gene ral latitudinal cline is apparent across the species range with spawning typically occurring earlie r in the spring/summer

at northern latitudes and later in southern latitude s (Healey 1991a).

On the spawning grounds, mate competition is intense with males competing to fertilize eggs and females competing for optimal nest site selec tion. Once fertilization occurs, female Chinook salmon bury the eggs in nests –termed “redds”- and guard the nests until their death, which generally occurs a couple days later to a couple weeks after spawning. A female generally deposits eggs in more than one depression with in a redd, excavating stream rock as she moves upstream, increasing the size of her re dd until all eggs are deposited.

Size and age at maturity is partially under genetic control, but can be influenced by environment and migration behavior (Roni and Quinn 1995). Generally, ocean-type salmon are at sea longer than their stream-type counterparts and tend to be larger in size at sp awning. Body size can be important in determining reproductive success in terms of nest selection and mating competition (Foote 1990). Chinook salmon age at maturity range s from 1 to 7 years with most returning to spawn between 3 and 4 years of age.

Habitat. The time necessary for egg incubation until emergence of alevins in freshwater varies among basins and among years within a basin, and is closely correlated to water temperatures such that low temperatures can prolong incuba tion. Incubation generally takes a couple of months or more. Alevin (also called “yolk-sa c” fry) remain buried until their yolk-sac is

significant units (ESU)” which NMFS uses to identify distinct population segments (DPS) of Pacific salmon. Any ESU or DPS is a “species” for the purposes of the ESA.95 absorbed, at which time they become free swimmi ng fry. Egg to fry survival can also vary widely across basins, years, and habitat conditions within a basin. In ge neral, the survival of eggs and alevin, and the fitness of emerging fr y are affected by sedime nt loading, intergravel water flow and dissolved oxygen levels, gravel composition, spawn timing, floods, redd and spawner density, and water temperatures.

Once emerged, fry behavior varies among popula tions and among individuals within races. Some juvenile Chinook salmon rear in freshwater for a few weeks to a few years, others move immediately downstream to coasta l waters where they rear in estuaries for a few weeks to months, while others migrate directly to o cean waters. Stream-type Chinook salmon do not migrate to sea until the spring following emerge nce, and ocean-type Chinook salmon migrate to the ocean within their first year. Generally, most fry move at night probably to reduce detection by predators, although some fish will move downstream during daylight. Not all movement is volitional as stream flows often displace fry to downstream areas after emergence. Densitydependent factors such as space, prey, or stream flows may influence the outmigration behavior of individual juvenile Chinook salmon.

While in fresh water, juvenile Chinook salmon ar e often found in the lowe r reaches of a river near its estuary, where they inhabit river margins in areas of shallow water, near woody debris,

or other areas of low water veloc ity. As juveniles grow in size, they tend to move away from the shoreline to deeper waters where the velocity is higher (Healey 1991a). The transformation from the freshwater fry/parr juvenile stage to smolt involves multiple physiological changes including increases in: body silvering, hypoosmotic regulatory capability, salinity tolerance and preference, growth rate, oxygen consumption, ammonia produ ction, endocrine activity (e.g., activation of thyroid, interregnal and pituitary growth hormone cells), and gill Na + , K + -ATPase activity. At the same time, body condition declines (Wedemeyer et al. 1980). Several factors can affect the smoltification process, not only at the interface between freshwater and saltwater, but higher in the watershed as the process of transformation begi ns long before fish enter saltwaters, including exposure to heavy metals and elevat ed water temperatures (Wedemeyer et al. 1980).

Life at sea varies according to population, race, a nd age-class. Chinook salmon tend to remain at sea between 1 and 6 years, with most fish returnin g to freshwater after 2-4 years at sea. Fishery

catches indicate that ocean- and stream-type fish exhibit divergen t migratory pathways while in the ocean (Healey 1983, 1991a). Ocean-type Chinook salmon tend to be found along the coastline, whereas stream-type Chinook salmon ar e found in the open ocean far from the coast (Healey 1983, 1991a). Juvenile Chinook along the Pacific northwest, Canada, and Alaska tend to remain within roughly 55 km (most with in 28 km) of the coast (NPFMC 1990, PFMC 2000). However, Chinook generally remain within 320 km of the coast (NPFMC 1990). Concentrations are known to occur around transient upwelling fe atures (PFMC 2000). I ndividuals hatching in rivers south of Cape Blanco, Oregon tend to stay south of this point, while those outmigrating north of this point tend to move north into the Gulf of Alaska through co astal migratory corridors (PFMC 2000).

Generally, Chinook salmon outmigrants (termed smo lts) are about 5-13 cm long when they enter saline (often brackish) waters. The process of smoltification is phys iologically demanding, involving osmoregulation—the maintenance of osmotic pressure as the fish enters waters of96 increased salinity, which is necessary to main tain body fluid concentration and composition to

maintain homeostasis. Smaller fish tend to remain closer to shore, while larger fish will enter

marine waters. The distribution of fish in th is stage does not appear to be correlated with salinities, as most fish can survive immediate transfer from freshwater to saltwater (~30 ppm salinity; (Healey 1991a). Once in the ocean, juveniles occupy waters 30-70 m deep and frequently associate with bottom topography (PFMC 2000). Although temperature ranges vary from 1° to 15° C, few individuals are f ound in waters below 5° C (MBC 1987, PFMC 2000). Chinook salmon originating from the same freshwat er region have similar age-dependent marine distributions (even if associated with different runs), which are di stinct from the distributions of Chinook from other freshwater regions (Weitkamp 2009). This distribution does not appear to be influenced by oceanographic variability (Weitkamp 2009). Older individuals appear to disperse more broadly than their younger count erparts (Weitkamp 2009). Adults do not appear to exhibit a preference in substrates and may be found down to 250 m (Beauchamp et al. 1983).

Feeding. Chinook salmon feed on a variety of prey organisms depending upon life stage. Adult oceanic Chinook salmon eat small fish, amphipods, and crab megalops (Brodeur et al. 2010)(Healey 1991). Fish, in particular herring, make up the largest portion of an adult Chinook salmon’s diet. In estuaries, Chinook salmon sm olts tend to feed on the chironomid larvae and

pupae Daphnia

, Eogammarus

, Corphium and Neomysis, as well as juvenile herring, sticklebacks

and other small fish. In freshwater, Chinook salmon juvenile s feed on adult a nd larval insects including terrestrial and aquatic insects such as dipterans, beetles, stoneflies, chironomids, and plecopterans (Healey 1991a). During the fi rst year of marine life, Puget Sound Chinook transition from nearshore foraging on insects and amphipods during June to offshore foraging on crab larvae and fish, such as herri ng during July through September (Duffy et al. 2010).

Natural threats. Chinook salmon are exposed to high rates of natural predation during freshwater rearing and migration stages, as well as during ocean migration. In general, Chinook salmon are prey for pelagic fishes, birds, and ma rine mammals, including ha rbor seals, sea lions,

and killer whales. There have been recent concer ns that the increasing size of tern, seal, and sea lion populations in the Pacific northwest may have reduced the survival of some salmon populations; 10% of salmonid smolts are eaten by Caspian terns and double-crested cormorants annually in the Columbia River estuary (NMFS 2011l). Invasive fishes also threaten the survival and recovery of Pacific salmonids by competing directly for resources, altering food webs and trophic structures, and altering evolut ionary trajectories (NMFS 2011l).

Anthropogenic threats. Chinook salmon have declined under the combined effects of fishery over-harvest; competition from fish raised in hatcheries and native and non-native exotic species; dams that block their migrations and alter ri ver hydrology; gravel mining that impedes their migration and alters the dynamics (hydrogeomorphol ogy) of the rivers an d streams that support juveniles; water diversions that deplete water levels in rivers and st reams; destruction or degradation of riparian habitat th at increase water temperatures in rivers and streams sufficient to reduce the survival of juvenile Chinook salm on; and land use practices (logging, agriculture, urbanization) that destroy wetland and riparian ecosystem s while introducing sediment, nutrients, biocides, metals, and other pollutants into surf ace and ground water and degrade water quality in the freshwater, estuarine, and coastal ecosy stems throughout the Pacific northwest (Buhle et al. 2009). Of several habitat factors for spring-r un Columbia River Chinook salmon, reductions in97 fine sediments may be particularly importa nt to changes in population numbers (Honea et al. 2009).

Population declines have resulted from several human-mediated causes, but the greatest negative influence has likely been the esta blishment of waterway obstructions such as dams, power plants, and sluiceways for hydropower, agricu lture, flood control, and water storage. These structures have blocked salmon migration to spawning habita t or resulted in direct mortality and have

eliminated entire salmon runs as a result. While so me of these barriers remain, others have been reengineered, renovated, or removed to allow for surviving runs to access former habitat, but success has been limited. These types of barri ers alter the natural hydrograph of basins, both upstream and downstream of the stru cture, and significantly reduce th e availability and quality of spawning and rearing habitat (Hatten and Tiffan 2009). Many st reams and rivers, particularly in urban or suburban areas, suffer from streamsi de development, which contributes sediment, chemical pollutants from pesticide applications and automobile or industr ial activities, altered stream flows, loss of streamside vegetation and allochthonous materials to name a few. These factors can directly cause mortal ity, reduce reproductive success, or affect the health and fitness of all salmon life stages.

Artificial propagation of hatchery fish has had profound consequences on the viability of some natural salmon populations, but ther e are potential benefits to th e artificial production of salmon as well. Adverse effects of artificial propaga tion include: a decline in the natural population from the taking of wild brood stoc k for artificial propagation, the genetic erosion of populations (introgression, hybridization), an increase incidence of disease to and increased rates of competition with and predation on naturally spaw ned salmon populations. Potential benefits to artificial propagation include th e bolstering of the numbers of naturally spawning fish in the short-term, the conservation of genetic resources , and guarding against th e catastrophic loss of naturally spawned populations at cr itically low abundance levels.

Fishing for salmon has also negatively impacted salmon populations. Fishing reduces the

number of individuals within a population and can lead to un even exploitation of certain populations and size classes (M undy 1997, Reisenbichler 1997). Ta rgeted fishing of larger

individuals results in excludi ng the most fecund individuals from spawning (Reisenbichler 1997). Genetic changes that promote smaller bo dy sizes have occurred in heavily exploited populations in response to si ze-selective harvest pressu res (Mundy 1997, Reisenbichler 1997,

Swain et al. 2007). Age at maturity can also be accel erated by fishing pressure (Reisenbichler 1997). Pacific salmon species are exposed to a number of contaminants throughout their range and life history cycle.

Exposure to pollution is also of significant concern for all life stag es, but is likely particularly significant for freshwater life stages. Organi c pollutants, especially PCBs, DDT and its congeners, pesticides, and endocri ne disruptor,s are pa rticularly concerning. These chemicals can inhibit smell, disrupt reproductive behavi or and physiology, impair immune function, and lead to mortality through impairment of wate r balance when traveling between fresh- and saltwater systems (Varanasi et al. 1993a, Varanasi et al. 1993b). Diffuse and extensive population centers contribute increase contam inan t volumes and variety from such sources as wastewater treatment plants and sprawling development. Urban runoff from impervious surfaces98 and roadways often contains oil, copper, pest icides, PAHs, and other chemical pollutants and flow into surface waters. Point and nonpoint pollution sources entering rivers and their tributaries affect water quality in available spawning and rearing habitat for salmon. Juvenile

salmonids that inhabit urban watersheds often carry high contaminant burdens, which is partly attributable to the biological transfer of contaminants through the food web (Brown et al. 1985,

Stein et al. 1992, Varanasi et al. 1993a).

Sacramento River winter-run Chinook salmon

Distribution. The Sacramento River winter-run Chi nook salmon ESU includes all naturally spawned populations of winter-run Chinook salmon in the Sacramento River and its tributaries in California. Four natural, independent populations once exis ted (NMFS 2011g). Two artificial propagation programs are included in this ESU: winter-run Chinook salm on from the Livingston Stone National Fish Hatchery, and winter-run Chinook salmon in a captive broodstock program maintained at the Livingston Stone National Fish Hatchery and the University of California Bodega Marine Laboratory. These artificially propagated populations are no more divergent

relative to the local na tural populations than would be e xpected between closely related populations within this ESU.

This ESU consists of a single spawning populatio n that enters the Sacr amento River and its

tributaries in California from November to June and spawns fr om late April to mid-August, with a peak from May to June (Table 7). Sacramen to River winter-run Chinook salmon historically occupied cold, headwater streams, such as the upper reaches of the L ittle Sacramento, McCloud, and lower Pit Rivers. Young winter-run Chi nook salmon venture to sea in November and December, after only four to seven months in fresh water (Groot et al. 1991).

Table 7. Sacramento River winter-run Chinook salmon abundance and selected measures of population viability.

Population Historical

Abundance a Mean number of Spawners (Range) b Percent Hatchery

Contribution Population growth rate ( λ ) c

Sacramento River winter-run 200,000 2,191 (364-65,683) <10 0.97 (0.87, 1.09)

a Historical abundance for the total ESU based on commer cial fishery landings in the 1870s (Fisher 1994). Individual river estimates of historical abundance not provided.

b Recent geometric mean number of spawners from Good et al. 2005.

c Lambda value reported by Good et al. 2005. The 90% confidence interval s are noted in parentheses.

Status and trends. NMFS listed Sacramento River winter-run Chinook salmon as endangered on January 4, 1994 (59 FR 440), and reaffirmed their status as endangered on June 28, 2005 (70 FR 37160), because dams restrict access to a small fraction of their histor ic spawning habitat and the habitat remaining to them is degraded. Sacramento River winter -run Chinook salmon consist of a single self-sustaining population which is entirely dependent upon the provision of suitably cool water from Shasta Reservoir during pe riods of spawning, incubation and rearing.

Construction of Shasta Dams in the 1940s elim inated access to historic spawning habitat for

winter-run Chinook salmon in the basin. Winte r-run Chinook salmon were not expected to

survive this habitat alteration (Moffett 1949). However, cold water releases from Shasta Dam

have created conditions suitabl e for winter Chinook salmon for roughly 60 miles downstream from the dam. As a result the ESU has been reduced to a single spawning population confined to99 the mainstem Sacramento River below Keswic k Dam, although some adult winter-run Chinook salmon were recently observed in Battle Creek, a tributary to the upper Sacramento River.

Quantitative estimates of run-size are not availabl e for the period before 1996, the completion of

Red Bluff Diversion Dam. Howe ver, winter-runs may have been as large as 200,000 fish based upon commercial fishery records from the 1870s (Fisher 1994). The California Department of Fish and Game estimated spawning escapement of Sacramento River winter-run Chinook salmon at 61,300 (60,000 in the mainstem, 1,000 in Battle Creek, and 300 in Mill Creek) in the early 1960s. During the first 3 years of operation of th e county facility at the Red Bluff Diversion Dam (1967 to 1969), the spawning run of wint er-run Chinook salmon averaged 86,500 fish. From 1967 through the mid-1990s, the population declined at an average rate of 18% per year, or roughly 50% per generation. The population reached critically low levels during the drought of 1987 to 1992; the 3-year average run size for th e period of 1989 to 1991 was 388 fish. Based on

the Red Bluff Diversion Dam c ounts, the population has been gr owing rapidly since the 1990s. Mean run size from 1995-2000 has been 2,191, but have ranged from 364 to 65,683 (Good et al. 2005b). Good et al. 2005 estimated that the short term trend is 0.26, while the population growth rate is still less than 1 (Table 13). The draft recovery goal fo r the ESU is an average of 10,000 female spawners per year and a population growth rate >1.0, calculated ov er 13 years of data (Good et al. 2005b). Since 2004, escapement in the central Valley has fallen dramatically from nearly 16,000 in 2004 to about 1,500 in 2010; abundance peaked in 2006 (NMFS 2011g). Abundance has fallen throughout the ESU since 2000 (NMFS 2011g).

Critical habitat. NMFS designated critical habitat fo r Sacramento River winter-run Chinook salmon on June 16, 1993 (58 FR 33212). The following areas consisting of the water, waterway bottom, and adjacent riparian zones: the Sacr amento River from Keswick Dam, Shasta County (river mile 302) to Chipps Island (river mile 0) at the westward margin of the Sacramento-San Joaquin Delta, and other specified estuarine wate rs. These areas are important for the species’ overall conservation by protec ting quality growth, reproduc tion, and feeding. Factors contributing to the downward trends in this ESU include reduced acce ss to spawning/rearing habitat, possible loss of geneti c integrity through population bottl enecks, inadequately screened diversions, predation at artif icial structures and by nonnativ e species, pollution from Iron Mountain Mine and other sources, adverse flow conditions, high summer water temperatures, unsustainable harvest rates, passage problems at various structures, and vulnerability to drought (Good et al. 2005b).

California coastal Chinook salmon

Distribution. The California Coastal Chinook salmon ESU includes all naturally spawned populations of Chinook salmon from rivers and streams south of the Klamath River to the Russian River, Califorian. Seven artificial pr opagation programs are part of this ESU: The Humboldt Fish Action Council (Freshwater Cr eek), Yager Creek, Redwood Creek, Hollow Tree, Van Arsdale Fish Station, Mattole Salmon Group, and Mad River Hatchery fall-run Chinook hatchery programs. These artificially propagate d populations are no more divergent relative to the local natural populations than would be expe cted between closely re lated populations within

this ESU.

California C oastal Chinook salmon are a fall-run, ocean-type fish. A spring-run (river-type)100 component existed historically, but is now considered extinct (Bjorkstedt et al. 2005). Table 8 identifies populations within the California Coastal Chinook salmon ESU, their abundances, and

the relative contribution of ar tificially propagated fish to th e population. Although difficult to identify, it appears that 17 populations comprise the ESU (NMFS 2011a).

Table 8. California coastal Chinook populations and selected measures of population viability.
Population Historical

Abundance a Mean Number of Spawners

(Range) b Percent Hatchery

Contribution c Long-term Trend d

Freshwater Creek 22 (13-22) 30-70 0.137 (-0.405, 0.678)

Eel River 17,000-55,000 ~30

Mainstem Eel River 13,000 Sprowl Creek 43 (43-497) -0.096 (-0.157, - 0.034)

Tomki Creek 61 (13-2233) -0.199 (-0.351, - 0.046)

Van Duzen River 2,500

Middle Fork Eel River 13,000

South Fork Eel River 27,000

North Fork Eel River Upper Eel River

Redwood Creek 1,000-5,000

Mad River 1,000-5,000 Canyon Creek 73 (19-103) 0.0102 (-0.106, 0.127)

Bear River 100

Mattole River 1,000-5,000 ~17

Russian River 50-500 ~0 Humbolt Bay tributaries 40

Tenmile to Gualala 0 Small Humboldt County rivers 1,500 0

Rivers north of Mattole River 600 0 Noyo River 50 0

a Historical abundance estimates based on professional op inion and evaluation of hab itat conditions (reported in

Good et al. 2005).

b 5-year (1997-2001) geometric mean number of counts of adults (quasi-systematic surveys of spawners – Canyon, Tomki, and Sprowl creeks; returning spaw ners at Freshwater Creek weir).

c Hatchery production in this ESU is at low levels, aimed at supplementing depressed runs. Operational procedures and low production suggest that the ESU may not be at substantial risk of degrad ed genetic integrity (Good et al. 2005).

d Long-term trends were calculated using the entire available data set (see Good et al. 2005). The 90% confidence intervals are noted in parentheses.

Status and trends. NMFS listed California Coastal Chinook salmon as threatened on September 16, 1999 (64 FR 50393), and they retain ed their threatened status on June 28, 2005 (70 FR 37160). California Coastal Chinook salmon we re listed due to the combined effect of dams that prevent them from reaching spawni ng habitat, logging, agricultural activities, urbanization, and water withdrawals in the ri ver drainages that suppor t them. Historical estimates of escapement, based on professional opinion and evaluation of habitat conditions, suggest abundance was roughly 73,000 in the early 1960s with the majority of fish spawning in the Eel River (CDFG, 1965 #272 in Good et al. 2005b). The species exists as small populations with highly variable cohort si zes. Sproul Creek abundance (p art of the lower Eeel River101 population) has been on a positive trend since 19 95, but remains negative over longer time series (NMFS 2011a). The Russian River probably contains some natural producti on, but the origin of

those fish is not clear because of a number of intr oductions of hatchery fish over the last century. Average number of returning spawners in th e Russian River has slow ly declined since 2001 (NMFS 2011a). The Eel River contains a substa ntial fraction of the remaining Chinook salmon spawning habitat for this species. Since its original listing and st atus review, little new data are available or suitable for analyzi ng trends or estimating changes in this population’s growth rate (Good et al. 2005b). Adult spawning returns in Pr airie Creek (part of the Redwood Creek population) have been on the decline since th ey were first monitored in 1998 (NMFS 2011a).

Long-term trends in Freshwater Creek were positive until 2008 and 2009, when only two fish were caught in weirs (NMFS 2011a). In Canyon Creek, although abundance trends are only slightly different than zero, th e trend is positive (Table 14). Long-term trends in Sprowl and Tomki creeks (tributaries of the Eel River), however, are negative. Good et al. (2005b) caution making inferences on the basin-wide status of these populations as they may be weak because the data likely include unquantifie d variability due to flow-related changes in spawners’ use of mainstem and tributary habitats. Unfortunately, none of the available data is suitable for analyzing the long-term trends of the ESU or estimating the population growth rate.

Spring-run Chinook appear to be extirpated from the north-caostal and north mountain interior strata of the ESU, representing a considerable loss of diversity (NMFS 2011a). In addition, all populations south of the Mattole River to the Russian River has resulted in the entire northcentral coastal stratum being unoccupied and the central coastal stratum has only one declining population (NMFS 2011a).

Critical habitat. NMFS designated critical habitat fo r California Coastal Chinook salmon on September 2, 2005 (70 FR 52488). Specific geogra phic areas designated include the following CALWATER hydrological units: Redwood Creek, Trinidad, Mad River, Eureka Plain, Eel River, Cape Mendocino, Mendocino Co ast, and the Russian River. These areas are important for the species’ overall conserva tion by protecting quality growt h, reproduction, and feeding. The critical habitat designation for this ESU identifies primary constituent elements that include sites necessary to support one or more Chinook salmon life stages. Specific sites include freshwater spawning sites, freshwater reari ng sites, freshwater migration co rridors, nearshore marine habitat and estuarine areas. The physical or biological featur es that characterize these sites include water quality and quantity, natural cover, forage , adequate passage c onditions, and floodplain connectivity. The critical habi tat designation (70 FR 52488) cont ains additional details on the sub-areas that are included as part of this designation, and the areas th at were excluded from designation.

In total, California Coastal Chinook salmon occ upy 45 watersheds (freshwa ter and estuarine). The total area of habitat designa ted as critical includes about 1,500 miles of stream habitat and about 25 square miles of estuarine habitat, mo stly within Humboldt Bay. This designation

includes the stream channels with in the designated stream reaches, and includes a lateral extent as defined by the ordinary high water line. In areas where the ordinary high-water line is not

defined the lateral extent is defined as the bankfu ll elevation. In estuarine areas the lateral extent

is defined b y the extreme high water because extreme high tide areas encompass those areas102 typically inundated by water and regularly occu pied by juvenile salmon during the spring and summer, when they are migrating in the nearshore zone and rely ing on cover and refuge qualities provided by these habitats, and wh ile they are foraging. Of the 45 watershed reviewed in NMFS'

assessment of critical habitat fo r California Coastal Chinook salmon, eight watersheds received a low rating of conservation value, 10 received a medium rating, and 27 received a high rating of conservation value for the species.

Critical habitat in this ESU consists of limited quantity and quality summ er and winter rearing habitat, as well as marginal spawning habitat. Compared to historical conditions, there are fewer pools, limited cover, and reduced habitat complexit y. The limited instream cover that does exist is provided mainly by large cobble and overhan ging vegetation. Instream large woody debris, needed for foraging sites, cover, and velocity refu ges is especially lacking in most of the streams throughout the basin. NMFS has determined that th ese degraded habitat conditions are, in part, the result of many human-induced factors affecting critical hab itat including dam construction, agricultural and mining ac tivities, urbanization, stream cha nnelization, water diversion, and logging, among others.

Lower Columbia River Chinook salmon

Distribution. The Lower Columbia River Chinook sa lmon ESU includes all naturally spawned populations of Chinook salmon from the Columbia Ri ver and its tributaries from its mouth at the Pacific Ocean upstream to a transitional point between Washington and Oregon, east of the Hood River and the White Salmon River, and includes the Willamette River to Willamette Falls, Oregon, exclusive of spring-run Chinook salmon in the Clackamas River (Figure 6). Seventeen artificial propagation programs are part of this ESU. These artificially propagated populations are no more divergent relative to the local natural populations th an would be expected between closely related populations within this ESU.103 Figure 6. Population boundaries, dams, accessible areas, and ex tirpated reaches of lower Columbia River Chinook distribution, incorporating critical habitat boundaries (dark black line) and fall versus spriung runs. Taken from NMFS (2011).

Lower Columbia River Chinook salmon have three life history types: early fa ll runs (tules), late fall runs (brights), and spring-runs. Spring and fall runs have been designat ed as part of a Lower

Columbia River Chinook salmon ESU. The Cowlitz, Kalama, Lewis, White Salmon, and Klickitat Rivers are the major river systems on the Washington side, and the lower Willamette and Sandy Rivers are foremost on the Oregon side. The eastern boundary for this species occurs at Celilo Falls, which corresponds to the edge of the drier Columbia Basin Ecosystem and historically may have been a barrier to salmon migration at certain times of the year. The predominant life history type for this species is the fall-run. Fall Ch inook typically enter the Columbia River in August through October to spaw n in the mainstems of large rivers (Kostow 1995). Spring Chinook enter freshwater in March through June to spawn in upstream tributaries and generally emigrate from freshwater as yearlings.

Status and trends. NMFS originally listed Lower Columbia River Chinook salmon as threatened on March 24, 1999 (64 FR 14308); NMFS r eaffirmed the threatened status of Lower Columbia River Chinook salmon on June 28, 2005 (70 FR 37160). Of 32 historical populations, 30 remain and 28 are considered at high ri sk of extinction (NMFS 2011d). Spring-fun individuals are at partic ularly high risk, as they have been cut off from extensive spawning areas104 by dams (NMFS 2011d). Historical records of Chinook salmon abundance are sparse, but cannery records suggest a peak run of 4.6 millio n fish (43 million pounds) in 1883 (Lichatowich 1999). Although fall-run Chinook salmon are still pr esent throughout much of their historical range, they are still subject to large-scale hatchery productio n, relatively high harvest, and extensive habitat degradation. Th e Lewis River late-fall-run populat ion is the healthiest and has a reasonable probability of be ing self-sustaining. Abundances largely declined during 1998-

2000 and trend indicators for most populations are negative, especi ally if hatchery fish are assumed to have a reproductive su ccess equivalent to that of na tural-origin fish (see Table 9).

Table 9. Lower Columbia River Chinook salmon life histories, populations and selected measures of population viability.

Life history Population Historical

abundance a Mean number of spawners

(range) b Percent hatchery

contribution c Long-term median growth rate ( λ ) d

Fall run Youngs Bay Grays River 2,477 99 38 0.944, 0.844 Big Creek

Elochoman River 676 68 1.037, 0.800

Clatskanie River e 50 (34-74) 0.99 Mill, Abernathy, and Germany Creeks 734 47 0.981, 0.829

Scappoose Creek

Coweeman River 4,971 274 0 1.092, 1.091 Lower Cowlitz River 53,956 1,562 62 0.998, 0.682 Upper Cowlitz River 5,682 Toutle River 25,.392

Kalama River 22,455 2,931 67 0.973, 0.818

Salmon Creek and Lewis

River 47,591 f 256 0 0.984, 0.979

Clackamas River 40 Washougal River 7,518 3,254 58 1.025, 0.815 Sandy River 183 Columbia Gorge-lower

tributaries Columbia Gorge-upper

tributaries 2,363 136 (Wind River only) 13 (Wind River only 0.959, 0.955

Hood River 18 Big White Salmon River 334 21 0.963, 0.945

Late fall (bright) Sandy River e 3085 (2337- 4074) 3 0.997

North Fork Lewis River 7,841 13 0.968, 0.948 Spring

run Upper Cowlitz River

Cispus River 1,787

Tilton River Toutle River 2,901 Kalama River 4,178 98

Lewis River 347 Sandy River e 297 (202-438) 0.961 Big White Salmon River

Hood River 51

a Historical abundance for various rivers was calculated using the Ecosystem and Diagnosis Treatment (EDT) model, which attempts to predict population perf ormance based on reach-sp ecific habitat attributes. Estimates are provided105 as a means of comparing the histori cal abundance of populations relative to current abundance. See Good et al. (2005) for a discussion about the uncertai nty associated with these estimates.

b Recent geometric mean number of spawners as reported in Good et al. 2005

c Recent average hatchery-origin spaw ners (%) as reported by Good et al. 2005. Natural-origin spawners are those that had parents that spawned in the wild, as opposed to hatchery-origin fish, whose parents were spawned in a hatchery.

d The long-term median growth rate ( λ ) is an estimate of the natural growth rate after accounting for hatchery-origin spawners. The two values are estimates under two hypotheses about the reproductive success of hatchery origin spawners. Hatchery fish are assumed to have zero reproductiv e success in the first estimate. In the second estimate hatchery fish are assumed to have the same relative reprodu ctive success as natural-origin fish. Growth rates were not calculated for all populations, as adeq uate data were not available (see Good et al. 2005 for 95% confidence intervals on growth estimates).

d Values for these populations are reported in McElhany et al. 2007, and represent estimates based on the total available data series, which varies by population.

fe Combined estimate of Lewis River fall run (East Fork only) and Lewis River brights (Good et al. 2005)

Most populations for which data are availabl e have a long-term declining population trend (Table 15). Currently, the spatial extent of popul ations in the Coastal and Cascade fall runs are similar to their respective hist orical conditions. New data in clude spawner abundance estimates through 2001, new estimates of the fraction of ha tchery spawners, and harvest estimates. In addition, estimates of historical abundance ha ve been provided by the Washington Department

of Fish and Wildlife. The Willamette/Lower Co lumbia River Technical Review Team estimated that 8-10 historic populations have been extirpat ed, most of them spri ng-run populations. Near loss of that important life hi story type remains an important concern. Although some natural production currently occurs in 20 or so populations, only one exceeds 1,000 spawners. Almost all spring-runs are at very high risk of extinction. High hatche ry production continues to pose genetic and ecological risks to natural populations and to mask their performance for Coastal, Cascade, and Gorge fall run populations. Most Lower Columbia River populations have not seen increases in recent years as pronounced as those that have occurred in many other geographic areas. Although populations increase d through most of the past decade, recent declines leave abundance near where it was in 2001 (NMFS 2011d).

Critical habitat. NMFS designated critical habitat for Lower Columbia River Chinook salmon on September 2, 2005 (70 FR 52630). Designated cri tical habitat includes all Columbia River estuarine areas and river reaches proceeding upstr eam to the confluence with the Hood River as well as specific stream reaches in a number of tr ibutary subbasins. These areas are important for the species’ overall conserva tion by protecting quality growt h, reproduction, and feeding. The critical habitat designation for this ESU identifies primary constituent elements that include sites necessary to support one or more life stages. Sp ecific sites include fres hwater spawning sites, freshwater rearing sites, freshw ater migration corridors, nearshor e marine habitat and estuarine areas. The physical or biological features thatcharacterize these sites include water quality and quantity, natural cover, forage, adequate pass age conditions, and floodplain connectivity. Of 52

subbasins reviewed in NMFS' assessment of cr itical habitat for the Lower Columbia River Chinook salmon ESU, 13 subbasins were rated as having a medium conservation value, four were rated as low, and the remaining subbasins (35), were rated as having a high conservation

value to the ESU. Factors contributing to the downward trends in this ESU are hydromorphological changes resulting from hydropower development, loss of tidal marsh and swamp habitat, and degraded freshwater and marine habitat from indus trial harbor and port development, and urban development. Limiting factors identified for this species include106 reduced access to spawning/reari ng habitat in tributaries, hatc hery impacts, loss of habitat diversity and channel stability in tributaries, excessive fine sediment in spawning gravels, elevated water temperature in tributaries, and harvest impacts. On the mainstem of the Columbia River, hydrop ower projects, includi ng the Federal Columbia River Hydropower System (FCRPS) , have significantly degraded salmon and steelhead habitats (Bottom et al. 2005, Fresh et al. 2005, NMFS 2006d). The series of dams and reservoirs that make up the FCRPS block an estimated 12 million cubic yards of debris and sediment that would otherwise naturally flow down the Columbia River and replenish shorelines along the Washington and Oregon coasts.

Industrial harbor and port devel opment are also significant influe nces on the Lower Willamette and Lower Columbia rivers (Bottom et al. 2005, Fresh et al. 2005, NMFS 2006d, LCFRB 2010). Since 1878, 100 miles of river chan nel within the mainstem Colu mbia River, its estuary, and Oregon’s Willamette River have been dredged as a navigation channel by the ACOE. Originally dredged to a 20-foot minimum depth, the Federal navigation ch annel of the Lower Columbia River is now maintained at a depth of 43 feetand a width of 600 feet. The Lower Columbia River supports five ports on the Washington St ate side: Kalama, Longview, Skamania County, Woodland, and Vancouver. In addition to loss of riparian habitat, and disruption of benthic habitat due to dredging, high levels of several sedi ment chemicals, such as arsenic and polycyclic aromatic hydrocarbons (PAHs), have been identifie d in Lower Columbia River watersheds in the vicinity of the ports and asso ciated industrial facilities.

The most extensive urban development in the Lo wer Columbia River subbasin has occurred in

the Portland/Vancouver area. Outside of this major urban area , the majority of residences and businesses rely on septic systems. Common wate r quality issues with urban development and residential septic systems include higher water temperatures, lowered dissolved oxygen, increased fecal coliform bacteria, and increased ch emicals associated with pesticides and urban

runoff.

The Columbia River estuary has lost a significant amount of th e tidal marsh and tidal swamp habitats that are critical to juvenile salmon and steelhead, particularly small or ocean-type species (Bottom et al. 2005, Fresh et al. 2005, NMFS 2006d, LCFRB 2010). Edges of marsh areas provide sheltered habitats for juvenile salmon and steelhead where food, in the form of amphipods or other small invertebrates which feed on marsh detritus, is plentiful, and larger predatory fish can be avoided. Historically, floodwaters of th e Columbia River inundated the margins and floodplains along the estuary, allowi ng juvenile salmon and steelhead access to a wide expanse of low-velocity marshland and tidal channel habitats. In ge neral, the riverbanks were gently sloping, with ripari an and wetland vegetation at the higher elevations of the river floodplain becoming habitat for salmon and steelh ead during flooding river discharges or flood

tides. Sherwood et al. (1990) estimated that the Columbia Ri ver estuary lost 20,000 acres of tidal swamps, 10,000 acres of tidal marshes, and 3,000 acres of tidal flats between 1870 and 1970. This study further estimated an 80% reduction in emergent vegetation production and a 15% decline in benthic algal production.

Habitat and food-web changes with in the estuary, and other fact ors affecting salmon population structure and life histories, have altered the estuary’s capac ity to support juvenile salmon107 (Bottom et al. 2005, Fresh et al. 2005, NMFS 2006d, LCFRB 2010). Di king and filling activities have reduced the tidal prism and eliminate em ergent and forested wetlands and floodplain habitats. These changes likely have reduced th e estuary’s salmon-rearing capacity. Moreover,

water and sediment in the Lower Columbia River a nd its tributaries have t oxic contaminants that

are harmful to fish and wildlif e (LCREP 2007). Contaminants of concern include dioxins and furans, heavy metals, polychlori nated biphenyls (PCBs) and organoc hlorine pesticides such as DDT. Simplification of the population structure a nd life-history diversity of salmon possibly is yet another important factor aff ecting juvenile salmon viability. Re storation of estuarine habitats, particularly diked emergent a nd forested wetlands, reduction of avian predation by terns, and flow manipulations to restore hi storical flow patterns may have begun to enhance the estuary’s productive capacity for salmon, although historic al changes in population structure and salmon life histories may prevent salmon from making full use of the produc tive capacity of estuarine

habitats.

Species-specific threats and limitations to recovery. Lower Columbia River Chinook salmon

face numerous hindrances to their survival and re covery, including degraded estuarine and nearshore marine habitat resulting from cumulative impacts of land use and flow management by the Columbia River hydropower system and floodplai n connectivity and function, channel structure and complexity, riparian areas, stream substrate, stream flow, and water quality have been degraded as a result of cumu lative impacts of agriculture, fore stry, and development (LCFRB 2010, NMFS 2011d). Further habitat-ba sed impacts to the species re sult from an altered flow regime and Columbia River plume has altered the temperature regime and estuarine food web, and has reduced ocean productivity, reduced access to off-channel rearing habitat in the lower Columbia River, and reduced pr oductivity resulting from sediment and nutrient-related changes in the estuary. Also hampering the species is reduced access to spawni ng and rearing habitat mainly as a result of tributary hydropower projec ts, hatchery-related effect s, and harvest-related effects on fall Chinook salmon, juvenile fish strandings that result from ship wakes, and contaminants affecting fish health and reproduction.

Central Valley spring-run Chinook salmon

Distribution. The Central Valley spring-run Chi nook salmon ESU includes all naturally spawned populations of spring-run Chinook salmon in the Sacramento River and its tributaries in California. This ESU includes one artificial pr opagation program, the Feather River Hatchery spring-run Chinook salmon program. This ar tificially propagated population is no more divergent relative to the local natural populations than would be expected between closely related populations within this ESU. Only three extant populations remain from 18 or 19 original populations (NMFS 2011b)(Table 10).

Central Valley spring-run Chinook salmon ESU includes Chinook salmon entering the Sacramento River from March to July and sp awning from late August through early October, with a peak in September. Spring-run fish in the Sacramento River exhibit an ocean-type life history, emigrating as fry, sub-yearlings, and yearlings. Central Valley spring-run Chinook salmon require cool freshwater whil e they mature over the summer.108 Table 10. Central Valley spring-run Chinook salmon populations and selected measures of population viability.

Population Historical

Abundance a Mean Number of Spawners (Range) b Percent Hatchery

Contribution c Mean Annual Population Growth Rate ( λ ) d
Butte Creek spring-run 4,513 (67-4,513) 1.30 (1.09-1.60) Deer Creek spring-run 1,076 (243-1,076) 1.17 (1.04-1.35) Mill Creek spring-run 491 (203-491) 1.19 (1.00-1.47)

a Historical abundance for the total ESU, based on gillnet fishery catches, is estimated at about 700,000 (Fisher 1994). Individual river estimates of historical abundance not provided.

b Recent geometric mean number of spawners as reported by Good et al. 2005. Note the curr ent geometric mean for Butte, Deer and Mill creeks are also the maximum means.

c Between 1967 and 1999 the Feather River Hatchery released between less than 1 million to as much as 5.5 million spring-run Chinook salmon in any given year. Returns ranged from less than 1,000 spawners to about 7,000 in the late 1980s (see Good et al. 2005). No other hatchery data reported.

d The λ calculation, provided by Good et al. 2005, is an estimate of the populatio n growth rate. The 90% confidence intervals are noted in parentheses.

Status and trends. NMFS originally listed Central Valley spring-run Chinook salmon as threatened on September 16, 1999 (64 FR 50393), a cl assification this species retained on June 28, 2005 (70 FR 37160). This species was listed becau se dams isolate them from most of their historic spawning habitat and the habitat remaining to them is de graded. Historically, spring-run Chinook salmon were predominant throughout th e Central Valley occupying the upper and middle reaches (1,000 to 6,000 feet) of the San Jo aquin, American, Yuba, Feather, Sacramento, McCloud and Pit Rivers, with smaller populations in most tributaries with sufficient habitat for over-summering adults (Stone 1874{R utter, 1904 #1367, Clark 1929).

The Central Valley drainage as a whole is estimated to have supported spring-run Chinook salmon runs as large as 700,000 fish between the late 1880s and the 1940s (Fisher 1994) , although these estimates may reflect an alrea dy declining population, in part from the commercial gillnet fishery that occurred in this ESU (Good et al. 2005b). Before construction of Friant Dam, nearly 50,000 adults were counted in the San Joaquin River alone {Fry, 1961 #1049}. Following the completion of Friant Dam, the native population from the San Joaquin River and its tributaries (i.e., the Stanislaus and Mokelumne Rivers) was extirpated. Spring-run Chinook salmon no longer exist in the American River due to th e operation of Folsom Dam. Naturally spawning populations of Central Va lley spring-run Chinook salmon currently are

restricted to accessible reaches of the upper Sacramento River, Antelope Creek, Battle Creek, Beegum Creek, Big Chico Creek, Butte Creek, Cl ear Creek, Deer Creek, Feather River, Mill Creek, and Yuba River (CDFG 1998). Since 1969, the Central Valley spring-run Chinook salmon ESU (excluding Feather River fish) has displayed broad fluctuations in abundance ranging from 25,890 in 1982 to 1,403 in 1993 (CDFG unpublished data in Good et al. 2005b).

The average abundance for the ESU was 12,499 for the period of 1969 to 1979, 12,981 for the period of 1980 to 1990, and 6,542 for the period of 1991 to 2001. In 2003 and 2004, total run size for the ESU was 8,775 and 9,872 adults resp ectively, well above the 1991 to 2001 average. Evaluating the ESU as a whole, however, mask s significant changes that are occurring among populations that comprise the ESU (metapopulati on). For example, the mainstem Sacramento River population has undergone a significant de cline while the abundance of many tributary populations increased. Average abundance of Sacramento River mainstem spring-run Chinook salmon recently declined from a high of 12,107 for the period 1980 to 1990, to a low of 609 for109 the period 1991 to 2001, while the average abundan ce of Sacramento Rive r tributary populations increased from a low of 1,227 to a hi gh of 5,925 over the same periods.

Abundance time series data for Mill, Deer, Butte, and Big Chico cr eeks spring-run Chinook salmon confirm that population increases seen in the 1990s have continued through 2001 (Good

et al. 2005b). Habitat improvements, including the removal of several small dams and increases in summer flows in the watersheds, reduced o cean fisheries, and a favorable terrestrial and marine climate, have likely contributed to this. Up until 2001, all three spring-run Chinook salmon populations in the Central Valley have long-and short-term positi ve population growth. However, since 2001, escapement has declined for all three populations, po ssibly resulting from 2007-2009 drough conditions in California’s central va lley and/or unusual ocean temperatures in 2005 and 2006 (NMFS 2011b). An execption to this is the continued abundance increase at Battle and Clear Creeks (NMFS 2011b). Althoug h the populations are small, Central Valley spring-run Chinook salmon have some of the hi ghest population growth rates in the Central Valley.

Critical habitat. NMFS designated critical habitat for Central Valley spring-run Chinook salmon on September 2, 2005 (70 FR 52488). Specifi c geographic areas designated include the following CALWATER hydrological units: Teha ma, Whitmore, Redding, Eastern Tehama, Sacramento Delta, Valley-Putah-Cache, Marysv ille, Yuba, Valley-American, Colusa Basin, Butte Creek, and Shasta Bally hydrological units. These areas are important for the species’ overall conservation by protecting quality growth, reproduction, and f eeding. The critical habitat designation for this ESU identifies primary consti tuent elements that include sites necessary to support one or more Chinook salmon life stages . Specific sites include freshwater spawning sites, freshwater rearing sites, freshwater migration corridors , nearshore marine habitat and estuarine areas. The physical or biological feat ures that characterize these sites include water quality and quantity, natural cover, forage , adequate passage conditions, and floodplain connectivity. The critical habi tat designation (70 FR 52488) cont ains additional details on the sub-areas that are included as part of this designation, and the areas th at were excluded from designation.

In total, Central Valley spring-run Chinook salmon occupy 37 watersheds (freshwater and estuarine). The total area of habitat designated as critical in cludes about 1,100 miles of stream habitat and about 250 square mile s of estuarine habitat in the San Francisco-San Pablo-Suisun Bay complex. This designation includes the stre am channels within the designated stream reaches, and includes a lateral extent as defined by the ordinary high water line. In areas where the ordinary high-water line is no t defined the lateral extent is de fined as the bankfull elevation.

In estuarine areas the lateral extent is defi ned by the extreme high water because extreme high tide areas encompass those areas typically inundate d by water and regularly occupied by juvenile salmon during the spring and summer, when they ar e migrating in the near shore zone and relying on cover and refuge qualities provided by these hab itats, and while they are foraging. Of the 37

watersheds reviewed in NMFS' assessment of critical habitat for Central Valley spring-run Chinook salmon, seven watersheds received a low ra ting of conservation valu e, three received a medium rating, and 27 received a high rating of conservation value for the species.

Upper Columbia River Chinook salmon110 Distribution. The Upper Columbia River spring-run Chinook salmon ESU includes all naturally spawned populations in all rive r reaches accessible to Chinook salmon in Columbia River tributaries upstream of Rock Island Dam and downstream of Chief Joseph Dam in Washington, excluding the Okanogan River (Figure 7). Six arti ficial propagation programs are part of this ESU. These artificially propa gated populations are no more di vergent relative to the local natural populations than would be expected between closely related populations within this ESU. Spring-runs currently spawn in only three river basins above Ro ck Island Dam: the Wenatchee, Entiat, and Methow Rivers. Table 11 iden tifies ESU populations, their abundances, and estimates of the proportion of hatchery fish that contribute to run sizes.

Figure 7. Population boundaries, dams, accessible areas, and ex tirpated reaches of upper Columbia River Chinook distribution, incorporating critical habitat boundaries (dark black line). Taken from NMFS (NMFS 2011i).

Upper Columbia River spring-run Chinook salmon begin returning to the Columbia in early spring and enter upper Columbia tributaries from April through July, with a peak in mid-May. Run timing is later when river temperature is cooler and flow is high; the corollary is also true (Keefer et al. 2008). After migration, individuals hold in freshwater tributaries until spawning in late summer, peaking in mid- to late August. J uvenile spring-runs remain in freshwater for a full year before emigrating to saltwater in the spring of their second year.111 Table 11. Upper Columbia River Chinook salmon populations and selected measures of population viability.
Population Mean number of spawners (range) a Percent hatchery

contribution b Current short-term trend (previous) c

Methow River 680 (79-9,904) 59 +2.0 (-15.3)

Methow mainstem 161 redds (17-2,864) 59 +6.5 Twisp River 58 redds (10-369) 54 -9.8 (-27.4)

Chewuch River 58 redds (6-1,105) 41 -2.9 (-28.1) Lost/Early Winter creeks 12 (3-164) 54 -14.1 (-23.2 d ) Entiat River 111 (53-444) 42 -1.2 (-19.4) Wenatchee River 470 (119-4,446) 42 -1.5 (-37.4) Chiwawa River 109 redds (34-1,046) 47 -0.7 (-29.3) Nason Creek 54 redds (8-374) 39 -1.5 (-26.0) Upper Wenatchee River 8 redds (0-215) 66 -8.9 White River 9 redds (1-104) 8 -6.6 (-35.9) Little Wenatchee River 11 redds (3-74) 21 -25.8 (-25.8)

a 5-year geometric mean number of spawners unless otherwise noted; Includes hatchery fish. Range denoted in parentheses. Means calculated from years 1997 to 2001, except Lost/Early Winter creeks did not include 1998 as no data was available. Data reported in Good et al. 2005.

b Percent hatchery-origin from 1987-1996, and reported in Good et al. 2005.

c Current trend – percent/year – from years 1997 to 2001. Previous trend, noted in parentheses, from 1987-1996.

From Good et al. 2005.

d Lost River data only.

Status and trends. NMFS listed Upper Columbia River spring-run Chinook salmon as endangered on March 24, 1999 (64 FR 14308), and rea ffirmed their status as endangered on June 28, 2005 (70 FR 37160), because they had been reduced to small populations in three watersheds. Based on redd count data series, sp awning escapements for th e Wenatchee, Entiat, and Methow rivers have declined an average of 5.6, 4.8, and 6.3% per ye ar, respectively, since 1958. In the most recent 5-year geometri c mean (1997-2001), spawning escapement for naturally produced fish was 273 for the Wena tchee population, 65 for the Entiat population, and 282 for the Methow population, only 8-15% of the minimum abundance thresholds, although escapement increased substantially in 2000 and 2001 in all three river systems. Based on 1980- 2004 returns, the average annual growth rate fo r this ESU is estimated at 0.93 (meaning the population is not replacing itself) (Fisher a nd Hinrichsen 2006). Assuming that population growth rates were to continue at 1980 to 2004 levels, Upper Columbia River spring-run Chinook salmon populations are projected to have very high probabilities of extinction within 50 years. Population viability analyses for this species (using the Dennis Model) suggest that these Chinook salmon face a significant risk of extincti on: a 75 to 100% probability of extinction within 100 years (given return rates for 1980 to present). Sin ce 2005, abundance estimates have increased throughout the ESU, a lthough a large portion of returns are hatchery origin fish (NMFS 2011i).

Hatchery influence and genetic diversity are sign ificant issues for the continued survival of Upper Columbia River Chinook salmon. This is a result of reduced genetic diversity from homogenization of populations that occurred under the Grand C oulee Fish Maintenance Project from 1939-1943. Stray hatchery fish and a high proportion of hatchery fish during spawning

have contributed to the hi gh genetic diversity risk.

Critical habitat. NMFS designated critical habitat fo r Upper Columbia River spring-run Chinook salmon on September 2, 2005 (70 FR 52630) . The designation in cludes all Columbia112 River estuaries and river reaches upstream to Ch ief Joseph Dam and several tributary subbasins. This designation includes the stream channels w ithin the designated stream reaches, and includes a lateral extent as defined by the ordinary high water line. In areas where the ordinary highwater line is not defined the late ral extent is defined as the ba nk full elevation. These areas are important for the species’ overall conservation by protecting quality growth, reproduction, and feeding. The critical habitat de signation for this ESU identifies primary constituent elements that include sites necessary to support one or more life stages. Specific s ites include freshwater spawning sites, freshwater rear ing sites, freshwater migratio n corridors, nearshore marine habitat, and estuarine areas. The physical or biological features that characterize these sites include water quality and quantity, natural cove r, forage, adequate passage conditions, and floodplain connectivity. The Upper Columbia River spring-run Chinook salmon ESU has 31 watersheds within its range. Five watersheds received a medium rating and 26 received a high rating of conservation value to the ESU. Th e Columbia River rearing/migration corridor downstream of the spawning range was rated as a high conservation value. Factors contributing

to the downward trends in this ESU include mainstem Columbia River hydropower system mortality, tributary riparian degradation and lo ss of in-river wood, altere d tributary floodplain and channel morphology, reduced tr ibutary stream flow and impaired passage, and harvest impacts.

Migratory habitat quality in th is area has been severely a ffected by the development and operation of the FCRPS dams and reservoirs in the mainstem Columbia River, Bureau of Reclamation tributary projects, and privately owned dams in the Snake and Upper Columbia river basins. For example, construction of Hells Canyon Dam eliminated access to several likely production areas in Oregon and Idah o, including the Burnt, Powder, Weiser, Payette, Malheur, Owyhee, and Boise river basins (Good et al. 2005b), and Grand Coulee and Chief Joseph dams completely block anadromous fish passage on the upper mainstem Columbia River. Hydroelectric development modified natural flow regimes, resulting in higher water temperatures, changes in fish community structur e leading to increased ra tes of piscivorous and avian predation on juvenile salmon and steelhead, and delayed migration for both adult and juveniles. Physical features of dams such as turbin es also kill migrating fi sh. In-river survival is inversely related to the number of hydropower projects encoun tered by emigrating juveniles.

Similarly, development and operation of extens ive irrigation systems and dams for water withdrawal and storage in tribut aries have drastically altered hydrological cy cles. A series of large regulating dams on the middle and upper De schutes River affect flow and block access to upstream habitat, and have extirpated one or mo re populations from the Cascades Eastern Slope major population (ICTRT 2003b). Similarly, opera tion and maintenance of large water reclamation systems such as the Umatilla Basin and Yakima Projects have significantly reduced flows and degraded water quality and physical habitat in this domain.

Many stream reaches designated as critical habita t are over-allocated under state water law, with more allocated water rights than existing str eamflow conditions can support. Withdrawal of water, particularly during low-flow periods that commonly overlap with agricultural withdrawals, often increases summer stream temp eratures, blocks fish migration, strands fish, and alters sediment transport (Spence et al. 1996). Reduced tributary stream flow has been113 identified as a major limiting factor for all listed salmon and steelhead species in this area except Snake River fall-run Chinook salmon and Snake River sockeye salmon (NMFS 2007c, 2011k).

Many stream reaches designated as critical habi tat are listed on the st ate of Oregon’s Clean Water Act section 303(d) list for water temperature. Many areas that were historically suitable rearing and spawning habitat ar e now unsuitable due to high summer stream temperatures. Removal of riparian vegetation, alteration of natural stream morphol ogy, and withdrawal of water for agricultural or municipal use all c ontribute to elevated stream temperatures.

Contaminants such as insecticides and herbicides from agricultural runo ff and heavy metals from

mine waste are common in some areas of critical habitat.

Species-specific threats and limitations to recovery. Several habitat and direct-effect threats exist to hinder recovery of Upper Columbia River spring-r un Chinook salmon (Upper Columbia Salmon Recovery Board 2007, NMFS 2011i). Main stem Columbia River hydropower stop or disrupt upstream and downstream fish passage , ecosystem structure and function, flows, and water quality. Floodplain connectivity and function, channel structure and complexity, riparian areas and large woody debris recruitment, stream flow, and water quality have been degraded as a result of cumulative impacts of agriculture, fore stry, and development. Past introductions and persistence of non-native (exotic) fish species continues to affect habitat conditions for listed species. Degraded estuarine and nearshore marine habitat as well as harvest in Columbia River

fisheriesare additional factor s in the species decline.

Puget Sound Chinook salmon

Distribution. The Puget Sound Chinook salmon ESU includes all naturally spawned populations of Chinook salmon from rivers and streams flowing into Puget Sound including the Straits of Juan De Fuca from the Elwha River, eastward, with rivers a nd streams flowing into Hood Canal, South Sound, North Sound and the Stra it of Georgia in Washington (Figure 8). Twenty-six artificial propagation programs are pa rt of the ESU. These artificially propagated populations are no more divergen t relative to the local natura l populations than would be expected between closely related populations within this ESU.114 Figure 8. Population boundaries, dams, accessible areas, and extirpated reaches of Puget Sound Chinook distribution, incorporating critical habitat boundaries (dark black line). Taken from NMFS (NMFS 2011f).

The Puget Sound ESU is comprised of 31 histor ical populations, of which 22 or more are

believed to be extant and nine are considered extinct. Table 12 identifies the current populations within the ESU for which there are data, and thei r recent abundance as well as long-term trends.

Chinook salmon in this area generally have an “ocean-type” life history. Puget Sound populations include both early-returning and late-returning spawners described by Healey (1991b). However, within these generalized behavioral forms, significan t variation occurs in residence time in freshwater and estuarine environments. For example, Hayman et al. (1996) described three juvenile Chinook sa lmon life histories with varying residency times in the Skagit River system in northern Puget Sound. Chinook salmon utilize nearshore Puget Sound habitats year-round, although they can be far from their natal river systems (Brennan et al. 2004).

Table 12. Puget Sound Chinook salmon populations and selected measures of population viability.
Population Historical

abundance a Mean number of

spawners (natural-origin) b Percent hatchery

contribution

(range) c λ (+/- SE) d

Nooksack-North Fork 26,000 1,538 (125) 91 (88-95) 0.75 (0.07)

Nooksack-South Fork 13,000 338 (197) 40 (24-55) 0.94 (0.05) Lower Skagit 22,000 2,527 (2,519) 0.2 (0-0.7) 1.05 (0.09) Upper Skagit 35,000 9,489 (9,281) 2 (2-3) 1.05 (0.06) Upper Cascade 1,700 274 (274) 0.3 1.06 (0.05)

Lower Sauk 7,800 601 (601) 0 1.01 (0.12) Upper Sauk 4,200 324 (324) 0 0.96 (0.06)115 Population Historical

abundance a Mean number of

spawners (natural-origin) b Percent hatchery

contribution

(range) c λ (+/- SE) d

Suiattle 830 365 (365) 0 0.99 (0.06) Stillaguamish-North Fork 24,000 1,154 (671) 40 (13-52) 0.92 (0.04) Stillaguamish-South Fork 20,000 270 0.99 (0.02)* Skykomish 51,000 4,262 (2,392) 40 (11-66) 0.87 (0.03) Snoqualmie 33,000 2,067(1,700) 16 (5-72) 1.00 (0.04) North Lake Washington 331 1.07 (0.07)* Cedar 327 0.99 (0.07)* Green 8,884 (1,099) 83 (35-100) 0.67 (0.06)* White 844 1.16 (0.06)* Puyallup 33,000 1,653 0.95 (0.06)* Nisqually 18,000 1,195 1.04 (0.07)* Skokomish 1,392 1.04 (0.04)* Dosewallips 4,700 48 1.17 (0.10)*

Duckabush 43

Hamma Hamma 196

Mid Hood Canal 311 Dungeness 8,100 222 1.09 (0.11)* Elwha 688 0.95 (0.11)*

a Estimated total historical abundance for this ESU was ab out 700,000 fish, but is not meant to reflect a summation

of individual river historic estimates. Individual river estimates of historical abundance are based on an EDT analysis as reported in Good et al. 2005.

b 5-year geometric mean number of spawners (hatchery plus natural) for years 1998-2002. Geometric mean of natural origin spawners noted in parentheses. From Good et al. 2005.

c Percent hatchery-origin from 1997-2001. Estimates are from the TRT database and reported in Good et al. 2005.

d Short-term median population growth rate estimates assume that the repr oductive success of naturally spawning hatchery fish is equivalent to that of natural origin fish. Except estimates noted * where an estimate of the fraction of hatchery fish is not available then λ represents hatchery fish + natural-origin spawners. Data years used for calculation 1990-2002 (Good et al. 2005).

Status and trends. The NMFS listed Puget Sound Chinook salmon as threatened in 1999 (64 FR 14308); that status was reaffirmed on June 28, 2005 (70 FR 37160). This ESU has lost 15 spawning aggregations (nine from the early-run t ype) that were either independent historical populations or major components of the remaini ng 22 existing independent historical populations identified (Good et al. 2005b). The disproportionate loss of early-run life history diversity represents a significant loss of the evolutionary legacy of the ESU.

Data reported by Good et al. (2005) indicate that long term tre nds in abundance for this ESU are split with about half of the populations declining, and the other half increasing. In contrast, the short-term trend for four populat ions is declining. The overa ll long-term trend in abundance indicates that, on average, populations are just replacing themselves. Estimates of the short-term median population growth rate ( λ )(1990-2002) indicate an even sp lit between populations that are growing and those that ar e declining, although estimates would be lower for several populations if the fraction of naturally spawning hatchery fish were available for all populations within the ESU. For available data, when λ is calculated assuming that hatchery fish have the equivalent success of natural spaw ners then the largest estimate d decline occurs in the Green River. Populations with the largest positive s hort and long-term trends include the White River and the North Fork Nooksack River (Good et al. 2005b). Lambda for the Skagit River, which produces the most Chinook salmon in this ESU, has increased slightly. Overall, the recent116 analysis by Good et al. (2005) illustrated that there has not be much change in this ESU since

NMFS’ first status review (Busby et al. 1996a). Individual populat ions have improved, while others have declined. However, the lack of information on the fraction of naturally spawning, hatchery-origin fish for 10 of the 22 populations within this ESU limits our understanding of the trends in naturally spawning fish for a large por tion of the ESU. Natural-origin pre-harvest recruit escapements remained fa irly constant from 1985-2009 (Ford et al. 2010). Returns (preharvest run size) from the natural spawners were highest in 1 985, declined through 1994, remained low through 1999, increased in 2000 and again in 2001, and have declined through 2009, with 2009 having the lowest returns since 1 997. Productivity in the five-year period from 2005-2009 has been the lowest in the ESU for any five-year period since 1985 and diversity of the populations has continued to decline; presentl y it is the lowest in the last 25 years (NMFS 2011f). Based on a Shannon Diversity Index at the ESU level, diversity is declining (due primarily to the increased abundance of retu rns to the Whidbey Basin region) for both distribution among populations and among regions (Ford et al. 2010).

The estimated total run size of Chinook salmon in Puget Sound in the early 1990s was 240,000 fish, representing a loss of nearly 450,000 fish from historic numb ers. During a recent 5-year period, the geometric mean of natural spawne rs in populations of Puget Sound Chinook salmon ranged from 222 to just over 9,489 fish. Most p opulations had natural spawners numbering in the hundreds (median recent natural escapement is 766), and of the six populations with greater than 1,000 natural spawners, only two have a low fr action of hatchery fis h. The populations with the greatest estimated component of hatchery fi sh tend to be in mid- to southern Puget Sound, Hood Canal, and the Strait of Juan de Fuca region s. Estimates of the historical equilibrium abundance, based on pre-European settleme nt habitat conditions, range from 1,700 to 51,000 potential spawners per populati on. The historical estimates of spawner capacity are several orders of magnitude higher than spawner abundances currently observed throughout the ESU (Good et al. 2005b).

Critical habitat. The NMFS designated critical habi tat for Puget Sound Chinook salmon on September 2, 2005 (70 FR 52630). The specific geographic area includes portions of the Nooksack River, Skagit River, Sauk River, Sti llaguamish River, Skykomish River, Snoqualmie River, Lake Washington, Green River, Puyallup River, White Ri ver, Nisqually River, Hamma

Hamma River and other Hood Canal watershe ds, the Dungeness/Elwha Watersheds, and

nearshore marine areas of the St rait of Georgia, Puget Sound, Hood Canal, and the Strait of Juan de Fuca. This designation includes the stream channels within the designated stream reaches, and includes a lateral extent as defined by the ordinary high water line. In areas where the ordinary high water line is not defined the lateral extent is defi ned as the bank full elevation.

The designation for this ESU include s sites necessary to support one or more life stages. These areas are important for the species’ overall conservation by protec ting quality growth, reproduction, and feeding. Specific primary cons tituent elements include freshwater spawning sites, freshwater rearing sites, freshwater migration corridors, nearshore marine habitat, and estuarine areas. The physical or biological feat ures that characterize these sites include water quality and quantity, natural cover, forage , adequate passage conditions, and floodplain connectivity. Of 49 subbasins reviewed in NMFS' assessment of critical habitat for the Puget Sound ESUs, nine subbasins were ra ted as having a medium conservation value, 12 were rated as117 low, and the remaining subbasins (40), where the bulk of Federal lands occur for this ESU, were rated as having a high conservation value to P uget Sound Chinook salmon. Factors contributing

to the downward trends in this ESU are hydromorphological change s (such as diking, revetments, loss of secondary channels in floodp lains, widespread bloc kages of streams, and changes in peak flows), degraded freshwater and marine habitat affected by agricultural activities and urbanization, and upper river tr ibutaries widely affected by poor forest practices. Changes in habitat quantity, availabi lity, diversity, flow, temperature, se diment load, and channel stability are common limiting factors in areas of critical habitat.

Landslides can occur naturally in steep, forested lands, but inappropriate land use practices likely have accelerated their frequency and the amount of sediment delivered to streams. Fine sediment from unpaved roads has also contributed to str eam sedimentation. Unpave d roads are widespread on forested lands in the Puget Sound basin, and to a lesser extent, in rural residential areas. Historical logging removed most of the ripari an trees near stream channels. Subsequent agricultural and urban conversion permanently altered riparian vegetation in the river valleys, leaving either no trees, ora thin band of trees. The ripari an zones along many agricultural areas are now dominated by alder, invasive canary gr ass and blackberries, and provide substantially reduced stream shade and large wood recruitment (SSPS 2007).

Diking, agriculture, revetments, railroads and roads in lower stream reaches have caused significant loss of secondary channels in major valley floodplains in this region. Confined main channels create high-energy peak flows that re move smaller substrate particles and large wood. The loss of side-channels, oxbow lakes, and backwa ter habitats has resulted in a significant loss of juvenile salmonid rearing and refuge habitat. When the water level of Lake Washington was lowered 9 feet in the 1910s, thousands of acr es of wetlands along the shoreline of Lake Washington, Lake Sammamish and the Sammamish River corridor were drained and converted to agricultural and urban uses. Wetlands play an important role in hydrologic processes, as they store water which ameliorates high and low flows. The interchange of su rface and groundwater in complex stream and wetland systems helps to moderate stream temperatures. Forest wetlands are estimated to have diminished by one-thi rd in Washington St ate ((FEMAT) 1993, Spence et
al. 1996, SSPS 2007).

Loss of riparian habitat, elevated water temper atures, elevated levels of nutrients, increased nitrogen and phosphorus, and higher levels of tu rbidity, presumably from urban and highway runoff, wastewater treatment, fa iling septic systems, and agricultu re or livestock impacts, have been documented in many Puget Sound tributaries (SSPS 2007).

Peak stream flows have increased over time due to paving (roads and parking areas), reduced percolation through surface soil s on residential and agricultural lands, si mplified and extended drainage networks, loss of wetla nds, and rain-on-snow events in higher elevation clear cuts (SSPS 2007). In urbanized Puget Sound, there is a strong association between land use and land cover attributes and rates of coho spawner morta lity likely due to runoff containing contaminants emitted from motor vehicles (Feist et al. 2011).

Juvenile mortality occurs in unscreened or inadequately screened diversions. Water diversion ditches resemble side channels in which j uvenile salmonids normally find refuge. When118 diversion headgates are shut, access back to the ma in channel is cut off and the channel goes dry. Mortality can also occur with inadequately scr eened diversions from impingement on the screen, or mutilation in pumps where gaps or oversized screen openings allow juveniles to get into the system (WDFW 2009). Blockages by dams, water diversions, and shifts in flow regime due to hydroelectric development and flood control proj ects are major habitat problems in many Puget Sound tributary basins (SSPS 2007).

The nearshore marine habitat has been exte nsively altered and ar mored by industrial and residential development near the mouths of many of Puget Sound’s tributaries. A railroad runs along large portions of the eastern shoreline of Puget Sound, elim inating natural cover along the shore and natural recruitment of beach sand (SSPS 2007).

Dams constructed for hydropower generation, i rrigation, or flood contro l have substantially affected Puget Sound Chinook salmon populations in a number of river systems. The construction and operation of dams have bloc ked access to spawning a nd rearing habitat ( e.g ., Elwha River dams block anadrom ous fish access to 70 miles of potential habitat) changed flow patterns, resulted in elevated temperatures a nd stranding of juvenile migrants, and degraded downstream spawning and rearing habitat by reducing recruitment of spawning gravel and large wood to downstream areas (SSPS 2007)). These actions tend to promote downstream channel incision and simplification (Kondolf 1997), limiting fish habitat. Water withdrawals reduce available fish habitat and alter sediment transport. Hydropower projects of ten change flow rates, stranding and killing fish, and reducing aquatic invertebrate (f ood source) productivity (Chappell 1980).

Species-specific threats and limitations to recovery. Residential and commercial development has reduced the amount of functioning nearshor e and estuarine habitat available for salmon rearing and migration (SSPS 2007, NMFS 2011f). The loss of mudflats, eelgrass meadows, and macroalgae further limits salmon foraging and re aring opportunities in ne arshore and estuarine areas. Floodplain connectivity a nd function, channel structure a nd complexity, riparian areas and large wood supply, stream substrate, and wate r quality have also been degraded for adult spawning, embryo incubation, and re aring as a result of cumulati ve impacts of agriculture, forestry, and development. Total fishery exploita tion rates have decreased 14 to 63% from rates in the 1980s, but weak natural-origin Chinook salmon populations in P uget Sound still require enhanced protective measures to reduce the risk of overharvest in Chinook salmon-directed

fisheries.

Snake River fall-run Chinook salmon Life history types. Although Snake River fall-run Chinook were originally listed under the assumption that all individuals migrated to the ocean at age 0 (wintering in the ocean), studies through the 1990s demonstrate that two dis tinct life historie s actually exist (Connor et al. 2005b).

Some individuals overwinter in the Lower Granite Reservoir and enter the ocean at age 1 and roughly twice the length of individuals en tering the Pacific Ocean at age 0 (Connor et al. 2005b).

Fish of both life histories return at roughly the same time, with 41% of wild and 61% of hatchery fish being the reservoir type (Connor et al. 2005b).

Distribution. The Snake River fall-run Chinook salmon ESU includes all naturally spawned119 populations of fall-run Chinook salmon in the ma instem Snake River below Hells Canyon Dam, and in the Tucannon River, Grande Ronde River, Imnaha River, Salmon River, and Clearwater River subbasins (Figure 9). Four artificial propagation programs ar e part of this ESU. These artificially propagated populations are no more divergent relative to the local natu ral populations than would be expected between closely related populations with in this ESU.

Figure 9. Population boundaries, accessible areas, extirpated reaches, and dams of Snake River fall run Chinook distribution, incorporating critical habitat boundar ies (dark black line). Taken from NMFS (2011l).

Historically, the primary fall-run Chinook sa lmon spawning areas occurred on the upper mainstem Snake River (Connor et al. 2005a). A series of Snake River dams blocked access to the upper reaches, which significan tly reduced spawning and reari ng habitat. Consequently, salmon now reside in waters that are generally co oler than pre-dam habitats. Currently, natural spawning occurs at the upper end of Lower Gr anite Reservoir to Hells Canyon Dam, the lower reaches of the Imnaha, Grande Ronde, Clearwa ter, and Tucannon rivers, and small mainstem

sections in the tailraces of the lo wer Snake River hydroelectric dams.

Adult Snake River fall-run Chinook salmon enter th e Columbia River in July and August, and spawning occurs from October through November. Juveniles emerge from gravel in March and April of the following year, moving downstream from natal spawning and early rearing areas from June through early fall. Prior to dam construction, fall Chinook salmon were primarily ocean-type (migrated downstream and reared in the mainstem Snake River during their first year). However, today both an ocea n-type and reservoir-type occur (Connor et al. 2005a). The reservoir-type juveniles overwinter in pools crea ted by dams before migrating to sea; this120 response is likely due to early development in co oler temperatures, which prevents rapid growth. Phenotypic characteristics have shifted in appa rent response to environmental changes from hydroelectric dams (Connor et al. 2005a). Migration downstream appears to be influenced by flow velocity within both rive r and reservoir systems (Tiffan et al. 2009).

Status and trends. The NMFS originally listed Snake River fall-run Chinook salmon as endangered in 1992 (57 FR 14653) but reclassified their status as threatened on June 28, 2005 (70 FR 37160). Estimated annual returns for the period 1938-1949 were 72,000 fish, and by the 1950s, numbers had declined to an annual aver age of 29,000 fish (Bjornn and Horner 1980).

Numbers of Snake River fall-run Chinook salmon continued to decline during the 1960s and 1970s as approximately 80% of thei r historic habitat was eliminated or severely degraded by the construction of the Hells Canyon complex (1958 -1967) and the lower Snake River dams (1961- 1975). Counts of natural-origin adults at Lower Granite Dam were 1,000 fish in 1975, and ranged from 78 to 905 fish (with an average of 489 fish) over the ensuing 25-year period (Good

et al. 2005b). Numbers of natural-origin individuals have increased over the last few years, with estimates at Lower Granite Dam of 2,652 fi sh in 2001, 2,095 fish in 2002, and 3,895 fish in 2003. Presently, hatcheries contribute rouighly 78% of the natural spawning population (NMFS 2011l).

Snake River fall-run Chinook salmon have exhi bited an upward trend in returns over Lower

Granite Dam since the mid 1990s. Returns classifi ed as natural-origin spawners exceeded 2,600 fish in 2001, compared to a 1997-2001 geometric m ean natural-origin count of 871 (35% of the proposed delisting abundance crite ria of 2,500 natural spawners averaged over 8 years). Both the long- and short-term trends in natural returns are positive. The latest 10-year geometric mean from 199-2008 is just over 2,200 adults (NMFS 2011l). In addition, the ESU’s productivity has improved to 1.28 (1990-2011) from 1.07 (1983-2003)(N MFS 2011l). Harvest impacts declined after listing and have remained relatively cons tant in recent years. Mainstem conditions for subyearling Chinook migrants from the Snake Ri ver have generally improved since the early 1990s. The hatchery component, derived from outs ide the basin, has decreas ed as a percentage

of the run at Lower Granite Dam from the 1998/ 99 status reviews (5-year average of 26.2%) to 2001 (8%). This reflects an in crease in the Lyons Ferry comp onent, systematic removal of marked hatchery fish at the Lower Granite trap, and modifications to the Umatilla supplementation program to increase homing of release groups. Hatcheries stocking fish produce genetic affects in the popul ation due to three major compone nts: natural-origin fish (which may be progeny of hatchery fish), retu rns of Snake River fish from the Lyons Ferry Hatchery program, and strays from hatchery programs outside the Snake River. Current models predict a 5-25% probability of extin ction within 100 years (NMFS 2011l).

Critical habitat. The NMFS designated critical habita t for Snake River fall-run Chinook salmon on December 28, 1993 (58 FR 68543). This cr itical habitat encompasses the waters, waterway bottoms, and adjacent riparian zones of specified lakes and river reaches in the

Columbia River that are or we re accessible to listed indi viduals (except reaches above impassable natural falls, and Dworshak and Hells Canyon Dams). These areas are important for the species’ overall conserva tion by protecting quality grow th, reproduction, and feeding. Adjacent riparian zones are define d as those areas within a horizontal distance of 300 feet from the normal line of high water of a stream channe l or from the shoreline of a standing body of121 water. Designated critical habitat includes the Columbia River from a straight line connecting the west end of the Clatsop jetty (Oregon side) and the west end of the Peacock jetty (Washington side) and including all river reaches from the estuary upstream to the confluence of the Snake River, and all Snake River reaches upstream to Hells Canyon Dam. Critical habitat

also includes several river reaches presently or historically acce ssible. Limiting factors include: mainstem lower Snake and Columbia hydrosyste m mortality, degraded water quality, reduced spawning and rearing habitat due to mainstem lower Snake River hydropower system, harvest impacts, impaired stream flows, barriers to fish passage in tributaries, excessive sediment, and altered floodplain and ch annel morphology (NMFS 2005b).

Species-specific threats and limitations to recovery. Floodplain connectivity and function, and channel structure and complexity have been degraded as a re sult of cumulative impacts of agriculture, forestry, and development have hindered Snake River fall-run Chinook salmon recovery, as has degraded fres hwater, estuarine, and marine habitat, harvest and hjatchery effects, as well as mainstem Columbia Ri ver and Snake River hydropower facilities (NMFS 2011l).

Snake River spring/summer-run Chinook salmon

Distribution. The Snake River spring/summer run Ch inook salmon ESU includes all naturally spawned populations of spring/summer-run Chinook salmon in the mainstem Snake River and the Tucannon River, Grande Ronde River, Imnaha River, and Salmon River subbasins (Figure 10). Fifteen artificial propagati on programs are part of the ESU. These artificially propagated populations are no more divergen t relative to the local natura l populations than would be expected between closely related populations with in this ESU. The Interior Columbia Basin Technical Recovery Team identified 32 populat ions in five major population groups (Upper Salmon River, South Fork Salmon River, Middle Fork Salmon River, Grande Ronde/Imnaha, Lower Snake Mainstem Tributaries) for this sp ecies. Historic populations above Hells Canyon Dam are considered extinct (ICBTRT 2003) Table 13 identifies extant p opulations within the Snake River spring/summer Chinook salmon ES U, their abundances, and the relative contribution of hatchery fish.122 Figure 10. Population boundaries, accessible areas, extirpated r eaches, and dams of Snak e River spring/summer run Chinook distribution, incorporating critical habitat boundaries (dark black line). Taken from NMFS (2011l).

Snake River spring/summer-run Chinook salmon have a stream-type life history. Spawning occurs in late summer and early fall and eggs in cubate over winter and hatch in late winter and early spring of the following year. Juveniles mature in the river fo r one year before migrating to the ocean in the spring of their second year. La rger outmigrants have a higher survival rate during outmigration (Zabel and Williams 2002, Zabel and Achord 2004). Depending on tributary and the specific habi tat conditions, juveniles may migrate widely from natal reaches into alternative summer-rearing or overwintering areas. Spawners return to spawn primarily as 4- and 5-year-olds after 2 to 3 y ears in the ocean. A small fracti on return as 3-year-old “jacks” (although sexually mature upon return, these fish are smaller in body and 1-2 years younger than most males on the spawning ground).

Table 13. Snake River spring/summer Chinook salmon populations and selected measures of population viability (rpm = redds per mile).
Current populations , Mean number of spawners (range) a Percent hatchery

contribution b Short-term trend (previous) c

Tucannon River 303 (128-1,012) 76 -4.1 (-11.0)

Wenaha River 225 (67-586) 64 -9.4 (-23.6)

Wallowa River 0.57 redds (0-29) 5 11.5

Lostine River 34 redds (9-131) 5 12.7

Minam River 180 (96-573) 5 3.3 (-14.5)

Catherine Creek 50 (13-262) 56 -25.1 (-22.5) Upper Grande Ronde River 46 (3-336) 58 -9.4123 Current populations , Mean number of spawners (range) a Percent hatchery

contribution b Short-term trend (previous) c

South Fork Salmon River 496 redds (277-679) 9 1.1 (-13.6)

Secesh River 144 redds (38-444) 4 9.8

Johnson Creek 131 redds (49-444) -1.5 Big Creek spring run 53 (21-296) 5.4 (-34.2) Big Creek summer run 5 redds (2-58) 1.7 (-27.9)

Loon Creek 27 redds (6-255) 12.2

Marsh Creek 53 (0-164) -4.0 Bear Valley/Elk Creek 266 (72-712) 6.2

North Fork Salmon River 5.6 redds (2-19)

Lemhi River 72 redds (35-216) 12.8 (-27.4)

Pahsimeroi River 161 (72-1,097) 12.8 East Fork Salmon spring run 0.27 rpm (0.2-1.41) -5.7 East Fork Salmon summer run 1.22 rpm 0.35-5.32) 0.9 (-32.9)
Yankee Fork spring run 0 rpm -6.3

Yankee Fork summer run 2.9 redds (1-18) 4.1 Valley Creek spring run 7.4 redds (2-28) 14.9 (-25.9) Valley Creek summer run 2.14 rpm(0.71-9.29) 5.8 (-29.3 Upper Salmon spring run 69 redds (25-357) 5.3 Upper Salmon summer run 0.24 rpm (0.07-0.58) -3.3

Alturas Lake Creek 2.7 redds (0-18) 10.2

Imnaha River 564 redds (194-3,041) 62 12.8(-24.1) Big Sheep Creek 0.25 redds (0-1) 97 0.8

Lick Creek 1.4 redds (0-29) 59 11.7

a All data reported in Good et al. 2005. Except where noted values represent the recent geometric mean number of spawners. RPM =redds per mile.

bc Reported in Good et al. 2005.

c For details on data series used in calculating the population’s short term trend see Good et al. 2005.

Status and trends. The NMFS originally listed Snak e River spring/summer-run Chinook salmon as threatened on April 22, 1992 (57 FR 14653), and reaffirmed their st atus as threatened on June 28, 2005 (70 FR 37160). Although direct es timates of historical annual Snake River spring/summer Chinook salmon returns are not avai lable, returns may have declined by as much

as 97% between the late 1800s and 2000. Accord ing to Matthews and Waples (1991), total annual Snake River spring/summer Chinook sa lmon production may have exceeded 1.5 million

adult fish in the late 1800s. Total (natural plus hatchery origin) returns fell to roughly 100,000 spawners by the late 1960s and were below 10,000 by 1980 (Fulton 1968). Between 1981 and 2000, total returns fluctuated between extremes of 1,800 and 44,000 fish. The 2001 and 2002

total returns increased to over 185,000 and 97,184 adults, respectively. The 1997-2001 geometric mean total return for the summer r un component at Lower Granite Dam was slightly more than 6,000 fish, compared to the geomet ric mean of 3,076 fish for the years 1987-1996. The 2002-2006 geometric mean of the combined Chinook salmon runs at Lower Granite Dam was over 18,000 fish. However, it is important to note that over 80% of the 2001 return and over 6% of the 2002 return origin ated in hatcheries (Good et al. 2005b). Good et al. (2005b) reported that risks to individual populati ons within the ESU may be greater than the extinction risk for the

entire ESU due to low levels of annual abundan ce and the extensive pro duction areas within the Snake River basin. Recent spawner returns have been higher, but low natural spawner abundance and the ability of the populations to c ope with low ocean survival remain concerns (NMFS 2011l).124 Critical habitat. The NMFS designated critical habita t for Snake River spring/summer-run Chinook salmon on October 25, 1999 (64 FR 57399). This critical habitat encompasses the waters, waterway bottoms, and adj acent riparian zones of specified lakes and river reaches in the

Columbia River that are or were accessible to listed Snake Ri ver salmon (except reaches above impassable natural falls, and Dworshak and Hells Canyon Dams). Adjacent riparian zones are

defined as those areas within a horizontal distan ce of 300 feet from the normal line of high water

of a stream channel or from the shoreline of a standing body of water. De signated critical habitat

includes the Columbia River from a straight line connecting the west end of the Clatsop jetty (Oregon side) and the west end of the Peacock jetty (Washington side) a nd including all river reaches from the estuary upstream to the conflu ence of the Snake River, and all Snake River reaches upstream to Hells Canyon Dam; the Palous e River from its confluence with the Snake River upstream to Palouse Falls, the Clearwater River from its c onfluence with the Snake River upstream to its confluence with Lolo Creek; the No rth Fork Clearwater River from its confluence with the Clearwater river upstream to Dworshak Da m. Critical habitat also includes several river reaches presently or historically accessible. Th ese areas are important for the species’ overall conservation by protecting quality growth, reprodu ction, and feeding. Limiting factors identified for this species include hydrosystem mortality, reduced stream flow, altered channel morphology and floodplain, excessive fine sediment, a nd degraded water quality (NMFS 2006i).

Species-specific threats and limitations to recovery. As with other listed salmonids, floodplain connectivity and function, channel struct ure and complexity, riparian areas and large wood supply, stream substrate, elevated water temp erature, stream flow, and water quality have been degraded as a result of cumulative impacts of agriculture, forestry, and development to the degree that species recovery is impaired (NM FS 2006i). Additional threats to recovery exist

from mainstem Columbia River and Snake Ri ver hydropower, predation, as well as harvestrelated effects.

Upper Willamette River Chinook salmon

Distribution. The Upper Willamette River Chinook salmon ESU includes all naturally spawned populations of spring-run Chinook salmon in the Clackamas River and in the Willamette River, and its tributaries, above Willamette Falls, Or egon (7 populations in all)(NMFS 2011j)(Figure 11). Seven artificial propagati on programs are part of the ESU. These artificially propagated populations are no more divergen t relative to the local natura l populations than would be expected between closely related po pulations within this ESU.125 Figure 11. Population boundaries, dams, accessible areas, and extirpated r eaches, of upper Willamette River Chinook distribution, incorporating critical habitat boundaries (dark black line). Taken from NMFS (2011).

Upper Willamette River Chinook salmon occupy the Willamette River and its tributaries. All spring-run Chinook salmon in the ESU, except th ose entering the Clackamas River, must pass Willamette Falls. In the past, this ESU incl uded sizable numbers of spawning salmon in the Santiam River, the middle fork of the Willamette River, and the McKenzie River, as well as smaller numbers in the Molalla River, Calapooia River, and Albiqua Cree k. Historically, access above Willamette Falls was restricted to spring when flows were high. In autumn, low flows prevented fish from ascending past the falls. The Upper Willamette spring-run Chinook salmon are one of the most genetically distinct Chinook salmon groups in the Columbia River Basin.

Individuals enter the Columbia River and its estuary earlier than ot her spring Chinook salmon ESUs (Meyers et al. 1998). Fall-run Chinook salmon spawn in the Upper Willamette but are not considered part of the ESU because they are not native.

Status and trends. The NMFS originally listed Upper Willamette River Chinook salmon as threatened on March 24, 1999 (64 FR 14308), and r eaffirmed their status on June 28, 2005 (70 FR 37160). The total abundance of adults (hatchery-origin plus natural-origin fish) passing Willamette Falls has remained relatively steady over the past 50 years (ranging from approximately 20,000 to 70,000 fish), but it is an order of magnitude below the peak abundance levels observed in the 1920s (approximatel y 300,000 adults). Although the Clackmas population consisted of roughly 12,000 spawners in 2004, that number dropped to 2,000 in 2009 and 2010, with a geometric mean of 850 individuals from 2006-2011 (NMFS 2011j) . McKenzie River126 returns had been increasing in abundance until peaking in 2004, but have since dropped to historical levels (NMFS 2011j). Until recent year s, interpretation of abundance levels has been confounded by a high but uncertain fraction of hatc hery-produced fish. Although the number of adults crossing Willamette Falls is in the same range (about 20,000-70,000 adults) it has been for the last 50 years, a large fraction of these are hatchery produced. Estimates of the percentage of hatchery fish range according to tributary, but exceed 70% (Good et al. 2005b). The Calapooia River is estimated to contain 100% hatchery fish. Insufficient information on hatchery production in the past prevents a meaningful anal ysis of the population tre nd; therefore no formal trend analysis is available.

Most natural spring Chinook salmon populations of the Upper Willamette River are likely extirpated or nearly so, with only one remaini ng naturally reproducing pop ulation identified in the McKenzie River. Unfortunately, recent short- term declines in abundance suggest that this population may not be se lf-sustaining (Meyers et al. 1998, Good et al. 2005b). Abundance in this population has been relatively low (low thousands) with a subs tantial number of these fish being of hatchery origin. The population increased substantially from 2000 to 2003, probably

due to increased survival in the ocean. Future oceanic survival rates ar e unpredictable, and the likelihood of long-term sustainability for this popul ation has not been determined. Of concern is that a majority of the spawning habitat and approximately 30-40% of total historical habitat are no longer accessible because of dams (Good et al. 2005b). Individuals from the ESU migrate far north and are caught incidentally in ocean fish eries, particularly o ff southeast Alaska and northern Canada, and in the mainstem Colu mbia and Willamette rivers during spring.

Critical habitat. The NMFS designated critical habita t for Upper Willamette River Chinook salmon on September 2, 2005 (70 FR 52630). Critic al habitat includes defined areas within subbasins of the middle fork Willamette River, upper Willamette River, McKenzie River, Santiam River, Crabtree Creek, Molalla River, an d Clackamas River. This designation includes the stream channels within the designated stream reaches, and includes a lateral extent as defined by the ordinary high water line. In areas where the ordinary high-water li ne is not defined the

lateral extent is defined as the bank full elevatio n. The critical habitat designation for this ESU identifies primary constituent elements that in clude sites necessary to support one or more Chinook salmon life stages. Specific sites include freshwater spawning and rearing sites, freshwater migration corridors. The physical or biologi cal features that ch aracterize these sites include water quality and quantity, natural cove r, forage, adequate passage conditions, and floodplain connectivity. Of 65 subb asins reviewed in NMFS’ assessment of ESU critical habitat, 19 subbasins were rated as having a medium cons ervation value, 19 were rated as low, and the 27 remaining subbasins were rated as having a high conservation value. Federal lands were generally rated as having high conservation value to the species’ spawning and rearing. Factors contributing to the downward trends in this ESU include reduced acce ss to spawning/rearing habitat in tributaries, hatchery impacts, altere d water quality and temper ature in tributaries, altered stream flow in tributaries, and lost or degraded floodplain connectivity and lowland

stream habitat.

Land m anagement activities have severely degraded stream habitat conditi ons in the Willamette

River mainstem above Willamette Falls and asso ciated subbasins. In the Willamette River mainstem and lower sub-basin mainstem reaches, high density urban development and127 widespread agricultural effects have reduced aqua tic and riparian habitat quality and complexity, and altered sediment and water quality and quantity, and watershed processes. The Willamette River, once a highly braided river system, has been dramatically simplified through channelization, dredging, and othe r activities that have reduced rearing habitat by as much as 75%. In addition, the construction of 37 dams in the basin blocked access to more than 435 miles of stream and river spawning habitat. The dams alter the temperature regime of the Willamette River and its tributaries, affecting the timing and development of naturally-spawned eggs and fry. Agriculture, urbanization, and gravel mining on the valley floor logging in the Cascade and Coast Ranges contribute to increased erosion and sediment loads throughout the basin.

The mainstem Willamette River has been channe lized and stripped of large wood. Development began to encroach on the riparian forest be ginning in the 1870s (Sedell and Froggatt 1984). Gregory et al. (2002d)(calculated that the total main stem Willamette Ri ver channel area decreased from 41,000 to 23,000 acres between 1895 a nd 1995. They noted that the lower reach,

from the mouth of the river to Newberg (RM 50), is confined with in a basaltic trench, and that due to this geomorphic constraint, less channel area has been lost than in upstream areas. The middle reach from Newberg to Albany (RM 50 to 120) incurred losses of 12% primary channel area, 16% side channels, 33% alcoves, and 9% islands. Even greater changes occurred in the upper reach, from Albany to Eugene (RM 187). There, approximately 40% of both channel length and channel area were lost, along with 21% of the primary channel, 41% of side channels, 74% of alcoves, and 80% of island areas.

The banks of the Willamette River have more than 96 miles of revetments; approximately half were constructed by the ACOE. Generally, the revetmen ts were placed in the vicinity of roads or on the outside bank of river bends, so that while only 26% of the total length is revetted, 65% of the meander bends are revetted (Gregory et al. 2002c). The majority of dynamic sections have been armored, reducing adjustments in channel bed and sediment storage by the river, and thereby diminishing both the complexity a nd productivity of aqua tic habitats (Gregory et al. 2002b).

Riparian forests have diminished considerably in the lower reaches of the Willamette River (Gregory et al. 2002a). Sedell and Froggatt (1984) noted that agriculture and cutting of streamside trees were major agents of change for riparian vegetation, along with snagging of large wood in the channel. The reduced shoreline, fewer and smaller snags, and reduced riparian forest comprise large functional losses to the ri ver, reducing structural features, organic inputs from litter fall, entrained allochthonous material s, and flood flow filtering capacity. Extensive changes began before the major dams were bui lt, with navigational and agricultural demands dominating the early use of the river. The once expansive forests of the Willamette River floodplain provided valuable nutrients and orga nic matter during flood pulses, food sources for macroinvertebrates, and slow-wat er refugia for fish during flo od events. These forests also cooled river temperatures as the ri ver flowed through its many channels.

Gregory et al. (2002a) described the changes in riparian vegetation in river reaches from the mouth to Newberg, from Newberg to Albany, and from Albany to Eugene. They noted that the riparian forests were formerly a mosaic of br ush, marsh, and ash tree openings maintained by annual flood inundation. Below the City of Newb erg, the most noticeable change was that128 conifers were almost eliminated. Above Newberg, the formerly hardwood-dominated riparian forests along with mixed forest made up less than half of the riparian vegetation by 1990, while agriculture dominated. This conversion has re duced river shading and the potential for recruitment of wood to the river, reducing ch annel complexity and the quality of rearing, migration and spawning habitats.

Hyporheic flow in the Willamette River has be en examined through discharge measurements and found to be significant in some areas, partic ularly those with gr avel deposits (Wentz et al. 1998,

Fernald et al. 2001). The loss of channel complexity a nd meandering that fosters creations of gravel deposits decreases the potential for hyporheic flows, as does gravel mining. Hyporheic flow processes water and affect s its quality on reemerging into the main channel, stabilizing variations in physical and chemical water char acteristics. Hyporheic flow is important for ecological functions, some aspects of water quali ty (such as temperature and dissolved oxygen), and some benthic invertebrate life stages. Alcove habitat, which has been limited by channelization, combines low hydraulic stress and high food availability w ith the potential for hyporheic flows across the steep hydraulic gradients in the gravel separating them from the main channel (Fernald et al. 2001).

Species-specific threats and limitations to recovery. Upper Willamette River Chinook recovery is specifically being hindered by severa l factors, including reduc ed access to spawning and rearing habitat because of tributary dams, de graded freshwater habita t, especially floodplain connectivity and function, channel structure and complexity, and riparian areas and large wood recruitment as a result of cumulative impacts of agriculture, forestry, and development, degraded water quality and altered temperature as a resu lt of both tributary dams and the cumulative impacts of agriculture, forestry, and urban deve lopment, hatchery-related effects, and ccean harvest rates of approximately 30 pe rcent (NMFS 2011j, ODFW and NMFS 2011). Significantly, anthropogenic introductions of non- native species and out-of-ESU races of salmon or steelhead have increased predation on, and competition with, native Upper Willamette River Chinook salmon to the point of hinde ring recovery of the species.

Chum salmon Description of the species. Chum salmon are more widely di stributed than other salmon and may have at one time made up nearly 50% of th e Pacific salmon biomass in the Pacific Ocean (Salo 1991). Historically, chum salmon were distributed throughout th e coastal regions of western Canada and the United States, as far sout h as Monterey Bay, California, to the Arctic

coast and east to the Mackenzie Ri ver, in the Beaufort Sea. They also ranged in Asia from Korea

to the Arctic coast of Russia and west to the Lena River. Presently, major spawning populations

on the west coast of the United States are f ound only as far south as Tillamook Bay on the northern Oregon coast. In this section of the Opinion, we disc uss the distribution, status, and critical habitats of the two listed species of threatened chum salmon separately; however, because chum salmon in the wild are virtually in distinguishable between listed ESUs, and are the same biological species sharing the same generalized life history, we begin this section describing those character istics common across ESUs.

Chum salmon exhibit obligatory anadromy (the re are no recorded landlocked or naturalized freshwater populations), and like Chinook salmo n, chum salmon are semelparous (die after one129 spawning event). Their general life cycle spans fresh and mari ne waters, although chum salmon

are more marine oriented than ot her Pacific salmon in that they spend very little time rearing in freshwater. Chum salmon spend 2-5 years in f eeding areas in the northeast Pacific Ocean, which is a greater proportion of their li fe history than other Pacific sa lmonids. Chum salmon distribute throughout the North Pacific Ocean and Bering Sea, although Nort h American chum salmon (as opposed to chum salmon originating in Asia), rarely occur west of 175° E longitude. North American chum salmon migrate north along the co ast in a narrow coastal band that broadens in southeastern Alaska, although some data sugge st that Puget Sound chum, including Hood Canal summer run chum, may not make extended migra tions into northern British Columbian and Alaskan waters, but instead may travel directly offshore into the north Pacific Ocean.

Habitat. Chum salmon are found in freshwater to eu ryhaline water at depths ranging from the surface to 250 m, although juveniles are primar ily epipelagic and are found from the surface down to 95 m and within 36 km of shore (Emmett et al. 1991c, Salo 1991). Chum salmon are found at a wide range of temperatures from 3° to 22° C but prefer temperatures from 8.3° to 15.6° C (Pauley et al. 1988). Juveniles migrat e within the Gulf of Al aska coastal belt during their first summer at sea (Salo 1991). Maturing individuals are al so distributed widely in the Gulf of Alaska during sp ring and summer (Salo 1991).

Chum salmon usually spawn in the lower reach es of rivers, with redds usually dug in the

mainstem or in side channels of rivers fr om just above tidal influence to roughly 80 km upstream. Juveniles outmigrate to seawater almo st immediately after emerging from the gravel that covers their redds (Salo 1991). This ocean -type migratory behavior contrasts with the stream-type behavior of some other species in the genus Oncorhynchus (e.g., coastal cutthroat trout, steelhead, Coho salmon, and most types of Chinook and sockeye salmon), which usually migrate to sea at a larger size, after months or years of freshwater rearing. This means that survival and growth in juvenile chum salm on depend less on freshwater conditions (unlike stream-type salmonids which depend heavily on fre shwater habitats) than on favorable estuarine

conditions. Another be havioral difference between chum salmon and species that rear extensively in freshwater is that chum salm on form schools, presumably to reduce predation (Pitcher 1986), especially if their movements ar e synchronized to swamp predators (Miller and Brannon 1982).

Reproduction. Spawning migrations generally occur in the summer and fall; the precise spawn timing and migration varies across populations. Stream flows and wa ter temperatures can influence stream entry. Sexual differences in the timing of returns to spawning grounds are apparent, with males generally arriving early an d females later in the run. Once on the spawning grounds mate competition is intense with male s competing to fertilize eggs and females competing for optimal nest site selection. Size and age at maturity is partially under genetic control, but can be influenced by environmen t and migration behavior . Generally, spawning runs consist of fish between 2 and 5 years of age, and like Chinook salmon, chum females will build large redds that consist of four or five egg pockets laid in succession. Chum salmon fecundity is highly variable, and is correlated with body size and region (l atitudinal trends are evident with northern population having lower ab solute and relative fecundities)(Salo 1991).

Size and age at maturity is partially under genetic control, but can be influenced by environment130 and migration behavior. Generall y, spawning runs consist of fish between 2 and 5 years of age, and like Chinook salmon, chum females will build la rge redds that consist of 4 or 5 egg pockets (Salo 1991). Chum salmon fecundity is highly variable, and is corre lated with body size and region (latitudinal trends are ev ident with northern population havi ng lower absolute and relative fecundities; (Salo 1991).

The time necessary for egg incubation until emer gence of alevins in freshwater varies among basins and among years within a basin, and is closel y correlated to water temperatures such that low temperatures prolong incubation. Egg and al evin survival, and the fitness of emerging fry are affected by sediment loading, intergravel wa ter flow and dissolved oxygen levels, gravel composition, spawning time and de nsity, and water temperatures.

Once they emerge from their gravel nests, c hum salmon fry outmigrate to seawater almost immediately (Salo 1991).

Feeding. Generally, chum fry emigrate to estuarie s between March and May where they forage on epibenthic and neritic food resources. As f ood resources decline and the fish grow, they move further out to forage on pelagic and nekt onic organisms (Simenstad and Salo 1982, Salo 1991). The timing of juvenile en try into seawater is commonly correlated with nearshore warming and associated plankton blooms (Groot and Margolis 1991) . General migratory studies indicate that chum salmon in their first year of life will typically maintain a coastal migratory pattern although the pattern is va riable as they mature at sea. At sea, chum salmon feed on pteropods, euphausiids, amphipods, fish, and squid larvae (Salo 1991). Chum salmon spend two to five years in feeding areas in the northeast Pa cific Ocean, which is a gr eater proportion of their life history than other Pacific salmonids.

Threats. Chum salmon are exposed to high rates of natural predation at each life stage, particularly during migration. Mortality at or prior to emergence is significant because eggs develop in the interstitial spaces in the stream gravel; storm surges that redeposit gravel and wash out eggs or introduce silt to the interstitial sp aces can reduce egg survival. Other factors that reduce egg survival and larvae developmen t include low dissolved oxygen, poor percolation, and extreme cold or warm temperatures. In fres hwater, fry fall prey to older salmon and other trout, as well as birds, sculpin, and various mammals; 10% of salmonid smolts are eaten by Caspian terns and double-crested cormorants annually in the Columbia River estuary (NMFS 2011l). Invasive fishes also threaten the su rvival and recovery of Pacific salmonids by competing directly for resources, altering food webs and trophic structures, and altering evolutionary traject ories (NMFS 2011l).

Chum salmon have declined under the combined effects of overharvests in fisheries; competition

from fish raised in hatcheries and native a nd non-native exotic species; dams that block their migrations and alter river hydrology; gravel mini ng that impedes their mi gration and alters the dynamics (hydrogeomorphology) of the rivers an d streams that support juveniles; water diversions that deplete water leve ls in rivers and streams; destru ction or degradation of riparian habitat that increase water temperatures in rivers and streams sufficient to reduce the survival of juvenile chum salmon; and land use practices (l ogging, agriculture, urbani zation) that destroy wetland and riparian ecosystems while introducin g sediment, nutrients, biocides, metals, and131 other pollutants into surface and ground water an d degrade water quality in the freshwater, estuarine, and coastal ecosystems throughout the Pacific northwest.

Columbia River chum salmon

Distribution. The Columbia River chum ESU includes all naturally-spawn ed populations of

chum salmon in the Columbia River and its tr ibutaries in Washington and Oregon (Figure 12). Three artificial propagation programs are part of the ESU. These artificially propagated populations are no more divergen t relative to the local natura l populations than would be expected between closely related po pulations within this ESU.

Figure 12. Population boundaries, dams, accessible areas, an d extirpated reaches of Columbia River chum distribution, incorporating critical habitat boundar ies (dark black line). Taken from NMFS (2011).

Most of the chum within this ESU return to northern tributaries of the Columbia River (in Washington State), primarily the Grays River, in areas immediat ely below Bonneville Dam, and in smaller numbers under the I-205 bridge near Vancouver. Chum populations that formerly occupied tributaries on the south bank of the Co lumbia (in Oregon) are c onsidered extirpated or nearly so. Observers have documented spawning over multiple years in the mainstem Columbia River, near McCord Creek and Multnomah Falls in Oregon, although the numb er of spawners in these areas are generally quite low (McElhany et al. 2007). Chum salmon return to the Columbia River in late fall (m id-October to December).132 Status and trends. The NMFS listed Columbia River chum salmon as threatened on March 25, 1999, and reaffirmed their status on June 28, 2005 (71 FR 37160). Chum salmon in the

Columbia River once numbered in the hundreds of thousands of adults and were reported in almost every river in the Lower Columbia Rive r basin, but by the 1950s most runs disappeared (Rich 1942, Marr 1943, Fulton 1970). The total number of chum salmon returning to the Columbia River in the last 50 y ears has averaged a few thousand pe r year, with returns limited to a very restricted portion of the historical range. Significant spaw ning occurs in only two of the 17 historical populations (Table 14), meaning that 88% of th e historical populations are extirpated, or nearly so. The two remaini ng populations are the Gray s River and the lower Columbia Gorge tributaries (Good et al. 2005b). Both long- and short-term trends for Grays River abundance are negative, but the current trend in abundance for the lower Columbia Gorge tributaries is slightly positive. Chum salmon app ear to be extirpated from the Oregon portion of this ESU. In 2000, ODFW conducted surveys to determine the abundance and distribution of chum salmon in the Columbia River, and out of 30 sites surveyed, only one chum salmon was

observed.

Table 14. Columbia River chum salmon populations and selected measures of population viability.
Current populations Historical

abundance a Recent spawner

abundance Short-term median growth rate ( λ ) c
Youngs Bay Gray’s River 7,511 331/704 b 1.043 (0.957-1.137) Big Creek

Elochoman River

Clatskanie River Mill, Abernathy, and Germany Creeks Scappoose Creek Cowlitz River 141,582 Kalama River 9,953 Lewis River 89,671

Salmon Creek

Clackamus River Sandy River Washougal River 15,140 Lower gorge tributaries >3,141 425 b 0.984 (0.883-1.096) Upper gorge tributaries >8,912

a Estimated total historical abundance for this ESU was ab out 283,421 fish, but is not meant to reflect a summation

of individual river historic estimates. Individual river estimates of historical abundance are based on an EDT analysis using equilibrium abundance under historical conditions. All data are reported in Good et al. 2005.

b Two different time series estimate are available but based on overlapping years. The first estimate is based on 1996-2000 data, while the second is based on 1996-2000 data.

c The λ calculation is an estimate of what the natural growth rate would have been after accounting for hatcheryorigin spawners. Two different estimates are available for the Grays River population; the Rawlings estimate (depicted in the table above) is believed to be more accura te. Other estimates, long- and short-term trends, suggest the population is declining (see Good et al. 2005).

Few Columbia River chum salmon have been ob served in tributaries between The Dalles and Bonneville dams. Surveys of the White Salmon River in 2002 found one male and one female carcass, with no evidence of spawning (Ehlke and Keller 2003). Chum salmon were not observed in any upper Columbia Gorge tri butaries during the 2003 and 2004 spawning ground surveys. Most Columbia River chum populations have been functionally extirpated or are133 presently at very low abundance levels. Outm igrating chum fry were observed moving past Bonneville Dam in 2010 for the first time (NMFS 2011d).

Historically, Columbia River chum salmon supported a large commercial fish ery in the first half of this century, which landed more than 500,000 fi sh per year as recently as 1942. Commercial catches declined beginning in the mid-1950s, and in later years rarely exceeded 2,000 per year. During the 1980s and 1990s, the combined abund ance of natural spaw ners for the lower Columbia Gorge, Washougal, and Grays Rive r populations was below 4,000 adults. In 2002, however, the abundance of natural spawners exhibite d a substantial increase at several locations (estimate of natural spawners is approximatel y 20,000 adults). The cause of this dramatic increase in abundance is unknown. However, l ong- and short-term productivity trends for populations are at or below replace ment. The loss of off-channel habitat and the extirpation of approximately 17 historical populations increase this species’ vulnerability to environmental variability and catastrop hic events. Overall, the populations that remain have low abundance, limited distribution, and poor connectivity (Good et al. 2005b).

Critical habitat. The NMFS designated critical habitat fo r Columbia River chum salmon on September 2, 2005 (70 FR 52630). The designate d includes defined areas in the following subbasins: Middle Columbia/Hood, Lowe r Columbia/Sandy, Lewis, Lower Columbia/Clatskanie, Lower Cowlitz, Lower Co lumbia subbasin and river corridor. This designation includes the stream ch annels within the designated st ream reaches, and includes a lateral extent as defined by the ordinary high water line. In ar eas where the ordinary high-water

line is not defined the lateral extent is defined as the bank full elevation.

The critical habitat designation for this ESU iden tifies primary constituent elements that include sites necessary to support one or more chum salmon life stages. These areas are important for the species’ overall conservati on by protecting quality growth, reproduction, and feeding and are rated as having high conservation value to the species. Columbia River chum salmon have primary constituent elements of freshwater spaw ning, freshwater rearing, freshwater migration, estuarine areas free of obstruction, nearshore ma rine areas free of obstructions, and offshore marine areas with good water quality. The physical or biological features that characterize these sites include water quality and qua ntity, natural cover, forage, adequate passage conditions, and floodplain connectivity. Of 21 subbasins reviewed in NMFS' assessment of critical habitat for the ESU, three subbasins were rated as having a medium conservation value, no subbasins were rated as low, and 18 subbasins, were rated as having a high conservation value. The major factors limiting recovery are altere d channel form and stability in tributaries, excessive sediment in tributary spawning gravels, a ltered stream flow in tributaries and the mainstem Columbia River, loss of some tributary habitat types, a nd harassment of spawners in the tributaries and

mainstem.

Species-specific threats and limitations to recovery. Lower Columbia River Chinook salmon

face numerous hindrances to their survival and re covery, including degraded estuarine and nearshore marine habitat resulting from cumulative impacts of land use and flow management by the Columbia River hydropower system and floodplai n connectivity and function, channel structure and complexity, riparian areas, stream substrate, stream flow, and water quality have been degraded as a result of cumulative impact s of agriculture, hydropower, forestry, and134 development (LCFRB 2010, NMFS 2011d). Further habitat-based impacts to the species result from an altered flow regime and Columbia Rive r plume has altered the temperature regime and estuarine food web, reduced water quality, and has reduced ocean productivity, reduced access to off-channel rearing habitat in the lower Columbia River , loss of access and loss of some habitat types as a result of passage barriers such as roads and railroads, reduced access to off-channel rearing habitat in the lower Columbia River, and reduced productivity resulting from sediment and nutrient-related changes in the estuary. Al so hampering the species is reduced access to spawning and rearing habitat main ly as a result of tributary hydropower projects, current or potential predation from hatchery-origin sa lmonids, including coho salmon, juvenile fish strandings that result from ship wakes, and cont aminants affecting fish health and reproduction.

Hood Canal summer-run chum salmon

Distribution. The Hood Canal summer-run chum salm on ESU includes all naturally spawned populations of summer-run chum salmon in Hood Ca nal and its tributaries as well as populations in Olympic Peninsula rivers between Hood Canal and Dungeness Bay, Washington (64 FR 14508)(Figure 13) from mid-September to mid- October (WDF (Washington Department of Fisheries) 1993), but may enter natal rivers in late August. Eight artificial propagation programs are considered to be part of the ESU. The NMFS determined that these artificially propagated populations are no more divergent relative to the local natural population(s) than what would be expected between closely related natural popula tions within the species. Table 15 identifies populations within the Hood Canal summer-run chum salmon ESU, their abundances, and hatchery input.

Figure 13. Population boundaries, dams, accessible areas, and extirpated reaches of Hood Canal summer-run chum distribution, incorporating critical habitat boundaries (dark black line). Taken from NMFS (NMFS 2011f).135 Table 15. Hood Canal summer-run chum populations and selected measures of population viability.
Populations a 1999-2002 mean escapement (range) Percent hatchery

contributions (1995-2001) λ (+/- SE)

Jimmycomelately Creek 10 (1-192) 0.85 (0.16)

Salmon/Snow creeks 1,521 (463-5,921) 0-69 1.23 (0.10) Big/Little Quilcene rivers 4,512 (3,065-6,067) 5-51 1.39(0.22) Lilliwaup Creek 13 (1-775) 1.19 (0.44)

Hamma Hamma River 558 (173-2,260) 1.3 (0.19)

Duckabush River 382 (92-942) 1.1 (0.17) Dosewallips River 919 (351-1,627) 1.17 (0.24)

Union River 1.15 (0.10)

Chimacum Creek* 198 (0-903) c 100 Big Beef Creek* 17 (0-826) c 100 Dewatto Creek* 9 (2-32) d

a All data is reported in Good et al. 2005. *Denotes extinct populations that have recently had some natural recolonization or have been seeded with hatchery fish.

On average Hood Canal chum salmon reside in es tuaries for 23 days; daily tidal migrations have not been observed, but prey availability does influence movement patt erns (Bax 1983). Upon leaving their natal estuaries, i ndividuals generally migrate thro ugh Hood Canal and into the main body of Puget Sound.

Status and trends. The NMFS listed Hood Canal summer-run chum salmon as threatened on March 25, 1999 (64 FR 14508), and reaffirmed th is status on June 28, 2005 (70 FR 37160). Historically, Hood Canal summer-run chum salmon comprised an estimated 16 populations; only eight extant populations remain (Good et al. 2005b). Most of the extirpated populations historically occurred on the eastern side of Hood Canal, wh ich is cause for concern over the current spatial structure of this ESU. The widespread loss of estuary and lower floodplain habitat is a continuing threat to ESU spatial structure and connectivity.

Although some population returns showed modest improvements in 2000, with upward trends continuing in 2001 and 2002, the recent 5-year mean abundance is variable, ranging from one fish to nearly 4,500 fish in the Big/Little Quil cene rivers. Hood Canal summer-run chum are the focus of an extensive rebuilding program de veloped and implemented since 1992 by state and tribal comanagers. Two populations (the comb ined Quilcene and Union River populations) are

above the conservation thresholds establishe d by the rebuilding plan. However, most populations remain depressed. Estimates of the fraction of naturally sp awning hatchery fish exceed 60% for some populations, indicating that reintroduction programs are supplementing the numbers of total fish spawning naturally in stream s. Long-term trends in productivity are above replacement for only the Quilcene and Union River populations. Buoyed by recent increases, seven populations are exhibiting short-term produc tivity trends above replacement. Although the 1994-2004 productivity trend for most populations wa s increasing, that trend generally reversed from 2005-2009 (NMFS 2011f).

Of the eight programs releasing indi viduals considered to be part of the ESU, six of the programs are supplementation programs implemented to preserve and increase the abundance of native populations in their natal watersheds. The NM FS’ assessment of the effects of artificial136 propagation on ESU extinction risk concluded that these hatche ry programs collectively do not substantially reduce the extinction risk of the ES U. The hatchery programs are reducing risks to ESU abundance by increasing total ESU abundance as well as the number of naturally spawning

individuals.

Critical habitat. The NMFS designated critical habita t for Hood Canal summer-run chum salmon on September 2, 2005 (70 FR 52630). The specific geographic area includes the Skokomish River, Hood Canal subbasin (Hamma Hamma and Dosewallips rivers and others), the Puget Sound subbasin, Dungeness/Elwha subbas in, and nearshore marine areas of Hood

Canal and the Strait of Juan de Fuca from the lin e of extreme high tide to a depth of 30 meters.

This includes a narrow nearshore zone from the extreme hightide to mean lower low tide within several Navy security/restricted zones. This also includes about 8 miles of habitat that was unoccupied at the time of the designation (Fin ch, Anderson and Chimacum creeks; 69 FR 74572; 70 FR 52630), but has recently been re-seeded. Chimacum Creek, however, has been naturally recolonized since at least 2007 (T. Johnson, pe rs. comm., Jan. 2010). The designation for Hood Canal summer-run chum, like others made at this tim e, includes the stream channels within the designated stream reaches, and includes a lateral extent as defined by the ordinary high water line. In areas where the ordinary high-water line is not defined the lateral ex tent is defined as the

bank full elevation.

The specific primary constituent elements identified for Hood Canal summer-run chum salmon are areas for spawning, freshwater rearing and migration, estuarine area s free of obstruction,

nearshore marine areas free of obstructions, and offshore marine areas with good water quality. The physical or biological features that char acterize these sites incl ude water quality and quantity, natural cover, forage, adequate pass age conditions, and floodplain connectivity. Of 17

subbasins reviewed in NMFS' assessment of critical habitat for the ESU, 14 subbasins were rated as having a high conservation value, while only thr ee were rated as having a medium value to the conservation. These areas are important fo r the species’ overall conservation by protecting quality growth, reproduction, and feeding. Limiti ng factors identified for this species include degraded floodplain and mainstem river channel structure, degraded es tuarine conditions and loss of estuarine habitat, ripari an area degradation and loss of in-river wood in the mainstems, excessive sediment in spawning gravels, a nd reduced stream flow in migration areas.

Degradation of the near-shore envi ronment has occurred in the southeastern areas of Hood Canal in recent years, resulting in late summer ma rine oxygen depletion and significant fish kills. Circulation of marine waters is naturally lim ited, and partially driven by freshwater runoff,

which is often low in the late summer. Howe ver, human development has increased nutrient loads from failing septic systems along the shor eline, and from use of nitrate and phosphate

fertilizers on lawns and farms. Shoreline reside ntial development is widespread and dense in many places. The combination of highways and dense residential development has degraded certain physical and chemical characteristics of the n ear-shore environment (Brewer et al. 2007, SSPS 2007).

Species-specific threats and limitations to recovery. Nearshore and estuarine habitat throughout the range of the species has b een altered by human activities (Brewer et al. 2007, NMFS 2011f). Nutrient loading ha s lowered dissolved oxygen concen trations, which can kill or137 stress marine organisms, including salmon. Residential and commercial development has reduced the amount of functioning habitat available for salmon r earing and migration. The loss of mudflats, eelgrass meadows, and macroalg ae further limits salmon foraging and rearing opportunities in nearshore and estu arine areas. In addition, fl oodplain connectivity and function, channel structure and complexit y, riparian areas and large wood supply, stream substrate, and stream flow have been degraded as a result of cumulative impacts of ag riculture, forestry, and development.

Coho salmon Description of the species. Coho salmon occur naturally in most major river basins around the

North Pacific Ocean from central Ca lifornia to northern Japan (Laufle et al. 1986). The typical life history of Coho salmon is similar to most of the other large bodied Pacific salmonids, in so much as adult fish spawn in the fall and wint er, young emerge in the spring, rear in freshwater and saltwater and return to spawn as adults. Sympatric in many river basins with Chinook, chum, sockeye, and pink salmon, partitioning occu rs through the species’ use of different areas of a river for reproduction and rearing, and the le ngth of time they spend in these ecosystems. For instance, Chinook salmon spawn in fast flow ing mainstem riverine reaches with large substrate; sockeye salmon spawn in rivers and lakes, and chum salmon spawn in mid- to lower reaches of rivers and have been observed spawning in areas of tidal influence. Coho salmon characteristically spawn in tributaries and sl ow-flowing shallow creeks in tributaries with gradients of 3% or less, which may be fed by cool groundwater sources, and are often widely dispersed within watersheds. Adult Coho salmon may remain in freshwater three or more months before spawning, with early migrants often moving fart her upstream (Sandercock 1991).

Most Coho salmon enter rivers between Septembe r and February, but entry is influenced by discharge and other factors. In many river systems, Coho salmon are unable to enter the rivers until sufficiently strong flows open passages and pr ovide sufficient depth. First fall freshets combined with high tides trigger the upstream migration of Coho salmon in many basins. Until then, if river flows are low or warm summer te mperatures persist, fish may congregate in pools

near the mouth of the river or natal stream until conditions change. Typically Coho salmon spawn from November to January, although ther e are many exceptions throughout their range. Spawning duration usually spans about three months in most basins, with individual fish actively spawning for several days to weeks. Spawning o ccurs in a few third-order streams, but most spawning activity occurs in fourth - and fifth-order streams. As with other P acific salmon, Coho salmon fecundity varies with the size of the fish and latitudinally with Coho salmon in northern climes generally exhibiting higher rates of fecundity (Sandercock 1991). Most Coho salmon mature and spawn at age 3, although there are exce ptions; in many basins in the northern portion of the species range Coho salmon spawn at age 2.

Habitat. Coho salmon are found in depths ranging from the surface to 250 m, but individuals in the open ocean generally stay w ithin 30 m of the surface (Emmett et al. 1991c). Juveniles occur at even shallower depths (<10 m)(PFMC 2000). Juveniles are also found closer to shore; generally within 74 km (PFMC 2000, NMFS-NWR 2005) . However, adults have been tracked well beyond the EEZ off Oregon (PFMC 2000). In y ears of weak upwelling, individuals tend to concentrate over submarine canyons and areas of more permanent upwelling, while strong upwelling years result in more dispersed stocks. Acceptable temperature regimes run from 4° to138 15.2° C, but optimal range is between 8° and 12° C (Emmett et al. 1991c).

Rates of incubation are largely temperature depe ndent: colder water temperatures will slow development. Generally, in optimal temperatur es eggs incubate for about 35-50 days, and fry start emerging from the gravel two to three weeks after hatching. Incubation and emergence success are also influenced by dissolved oxygen levels, sediment loading, and scouring high flows. Following emergence, fry aggregate and move to shallow areas near the stream banks.

Most Coho salmon rear in freshwater for about 15-18 months. As fry grow, they disperse upand downstream to establish and defend territories. Juvenile rearing usually occurs in tributaries with gradients of 3% or less, although they ma y move to streams with gradients of 4-5%.

Juvenile Coho salmon are often found in small str eams less than five feet wide, and may migrate

considerable distances to rear in lakes and off-channel ponds. During the summer, fry prefer pools featuring adequate cover such as large woody debris, undercut banks, and overhanging vegetation. Overwintering tends to occur in larger pools, backwater areas, and off stream channels and ponds (e.g., wall-based cha nnels that are groun dwater fed).

At not quite 2 years of age, Coho salmon w ill migrate downstream where they undergo the physiological transition to salt wate r. The outmigration of smolts begins as early as February and may continue through the summer and fall, with peak outmigration often between March and June, although this varies among basins and environmental conditions (Sandercock 1991). Several weeks are spent in coastal waters prio r to northward migration (PFMC 2000). This is particularly true for Coho orig inating from Oregonian streams, whose northward movement is generally delayed by strong southerly currents which weaken in the winter months (PFMC 2000). Once in the ocean, Coho salmon generall y migrate north along the coast in a narrow

coastal band that broadens in southeastern Alaska. During this migration, juvenile Coho salmon

tend to occur in both coastal and offshore wate rs. During spring and summer, Coho salmon will forage in waters between 46o N, the Gulf of Alaska, and along Alaska’s Aleutian Islands (PFMC 2000).

Feeding. Coho salmon are opportunistic feeders. While at sea, Coho salmon tend to eat fish, including herring, sand lance, sticklebacks, sardines, shrimp and surf smelt (Emmett et al. 1991c). While in estuaries and in freshwater Coho salmon are significant predators of Chinook, pink, and chum salmon, as well as aq uatic and terrestrial insects. Smaller fish, such as fry, eat chironomids, plecopterans, and othe r larval insects, and typically use visual cues to find their prey. Juveniles appear to prefer to feed in upwelled oceanic waters, although they are also present in eddy systems (Pool et al. 2008). Oceanic juveniles commonly feed upon euphausiids, chaetognaths, and decapod megalopae in these locations (Pool et al. 2008, Brodeur et al. 2010).

Threats . Coho salmon, like other salmon, are exposed to high rates of natura l predation at each life stage. Most mortality, howev er, occurs in freshwater. Wint er mortality may be significant for Coho salmon because they overwinter in fres hwater, where they can be swept downstream from freshets or eaten by racc oon, cutthroat trout, or other sma ll animals. Once Coho reach the ocean, survival is high (Sandercock 1991). In fres hwater, fry fall prey to older steelhead and other trout, as well as birds, sc ulpin, and various m ammals; 10% of salmonid smolts are eaten by Caspian terns and double-crested cormorants annually in the Columbia River estuary (NMFS 2011l). Invasive fishes also threaten the su rvival and recovery of Pacific salmonids by139 competing directly for resources, altering food webs and trophic structures, and altering evolutionary traject ories (NMFS 2011l).

Coho salmon have declined under the combined e ffects of overharvests in fisheries; competition

from fish raised in hatcheries and native a nd non-native exotic species; dams that block their migrations and alter river hydrology; gravel mini ng that impedes their mi gration and alters the dynamics (hydrogeomorphology) of the rivers an d streams that support juveniles; water diversions that deplete water leve ls in rivers and streams; destru ction or degradation of riparian habitat that increase water temperatures in rivers and streams sufficient to reduce the survival of juvenile Coho salmon; and land use practices (l ogging, agriculture, urbani zation) that destroy wetland and riparian ecosystems while introducin g sediment, nutrients, biocides, metals, and other pollutants into surface and ground water and degrade water quality in the fresh water, estuarine, and coastal ecosyste ms throughout the species range.

Central California Coast Coho salmon

Distribution. The Central California Coast coho salmon ESU extends from Punta Gorda in northern California south to and including the San Lorenzo River in central California (Weitkamp et al. 1995a). The ESU includes all naturally spawned populations of coho salmon

from Punta Gorda in northern California south to and including the San Lorenzo River in central California, as well as populations in tributaries to San Franci sco Bay, excluding the SacramentoSan Joaquin River system. Two natural populati ons have been genetically identified (NMFS 2011a). Four artificial propagation programs ar e part of the Central California Coast coho salmon ESU: the Don Clausen Fish Hatchery Captive Broodstock Program, Scott Creek/King although 12 independent and 63 dependent populati ons once existsed (NMFS 2011a). Fisher Flats Conservation Program, Scott Creek Captiv e Broodstock Program, an d the Noyo River Fish Station egg-take Program (recently terminated) coho hatchery programs. These artificially propagated populations are no more divergent re lative to the local na tural populations than would be expected between closely rela ted populations within this ESU.

Coho salmon in this ESU enter rivers to spawn very late (peaking in Janu ary), with little time spent in fresh water between river entry and spawning. This compressed adult freshwater residency appears to coincide w ith the single, brief peak of ri ver flow characteristic of this region.

Status and trends. NMFS originally listed the central California coast coho salmon ESU as threatened on October 31, 1996 (61 FR 56138) and la ter reclassified their status to endangered June 28, 2005 (70 FR 37160). Information on the abundance and productivity trends for the naturally spawning component of the central California coast coho ESU is extremely limited. There are no long-term time series of spawner ab undance for individual rive r systems. The best available data are from Lagunitas Creek, where re d data have been collected since 1997 (NMFS 2011a). Here, red abundance has been declining since a peak in 2004 (NMFS 2011a). Historical estimated escapement for this ESU is 56,100 fo r 1963, and more recent es timates suggest the ESU dropped to about one-fourth that si ze by the late 1980s and early 1990s (Good et al. 2005c).

Where data are available, analyses of juveni le coho presence-absence information, juvenile density surveys, and irregular adult counts for the South Fork Noyo River indicate low140 abundance and long-term downward trends for the naturally spawning populations throughout the ESU (Good et al. 2005c, NMFS 2011a). Some dependent populations have recent estimates since 1999 or 2000, including Pudding Creek (495 spawners), Casper Creek (155 spawners), Little Creek (40 spawners), Redwood Creek (0 to 93 redds and on a declining trend), and Scott Creek (declining trend since 2003- 2004 with a couple to few hundred spawners in these years; almost all were hatchery origin) (NMFS 2011i). Improved ocean conditions coupled with

favorable stream flows and harves t restrictions have contributed to increased returns in 2001 in streams in the northern portion of the ESU, as i ndicated by an increase in the observed presence of fish in historically occupied streams. Data are particularly l acking for many river basins in the southern two-thirds of the ESU where naturally spawning populations are considered to be at the greatest risk. The extirpation or near extirpation of natural co ho salmon populations in several major river basins, and across most of the southe rn historical range of the ESU, represents a significant risk to ESU spatia l structure and diversity (Good et al. 2005c).

Artificial propagation of coho salmon within the Central California Coast ESU has declined since the ESU was listed in 1996 though it con tinues at the Noyo River and Scott Creek facilities, and two cap tive broodstock populations have recen tly been established. Genetic diversity risk associated with out-of-basin transfers appears to be minimal, but diversity risk

from domestication selection a nd low effective population sizes in the remaining hatchery programs remains a concern. An out-of-ESU artificial propagation program for coho was operated at the Don Clausen hatchery on th e Russian River through the mid 1990s, but was terminated in 1996. Termination of this progra m was considered by the biological review team as a positive development for naturally produced coho in this ESU.

For the naturally spawning component of the ESU, the biological review team found very high risk of extinction for the abundance, productivity , and spatial structure of the Viable Salmon Population (VSP) parameters and co mparatively moderate risk with respect to the diversity VSP parameter. The lack of direct estimates of the performance of the naturally spawned populations in this ESU, and the associated uncertainty th is generates, was of specific concern to the biological review team. Informed by the VSP ri sk assessment and the associated uncertainty, the strong majority opinion of the biological review team was that the naturally spawned component

of the Central California Coast coho ESU was “i n danger of extinction.” The minority opinion was that this ESU is “likely to become endange red within the foreseeable future” (70 FR 37160). Accordingly, NMFS upgraded the status of centr al California coast coho ESU to endangered on June 28, 2005 (70 FR 37160).

Central California Coast coho salmon populations con tinue to be depressed relative to historical numbers. Strong indications show that breedi ng groups have been lost from a significant percentage of historical stream range. Overall, available population trends indicate poor adult

returns within the ESU between 2006 and 2010 (NMFS 2011i). A number of coho populations in the southern portion of the range appear to be either extinct or nearly so, including those in Gualala, Garcia, and Russian rivers, as well as smaller coastal stream s in and south of San Francisco Bay (Good et al. 2005c, NMFS 2011a). However, juveniles were recently identified in Soquel Creek south of the San Lorenzo River, indicating a recoloniza tion of that area in 2007- 2008 (NMFS 2011a).141 Critical habitat. NMFS designated critical habitat for central California coast coho salmon on May 5, 1999 (64 FR 24049). The designation enco mpasses accessible reaches of all rivers (including estuarine areas and riverine reaches) between Punta Gorda and the San Lorenzo River (inclusive) in California, in cluding two streams entering Sa n Francisco Bay: Arroyo Corte

Madera Del Presidio and Corte Ma dera Creek. This critical ha bitat designati on includes all waterways, substrate, and adjacent riparian zone s of estuarine and riverine reaches (including off-channel habitats) below longstanding naturally impassable barriers (i.e. natural waterfalls in existence for at least several hundred years). Th ese areas are important for the species’ overall conservation by protecting grow th, reproduction, and feeding.

Lower Columbia River coho salmon

Distribution. The lower Columbia River Coho salmon ESU includes all naturally spawned populations of Coho salmon in the Columbia River and its tribut aries in Washington and Oregon, from the mouth of the Columbia up to and in cluding the Big White Salmon and Hood Rivers, and includes the Willamette River to Willamette Falls, Oregon (Figure 14). Twenty-five artificial propagation progra ms are part of this ESU.

Figure 14. Population boundaries, dams, accessible areas, and extirpated reaches of lower Columbia River coho distribution, incorporating critical habitat boundar ies (dark black line). Taken from NMFS (2011).

Two distinct runs distinguished by the timing of a dult returns to freshwater (early returners and later returners) occur within the ESU. Early returning adults generally migrate south of the Columbia River once they reach the ocean, returning to freshwater in mid-August and to spawning tributaries in early Sept ember. Peak spawning of early returning adults occurs from mid-October to early November. Late returni ng adults exhibit a northern oceanic distribution,142 returning to the Columbia River from late Se ptember through December, and enter tributaries from October through January. Most late re turn adults spawn be tween November through January, although some spawn in February and as late as March (LCFRB 2004). Almost all

Lower Columbia River ESU Coho salmon females and most males spawn at 3 years of age.

Status and trends. The NMFS listed Lower Columbia River Coho salmon as endangered on June 28, 2005 (70 FR 37160). Twenty four historic al populations occur in the ESU; 21 are at high risk of extinction while th e other three are fairing some what better (NMFS 2011d). The vast majority (over 90%) of the historic population in the Lower Columbia River Coho salmon ESU appear to be either extirpated or nearly s o. Recent counts of natu ral-origin spawners and the recent fraction of hatchery-origin spawners are noted in Table 16, where available.

Table 16. Lower Columbia River Coho salmon populations and selected measures of population viability.

River 2002 spawner
count a Geometric

mean abundance

2000-2002 b Percent hatchery

contributions c Long-term median growth rate ( λ ) d

Youngs Bay and Big Creek 4,473 91 Grays River

Elochoman River

Clatskanie River 229 60 Mill, Germany, and Abernathy

creeks Scappoose Rivers 458 0 Cispus River

Tilton River Upper Cowlitz River

Lower Cowlitz River

North Fork Toutle River

South Fork Toutle River

Coweeman River

Kalama River

North Fork Lewis River

East Fork Lewis River Upper Clackamas River 1,001 2,122 12 1.009 (0.898-1.177) Lower Clackamas River 2,402 78

Salmon Creek Upper Sandy River 310 643 0 1.012 (0.874- 1.172) Lower Sandy River 271 97 Washougal River Columbia River Gorge – lower

tributaries

White Salmon Columbia River Gorge – upper

tributaries 1,317 >65

Hood River

a All data are reported in Good et al. 2005. Spawner data from 2002 only.

b Geometric mean number of Coho salmon above the dams. * is a combined total for the upper and lower Clackamas River. Reported in Good et al. 2005

c Hatchery production likely dominates yearly returns for the ESU as a whole.

d The λ calculated estimates the natural growth rate after accounting for hatchery-origin spawners. The estimate provided above assumes that hatchery-origin spawners make no reproductive contribution. The λ for the Clackamas143 River is calculated with data spanning 1973-2002, and for the Sandy River covers 1977-2002. The Clackamas River value includes both early-run and late-run Coho salmon.

Only two populations of Coho salmon within th is ESU produce a sizeable number of naturally spawned fish, the upper Sandy River population a bove Marmot Dam and the Clackamas River population above the North Fork Dam. Excludi ng these, natural spawner abundance is below 500 individuals (NMFS 2011d). A survey from Mill/Germany/Abernathy creeks produced an estimate of 3,150 spawners in 2006 ; over half of these were ha tchery fish (NMFS 2011d). The long-term and short-term trends for Marmot Dam counts are both negative. The long-term median growth rate is slightly positive fo r both the Sandy and Clackamas rivers, but the confidence intervals for each are very wide indi cating there is a large amount of uncertainty. Both populations within the Sandy and Clackamas rivers have suffe red from recruitment failure a number of times over the past 15 years.

The most serious threat facing this ESU is the scarcity of naturally-produced spawners, with attendant risks associated with small populat ion, loss of diversity, and fragmentation and isolation of the remaining naturally-produced fi sh. Spatial structure has been substantially reduced by the loss of access to upper basins from tributary hydro development (i.e., Condit Dam on the Big White Salmon River and Powerdale Dam on the Hood River). The diversity of populations in all three areas has been eroded by la rge hatchery influences and periodically, low effective population sizes.

Critical habitat. The NMFS has not designated critical habitat for Lower Columbia River Coho

salmon.

Species-specific threats and limitations to recovery. Several direct and ha bitat-related factors inhibit the recovery of Lower Columbia River Coho salmon (LCFRB 2010, NMFS 2011d). Habitat impacts include degraded estuarine a nd near-shore marine habitat resulting from cumulative impacts of land use and flow mana gement by the Columbia River hydropower system, fish passage barriers that limit access to spawning a nd rearing habitats, floodplain connectivity and function, channe l structure and complexity, ri parian areas and large wood supply, stream substrate, stream flow, and water quality have been degraded as a result of cumulative impacts of agriculture, forestry, and development. Additiona l threats to habitat include an altered flow regime and Columbia River plume has altered the temperature regime and estuarine food web, and has reduced ocean productivity, reduced access to off-channel rearing habitat in the lower Columbia River, as well as reduced productivity resulting from

sediment and nutrient-related ch anges in the estuary. Other imp acts also include hatchery- and harvest-related effects, juvenile fish strandings that result from ship wakes, and contaminants affecting fish health and reproduction.

Oregon Coast coho salmon

Distribution. The Oregon Coast Coho salmon ESU include s all naturally spawned populations of Coho salmon in Oregon coastal streams south of the Columbia River and north of Cape Blanco (63 FR 42587; August 1998). One hatche ry population, the Cow Creek hatchery Coho salmon, is considered part of the ESU. Table 17 identifies populations within the Oregon Coast Coho salmon ESU, their abundances, and hatchery input.144 Table 17. Oregon Coast Coho populations and selected measures of population viability.

Basin a Mean Spawner

Abundance b 13-Year Spawner Trend (SE) c Percent Hatchery

Contribution d

Necanicum 1,889 1.16 9 (0.860) 2.9-6.4 Nehalem 18,741 1.206 (0.889) 0.5-26.0 Tillamook Bay 3,949 1.191 (1.084) 0-5.6 Nestucca 3,846 1.230 (1.015) 0-10.4 Siletz 2,295 1.070 (0.760) 1.8-100 Yaquima 3,665 1.204 (1.205) 0-37.5 Alsea 3,621 1.042 (0.960) 0-87.5 Siuslaw 16,213 1.120 (1.037) 0.3-11.1 Umpqua 24,351 1.182 (0.662) 2.1-8.3 Coos 20,136 1.088 (1.066) 0-1.9 Coquille 8,847 1.070 (0.649) 0-6.0

a Population structure is unclear. The above data reflects the assumption that spawners from major river basins are largely isolated, and each basin comprises a p opulation. All data are reported in Good et al. 2005.

b Recent 3-year geometric mean of natural-origin spawners.

c Data years 1990-2002.

d Data represents the range of percent hatchery c ontributions from 1998 through 2002 (from Jacobs et al. 2002, 2001,

and 2002 as cited in Good et al. 2005).

Status and trends. The Oregon coast Coho salmon ESU wa s listed as a threatened species under the ESA on February 11, 2008 (73 FR 7816), th e conclusion to a 13-year history of court

cases. The most recent NMFS status review for the Oregon Coast Coho ESU was conducted by the biological review team in 2003, which a ssessed data through 2002. The abundance and productivity of Oregon Coast Coho since the previous status review represented some of the best and worst years on record (NMFS 1997a). Yearly adult returns for the Oregon Coast Coho ESU were over 160,000 natural spawners in 2001 and over 260,000 in 2002, far exceeding the abundance observed for the past several decades . These increases in spawner abundance in 2000 to 2002 followed three consecutive brood years (the 1994-1996 brood years returning in 1997 to 1999, respectively) exhibiting recruitment failure (re cruitment failure is when a given year class of natural spawners fails to replace itself when its offspring return to the spawning grounds 3 years later). These 3 years of recruitment failure were the only such instances observed thus far in the entire 55-year abu ndance time series for Oregon Coast Coho salmon (although comprehensive population-level survey data have only been available since 1980). The 2000 to 2002 increases in natural spawner abundance oc curred in many populations in the northern portion of the ESU, which were the most depresse d at the time of the last review (NMFS 1997a). Although encouraged by the increase in spaw ner abundance in 2000-2002, the biological review team noted that the long-term trends in ESU productivity were still negative due to the low abundances observed during the 1990s.

Since the biological review team convened, th e total abundance of natu ral spawners in the Oregon Coast Coho ESU has declined each year (i.e., 2003 to 2006). The abundance of total natural spawners in 2006 (111,025 spawners) was approximately 43% of the recent peak abundance in 2002 (255,372 spawners). In 2003, ESU-le vel productivity (evaluated in terms of the number of spawning recruits resulting from spaw ners 3 years earlier) was above replacement, and in 2004, productivity was approximately at replacement level. However, productivity was below replacement in 2005 and 2006, and dropped to the lowest level si nce 1991 in 2006 (73 FR 7816).145 Preliminary spawner survey data for 2007 (the average peak number of spawners per mile observed during random Coho spawning surveys in 41 streams) suggest that the 2007 to 2008 return of Oregon Coast Coho is either (1) much reduced from abundance levels in 2006, or (2) exhibiting delayed run timing from previous ye ars. As of December 13, 2007, the average peak number of spawners per mile was below 2006 le vels in 38 of 41 surveyed streams (ODFW 2007

in 73 FR 7816). It is possible th at the timing of peak spawner abundance is delayed relative to previous years, and that increased spawner abundance in late December and January 2008 will compensate for the low levels observed thus far.

The recent 5-year geometric mean abundance (2002 to 2006) of approximately 152,960 total natural spawners remains well above that of a decade ago (approximately 52,845 from 1992 to 1996). However, the decline in productivity from 2003 to 2006, despite generally favorable

marine survival conditions and low harves t rates, is of con cern (73 FR 7816).

Critical habitat. The NMFS designated critical habitat for Oregon Coast Coho on February 11, 2008 (73 FR 7816). The designation includes 72 of 80 watersheds occupied by Oregon Coast Coho salmon, and totals about 6,600 stream miles including all or portions of the Nehalem, Nestucca/Trask, Yaguina, Alsea, Umpqua and Coquille basins. These areas are essential for feeding, migration, spawning, and rearing. The sp ecific primary constitu ent elements include: spawning sites with water and substrate quant ity to support spawning, incubation, and larval development; freshwater rearing sites with wa ter quantity and floodplain connectivity to form and maintain physical habitat conditions and su pport juvenile growt h, foraging, behavioral development (e.g., predator avoidance, compe tition), and mobility; freshwater migratory corridors free of obstruction with adequate wate r quantity and quality conditions; and estuarine,

nearshore and offshore areas free of obstructi on with adequate water quantity, quality and salinity conditions that support physiological trans itions between fresh- and saltwater, predator avoidance, foraging and othe r life history behaviors.

The historical disturbance regime in the central Oregon Coast Range was dominated by a mixture of high and low-severity fires, with a natural rotation of approximately 271 years. Oldgrowth forest coverage in the Oregon Coast Range varied from 25 to 75% during the past 3,000 years, with a mean of 47%, and never fell below 5% (Wimberly et al. 2000). Currently, the Coast Range has approximately 5% old-growth, almo st all of it on Federa l lands. The dominant disturbance now is logging on a cycle of approxima tely 30 to 100 years, with fires suppressed.

The State of Oregon (2005) completed an assessm ent of habitat conditio ns in the range of Oregon coast coho salmon in 2005. Oregon’s a ssessment mapped how streams with high intrinsic potential for coho salmon rearing ar e distributed by land ow nership categories. Agricultural lands and private industrial forests have by far the highest percentage of land ownership in high intrinsic potenti al areas and along all coho salm on stream miles. Federal lands have only about 20% of coho salmon stream mile s and 10% of high intrin sic potential stream reaches. Because of this distribution, activities in lowland agricultural areas are particularly important to the conservation of Oregon coast coho salmon.

The Oregon coast coho salmon assessment concluded th at at the scale of the entire domain, pools are generally abundant, although sl ow-water and off-channel ha bitat (which are important146 refugia for coho salmon during high winter flows) are limited in th e majority of streams when compared to reference streams in minimally-dis turbed areas. Amounts of large wood in streams are low in all four Oregon Department of Fish and Wildlife monitoring ar eas and land-use types

relative to reference conditions. Amounts of fine sediment are high in three of the four monitoring areas, and were comparable to reference conditions only on public lands. Approximately 62 to 91% of tidal wetland acres (depending on estimation pr ocedures) have been lost for functionally and potentially independent populations of coho salmon.

As part of the coastal coho salmon assessment, the Oregon Department of Environmental Quality analyzed the status and trends of water quality in the range of Oregon coast coho salmon using the Oregon water quality index, which is based on a combination of temperature, dissolved oxygen, biological oxygen demand, pH , total solids, nitrogen, total phosphates, and bacteria. Using the index at the species s cale, 42% of monitored sites ha d excellent to good water quality, and 29% show poor to very poor water quality. Within the four monitoring areas, the North

Coast had the best overall conditi ons (six sites in excellent or good condition out of nine sites), and the Mid-South coast had the poorest conditio ns (no excellent conditio n sites, and only two out of eight sites in good c ondition). For the 10-year peri od monitored between 1992 and 2002, no sites showed a declining trend in water qualit y. The area with the most improving trends was the North Coast, where 66% of th e sites (six out of ni ne) had a significant improvement in index scores. The Umpqua River basin, with one out of nine sites (11%) showing an improving trend, had the lowest number of improving sites.

Species-specific threats and limitations to recovery. Oregon coast coho salmon are impaired in their recovery by several an thropogenic factors, including floodplain connectivity and function, channel structure and complexity, ri parian areas and large wood supply, stream substrate, stream flow, and wate r quality that have been degr aded as a result of cumulative impacts of agriculture, forestry, in stream mining, dams, road crossi ngs, dikes, levees, etc. (Stout

et al. 2011). In addition, fish passage barriers lim it access to spawning and rearing habitats and adverse climate, altered past ocean/marine pr oductivity, and current ocean ecosystem conditions have favored competitors and predators and reduced salmon survival rates in freshwater rivers and lakes, estuaries, and marine environments

Southern Oregon/Northern California coast coho salmon

Distribution. Southern Oregon/Northern California coas t coho salmon consists of all naturally spawning populations of coho salmon that reside below long-term, naturally impassible barriers in streams between Punta Gorda, California and Cape Blanco, Oregon, as well as three artificial propagation programs: the Cole Rivers Hatche ry, Trinity River Hatc hery, and Iron Gate Hatchery coho hatchery programs. The three major river systems supporting Southern Oregon –

Northern Coastal California coas t coho are the Rogue, Klamath (inc luding the Trinity), and Eel

rivers.

Southern Oregon and Northern California coast c oho immigrate to natal rivers in September or October. River entry is much later south of the Klamath River Basin, occurring in November and December, as well as in basins south of the Klamath River to the Mattole River, California. River entry occurs from mid-December to mid- February in rivers fa rther south. Because

individuals enter rivers late, they spend much less time in th e river. Coho salmon adults spawn147 at age 3, spending just over 1 year in fres h water and a year and a half in the ocean.

Status and trends. Southern Oregon/Northern Californi a coast coho salmon were listed as threatened on May 7, 1997 (62 FR 24588); they retain ed that classification when their status was reviewed on June 28, 2005 (70 FR 37160). Sout hern Oregon/Northern California Coast coho salmon extend from Cape Blanco in southern Or egon to Punta Gorda in northern California (Weitkamp et al. 1995b). The status of coho salmon coast-wide, including the Southern Oregon/Northern California Coast coho salmon ES U, was formally assessed in 1995 (Weitkamp

et al. 1995b). Two subsequent status review updates have been published by NMFS, one addressing all West Coast coho salmon ESUs a nd a second specifically addressing the Oregon Coast Southern Oregon/Northern California Coast coho salmon ESUs (NMFS 1996, 1997a). In the 1997 status update, estimates of natural population abundance were based on very limited information. New data on presence/absence in northern California streams that historically supported coho salmon were even more disturbing than earlier results, in dicating that a smaller percentage of streams contained coho salmon comp ared to the percentage presence in an earlier study. However, it was unclear whether these ne w data represented actual trends in local extinctions or were biased by sampling effort.

Data on population abundance and trends are limited for the California porti on of this ESU. No regular estimates of natural spawner escapement are available. Historical point estimates of coho salmon abundance for the early 1960s and mid-1980s suggest that statewide coho spawning escapement in the 1940s ranged between 200,000 and 500,000 fish. Numbers declined to about 100,000 fish by the mid-1960s with about 43% originating from this ESU. Brown et al. (1994)

estimated that the California portion of this ESU was represented by about 7,000 wild and naturalized coho salmon (Good et al. 2005b). In the Klamath River, the estimated escapement has dropped from approximately 15,400 in the mid-1960s to about 3,000 in the mid 1980s, and more recently to about 2,000 (Good et al. 2005b). The second largest producing river in this ESU, the Eel River, dropped from 14,000, to 4,000 to about 2,000 during the same period. Historical estimates are considered “best guess es” made using a combination of limited catch statistics, hatchery records, and the personal ob servations of biologists and managers. Although quantitiative estimates of abundance in this ESU ar e rare, the best indica tions are that since 2005, populations within the ESU have continued to decline (NMFS 2011h).

Most recently, Williams et al. (2006) described the structure of historic populations of Southern Oregon/Northern California Coast coho salmon. They described thre e categories of populations: functionally independent populat ions, potentially independent populations and dependent populations. Functionally indepe ndent populations are populations capable of existing in

isolation with a minimal risk of extinction. Po tentially independent popul ations are similar but rely on some interchange with adjacent populations to maintain a low probability of extinction. Dependent populations have a high risk of extinct ion in isolation over a 100-year timeframe and rely on exchange of individuals from adjacent populations to maintain themselves.

Critical habitat. NMFS designated critical habitat fo r Southern Oregon/Northern California Coast coho salmon on May 5, 1999 (64 FR 24049). Crit ical habitat for this species encompasses

all accessible river reaches between Cape Blanco, Oregon, and Punta Gorda, California. Critical habitat consists of the water, substrate, and river reaches (including o ff-channel habitats) in148 specified areas. Accessible reaches are those within the historical range of the ESU that can still be occupied by any life stage of coho salmon. Of 155 historical streams for which data are available, 63% likely still support coho salmon. Th ese river habitats are important for a variety of reasons, such as supporting the feeding and growth of juveniles and serving as spawning habitat for adults. Limiting factors identified for this species include: loss of channel complexity, connectivity and sinuosity, loss of floodpl ain and estuarine habita ts, loss of riparian habitats and large in-river wood, reduced stream flow, poor water quality, temperature and excessive sedimentation, and unscreened dive rsions and fish passage structures. The Elk River flows through Curry County, and dr ains approximately 92 square miles (or 58,678 acres)(Maguire 2001). Historical logging, mining, a nd road building have degraded stream and riparian habitats in the Elk River basin. Lim iting factors identified for salmon and steelhead production in this basin include sp arse riparian cover, especially in the lower reaches, excessive fine sediment, high water temperatures, a nd noxious weed invasions (Maguire 2001).

The Rogue River drains approximately 5,160 square miles within Curry, Jackson and Josephine counties in southwest Oregon. The mainstem is about 200 miles long and traverses the coastal mountain range into the Cascades. The Rogue River estuary has been modified from its historical condition. Jetties were built by the Army Corp s of Engineers in 1960, which stabilized and deepened the mouth of the river. A dike that extends from the south shore near Highway 101 to the south jetty was completed in 1973. This dike created a backwater for the large shallow area that existed here, which has been developed into a boat basin and marina, eliminating most of the

tidal marsh.

The quantity of estuary habitat is naturally limited in the Rogue River. The Rogue River has a drainage area of 5,160 square miles, but the es tuary at 1,880 acres is one of the smallest in Oregon. Between 1960 and 1972, approximately 13 acres of intertidal and 14 acres of subtidal land were filled in to build the boat basin dike, the marina, north shore ri prap and the other north shore developments (Hicks 2005). Jetties constructed in 1960 to st abilize the mout h of the river and prevent shoaling have altered the Rogue Ri ver, which historically formed a sill during summer months (Hicks 2005).

The Lower Rogue Watershed Council’s watershed analysis (Hicks 2005) lists factors limiting fish production in tributaries to Lower Rogue River watershed. The list includes water temperatures, low stream flows, riparian fo rest conditions, fish pa ssage and over-wintering habitat. Limiting factors identified for the Upper Rogue River basin include fish passage barriers, high water temperatures, insuffici ent water quantity, lack of large wood, low habitat complexity, and excessive fine sediment (RBCC 2006).

The Chetco River estuary has been significantly modified from its historical condition. Jetties were erected by the Army Corps of Engineers in 1957, which stabilized and deepened the mouth of the river. These jetti es have greatly altered the mouth of the Chetco River and how the estuary functions as habitat for salmon migrating to the ocean. A boat basin and marina were built in the

late 1950s and eliminated most of the functional tidal marsh. The structures eliminated shallow water habitats and vegetation in favor of banks stabilized with riprap. Since then, nearly all remaining bank habitat in the estuary has been stabilized with riprap. The factors limiting fish production in the Chetco River appear to be hi gh water temperature caused by lack of shade,149 especially in tributaries, high rates of sedimentation due to ro ads, poor over-wintering habitat due to a lack of large wood in tributaries a nd the mainstem, and poor quality estuary habitat (Maguire 2001).

Species-specific threats and limitations to recovery. Numerous anthropogenic factors hinder recovery of Southern Oregon/Nort hern California Coast coho salmon (NMFS 2007c, 2011h). These include lack of floodplain and channel st ructure, impaired water quality, estuarine function, fish passage, adverse hatchery-related effects, predation, competition, and disease, altered sediment transport, degraded stream subs trate, altered hydrologic function due to altered amount and timing of river flows, and degraded riparian forest c onditions and large wood

recruitment.

Sockeye salmon Description of the species. Sockeye salmon are the second most abundant of the seven Pacific salmon species, and occur in the North Pacific and Arctic oceans and associated freshwater systems. This species ranges south as far as th e Sacramento River in California and northern Hokkaido in Japan, to as far north as far as Bathur st Inlet in the Canadian Arctic and the Anadyr River in Siberia (Burgner 1991). The largest populations, and hence the most important commercial populations, occur north of the Columbia River.

Sockeye salmon exhibit a very diverse life history, characteristically using both riverine and lake habitat, exhibiting both freshwater resident and anadromous forms. The vast majority of sockeye

salmon are anadromous fish that make use of lacu strine habitat for juvenile rearing. These “laketype” fish typically spawn in the outlet streams of lakes and occasionally in the lakes themselves. Juvenile sockeye salmon will then use the lake environment for rearing for up to 3 years before migrating to sea. After 1 to 4 years at sea, sockeye salmon will return to their natal lake to spawn. Some sockeye, however, spawn in rivers without lake habitat for juvenile rearing. Offspring of these riverine spawners tend to use the lower veloc ity sections of rivers as the juvenile rearing environment for 1 to 2 years, or may migrate to sea in their first year.

Sockeye salmon also have a wholly freshwater life history form, called kokanee (Burgner 1991). In some cases a single population will give rise to both the anadromous and freshwater life history form. While in freshwater juvenile s of both life history types prey primarily upon insects. The presence of both life history t ypes may be related to the energetic costs of outmigrating to sea, and the productivity of the lacu strine system they inhabit. In coastal lakes, where the migration to sea is relatively s hort and energetic cost s are minimal, kokanee populations are rare.

Habitat. Once smolts enter the Pacific Ocean, they di stribute widely across the North Pacific, generally above 40o N, where a current boundary is located. Season, temperature, salinity, life stage, age, size, availability of prey and population-of-origin are all factors that influence offshore movements (Burgner 1991). Sockeye tend to occupy the upper water column, within 30 m of the surface and preferably within 15 m and tend to be closer to the surface at night versus daytime (Manzer 1964, French et al. 1976, Pauley et al. 1989, Burgner 1991). Sockeye migrate several thousand miles in search of prey and are considered to travel continuously (Royce et al. 1968). Juveniles leaving freshwat er move north along the Alaskan coastal belt beginning in mid-150 May until early winter when they disper se into the Gulf of Alaska (Emmett et al. 1991b, NMFSAKR 2005).

Thermoclines may also influence vertical dist ribution, with fish only mingling between surface and deeper waters when the boundary temperature di fference is weak. Sockeye appear to prefer

cooler waters relative to ot her salmon species, but younger sa lmon may prefer warmer sea surface temperatures (optimally 15° C) than la rger, older fish (12° and 14° C), possibly an artifact of older fish being distributed further north (Pauley et al. 1989). Adult upstream migration may be blocked by temperatures above 21o C (McCullough 1999). However, temperatures below 21o C can st ress fish by increasing their su sceptibility to disease and elevating their metabolism (Brett 1979). Maturati on and timing of return to spawn by sockeye appears to be linked to water temperature, with gonad development increasing in late May through early July (Nishiymama 1984).

Feeding. While in freshwater, juveniles prey primarily upon insects. While at sea, sockeye prey upon a variety of organisms, including small fish (capelin, lantern fish, cod, sand lance, herring, and pollock), squid, crustacean larvae, krill, and other invertebrates (Foerster 1968, French et al. 1976, Wing 1977).

Reproduction. Spawning generally occurs in late summ er and autumn, but the precise time can vary greatly among populations. Ag e at maturity varies by region from 2 to 5 years, but is generally 2 to 4 years in Wash ington State (Burgner 1991). Male s often arrive earlier than females on the spawning grounds, and will persis t longer during the spawning period. Average fecundity ranges from about 2,000-2,400 eggs per female up to 5,000 eggs, depending upon the population and average age of the female. Fecund ity in kokanee is much lower and may range from about 300-2,000 eggs.

Incubation is a function of water temperatures , but generally lasts between 100 and 200 days (Burgner 1991). After emergence, fry move ra pidly downstream or upstream along the banks to the lake rearing area. Fry emerging from la keshore or island spawning grounds may simply move along the shoreline of the lake (Burgner 1991).

Threats. In freshwater, fry fall prey to older salmon and other trout, as well as birds, sculpin, and various mammals; 10% of salmonid smolts are eaten by Caspian terns and double-crested cormorants annually in the Columbia River estuary (NMFS 2011l). In the ocean, marine mammals and other fish prey on sockeye, but the extent of such predation is not well known.

Invasive fishes also threaten the survival and recovery of Pacific salmonids by competing directly for resources, altering food webs and tr ophic structures, and altering evolutionary trajectories (NMFS 2011l).

Ozette Lake sockeye salmon

Distribution. This ESU includes all naturally spawne d sockeye salmon in Ozette Lake, Ozette River, Coal Creek, and other tr ibutaries flowing into Ozette Lake, Washington (Figure 15). Composed of only one population, the Ozette La ke sockeye salmon ESU consists of five spawning aggregations or subpopulations whic h are grouped according to their spawning locations. The five spawning locations are Um brella and Crooked cr eeks, Big River, and151 Olsen’s and Allen’s beaches (NMFS 2009c).

Figure 15. Population boundaries of Lake Ozette sockeye di stribution, incorporating critical habitat boundaries (dark black line). Taken from NMFS (2011).

Adult Ozette Lake sockeye salmon enter Ozette La ke through the Ozette Ri ver from mid-April to mid-August, holding three to nine months in Ozet te Lake prior to spaw ning in late October through January. Sockeye salmon spawn primarily in lakeshore upwelling ar eas in Ozette Lake (particularly at Allen's Bay and Olsen's Beach), and in two tributaries Umbrella Creek and Big River. Minor spawning may occur below Ozette Lake in the Ozette River or in Coal Creek, a tributary of the Ozette River. Beach spawners are almost a ll age-4 adults, while tributary spawners are ages 3 and 5 (NMFS 2009c). Spawning occurs in the fall through early winter, with peak spawning in tributarie s in November and December. E ggs and alevins remain in the gravel until the fish emerge as fry in spring. Fry then migrate immediately to the limnetic zone in Ozette Lake, where the fish rear. After one year of rearing, in late spring, Ozette Lake sockeye salmon emigrate seaward as age-1+ smo lts, where they spend between 1 and 3 years in ocean before returning to freshwater.

Status and trends. The NMFS originally listed Ozette Lake sockeye salmon ESU as a threatened species in 1999 (64 FR 14528). This classification was retain ed on June 28, 2005 (70 FR 37160). This ESU includes all naturally spaw ned populations of sockeye salmon in Ozette Lake, Ozette River, Coal Cree k, and other tributaries flowing into Ozette Lake, Washington. Two artificial propagation programs are considered pa rt of this ESU. The NMFS considers these artificially propagated populations no more divergent relative to the local natural population than would be expected between closely rela ted natural populations (70 FR 37160).152 The historical abundance of Ozette Lake sock eye salmon is poorly documented, but may have been as high as 50,000 individuals (Blum 1988). The overall abundance of naturally–produced Ozette Lake sockeye salmon is belie ved to have declined substantiall y from historical levels. In the first study of lake escape ment of Ozette Lake sockeye salmon (Kemmerich 1945), the run size entering the lake was estimated at a level of several thousand fish. These counts appear to be roughly double the current mean lake abundance, considering th at they were likely conducted upstream from fisheries in or near to the Oze tte River. Makah Fisheries Management (MFM 2000) concluded that there appears to be a substantial decline in th e tribal catch of Ozette Lake sockeye salmon beginning in the 1950s and a simila r decline in the run si ze since the 1920s weir counts reported by Kemmerich (1945).

An analysis of total annual O zette Lake sockeye salmon abundan ce (based on adult run size data presented in Jacobs et al. 1996) indicates a trend in abundanc e averaging -2% per year over the period 1977 through 1998 (NMFS 1998c). The current tributary-based hatchery program was planned and initiated in response to the declinin g population trend identified for the Ozette Lake sockeye salmon population. The most recent (19 96 to 2003) run-size estimates range from a low of 1,609 in 1997 to a high of 5,075 in 2003, averaging approximately 3,600 sockeye per year (NMFS 2009c). For return years 2000 to 2003, the 4-year average abundance estimate was slightly over 4,600 sockeye. Because run-size esti mates before 1998 are likely to be even more unreliable than recent counts, and new counti ng technology has resulted in an increase in estimated run sizes; no statistical estimation of trends is reported. The current trends in abundance are unknown for the beach spawning a ggregations. Although overall abundance appears to have declined from historical leve ls, whether this resulted in fewer spawning aggregations, lower abundances at each aggregation, or both, is not known (Good et al. 2005b). Based on estimates of habitat carrying capacit y, a viable sockeye salmon population in Lake Ozette watershed would range be tween 35,500 to 121,000 spawners (Rawson et al. 2009).

There has been no harvest of Ozette Lake sock eye salmon for the past four brood-cycle years (since 1982). Prior to that time, ceremonial a nd subsistence harvests by the Makah Tribe were low, ranging from 0 to 84 fish per year. Harvest has not been an important mortality factor for the population in over 35 years. In addition, du e to the early river entry timing of returning Ozette Lake sockeye salmon (beginning in late Apr il, with the peak returns prior to late-May to mid-June), the fish are not intercepted in Cana dian and United States marine area fisheries directed at Fraser River sockeye salmon. Th ere are currently no known marine area harvest impacts on Ozette Lake sockeye salmon.

Overall abundance is substantiall y below historical levels (Good et al. 2005b). Declines in

abundance have been attributed to a combina tion of introduced species , predation, loss of tributary populations, a loss of quality of beach spawning habitat, temporarily unfavorable ocean conditions, habitat degradation, and ex cessive historical harvests (Jacobs et al. 1996). In the last few years the number of returning adults has increased, although some of thes e individuals are of hatchery origin. This produces uncertainty rega rding natural growth rate and productivity of the ESU’s natural component. In addition, genetic in tegrity has perhaps been compromised due to the artificial supplementation th at has occurred in this popul ation, since approximately one million sockeye have been released into the Oz ette watershed from the late 1930s to present153 (Kemmerich 1945, Boomer 1995).

Critical habitat. On September 2, 2005, the NMFS designated critical habitat for the Ozette Lake sockeye salmon ESU (70 FR 52630). The sp ecific geographic areas de signated as critical are the Hoh/Quillayute Subbasin, Ozette Lake an d the Ozette Lake watershed, and the Ozette River upstream to endpoints in Big River, Coal Creek, East Branch Umbrella Creek, the North and South Fork of Crooked Creek and several other tributaries. The specif ic primary constituent elements identified for Lake Ozette sockeye sa lmon are areas for spawni ng, freshwater rearing and migration, estuarine areas fr ee of obstruction, nearshore ma rine areas free of obstructions,

and offshore marine areas with good water quality. The physical or biological features that

characterize these sites include water quality and quantity, natura l cover, forage, and adequate passage conditions. Only one watershed supports this ESU, and it is rated as having a high

conservation value. This watershed is essential to the species’ overall conservation by protecting quality growth, reproduction, and feeding.

The Ozette Lake tri butary basin is 77 mi 2 and includes several large tributaries and numerous smaller tributaries. Currently, land ownershi p in the watershed is 73% private land, 15% Olympic National Park, 11% Washington State, and 1% Tribal. Natural disturbance in the watershed was dominated by wind and hydr ogeomorphic events, while contemporary disturbance additionally includes logging, road construction and maintenance, residential and agricultural development, stream channelization and direct and i ndirect stream wood clearance.

These activities alter stream flow patterns and elevate of sediment loads and increased sedimentation within drainage basins. Wood remo val has resulted in less hydraulic roughness, reduced instream water depths, and reduced back water effects on Lake Ozette, which has thus altered the entire hydraulic contro l on Lake Ozette levels and changed the in-river stagedischarge relationship. More recentl y, deposition of sediment origin ating from Coal Creek at the

lake outlet has further altered la ke and river levels (Haggerty et al. 2009b).

Private timber companies own approximately 93% of the four largest trib utary watersheds to Lake Ozette. Logging accelerated over the period of record, with 8.7% of the entire Ozette Lake basin clear-cut by 1953, increasing to 83.6% of the basin area clear-cut by 2003 (Haggerty et al. 2009b). Effects associated with logging depende d on stream size, gradient, and time after harvest. In high-energy coast streams, landslide s and debris torrents of ten modify steep slope

tributaries and the mainstem of creeks. Bank er osion also alters the stream channel on the alluvial flood plain. These effects are additive in the system and reduced the quality of spawning and rearing habitat for j uvenile salmonids (Hartman et al. 1996). Lower gradient streams typically will have an accumulation of sediment. Second-growth logged sections (1 2- 35 yr after logging), re-shaded by deciduous fo rest canopy, have lower biomass of trout and fewer predator taxa than old-growth sites (Murphy and Hall 19 81). Based on the quantity and quality of the physical and biological features, the conservatio n value of the Ozette Lake HUC5 watershed (#1710010102) for sockeye salmon is likely to be “high” (NMFS 2005e).

Species-specific threats and limitations to recovery. Floodplain connectivity and function, channel structure and complex ity, riparian areas and large wood supply, lake beach spawning habitat, and stream substrate have been degrad ed as a result of cumulative impacts of forest practices, agriculture, an d development (Haggerty et al. 2009a, USDC 2009a, Muir and Williams154 2011). Harbor seals and river ot ters, and predaceous non-native and native fish species, are reducing the abundance of adult fish that succe ssfully spawn, and the abundance of sockeye smolts escaping seaward from the watershed each year.

Snake River sockeye salmon

Distribution. Snake River sockeye salmon are unique compared to other sockeye populations: they spawn at a higher elevation (6,500 feet) and have a longer freshwater migration (approximately 900 miles) than any other so ckeye salmon population (Figure 16). Sockeye salmon in this ESU spawn in Redfish Lake in Idaho’s Stanley Basin (Bjornn et al. 1968, Foerster 1968). Stanley Basin sockeye salmon are separa ted by 700 or more river miles from two other extant upper Columbia River populations in the Wenatchee River and Okanogan River drainages. These latter populations return to lakes at substantia lly lower elevations (Wenatchee at 1,870 feet and Okanagon at 912 feet) and o ccupy different ecoregions. The Snake River sockeye salmon ESU includes all anadromous and residual sockeye salmon from the Snake River basin of Idaho, as well as hatchery indivi duals from the Redfish Lake Captive Broodstock Program.

Figure 16. Population boundaries, accessible areas, and extirpa ted reaches of Snake River sockeye distribution, incorporating critical habitat boundaries (dark black line). Taken from NMFS (2011l).

Status and trends. Snake River sockeye salmon were or iginally listed as endangered in 1991

and the NMFS retained that classification when their status was reviewed on June 28, 2005 (70 FR 37160). The only extant sockeye salmon populat ion in the Snake River basin at the time of listing was that in Redfish Lake. Other lakes in the Snake River basin historically supported155 sockeye salmon populations, including Wallowa La ke (Grande Ronde River drainage, Oregon), Payette Lake (Payette River drainage, Ida ho) and Warm Lake (South Fork Salmon River drainage, Idaho (Waples et al. 1997). These populations are now considered extinct. Although kokanee, a resident form of O. nerka , occur in numerous lakes in the Snake River basin, other lakes in the Stanley Basin, and sympatrically with sockeye in Redfish Lake, resident O. nerka were not considered part of the species at th e time of listing (1991). Subsequent to the 1991 listing, a residual form of sockeye residing in Re dfish Lake was identified. The residuals are non-anadromous, completing their entire life cycle in freshwater, but spawn at the same time and in the same location as anadromous sockeye sa lmon ESU. In 1993, the NMFS determined that residual sockeye salmon in Redfish Lake were pa rt of the Snake River sockeye salmon. Also, artificially propagated sockeye salmon from th e Redfish Lake Captive Propagation program are considered part of this sp ecies (70 FR 37160; June 28, 2005).

The NMFS has determined that this artificia lly propagated population is genetically no more than moderately divergent from the natural popu lation (NMFS 2005f). Five lakes in the Stanley Basin historically contained sockeye salmon: A lturas, Pettit, Redfish, Stanley and Yellowbelly (Bjornn et al. 1968). It is generally belie ved that adults were prevented from returning to the Sawtooth Valley from 1910 to 1934 by Sunbeam Da m. Sunbeam Dam was constructed on the Salmon River approximately 20 miles downstrea m of Redfish Lake. Whether Sunbeam Dam was a complete barrier to adult migration rema ins unknown. It has been hypothesized that some passage occurred while the dam was in plac e, allowing the Stanley Basin population or populations to persist (Bjornn et al. 1968, Waples et al. 1991).

Adult returns to Redfish Lake during the period 1954 through 1966 ranged from 11 to 4,361 fish (Bjornn et al. 1968). Sockeye salmon in Alturas Lake we re extirpated in the early 1900s as a result of irrigation diversions, although residual sockeye may still exist in the lake (Chapman and Witty 1993). From 1955 to 1965, the Idaho Departme nt of Fish and Game eradicated sockeye salmon from Pettit, Stanley, and Yellowbelly lake s, and built permanent structures on each of the lake outlets that prevented re-entry of anad romous sockeye salmon (Chapman and Witty 1993). In 1985, 1986, and 1987, 11, 29, and 16 sockeye, respectiv ely, were counted at the Redfish Lake weir (Good et al. 2005b). Only 18 natural origin sockey e salmon have returned to the Stanley Basin since 1987. During the fall of 1990, no fish were observed at Lower Granit Dam or entering the lake and only one fish was observed in each of the two previous years. The first adult returns from the captive br oodstock program returned to th e Stanley Basin in 1999. From 1999 through 2005, a total of 345 captive brood program adults that had migrated to the ocean returned to the Stanley Basin.

Recent annual abundances of natural origin so ckeye salmon in the Stanley Basin have been extremely low. No natural origin anadrom ous adults have returned since 1998 and the abundance of residual sockeye salmon in Redfis h Lake is unknown. This species is entirely supported by adults produced through the capti ve propagation program at the present time. Current smolt-to-adult survival of sockeye origin ating from the Stanley Basin lakes is rarely greater than 0.3% (Hebdon et al. 2004). The status of this ESU is extremely precarious, such that there was unanimous consent among the biol ogical review team members that the species remains in danger of extinction (Good et al. 2005 b). However, the most recent run abundances from 2008 and 2009 have been more robust, with 650 and 809 indivi duals, respectively,156 returning to the Sawtoot h Hatchery (NMFS 2011l).

Critical habitat. Critical habitat for these salmon was designated on December 28, 1993 (58 FR 68543), and encompasses the waters, waterway botto ms, and adjacent riparian zones of specified

lakes and river reaches in the Columbia River th at are or were accessible to listed Snake River salmon (except reaches above impassable natural falls, and Dworshak and Hells Canyon Dams). Adjacent riparian zones are defined as those areas within a horizontal distance of 300 feet from the normal line of high water of a stream channe l or from the shoreline of a standing body of water. Designated critical habitat includes the Columbia River from a straight line connecting the west end of the Clatsop jetty (Oregon side) and the west end of the Peacock jetty (Washington side) and including all river reaches from the estuary upstream to the confluence of the Snake River, and all Snake River reaches upstr eam to the confluence of the Salmon River; all Salmon River reaches to Alturas Lake Creek; St anley, Redfish, Yellow Belly, Pettit, and Alturas Lakes (including their inlet and outlet creeks); Alturas Lake Creek and that portion of Valley Creek between Stanley Lake Creek and the Salmon Rive r. Critical habitat also includes all river lakes and reaches presently or historically accessible to Snake River sockeye salmon. These

habitats are critical for the conservation of th e species because it provides spawning and juvenile rearing habitat, areas for juvenile growth and de velopment, and migration corridors for smolts to the ocean and adults to spawning habitat from th e Pacific Ocean. Limiting factors identified for Snake River sockeye include: reduced tributary stream flow, impaired tributary passage and blocks to migration, and mainstem Colu mbia River hydropower system mortality.

Species-specific threats and limitations to recovery. The key factor limiting recovery of Snake River sockeye salmon ESU is survival outside of the Stanley Basin. Por tions of the migration corridor in the Salmon River are impeded by water quality and temperature (IDEQ 2011). Increased temperatures may reduce the survival of adult sockeye returning to the Stanley Basin. The natural hydrological regime in the upper mainstem Salmon River Basin has been altered by water withdrawals. In most years, sockeye a dult returns to Lower Gr anite suffer catastrophic losses ( e.g. , >50% mortality in one year; (Reed et al. 2003)) before reaching the Stanley Basin, although the factors causing these losses have not been identified. In the Columbia and lower Snake River migration corridor, predation rates on juvenile sockeye salmon are unknown, but

terns and cormorants consume 12% of all salm on smolts reaching the estuary, and piscivorous

fish consume an estimated 8% of mi grating juvenile salmon (NMFS 2011l).

Steelhead Description of the species. Steelhead, the common name of the anadromous form of O. mykiss

, are native to Pacific Coast streams extending from Alaska south to north western Mexico (Moyle 1976a, Stolz and Schnell 1991, NMFS 1997b, Good et al. 2005a, Good et al. 2005b). The life history of this species varies considerably throughout its range. Generally, steelhead occur in two races: the stream-maturing t ype, summer steelhead, enters fres hwater in a sexually immature condition and requires several months in freshw ater to mature and spawn; and the oceanmaturing type, winter steelhead, enters freshw ater with well-developed gonads and spawns shortly after river entry. Variations in migr ation timing exist between populations, and some

river basins have both summer and winter steelhead, while ot hers only have one race.

There is a high degree of overlap in spawning tim ing between populations re gardless of run type157 (Busby et al. 1996d). Difficult field condi tions at that time of year and the remoteness of spawning grounds contribute to the relative lack of specific information on steelhead spawning. Unlike Pacific salmon, steelhead are iteroparous, or capable of spawning more than once before death (Nickelson et al. 1992, Busby et al. 1996d). Second-time spawners often make up about 70-85% of repeat spawners, with third time spawners make up 10-25% of repeats (Stolz and Schnell 1991). Iteroparity is more common among southern steelhead popul ations than northern populations (Busby et al. 1996d).

Habitat. Steelhead occur in marine waters from the surface down to 200 m in waters with temperatures up to 24° C, a lthough 10° C is optimum (Pauley et al. 1986).

Summer steelhead enter freshwater between May and October in the Pacific northwest (Nickelson et al. 1992, Busby et al. 1996d). They require cool, deep holding pools during summer and fall, prior to spawning (Nickelson et al. 1992). Summer steelhead migrate inland toward spawning areas, overwinter in the larger rivers, resume migr ation in early spring to natal streams, and then spawn in January and Fe bruary (Barnhart 1986, Meehan and Bjornn 1991,

Nickelson et al. 1992). Winter steelhead enter freshwater between November and April in the Pacific northwest (Nickelson et al. 1992, Busby et al. 1996d), migrate to spawning areas, and then spawn, generally in April and May (Barnha rt 1986). Some adults, however, do not enter some coastal streams until spring, just before spawning (Meehan and Bjornn 1991).

As with other salmonids, the larg er the fish, the more eggs pro duced. Egg and hatching success

are related to the conditions within the redd a nd time to hatching is temperature dependent. Fertilization to hatching is generally less than a month, after which newly hatched fish will

remain in the redd for another 2-3 weeks. In late spring and following yolk sac absorption, alevins emerge from the gravel and begin activel y feeding. After emerging from the gravel, fry usually inhabit shallow water al ong banks of perennial streams. Fry occupy stream margins (Nickelson et al. 1992). Summer rearing takes place prim arily in the faster parts of pools, although young-of-the-year are abunda nt in glides and riffles. Winter rearing occurs more uniformly at lower densities across a wide range of fast and slow habitat types. Some older juveniles move downstream to rear in larger tributaries and mainstem rivers (Nickelson et al. 1992).

Juvenile steelhead migrate little during thei r first summer and occupy a range of habitats featuring moderate to high water velo city and variable depths (Bisson et al. 1988). Steelhead

hold territories close to the substratum where fl ows are lower and sometimes counter to the main stream; from these, they can make forays up into surface currents to take drifting food (Kalleberg 1958). Juveniles rear in freshwat er from 1 to 4 years, then sm olt and migrate to the ocean in March and April (Barnhart 1986). Winter steelhead juveniles generally smolt after 2 years in freshwater (Busby et al. 1996d). Juvenile steelhead tend to m i grate directly offshore during their first summer from whatever point they enter th e ocean rather than mi grating along the coastal belt as salmon do. Steelhead typically reside in ma rine waters for 2 or 3 years prior to returning to their natal stream to spawn as 4- or 5-year ol ds; fish in the northern portion of the range may spend more time rearing in marine waters (Stolz and Schnell 1991).

Feeding. Juveniles feed primarily on insects (chi ronomids, baetid mayflies, and hydropsychid158 caddisflies (Merz 1994). Adults feed on aquatic and terrestrial insects, mollusks, crustaceans, fish eggs, minnows, and other small fishes (i ncluding greenling and othe r trout; (Chapman and Bjornn 1969, Stolz and Schnell 1991)). Survival at sm oltification is higher for larger fish than smaller ones; this is particularly true for individuals that grew larger earlier in life (Beakes et al. 2010).

Mortality. Steelhead mortality is high early in life a nd decreases with age. For example, Puget Sound steelhead leaving freshwater and estuarin e habitats experience 55- 86% survival to the point of reaching Hood Canal and 0-49% from Hood Canal to the Strait of Juan de Fuca, with survival increasing greatly upon entering the Pacific Ocean (Moore et al. 2010).

Threats. Steelhead are exposed to high rates of natu ral predation each life stage. The highest mortality occurs between the egg stage and smo lt outmigration, and is highest in the first few months following emergence from the redd (Stolz and Schnell 1991). In freshwater, fry fall prey to older steelhead and other trout, as well as birds, sculpin, and various mammals; 10% of salmonid smolts are eaten by Caspian terns a nd double-crested cormorants annually in the Columbia River estuary (NMFS 2011l). In the ocean, marine mammals and other fish prey on steelhead, but the extent of such predation is not well known. Inva sive fishes also threaten the survival and recovery of Pacific salmonids by competing directly for resources, altering food webs and trophic structures, and alteri ng evolutionary trajec tories (NMFS 2011l).

Steelhead have declined under th e combined effects of overharve sts in fisheries; competition

from fish raised in hatcheries and native a nd non-native exotic species; dams that block their migrations and alter river hydrology; gravel mini ng that impedes their mi gration and alters the dynamics (hydrogeomorphology) of the rivers an d streams that support juveniles; water diversions that deplete water leve ls in rivers and streams; destru ction or degradation of riparian habitat that increase water temperatures in rivers and streams sufficient to reduce the survival of juvenile steelhead; and land use practices (logging, agriculture , urbanization) that destroy wetland and riparian ecosystems while introducin g sediment, nutrients, biocides, metals, and other pollutants into surface and ground water and degrade water quality in the fresh water, estuarine, and coastal ecosyste ms throughout the species range.

Critical habitat. All steelhead critical habitat was published on September 2, 2005 (70 FR 52488). Critical habitat has been designated for all DPSs except Puget Sound steelhead. All

steelhead critical habitat includes the same PCEs for the same conservation reasoning:

 Freshwater spawning sites with water a nd substrate quantity to support spawning, incubation, and larval development.

 Freshwater rearing sites with water quanti ty and floodplain connectivity to form and maintain physical habitat conditions and s upport juvenile growth, foraging, behavioral development (e.g., predator avoidance, competition), and mobility. Specific features include forage supporting juvenile developmen t as well as natural cover such as shade, submerged and overhanging large wood, log jams and beaver dams, aquatic vegetation, large rocks and boulders, side channels, and undercut banks.

 Freshwater m igration corridors free of obstruction with wa ter quantity and natural cover such as submerged and overhanging large wood, aquatic vegetation, large rocks and159 boulders, side channels, and undercut banks to support mobility and survival. Without these features, juveniles cannot avoid high fl ows and predators, successfully compete, begin the behavioral and physiological changes needed for marine life, or out-migrate.

 Estuarine areas free of obstruction with wate r quality and salinity conditions supporting juvenile and adult physiological transitions be tween fresh- and saltwater. Estuaries must also include natural cover (submerged a nd overhanging large wood, aquatic vegetation, large rocks and boulders), side channels as we ll as prey for both juveniles and adults.

These features are essential to conservation because withoutthem juveniles cannot reach the ocean in a timely manner and use the variet y of habitats that allow them to avoid predators, compete successfully, and comple te the behavioral and physiological changes

needed for life in the ocean.

 Nearshore marine areas free of obstruction w ith water quality and quantity conditions and forage, including aquatic invertebrates and fishes, supporting growth and maturation; and

natural cover such as subm erged and overhanging large wo od, aquatic vegetation, large rocks and boulders, and side channels. No ar eas are specifically designated as critical habitat, but areas under this category are an important component to Oregon Coast Coho life history.

 Offshore marine areas with water quality conditions and forage, including aquatic invertebrates and fishes, supporting growth a nd maturation. These features are essential for conservation because without them juven iles cannot forage and grow to adulthood.

Central California coast steelhead

Distribution. The Central California Coast steelh ead DPS includes all naturally spawned anadromous steelhead populations below natural and manmade impassable barriers in California streams from the Russian River (inclusive) to Ap tos Creek (inclusive), a nd the drainages of San Francisco, San Pablo, and Suisun Bays eastward to Chipps Island at the confluence of the Sacramento and San Joaquin Rivers (Table 18). Tributary streams to Suisun Marsh including Suisun Creek, Green Valley Creek, and an unnamed tributary to Cordelia Slough (commonly referred to as Red Top Creek), excluding the Sa cramento-San Joaquin River Basin, as well as two artificial propagation programs.

Table 18. Central California Coast steelhead populations and selected measures of population viability.

Basin Historical

abundance Most recent spawner

abundance Hatchery abundance

contributions

Russian River 65,000 (1970) 1,750-7,000 (1994) Unknown

Lagunitas Unknown 400-500 (1990s) Unknown San Gregorio 1,000 (1973) Unknown Unknown

Waddell Creek 481 150 (1994) Unknown

Scott Creek Unknown <100 (1991) Unknown

San Vicente Creek 150 (1982) 50 (1994) Unknown

San Lorenzo River 20,000 <150 (1994) Unknown Soquel Creek 500-800 (1982) <100 (1991) Unknown Aptos Creek 200 (1982) 50-75 (1994) Unknown Total 94,000 2,400-8,125

The DPS is entirely composed of winter run fish. Adults migrate upstream from DecemberApril, and smolts emigrate between March-May (Shapovalov and Taft 1954, Hayes et al. 2008).

At the time of the 1996 status review and 1997 listing, little information was available on the160 specific demographics and life history characterist ics of steelhead in this DPS. While age at smoltification typically ranges from 1 to 4 years, recent studies by Sogard et al. (2009) that growth rates in Soquel Creek likely prevent j uveniles from undergoing smoltification until age 2.

Survival in freshwater reaches tends to be higher in summer and lower from winter through spring for year classes 0 and 1 (Sogard et al. 2009). Larger individuals also survive more readily than do smaller fish within year classes (Sogard et al. 2009). Greater movement of juveniles in freshwater has been observed in winter and spring versus summer and fall time periods, with smaller individuals more likely to move between stream areas (Sogard et al. 2009). Growth rates during this time have rarely been observed to exceed 0.3 mm per day and are highest in winter through spring, potentially due to higher water flow rates and greater food availability (Boughton et al. 2007, Hayes et al. 2008, Sogard et al. 2009).

Status and trends. The Central California Coast steelhead DPS was listed as a threatened species on August 18, 1997 (62 FR 43937) and was reaffirmed on January 5, 2006 (71 FR 834).

Table 19 identifies runs within the Central Calif ornia Coast steelhead DPS and their estimated

run sizes.

Table 19. Central California coast steelhead popu lations and their estimated abundances.

Basin Estimated abundance a Year

Russian River 65,000 1970 1,750-7,000 1994 Lagunitas 500 1994

400-500 1990s San Gregorio 1,000 1973

Waddell Creek 481 1933-1942

250* 1982

150* 1994

Scott Creek 400 1991

<100 1991

300 1994

San Vicente Creek 150* 1982

50* 1994 San Lorenzo River 20,000 Pre 1965 1,614 1977 >3,000* 1978

600 1979 3,000 1982

“few” 1991

<150* 1994 Soquel Creek 500-800* 1982

<100 1991

50-100* 1994 Aptos Creek 200* 1982

<100 1991

50-75* 1994

a A complete list of data sources is available in Good et al. 2005. According to Good et al. the basis for certain estimates is questionable (noted with an asterisk above).

Estimates of historical abundance are provided he re only for background, as the accuracy of the

estimates is unclear. An estimate of historical abundance for the total DPS is provided by CDFG at 94,000 fish. This estimate is based on a partial data set and “best professional judgment” (see161 (Good et al. 2005b) for a discussion). Ot her estimates of historical abundance are on a per river basis. According to Busby et al. (1996a), Shapovalov and Taft (1954) described an average of about 500 adults in Waddell Creek (Santa Cruz County) for the 1930s and early 1940s, whereas Johnson (1964) estimated a run size of 20,000 steelhead in the San Lorenzo River before 1965.

Most of the estimates for run sizes within th e DPS are more recent (Table 26). Two rivers thought to have contained the largest populations w ithin the DPS were the Russian River, and the

San Lorenzo River. Based on run size estimates fr om the 1990s, the Russian River is still likely the largest run within th e DPS, albeit estimates suggest th e population has declined between 90-

96% from the 1970 level.

No current estimates of total population size are ava ilable for this DPS, and consequently there is

no time series data available to evaluate the cen tral California coast stee lhead population trends. Rather, a general dearth of data on adult steelhead within the D PS led the biological review team to examine data collected on juvenile steelhead (Good et al. 2005b). In general, juvenile data is considered a poor indicator of the reproductive po rtion of the population as juvenile age classes exhibit greater mortality rates, which are closely tied to stochastic events, and may move widely within a basin (which may include intermixing with other populations). There is no simple relationship between juvenile and adult numbers (Shea and Mangel 2001). Nonetheless, there was not enough adult data upon which the biological review team could base an assessment of the population trends within the DPS. Therefore, the biological review team log-transformed and normalized juvenile survey da ta from a number of watersheds (presumed populations). As a result, the team derived trend estimates for fi ve populations: the San Lorenzo River, Scott Creek, Waddell Creek, Gazos Creek, and Re dwood Creek in Marin County (see (Good et al. 2005b) for a detailed discussion of the approach). All populations exhibi ted downward trends in abundance. Accordingly, provided th e juvenile data is representative of the true trend, this data suggests that there is an overall downward trend in abundance in the DPS.

In the most recent review of the status of this DPS, most members of th e biological review team (69%) considered this DPS likely to become endangered thus supporting the renewal of the

threatened status for central California coast st eelhead. Notably, 25% of the team voted that the DPS be upgraded to endangered status (voted the DPS as in dange r of extinction)(Good et al. 2005b). Abundance and productivity were of relativ ely high concern (as a c ontributing factor to risk of extinction), and spatial st ructure was also of concern.

Since the original status review, fishing re gulations have changed that probably reduces

extinction risk for Central California Coast stee lhead. Ocean sport harvest is prohibited, and ocean harvest is considered rare. Although freshw ater streams are closed to fishing year round, CDFG has identified certain streams as exceptions where they allow catch-and-release angling or summer trout fishing. In catch-a nd-release streams, all wild steelhead must be released

unharmed.

Critical habitat. Critical habitat was designated for the Central California Coast steelhead DPS on September 2, 2005 (70 FR 52488), and includes areas within the following hydrologic units: Russian River, Bodega, Marin Coastal, San Mate o, Bay Bridge, Santa Clara, San Pablo, and Big Basin. These areas are important for the spec ies’ overall conservati on by protecting quality growth, reproduction, and feeding hab itat. The critical habitat desi gnation for this ESU identifies162 primary constituent elements that include sites necessary to support one or more steelhead life stages. Specific sites include freshwater spawni ng sites, freshwater rear ing sites, freshwater migration corridors, nearshore marine habitat and estuarine areas. The physical or biological

features that characterize thes e sites include water quality and quantity, natural cover, forage, adequate passage conditions, and floodplain connectivity. The critical ha bitat designation (70 FR 52488) contains additional deta ils on the sub-areas that are included as part of this designation, and the areas that were excluded from designation.

In total, Central California Coast steelhead occupy 46 watersheds (freshwater and estuarine). The total area of habitat designa ted as critical includes about 1,500 miles of stream habitat and about 400 square miles of estuar ine habitat (principally Humboldt Bay). This designation

includes the stream channels with in the designated stream reaches, and includes a lateral extent as defined by the ordinary high water line. In areas where the ordinary high-water line is not

defined the lateral extent is defined as the bank full elevation. In estuarine areas the lateral extent is defined by the extreme high water be cause extreme high tide areas encompass those areas typically inundated by wate r and regularly occupied by j uvenile salmon during the spring and summer, when they are migrating in the ne arshore zone and relyi ng on cover and refuge qualities provided by these habitats, and while they are foraging. Of the 46 occupied watersheds

reviewed in NMFS' assessment of critical habita t, 14 watersheds received a low rating of conservation value, 13 received a medium rating, and 19 received a high rating of conservation value for the species.

California Central Valley steelhead

Distribution. California Central Valley steelhead occupy the Sacramento and San Joaquin Rivers and their tributaries, although they were once widespread throughout the Central Valley (Busby et al. 1996a, Zimmerman et al. 2008). Steelhead were found from the upper Sacramento and Pit River systems (now inaccessible due to Shasta and Keswick Dams), south to the Kings and possibly the Kern River systems (now inacce ssible due to extensive alteration from water diversion projects), and in both east- and west-side Sacramento River tributaries (Yoshiyama et
al. 1996). The present distribution has been grea tly reduced (McEwan and Jackson 1996a). The CACSS (1988) reported a reduction of steelhead habitat from 6,000 miles historically to 300 miles today. Historically, steelhead probably as cended Clear Creek past the French Gulch area, but access to the upper basin was blocked by Whiskeytown Dam in 1964 (Yoshiyama et al. 1996). Steelhead also occurred in the upper drai nages of the Feather, American, Yuba, and

Stanislaus Rivers which are now inacce ssible (McEwan and Jackson 1996a, Yoshiyama et al. 1996).

Existing wild steelhead populations in the Centra l Valley are mostly confined to the upper Sacramento River and its tributaries, including Antelope, Deer, and Mill Creeks and the Yuba River. Populations may exist in Big Chico a nd Butte Creeks and a few wild steelhead are produced in the American and Feather Rivers (McEwan and Jackson 1996a). Recent snorkel surveys (1999 to 2002) indica te that steelhead are pr esent in Clear Creek (Good et al. 2005a). Because of the large resident O. mykiss population in Clear Creek, st eelhead spawner abundance has not been estimated. Until recently, steelhead were thought to be extirpated from the San Joaquin River system. Recent monitoring has de tected small self-sustaining populations of steelhead in the Stanislaus, Moke lumne, Calaveras, and other str eam s previously thought to be163 devoid of steelhead (McEwan 2001). On the Stan islaus River, steelhead smolts have been captured in rotary screw traps at Caswell Stat e Park and Oakdale each year since 1995 (Demko and Cramer 2000). It is possibl e that naturally spaw ning populations exist in many other streams but are undetected due to lack of monitoring programs (IEPSPWT 1999). Coleman and Feather

River hatcheries were included in the DPS in 2006 (NMFS 2011c).

The Sacramento and San Joaquin Rivers offer th e only migration route to the drainages of the

Sierra Nevada and southern Cascade mountai n ranges for anadromous fish. The CDFG considers all steelhead in the Central Valley as winter steelhead, although “t hree distinct runs,” including summer steelhead, may have occurred th ere as recently as 1947 (McEwan and Jackson 1996a)(CDFG 1995). Steelhead in thes e basins travel extensive di stances in freshwater (some

exceed 300 km to their natal str eams), making these the longest freshwater migrations of any population of winter steelhead. The upper Sacr amento River essentially receives a single continuous run of steelhead from July through May, with peaks in September and February. Spawning begins in late December and can ex tend into April (McEwan and Jackson 1996b).

Status and trends. The NMFS originally listed California Central Valley steelhead as threatened in 1998; this status was reviewed and retained on January 5, 2006 (71 FR 834).

Historic run size is difficult to estimate given the paucity of da ta, but may have approached one to two million adults annually (McEwan 2001). By the early 1960s, the steelhead run size had declined to about 40,000 adults (McEwan 2001). Ov er the past 30 years, the naturally spawned steelhead populations in the uppe r Sacramento River have dec lined substantially. Hallock et al. (1961) estimated an average of 20,540 adult st eelhead occurred in the Sacramento River (upstream of the Feather River). Steelhead counts at Red Bluff Diversion Dam declined from an average of 11,187 for the period of 1967 to 1977, to an average of approximately 2,000 through the early 1990s, with an estimated total annual run size for the entire Sacramento-San Joaquin system at no more than 10,000 adults (based on Red Bluff Diversion Da m counts; (McEwan and Jackson 1996a, McEwan 2001). The five-year geometric mean, however, is just under 2,000 steelhead (Table 20), and the l ong-term trend suggests that th e population is declining.

Table 20. California Central Valley steel head and their long-term trend

. Population 5-year mean (min – max) a λ Long-term trend b

Sacramento River 1,952 (1,425-12,320) 0.95 (0.90,1.02) -0.09 (-0.13,-0.06)

a Refers to the period ending in 1993, when steelhead counts at Red Bluff Diversion dam ended. Data reported in

Good et al. 2005.

b 90% confidence limits in parentheses.

The only consistent data availa ble on steelhead numbers in th e San Joaquin River basin come from CDFG mid-water trawling samples collected on the lower San Joaquin River at Mossdale.

These data indicate a decline in steelhead numbers in the earl y 1990s, which have remained low through 2002 (Good et al. 2005a). In 2004, a total of 12 steelhead smolts were collected at Mossdale (Good et al. 2005a).

Based upon information since 2005, Central Vall ey steelhead populati ons have generally declined, with hatchery fish forming a larg er portion of populations (NMFS 2011c). Battle Creek surveys have documented relatively stable counts since 2005 after a precipitous drop for three years prior to that. Returns to Coleman ha tchery have varied widely from year to year,164 with no clear trend. An average of 151 redds have been found in Clear Creek from 2001 to 2010, with similar numbners (154) in American Rive r over nearly the same time frame. A very small number (40-70) of redds have been c ounted over the past three years in the Lower Mokelumne River, but the trend is on the rise. Returns to the Feather River Hatchery have plummeted from 679 to 86 fish from 2008 to 2010.

Reynolds et al. (1993) reported that 95% of salmonid ha bitat in California’s Central Valley has been lost, largely due to mining and water develo pment activities. They also noted that declines are “due mostly to water development, inadequate instream flows, rapid flow fluctuations, high summer water temperatures in streams immediat ely below reservoirs, diversion dams which block access, and entrainment of juveniles into un screened or poorly screen ed diversions.” Thus, overall habitat problems in this ESU relate primarily to water development resulting in inadequate flows, flow fluctuations, blockages, and entrainment into diversions (McEwan and Jackson 1996a). Other problems related to land use practices (agricultu re and forestry) and

urbanization have also contribute d to population declines. It is unclear how harvest has affected California’s Central Valley stee lhead, although it is lik ely a continuing thr eat. A CDFG creel census in 2000 indicated that most fish are caugh t and released, but due to the size of the catch and release fishery (more than 14,000 steelhead were caught and releas ed according to the survey) even a small amount of mo rtality in this fishery could cau se declines in the populations.

Critical habitat. The NMFS designated critical habitat fo r California Central Valley steelhead on September 2, 2005 (70 FR 52488). Specific geographic areas designated include the following CALWATER hydrological units: Teha ma, Whitmore, Redding, Eastern Tehama, Sacramento Delta, Valley-Putach-Cache, American River, Marysville, Yuba, Valley American, Colusa Basin, Butte Creek, Ball Mountain, Shata Bally, North Valley Floor, Upper Calaveras, Stanislaus River, San Joaquin Valley, Delta-M endota Canal, North Diablo Range, and the San Joaquin Delta. These areas are important fo r the species’ overall conservation by protecting quality growth, reproduction, and feeding habitat. The critical habitat designation for this ESU identifies primary constituent elements that in clude sites necessary to support one or more steelhead life stages. Specific sites include fres hwater spawning sites, freshwater rearing sites, freshwater migration corridors, ne arshore marine habitat and estuarine areas. The physical or biological features that characte rize these sites include water quali ty and quantity, natural cover, forage, adequate passage conditions, and fl oodplain connectivity. The critical habitat designation (70 FR 52488) contains additional details on the sub-areas that are included as part of this designation, and the areas th at were excluded from designation.

In total, California Central Valley steelhead oc cupy 67 watersheds (fresh water and estuarine). The total area of habitat designa ted as critical includes about 2,300 miles of stream habitat and about 250 square miles of estuarin e habitat in the San Francisco- San Pablo-Suisan Bay estuarine complex. This designation includes the stream ch annels within the designated stream reaches, and includes a lateral extent as defined by the ordinary high water line. In areas where the ordinary high-water line is not de fined the lateral extent is define d as the bank full elevation. In

estuarine areas the lateral extent is defined by the extreme high water because extreme high tide areas encompass those areas typically inundated by water and regularly occupied by juvenile salmon during the spring and summer, when they ar e migrating in the near shore zone and relying on cover and refuge qualities provided by these hab itats, and while they are foraging. Of the 67165 watersheds reviewed in the NMFS' assessment of critical habitat, seven watersheds received a low rating of conservation value, three received a medium rating, and 27 received a high rating of conservation value for the species.

Lower Columbia River steelhead

Distribution. Lower Columbia River steelhead include naturally produced steelhead returning to Columbia River tributaries on the Washington side between the Cowlit z and Wind rivers in Washington and on the Oregon side between the Willamette and Hood rivers, inclusive (Figure 17). In the Willamette River, the upstream boundary of this species is at Willamette Falls. This species includes both winter and summer steelhead. Two hatchery populations are included in this species. Table 21 identifie s the populations that comprise Lower Columbia River steelhead (23 historically) and summarizes several measur es available to charac terize population viability.

Figure 17. Population boundaries, dams, accessible areas, an d extirpated reaches of lower Columbia River steelhead distribution, incorporating critical habitat boundaries (dark black line) and incorporating summer versus winter runs. Taken from NMFS (2011).

Table 21. Lower Columbia River steelhead populations and select measures of population viability.

Life history Population Historical

abundance a Mean number of spawners Percent hatchery

contribution Median shortterm growth

rate ( λ ) b

Winter Cispus River

Tilton River 2,787 c 73 Upper Cowlitz River Lower Cowlitz River 1,672166 Life history Population Historical

abundance a Mean number of spawners Percent hatchery

contribution Median shortterm growth

rate ( λ ) b

Coweeman River 2,243 466 d 50 0.920, 0.787 South Fork Toutle River 2,627 504 d 2 0.933, 0.929 North Fork Toutle River 3,770 196 d 0 1.038, 1.038

Kalama River 554 726 d 0 0.984, 0.922

North Fork Lewis River 713 East Fork Lewis River 3,131

Salmon Creek Washougal River 2,497 323 d 0

Clackamas River 560 e 41 0.875, 0.830 Sandy River 977e 42 0.866, 0.782 Lower Columbia Gorge

tributaries 793

Upper Columbia Gorge

tributaries 243

Hood River 756 f 52 Summer Wind River 2,288 472 g 5 0.995, 0.903

Hood River 931 f 83 Unknown Washougal River 1,419 264 g 8 1.029, 0.960

East Fork Lewis River 422 434 g 25

North Fork Lewis River Kalama River 3,165 474 g 32 0.900, 0.664

a All data reported by Good et al. 2005. Estimate of historical abundance derived through EDT model associated with large uncertainty. Model also incorporates presently available habitat that was not historically available and

vice versa.

b λ

calculation assumed either hatchery fish fail to repr oduce or reproduce at the rate of wild individuals, respectively.

c Data from 2002.

d Data from 1998-2002.

e Data from 1997-2001.

f Data from 1996-2000.

g Data from 1999-2003.

Summer steelhead return sexuall y immature to the Columbia River from May to November, and spend several months in freshwater prior to sp awning. Winter steelhead enter freshwater from November to April, are close to sexual matu ration during freshwater entry, and spawn shortly

after arrival in their natal streams. Wher e both races spawn in the same stream, summer steelhead tend to spawn at higher el evations than the winter forms.

Status and trends. The NMFS listed Lower Columbia River steelhead as threatened on March 19, 1998 (63 FR 13347), and reaffirmed their stat us as threatened on January 5, 2006 (71 FR 834). The 1998 status review noted that this ESU is characterized by populations at low

abundance relative to historical levels, significant population declines since the mid-1980s, and widespread occurrence of hatchery fish in natu rally spawning steelhead populations. During this

review the NMFS was unable to id entify any natural populations that would be considered at low

risk.

All populations declined between 1980 and 2000, w ith sharp declines beginning in 1995. Those with adequate data for modeling are estim ated to have a high extinction risk (Good et al. 2005b). Abundance trends are generally ne gative, showing that most populat ions are in decline, although167 some populations, particularly summer run, have show n higher return in the last 2 to 3 years.

Historical counts in some of the larger tributaries (Cowlitz, Kalama, and Sandy Rivers) suggest the population probably exceeds 20,000 fish while in the 1990s fish abundance dropped to 1,000 to 2,000. Recent abundance estimates of natural- origin spawners range from completely extirpated for some populations above impassabl e barriers to over 700 for the Kalama and Sandy winter-run populations. A number of the populations have a substantial fraction of hatcheryorigin spawners in spawning areas, and are hypot hesized to be sustained largely by hatchery production. Exceptions are the Kalama, the Toutle, and East Fork Lewis wi nter-run populations. These populations have relatively low recent m ean abundance estimates with the largest being the Kalama (geometric mean of 728 spawne rs). Since 2000, abundance for most populations increased, peaking in 2004 before returning to lower levels near the historical mean (NMFS 2011d). No sustained improvement in abundance has been observed since this time (NMFS 2011d). However, hatchery releases have grea tly increased from two million to three million fish annually (NMFS 2011d). Currently, 16 of 23 populations are considered at high risk of

extinction.

Critical habitat. The NMFS designated critical habitat for Lower Columbia River steelhead on September 2, 2005 (70 FR 52630). Designated critical habitat includes the following subbasins: Middle Columbia/Hood subbasin, Lower Columb ia/Sandy subbasin, Lewis subbasin, Lower Columbia/Clatskanie subbasin, Upper Cowlitz s ubbasin, Cowlitz subbasin, Clackamas subbasin, Lower Willamette subbasin, and the Lower Columbia River corridor. These areas are important for the species’ overall conservation by protec ting quality growth, reproduction, and feeding habitat. The critical habitat designation for this DPS identifies primary constituent elements that include sites necessary to support one or more steelhead life stages. Specific sites include freshwater spawning sites, freshwater rearing si tes, freshwater migra tion corridors, nearshore marine habitat and estuarine areas. The physical or biological features that characterize these sites include water quality and qua ntity, natural cover, forage, adequate passage conditions, and floodplain connectivity. The cri tical habitat designation (70 FR 52630) contains additional description of the watersheds th at are included as part of this designation, and any areas specifically excluded fr om the designation.

In total, Lower Columbia River steelhead occu py 32 watersheds. The total area of habitat designated as critical includes a bout 2,340 miles of stream habitat. This designation includes the stream channels within the designated stream re aches, and a lateral extent as defined by the ordinary high water line. In areas where the or dinary high-water line is not defined, the lateral

extent is defined as the bank full elevation. Of the 32 watersheds reviewed in NMFS' assessment of critical habitat, two waters heds received a low rating of c onservation value, 11 received a medium rating, and 26 received a high rating of c onservation value for the species. Limiting

factors identified for Lower Columbia River st eelhead include: degrad ed floodplain and steam channel structure and function, reduced access to sp awning/rearing habitat, altered stream flow in tributaries, excessive sediment and elevated water temperatures in tr ibutaries, and hatchery impacts.

Species-specific threats and limitations to recovery. Several direct and ha bitat-related factors inhibit the recovery of Lower Columbia River Coho salmon (LCFRB 2010, NMFS 2011d). Habitat impacts include degraded estuarine a nd near-shore marine habitat resulting from168 cumulative impacts of land use and flow mana gement by the Columbia River hydropower system, fish passage barriers that limit access to spawning a nd rearing habitats, floodplain connectivity and function, channe l structure and complexity, ri parian areas and large wood supply, stream substrate, stream flow, and water quality have been degraded as a result of cumulative impacts of agriculture, forestry, and development. Additiona l threats to habitat include an altered flow regime and Columbia River plume has altered the temperature regime and estuarine food web, and has reduced ocean productivity, reduced access to off-channel rearing habitat in the lower Columbia River, as well as reduced productivity resulting from

sediment and nutrient-related ch anges in the estuary. Other imp acts also include hatchery- and harvest-related effects, juvenile fish strandings that result from ship wakes, avian and marine mammal predation in the lower mainstem Colu mbia River and estuary, and contaminants affecting fish health and reproduction.

Middle Columbia River steelhead

Distribution. The Middle Columbia River steelhead DPS includes all naturally spawned anadromous steelhead populations below natura l and manmade impassible barriers in Oregon and Washington drainages upstream of the Hood and Wind River systems, up to and including the Yakima River (61 FR 41541)(Figure 18). Steelhead from the Snake River Basin are excluded from this DPS. Seven artificial propa gation program are part of this DPS. These artificially propagated populations are considered no more divergen t relative to th e local natural populations than would be expected between closely related natural populations within the DPS.

Figure 18. Population boundaries, dams, accessible areas, an d extirpated reaches of lower Columbia River steelhead distribution, incorporating critical habitat boundaries (dark black line). Taken from NMFS (2011e).

Middle Columbia River steelhead occupy the in termontane region of the Pacific northwest,169 which includes some of the driest areas in th e region generally receiv ing less than 40 cm of rainfall annually. Major drainages in this ESU are the Deschutes, John Day, Umatilla, Walla Walla, Yakima, and Klickitat river systems. The area is generally ch aracterized by its dry climate and harsh temperature extremes. Almost all steelhead populations within this DPS are summer-run fish; the only exceptions are the only populations of inland winter steelhead, which

occur in the Klickitat River and Fifteenmile Creek (Busby et al. 1996d). According to Interior Columbia Basin Technical Recovery Team (ICTR T 2003a) this DPS is comprised of 16 putative populations in four major popula tion groups (Cascades Eastern Sl opes Tributaries, John Day River, Walla Walla and Umatilla Rivers, and Yakima River) and one unaffiliated independent population (Rock Creek). Table 22 lis ts extant (putative) populati ons that compose this DPS. There are two extinct populations in the Ca scades Eastern Slope major population group, the

White Salmon River and Deschutes Crooked Rive r above the Pelton/Round Butte Dam complex. Present population structure is delineated largely on the ba sis of geographical proximity, topography, distance, ecological similarities or differences. Additional genetic studies are needed to describe the DPS substructure, as we ll as the fine-scale ge netic structure of the populations within a particular basin (e.g., John Day River).

Table 22. Middle Columbia River steelhead populations and select measures of population viability.
Population a Major population

groups Mean number of spawners (range) b Percent hatchery

contribution c Long-term growth rate ( λ ) d

Klickitat River Cascade Eastern Slope 155 redds (97-261) Fifteenmile Creek Cascade Eastern Slope 2.87 rpm (1.3-6.0) 0 1.129

Deschutes River -

eastside Cascade Eastern Slope 13,455 (10,026- 21,457) 72 1.022,0.840,

0.942

Descutes River –

westside Cascade Eastern Slope

John Day lower

mainstem tributaries John Day River 1.4 rpm (0-5.4) 1.013

North Fork John Day John Day River Upper NF - 2.57 rpm (1.6-5.0) e 1.011

Lower NF - 3.52 rpm (1.5-8.8) 1.174

Middle Fork John Day John Day River 3.70 rpm (1.7-6.2) 0.966 South Fork John Day John Day River 2.52 rpm (0.9-8.2) 0.967 John Day upper

mainstem John Day River 2,122 (926-4,168) 4 0.975, 0.966

Rock Creek Unaffiliated Area Umatilla River Walla Walla & Umatilla 2,486 (1,480-5,157) 40 1.007, 0.969

Walla Walla Walla Walla & Umatilla Touchet River Walla Walla & Umatilla 345 (273-527) f 16 0.961, 0.939 Toppenish & Satus

Creek Yakima River

Naches River Yakima River Yakima River upper

mainstem Yakima River 1,801 (1,058-4,061) 3 1.009

a Population groups defined by the ICBTRT (2003).

b Values represent the 5-year geometric mean in spawners, re dds, or redds per mile (RPM). Values calculated from data series using years 1997-2001 or 1998-2001. See Good et al. (2005) for details.

c Hatchery production in the recent past and at present cons ists of locally-derived broodstock, although straying of production fish into the Deschutes River has been a persistent problem. Data from Good et al. 2005.170 d Multiple estimates for long-term growth ( λ ) presented for some populations representing two different assumptions on the contribution of hatchery fish to the natural production. Where two or more values are presented, the first value reflects the assumption that hatchery fish do not contribute to natural production, and the second value reflects the assumption that hatchery contribute to natural production at the same rate as natural-origin spawners. Deschutes River values are reflective of total population, not eastside only. The λ value is calculated from data (1980-1999) from Warm Springs area. Data series upon which va lues are calculated varies across basins. See Good et al. (2005) for details on the length and time of data series available by population.

Most Middle Columbia River steelhead smolt at 2 years of age and spend 1 to 2 years at sea prior to re-entering natal river systems. They may remain in such rivers for up to a year prior to spawning (Howell et al. 1985b). Within this ESU, the Klick itat River is unusual, as it produces

both summer and winter steelhead. The summer steelhead are dominated by year-class-two ocean steelhead, whereas most other rivers in this region produce about equal numbers of both age-one and age-two ocean steelhead. Factor s contributing to dec line include hydropower development and agriculture; these land uses impede or prevent migrations, alter water availability, and alter water chemistry and temperatures.

Status and trends. Middle Columbia River steelhead were listed as threatened in 1999 (64 FR 14517), and their status was r eaffirmed on January 5, 2006 (71 FR 834). The precise pre-1960 abundance of this species is unknown. Based upon the Washington Department of Fish and

Wildlife’s estimates of the historic run size for the Yakima River at 100,000 steelhead, Busby et
al. (1996d) surmised that total DPS abundance likely exceeded 300,000 returning adults. By 1993, the estimated 5-year average size (ending in 1993) of the Middle Columbia steelhead DPS was 142,000 fish (Busby et al. 1996d). Survey data collecte d between 1997 and 2001 indicates that several populations within the DPS have increased since th e last status review (Good et al. 2005b). However, the long-term annual population growth rate ( λ ) is less than one for most populations (see Table 29).

In contrast, short term trends in major areas were positive for 7 of the 12 areas with available data (Good et al. 2005b). Spawner numbers in the Yaki ma River, the Deschutes River and sections of the John Day River system were subs tantially higher compared to numbers surveyed between 1992 and 1997 (Good et al. 2005b). Similarly, spawner numbe rs substantially increased

in the Umatilla River and Fifteenmile Creek rela tive to annual levels in the early 1990s. Nonetheless, most populations remain below inte rim target levels. For instance, the Yakima

River returns are still substant ially below interim target levels of 8,900 (the current 5-year average is 1,747 fish) and estimated historical return levels. In fact, the majority of spawning occurs in only one tributary, Satus Creek (Berg 2001). Based on recent 5-year geometric means, only the Deschutes River exceeded interim target levels (Good et al. 2005b). While increases in short-term trends could suggest improvements with in the DPS, given that the average population growth rate across all streams is negative (0.98 as suming hatchery spawners do not contribute to production, and 0.97 assuming that both hatchery and natural-origin fish contribute equally) and evidence of large fluctuation in marine survival for the species, recent increases in population sizes must be viewed cautiously.

Recent data indicate varied trends among the population groups (NMFS 2011e). Monitored populations with the Cassades Eastern Slope Trib utaries group show declining trends since 2005, although natural origin spawners are composing a higher proportion of total retuns. On the contrary, John Day River populations seem to be improving. Populations in the Yakima River171 group also improved, although abundance remains very low. Two populations within the Umatilla/Walla Walla Rivers group improved, while one remained stable.

Critical habitat. The NMFS designated critical habitat for Middle Columbia River steelhead on September 2, 2005 (70 FR 52630). Designated critical habitat includes the following subbasins: Upper Yakima, Naches, Lower Yakima, Middle Columbia/Lake Wallula, Walla Walla, Umatilla, Middle Columbia/Hood, Klickitat, Upper John Day, North Fork John Day, Middle Fork John Day, Lower John Day, Lower Deschutes, Trout, and the Upper Columbia/Priest Rapids subbasins, and the Columbia River corridor. Th ese areas are important for the species’ overall conservation by protecting quality growth, reproduction, and feeding habitat. The critical habitat designation for this DPS identifies primary consti tuent elements that include sites necessary to support one or more steelhead life stages. Sp ecific sites include freshwater spawning sites, freshwater rearing sites, freshw ater migration corridors, nearshor e marine habitat and estuarine areas. The physical or biological features thatcharacterize these sites include water quality and quantity, natural cover, forage, adequate pass age conditions, and floodplain connectivity. The final rule (70 FR 52630) lists the watersheds th at comprise the designated subbasins and any areas that are specifically ex cluded from the designation.

In total, there are 114 watersheds within the range of Middle Columbia River steelhead. The total area of habitat designated as critical includes about 5,800 m iles of stream habitat. This designation includes the stream ch annels within the designated st ream reaches, and includes a lateral extent as defined by the ordinary high water line. In ar eas where the ordinary high-water

line is not defined the lateral exte nt is defined as the bank full elevation. Of the 114 watersheds

reviewed in NMFS' assessment of critical habi tat for Middle Columbia River steelhead, nine watersheds received a low rating of conserva tion value, 24 received a medium rating, and 81 received a high rating of conservation value for the species. Although pris tine habitat conditions are still present in some wilderness, roadless, and undeveloped areas, habitat complexity has been greatly reduced in many ar eas of designated critical hab itat for Middle Columbia River steelhead. Limiting factors iden tified for Middle Columbia Rive r steelhead include: hydropower system mortality, reduced stream flow, impaired passage, excessive sedi ment, degraded water quality, and altered channe l morphology and floodplain.

Species-specific threats and limitations to recovery. Floodplain connectivity and function, channel structure and complexity, riparian areas, fish passage, stream subs trate, stream flow, and water quality have been degraded as a result of cumulative impacts of agriculture, forestry, tributary hydro system activities, and development impair recovery of Middle Columbia River steelhead (NMFS 2009b, 2011e). Additional problems for recovery are posed by hatchery- and harvest-related effects, predation, competition, disease, degraded estuarine and nearshore habitats, and the influence of mainstemColumbia River hydropower facilities.

Upper Columbia River steelhead

Distribution. Upper Columbia River steelhead occupy the Columbia River Basin upstream from the Yakima River, Washington to the border between the US and Canada (Figure 19). This area includes the Wenatchee, Entiat, and Okanoga n Rivers (Table 23). All Upper Columbia

River steelhead are summer steelhe ad. Steelhead primarily use str eams of this region that drain the northern Cascade Mountains of Washington State. This species includes hatchery172 populations of summer steelhead fr om the Wells Hatchery because it probably reta ins the genetic resources of steelhead populat ions that once occurred above the Grand Coulee Dam. This species does not include the Skamania Hatchery stock because of its non-native genetic heritage.

Figure 19. Population boundaries, dams, accessible areas, an d extirpated reaches of upper Columbia River steelhead distribution, incorporating critical habitat boundaries (dark black line). Taken from NMFS (NMFS 2011i).

Columbia River steelhead life histories are complex. Adults return in late summer and early fall, with most migrating relatively qui ckly to their natal tributaries. A portion of the returning run overwinters in mainstem reservoirs, passing ove r upper- to mid-Columbia dams in April and May of the following year. Spawning occurs in th e late spring of the year following river entry. Juvenile steelhead spend one to se ven years rearing in freshwater be fore migrating to sea. Smolt outmigrations are predominantly year class two and three (juveniles). Most adult steelhead

return after from sea af ter one or two years.

Table 23. Upper Columbia River steelhead salmon populations, abundances, and hatchery contributions (Good et
al. 2005b).

Population Historical

abundance Most recent spawner

abundance Hatchery abundance

contributions

Wenatchee/Entiat rivers Unknown 1,899-8,036 71% Methow/Okanogan rivers Unknown 1,879-12,801 91%

Total Unknown 3,778-20,837

Status and trends. Upper Columbia River steelhead were or iginally listed as endangered in 1997 (62 FR 43937), after their status was reviewe d, they were reclassified to threatened on January 5, 2006 and then reinstated to endangered status per US District Court decision in June 2007 (62 FR 43937).173 Abundance estimates of returning naturally produced U pper Columbia River steelhead have been based on extrapolations from mainstem dam c ounts and associated sampling information (e.g., hatchery/wild fraction, age composition). Return s of both hatchery and naturally produced steelhead to the upper Columbia River have incr eased in recent years. The average 1997-2001 return through the Priest Rapids fish ladder was approximately 12,900 fish. The average for the previous five years (1992 to 1996) was 7,800 fis h. Abundance estimates of returning naturally produced Upper Columbia River steelhead have been based on extrapolations from mainstem dam counts and associated sampling information (e.g., hatchery/wild fraction, age composition). The natural component of the a nnual steelhead run over Priest Ra pids Dam increased from an average of 1,040 (1992-1996), repr esenting about 10% of the to tal adult count, to 2,200 (1997- 2001), representing about 17% of the adult coun t during this period of time (ICBTRT 2003). Since 2005, abundance estimates have increased in all four populations wi thin the ESU, although a large portion of returns are hatc hery origin fish (NMFS 2011i).

In terms of natural production, recent population abundances for th e Wenatchee, Entiat, and the Methow population remain well below the minimu m abundance thresholds developed for these populations (ICBTRT 2005). A fi ve-year geometric mean (1997 to 2001) of approximately 900 naturally produced steelhead returned to the We natchee and Entiat rivers (combined). Although

this is well below the minimum abundance thre sholds, it represents an improvement (an increasing trend of 3.4% per year). However, th e average percentage of natural fish for the recent five-year period dropped from 35% to 29%, co mpared to the previous status review. For the Methow population, the five-year geometric m ean of natural returns over Wells Dam was 358. Although this is well below the minimum ab undance threshold, it is increasing trend of 5.9% per year. In addition, the 2001 return (1,38 0 naturally produced spawners) was the highest single annual return in th e 25-year data series. However, the average percentage of wild origin spawners dropped from 19% for th e period prior to the 1998 status review to 9% for the 1997 to 2001 returns. This DPS is failing to meet viabil ity criteria in all four categories; productivity, abundance, spatial structure, and genetic diversity. Overall, l ong-term population growth is in decline. Returns are predominantly hatchery origin and all Upper Columbia River steelhead populations have reduced genetic diversity from homogenization of populations that occurred during the Grand Coulee Fish Maintenance project from 1939-1943, 1960, and 1981 (Chapman

et al. 1994). The ICBTRT has characterized the spa tial structure risk as “low” for the Wenatchee and Methow, “moderate” for the Entiat, and “high” for the Okanogan.

Critical habitat. Critical habitat was designated fo r this ESU on September 2, 2005 (70 FR 52488). It includes all Columbia River estuarine areas and river reaches upstream to Chief Joseph Dam and several tributary subbasins. Th e critical habitat desi gnation for this DPS identifies PCEs that include sites necessary to support one or more life st ages of steelhead. The DPS has 42 watersheds within its range. Three watersheds received a low rating, eight received a medium rating, and 31 rated a high conservation va lue to the DPS. In addition, the Columbia River rearing/migration corrido r downstream of the spawning range was rated as a high conservation value. Limiting factors include: (1) Mainstem Columbia River hydropower system mortality, (2) reduced tributary st reamflow, (3) tributary riparian degradation and loss of in-river wood, (4) altered tributary floodplain and channe l m orphology, and (5) excessive fine sediment and degraded tributary water quality.174 Species-specific threats and limitations to recovery. The recovery of Upper Columbia River steelhead is impaired by a variety of anthropoge nic factors, including im paired tributary fish passage, hatchery and harvest effects, comple tion, predation, disease, as well as fisheries management, including past introductions and persis tence of non-native (exo tic) fish species that

continue to affect habitat c onditions for listed species (Upper Columbia Salmon Recovery Board 2007, NMFS 2011i). Also significant is floodplain connectivity and function, channel structure and complexity, riparian areas and large woody debris recruitment, stream flow, and water quality have been degraded as a result of cu mulative impacts of agriculture, forestry, and development.

Puget Sound steelhead

Distribution. Puget Sound steelhead occupy river basins of the Strait of Juan de Fuca, Puget Sound, and Hood Canal, Washington (Figure 20). Included are rive r basins as far west as the

Elwha River and as far north as the Nooksack River. Puget Sound's fj ord-like structure may affect steelhead migration patterns; for example, some populations of Coho and Chinook salmon, at least historically, remained w ithin Puget Sound and did not migrat e to the Pacific Ocean itself. Even when Puget Sound steelhead migrate to the hi gh seas, they may spend considerable time as juveniles or adults in the pr otected marine environment of Puget Sound, a feature not readily

accessible to steelhead from other areas of the Pa cific northwest. This species is primarily composed of winter steelhead but includes se veral stocks of summer steelhead, usually in subbasins of large river systems and above seasonal hydrologic barriers. Li fe history attributes of Puget Sound steelhead (migration and spawn ti ming, smolt age, ocean age, and total age at first spawning) appear similar to th ose of other west coast steelhead.

Figure 20. Population boundaries, dams, accessible areas, an d extirpated reaches of Puget Sound steelhead175 distribution, incorporating critical habitat boundaries (dark black line). Taken from NMFS (NMFS 2011f).

Status and trends. The DPS was listed as a threatened species on May 11, 2007 (72 FR 26722). Run size was calculated in the early 1980s at about 100,000 winter-run fish and 20,000 summerrun fish. It is not clear what portion were hatchery fish, but a combined estimate with coastal steelhead suggested that roughly 70% of steelhead in ocean runs were of hatchery origin. The percentage in escapement to spawning grounds would be substantially lower due to differential harvest and hatchery rack returns. By the 1990s, total run size for four major stocks exceeded 45,000; roughly half of which was natural escapement.

Nehlsen et al. (1991) identified nine Puget Sound steelhead stocks at some degree of risk or concern, while the WDFW et al. (1993) estimated that 31 of 53 st ocks were of native origin and predominantly natural production. Their assessmen t of the status of these 31 stocks was 11 healthy, three depressed, one cr itical, and 16 of unknown status. Their assessment of the status of the remaining (not native/n atural) stocks was three health y, 11 depressed, and eight of

unknown status.

Of the 21 populations in the Puget Sound ESU reviewed by Busby et al. (1996d), 17 had declining and four increasing tr ends, with a range from 18% annual decline (Lake Washington winter-run steelhead) to 7% a nnual increase (Skykomish River winter-run steelhead). These trends were for the late-run naturally produced component of winter-run steelhead populations;

no adult trend data were availabl e for summer-run steelhead. Most of these trends were based on relatively short data series. The Skagit and Snohomish River wi nter-run populations have been approximately three to five times larger than the other populations in the DPS, with average annual spawning of approximately 5,000 and 3,000 total adult spawners, respectively. These

two basins exhibited mode st overall upward trends at the time of the Busby et al. (1996d) report. Busby et al. (1996d) estimated five-year average natura l escapements for streams with adequate data range from less than 100 to 7,200, with corresponding total run sizes of 550 to 19,800. NMFS (2011f) and Ford et al. (2010) reported that al l but a few small populations were currently declining at a 3-10% rate annually, with a high ri sk of extinction in the next 100 years (Table 24).

Table 24 . Puget Sound steelhead populations and risk of extinction (Ford et al. 2010).

Geographic Region (MPGs) Population (Watershed) Extinction Risk (probability of decline to 10% of its current estimated abundance)

Northern Cascades Samish River (winter) High—about 80% within 25 years Skagit River (winter) High—about 80% within 75 years. Snohomish River (winter) Moderately High—about 50% within 100 years Stillaguamish River (winter) High—about 90% within 60 years

Tolt River summer High—near ly 80% within 100 years Nooksack River (winter) Unable to calculate

South Puget Sound Lake Washington (winter) High—~ 90% within 40 years Green River (winter)r High—a bout 90% within 80 years Nisqually River (winter) High—about 80% within 40 years Puyallup River (winter) High—about 90% within 25

‐ 30 years White River (winter) High—about 90% within 50 years

South Sound Tributaries Unable to calculate176 (winter)

Olympic Elwha River (winter) Fairly High— ~ 90% within 40 years Dungeness River (winter) High—within 100 years (population too low to calculate %) Port Angeles (winter) High—nearly 80% within 100 years West Hood Canal (winter) Low— near zero with in 100 years East Hood Canal (winter) Low—about 30% within 100 years Skokomish River (winter) High—about 80% within 80 years

Critical habitat. Critical habitat is not currently designated for Puget Sound steelhead. However, factors for essential habitat are under ev aluation to designate futu re critical habitat.

Species-specific threats and limitations to recovery. Numerous impacts hinder the survival and recovery of Puget Sound steelhead (NMFS 2011f ). These include widespread declines in adult abundance (total run size), despite significant reductions in harvest in recent years and use of two hatchery steelhead stocks (Chambers Creek and Skamania) inconsistent with wild stock diversity throughout the DPS. Furt her impairment results from d eclining diversity in the DPS, including the uncertain but weak status of summer-run fish in the DPS, a reduction in spatial structure for steelhead in the DPS, and redu ced habitat quality through changes in river hydrology, temperature profile, downstream gravel re cruitment, and reduced movement of large woody debris. Further habitat-based threats incl ude increased flood frequency and peak flows during storms, reduced groundwater-driven summer flows in the lower reaches of many rivers and their tributaries in Puget Sound where urba n development has occurred, has resulted in gravel scour, bank erosion, and sediment deposition as well as dikes, hardening of banks with riprap, and channelization, which have reduced river braiding a nd sinuosity, have increased the likelihood of gravel scour and disl ocation of rearing juveniles.

Northern California steelhead

Distribution. Northern California steelhead includes stee lhead in California coastal river basins

from Redwood Creek south to the Gualala Rive r, inclusive. Table 25 identifies populations

within the Northern California Steelhead salmon ESU, their abundances, and hatchery input.

Both winter and summer steelhead occur in th is ESU, with immature half-pounders returning from the sea after only two to four months and ove rwintering in freshwater . These juveniles then outmigrate in the following spring.

Table 25. Northern California steelhead salmon populations, abundances, and hatchery contributions (Good et al. 2005b).

River Historical

abundance Most recent spawner

abundance Hatchery abundance

contributions

Redwood Creek 10,000 Unknown Unknown

Mad River 6,000 162-384 Unknown

Eel River 82,000 3,127-21,903 Unknown

Mattole River 12,000 Unknown Unknown

Ten Mile River 9,000 Unknown Unknown Noyo River 8,000 Unknown Unknown Big River 12,000 Unknown Unknown

Navarro River 16,000 Unknown Unknown

Garcia River 4,000 Unknown Unknown

Gualala River 16,000 Unknown Unknown Other Humboldt County streams 3,000 Unknown Unknown177 River Historical

abundance Most recent spawner

abundance Hatchery abundance

contributions Other Mendocino County streams 20,000 Unknown Unknown Total 198,000 Unknown

Status and trends. Northern California steelhead were listed as threatened on June 7, 2000 (65 FR 36074), and when their status was review ed on January 5, 2006 they retained that classification (71 FR 834). Long-term data sets are limited. Before 1960, estimates of

abundance were available from dam counts in the upper Eel River (Cape Horn Dam–annual average number of adults was 4,400 in the 1930 s), the South Fork Eel River (Benbow Dam– annual average number of adults was 19,000 in th e 1940s), and the Mad River (Sweasey Dam– annual average number of adults was 3,800 in th e 1940s). Estimates of steelhead spawning populations for many rivers in this DPS totale d 198,000 by the mid-1960s. At the time of the first status review, adult escapement trends coul d be computed on seven populations. Five of the seven populations exhibited declin es, while two exhibite d increases, with a range of almost 6% annual decline to a 3.5% increase. At the time, little information was available on the actual contribution of hatchery fish to natural spawning (Busby et al. 1996d).

More recent time serious data come from snor kel counts conducted on summer-run steelhead in

the Middle Fore Eel River. An estimate of lambda over the interval 1966-2002 was made and a

random-walk with drift m odel fitted using Bayesian assumptions. Good et al. (2005b) estimated

lambda at 0.98 with a 95% confidence interval of 0.93 and 1.04. The result is an overall downward trend in both the long and short term. Juvenile data were also recently examined. Both upward and downward tr ends were apparent (Good et al. 2005b).

Critical habitat. Critical habitat was designated for th is species on September 2, 2005 (70 FR 52488). The critical habitat design ation for this DPS identifies PCEs that include sites necessary to support one or more life stages of steelhead.

Snake River steelhead

Distribution. Snake River basin steelhead are an in land species that occupy the Snake River basin of southeast Washington, northeast Oregon, and Idaho (Figure 21). The historic spawning range of this species included the Salmon, Pahs imeroi, Lemhi, Selway, Clearwater, Wallowa, Grande Ronde, Imnaha, and Tucannon Rivers. Ta ble 26 identifies populations within ESU, their abundances, and hatchery input.178 Figure 21. Population boundaries, dams, accessible areas, an d extirpated reaches, of Snake River steelhead distribution, incorporating critical habitat boundar ies (dark black line). Taken from NMFS (2011l).

Snake River steelhead occupy a basin that is an nually warmer and drier than other steelhead

DPSs. Snake River Basin steelhead are generally classified as su mmer run, based on their adult run timing pattern. Sexually im mature adults enter the Columbia River from late June to October as A-run fish that spend one year in the ocean, or larger B-run fish that spend two years at sea. Adults typically migr ate upriver until they reach trib utaries from 3,300-6,600 feet above sea level where they spawn between March a nd May of the following year. Adult spawner survival is much higher for indivi duals returning in early to mid May than for those returning in mid June (Scheuerell et al. 2009). Unlike other anadromous salmonids, some adult steelhead survive spawning, return to the sea, and later return to spawn a second time. After hatching, juvenile Snake River steelhead t ypically spend two to three years in river before they smolt and migrate to the ocean.

Table 26. Snake River basin steelhead salmon populations, abundances, and hatchery contributions (Good et al. 2005b) [rpm = redds per mile]).

River Historical

abundance Most recent spawner

abundance Hatchery abundance

contributions

Tucannon River 3,000 257-628 26%

Lower Granite run Unknown 70,721-259,145 86%

Snake A run Unknown 50,974-25,950 85%

Snake B run Unknown 9,736-33,195 89%

Asotin Creek Unknown 0-543 redds Unknown179 River Historical

abundance Most recent spawner

abundance Hatchery abundance

contributions Upper Grande Ronde River 15,000 1.54 rpm 23% Joseph Creek Unknown 1,077-2,385 0%

Imnaha River 4,000 3.7 rpm 20% Camp Creek Unknown 55-307 0% Total 22,000 (min)

Managers classify summer steelhead runs into two groups based on ocean age and adult size when they return to the Columbia River. A -run steelhead are predomin ately one-year-old ocean fish. B-run steelhead are predominately two-yea r-old ocean fish. A-run populations are found in lower Clearwater River, upper Salmon River, lo wer Salmon River and their tributaries, the Grand Ronde River, Imnaha River, and possibly the Snake River’s mainstem tributaries below Hells Canyon Dam. B-run steelhead occupy f our major subbasins not occupied by A-run

steelhead: two on the Clearwate r River (Lochsa and Selway) and two on the Salmon River (Middle Fork and South Fork Salmon). Some natural B-run steelhead spawn in the mainstem section of the Clearwater River and its major tr ibutaries. There are alternative escapement objectives for B-run steelhead. B -run steelhead represent at least one-third and as much as threefifths of the production capacity of the D PS (10,000 from the Columbia River Fisheries Management Plan and 31,400 from Idaho).

Status and trends . Snake River steelhead were listed as threatened in 1997 (62 FR 43937), when their status was reviewed on January 5, 2006 they retained that clas sification (71 FR 834). The ICBTRT (2003) identified 23 populations in the following six major population groups in this species: Clearwater River, Grande Ronde River, Hells Canyon, Imnaha River, Lower Snake River, and Salmon River. Snake River Basin stee lhead remain spatially well distributed in each of the six major geographic areas in the Snake River basin (Good et al. 2005b). Environmental conditions are generally drier a nd warmer in these areas than in areas occupied by other steelhead species in the Pacific northwest. Sn ake River Basin steelhead were blocked from portions of the upper Snake River beginning in the late 1800s and culminating with the construction of Hells Canyon Dam in the 1960s. The Snake River Basin steelhead “B run” population levels remain pa rticularly depressed.

The paucity of information on adult spawning es capement for specific tributary production areas for Snake River Basin steelhead made a quantitati ve assessment of viability difficult. Annual

return estimates are limited to counts of the aggregate return over Lower Granite Dam, and spawner estimates for the Tucannon, Grande R onde, and Imnaha Rivers. The 2001 return over Lower Granite Dam was substantially higher relativ e to the low levels seen in the 1990s; the recent 5-year mean abundance (14,768 natural retu rns) was approximately 28% of the interim recovery target level (52,000 natu ral spawners). The 10-year aver age for natural-origin steelhead passing Lower Granite Dam between 1996 and 2005 wa s 28,303 adults. Parr densities in natural production areas, which are another indicator of population status, ha ve been substantially below estimated capacity for several decades. The Snake River supports approximately 63% of the total natural-origin production of steelhead in the Columbia River Basin. Genetic diversity is currently being depressed by the displacement of natural fish by hatchery fish (declining proportion of natural-origin spawners). Homogeni zation of hatchery stocks occur within basins, and some stocks exhibit high stray rates. Littl e change from the 2005 review was found in the180 2011 status review (NMFS 2011l).

Critical habitat. Critical habitat was designated for th is species on September 2, 2005 (70 FR 52488). The critical habitat design ation for this ESU identifies PCEs that include sites necessary to support one or more life stages of steelhead. The critical habitat de signation for this ESU identifies PCEs that include sites necessary to support one or more steelh ead life stages. Of the 291 fifth order streams reviewed in this DPS, 220 were rated as high, 44 were rated as medium,

and 27 were rated as low conservation value. Limiting factors identifie d for Snake River Basin steelhead include: (1) hy drosystem mortality, (2) reduced st ream flow, (3) altered channel morphology and floodplain, (4) excessive sediment , (5) degraded water quality, (6) harvest impacts, and (7) hatchery impacts (NMFS 2006i).

Species-specific threats and limitations to recovery. Snake River Basin steelhead recovery is limited by predation, impaired water quality and in creased temperature, harvest (particularly of B-run steelhead), genetic diversity stemming from relerase of out-of-population hatchery fish, impaired tributary fish passage (ICTRT 2006, NMFS 2011l). Also significant is floodplain connectivity and function, channe l structure and complexity, ri parian areas and large woody debris recruitment, stream flow , and water quality have been degr aded as a result of cumulative impacts of agriculture, forestry, and development.

Upper Willamette River steelhead

Distribution. Upper Willamette River steelhead occupy the Willamette River and its tributaries upstream of Willamette Falls (Figure 22). Table 27 identifies the four populations within the Upper Willamette River Steelhe ad salmon ESU, their abundances, and hatchery input.181 Figure 22. Population boundaries, dams, accessible areas, and extirpated r eaches, of upper Willamette River steelhead distribution, incorporating critical habitat boundaries (dark black line). Taken from NMFS (2011).

Table 27. Upper Willamette river steelhead populations, abundances, and hatchery contributions (Good et al. 2005b) [rpm = redds per mile]).

Population Historical

abundance Most recent spawner

abundance Hatchery abundance

contributions

Mollala Rivers Unknown 0.972 rpm Unknown

North Santiam River Unknown 0.963 rpm Unknown

South Santiam River Unknown 0.917 rpm Unknown Calapooia River Unknown 1.053 rpm Unknown

Total Unknown 5,819

This is a late-migrating winter group that enters freshwater in January and February (Howell et
al. 1985a). They do not ascend to their spawning areas until late March or April (Dimick and Merryfield 1945) and spawning occurs from April to June. The smolt migr ation past Willamette Falls also begins in early April and proceeds in to early June, peaking in early- to mid-May (Howell et al. 1985a). Smolts generally migrate thr ough the Columbia via Multnomah Channel

rather than the mouth of the Willamette River. Most spend two years in the ocean before reentering natal rivers to spawn (Busby et al. 1996b). Steelhead in the Upper Willamette River DPS generally spawn once or twice, although some may spawn three times. Repeat spawners are predominantly female and generally account for less than 10% of the total run size (Busby et
al. 1996b).

Status and trends. Upper Willamette River steelhead were listed as threatened in 1999 (64 FR182 14517) and when their status was reviewed on January 5, 2006 they retained that classification (71 FR 834). A major threat to Upper Willame tte River steelhead results from artificial production practices. Fishways built at Willamette Falls in 1885 have allowed Skamania-stock summer steelhead and early-migrating winter steelh ead of Big Creek stock to enter the range of Upper Willamette River steelhea d. The population of summer steelhead is almost entirely maintained by hatchery salmon, although natural-or igin, Big Creek-stock wi nter steelhead occur in the basin (Howell et al. 1985a). In recent years, releases of winter steelhead are primarily of native stock from the Santiam River system.

Steelhead in this DPS are depressed from historical levels (McElhaney et al. 2007). Although long-term growth is negative, short term abunda nce studies support a rece nt upturn in population numbers until 2002, when the population peaked and ha s since returned to hi storical levels (2008 abundance was 4,915 individuals passing Willa mette Falls and 2,110 in 2009)(McElhaney et al. 2007, NMFS 2011j). Spatial structure for the No rth and South Santiam populations has been substantially reduced by the loss of access to the upper North San tiam basin and the Quartzville

Creek watershed in the South Santiam subbasi n due to dam construction lacking passage facilities (McElhaney et al. 2007). Additionally, habitat in the Molalla subbasin has been reduced significantly by habitat degradation and in the Calapooi a by habitat degradation as well as passage barriers. Finally, the diversity of some populations has been eroded by small population size, the loss of access to hi storical habitat, legacy effect s of past winter-run hatchery releases, and the ongoing release of summer steelhead (McElhaney et al. 2007).

Critical habitat. Critical habitat was designated for th is species on September 2, 2005 (70 FR 52488). The designation includes all Columbia Rive r estuarine areas and river reaches upstream to the confluence with the Willamette River as well as specific steam reaches in the following subbasins: Upper Willamette, North San tiam, South Santiam, Middle Willamette, Molalla/Pudding, Yamhill, Tualatin, and Lower Willamette (NMFS 2005b). The critical habitat designation for this DPS identifies PCEs that in clude sites necessary to support one or more life stages of steelhead. Of 43 subba sins reviewed in NMFS' assessmen t of critical habitat for Upper Willamette River steelhead, 20 subbasins were ra ted as having a high conservation value, while six were rated as having a medium value and 17 were rated as having a low value to the

conservation of the DPS.

Land management activities have severely degraded stream habitat conditi ons in the Willamette

River mainstem above Willamette Falls and asso ciated subbasins. In the Willamette River mainstem and lower sub-basin mainstem reaches, high density urban development and widespread agricultural effects have reduced aqua tic and riparian habitat quality and complexity, and altered sediment and water quality and quantity, and watershed processes. The Willamette River, once a highly braided river system, has been dramatically simplified through channelization, dredging, and othe r activities that have reduced rearing habitat by as much as 75%. In addition, the construction of 37 dams in the basin blocked access to more than 435 miles of stream and river spawning habitat. The dams alter the temperature regime of the Willamette River and its tributaries, affecting the timing and development of naturally-spawned eggs and fry. Agriculture, urbanization, and gravel mining on the valley floor logging in the Cascade and Coast Ranges contribute to increased erosion and sediment loads throughout the basin.183 The mainstem Willamette River has been channe lized and stripped of large wood. Development began to encroach on the riparian forest be ginning in the 1870s (Sedell and Froggatt 1984).

Gregory et al. (2002d) calculated that the total main stem Willamette River channel area decreased from 41,000 to 23,000 acres between 1895 a nd 1995. They noted that the lower reach,

from the mouth of the river to Newberg (RM 50), is confined with in a basaltic trench, and that due to this geomorphic constraint, less channel area has been lost than in upstream areas. The middle reach from Newberg to Albany (RM 50 to 120) incurred losses of 12% primary channel area, 16% side channels, 33% alcoves, and 9% islands. Even greater changes occurred in the upper reach, from Albany to Eugene (RM 187). There, approximately 40% of both channel length and channel area were lost, along with 21% of the primary channel, 41% of side channels, 74% of alcoves, and 80% of island areas.

The banks of the Willamette River have more than 96 miles of revetments; approximately half were constructed by the ACOE. Gene rally, the revetments were placed in the vicinity of roads or on the outside bank of river bends, so that while only 26% of the total length is revetted, 65% of the meander bends are revetted (Gregory et al. 2002c). The majority of dynamic sections have been armored, reducing adjustments in channel bed and sediment storage by the river, and thereby diminishing both the complexity a nd productivity of aqua tic habitats (Gregory et al. 2002b).

Riparian forests have diminished considerably in the lower reaches of the Willamette River (Gregory et al. 2002a). Sedell and Froggatt (1984) noted that agriculture and cutting of streamside trees were major agents of change for riparian vegetation, along with snagging of large wood in the channel. The reduced shoreline, fewer and smaller snags, and reduced riparian forest comprise large functional losses to the ri ver, reducing structural features, organic inputs from litter fall, entrained allochthonous material s, and flood flow filtering capacity. Extensive changes began before the major dams were bui lt, with navigational and agricultural demands dominating the early use of the river. The once expansive forests of the Willamette River floodplain provided valuable nutrients and orga nic matter during flood pulses, food sources for macroinvertebrates, and slow-wat er refugia for fish during flo od events. These forests also cooled river temperatures as the ri ver flowed through its many channels.

Gregory et al. (2002a) described the changes in riparian vegetation in river reaches from the mouth to Newberg, from Newberg to Albany, and from Albany to Eugene. They noted that the riparian forests were formerly a mosaic of br ush, marsh, and ash tree openings maintained by annual flood inundation. Below the City of Newb erg, the most noticeable change was that conifers were almost eliminated. Above Newberg, the formerly hardwood-dominated riparian forests along with mixed forest made up less than half of the riparian vegetation by 1990, while agriculture dominated. This conversion has re duced river shading and the potential for recruitment of wood to the river, reducing ch annel complexity and the quality of rearing, migration and spawning habitats

Hyporheic flow in the Willamette River has be en examined through discharge measurements and found to be significant in some areas, partic ularly those with gr avel deposits (Wentz et al. 1998,

Fernald et al. 2001). The loss of channel complexity a nd meandering that fosters creations of184 gravel deposits decreases the potential for hyporheic flows, as does gravel mining. Hyporheic flow processes water and affect s its quality on reemerging into the main channel, stabilizing variations in physical and chemical water char acteristics. Hyporheic flow is important for ecological functions, some aspects of water quali ty (such as temperature and dissolved oxygen), and some benthic invertebrate life stages. Alcove habitat, which has been limited by channelization, combines low hydraulic stress and high food availability w ith the potential for hyporheic flows across the steep hydraulic gradients in the gravel separating them from the main channel (Fernald et al. 2001).

Species-specific threats and limitations to recovery. Habitat degredation and hatchery-related impacts are significant in hindering the rec overy of Upper Willamette River steelhead (Guimarães et al. 2011, ODFW and NMFS 2011). Floodpl ain connectivity and function, channel structure and complexit y, riparian areas and large wood recruitment, and stream flow have been degraded as a result of cumulative im pacts of agriculture, forestry, and development. Degraded water quality and alte red temperature as a result of both tributary dams and the cumulative impacts of agriculture, forestry, and ur ban development also has reduced the quality of available habitat. Spawning individuals have reduced access to spawning and rearing habitats mainly as a result of artificial barriers in sp awning tributaries. Ge netic effects of non-native summer steelhead hatchery programs and anthrop ogenic introductions of non-native species and

out-of-ESU races of salmon or steelhead have increased predation and competition on native

UWR steelhead.

South-Central California coast steelhead

Distribution. The South-Central California steelhead DPS includes all naturally spawned populations of steelhead (and their progeny) in streams from the Pa jaro River (inclusive) to, but not including the Santa Maria River, California

. Only winter steelhead are found in this ESU. Migration and spawn timing are simila r to adjacent steelhead populations.

Status and trends. South-Central California Coast steelh ead were listed as threatened in 1997, when their status was reviewed on January 5, 2006 they retained that clas sification (71 FR 834).

Historical data on the South-Central Californi a Coast steelhead DPS are limited. In the mid1960s the California Department of Fish and Game estimated that the ad ult population at about 18,000. We know of no recent estimates of the tota l DPS. However, five river systems, the Pajaro, Salinas, Carmel, Little Sur, and Big Sur, indicate that runs ar e currently less than 500

adults. Past estimates for these basins were al most 5,000 fish. Carmel River time series data indicate that the population declined by about 22% per ye ar between 1963 and 1993 (Good et al. 2005b). From 1991 the population increased from one adult, to 775 adults at San Clemente

Dam. Good et al. (2005b) thought that this recent increase seemed too great to attribute simply to improved reproduction and survival of the loca l steelhead population. Other possibilities were considered including that the substantial immigrati on or transplantation occu rred, or that resident trout production increased as a re sult of i mproved environmental c onditions within the basin.

Critical habitat. Critical habitat was designated for th is species on September 2, 2005 (70 FR 52488). The critical habitat design ation for this DPS identifies PCEs that include sites necessary to support one or more life stages of steelhead. The critical habitat de signation for this DPS identifies PCEs that include sites necessary to support one or more steelhead life stages.185 Southern California Steelhead

Distribution. Southern California steelhead occupy rive rs from the Santa Maria River to the US –Mexico border. Table 28 identifies populations within the Southern California Steelhead salmon ESU, their abundances, and hatchery inpu t. Migration and life history patterns of southern California steelhead are dependent on rainfall and streamflow (Moore 1980). Steelhead within this ESU can withstand higher temperat ures than more northerly populations and the relatively warm and productive waters of the Vent ura River permit more ra pid juvenile growth (Moore 1980).

Table 28. Southern California steelhead populations, abundances, and hatchery contributions (Good et al. 2005).

River Historical

Abundance Most Recent Spawner Abundance Hatchery Abundance

Contributions

Santa Ynez River 12,995-30,000 Unknown Unknown

Ventura River 4,000-6,000 Unknown Unknown Matilija River 2,000-2,500 Unknown Unknown

Creek River Unknown Unknown Unknown

Santa Clara River 7,000-9,000 Unknown Unknown Total 32,000-46,000 <500

Status and trends. Southern California steelhead were listed as endangered in 1997 (62 FR 43937), when their status was reviewed on January 5, 2006 they retained th at classification (71 FR 834). In many watersheds throughout Southe rn California, dams isolate steelhead from historical spawning and rearing habitats and alter the hydrology of th e basin (e.g., Twitchell

Reservoir within the Santa Maria River waters hed, Bradbury Dam within the Santa Ynez River watershed, Matilija and Casitas dams within th e Ventura River watershed, Rindge Dam within the Malibu Creek watershed). Based on combined estimates for the Santa Ynez, Ventura, and Santa Clara rivers, and Malibu Creek, an es timated 32,000 to 46,000 adult steelhead occupied this DPS. In contrast, less than 500 adults are estimated to occupy the same four waterways presently. The last estimated run size for st eelhead in the Ventura River, which has its

headwaters in Los Padres Nati onal Forest, is 200 adults (Busby et al. 1996c, Busby et al. 1996d).

Critical habitat. Critical habitat was designated for this species on September 2, 2005. The designation identifies PCEs that include sites necessary to support one or more steelhead life stages and, in turn, these sites contain the physical or biological features essential for the species

conservation.

Southern green sturgeon Description of the species. The species is divided into two ge netically distinct but physically indistinguishable clades: a Northern DPS w hose populations are rela tively healthy, and a

Southern DPS that has under gone significant decline (Adams et al. 2007). Only the Southern DPS of green sturgeon is listed under the ESA.

Distribution. Green sturgeon occur along the west coast of North America from Mexico to the Bering Sea (Adams et al. 2002, Moyle 2002a, Colway and Stevenson 2007), as well in

freshwater rivers and estuaries in between. Bo th northern DPS and southern DPS green sturgeon occupy coastal estuaries and coastal marine waters from southern California to Alaska, including Humbolt Bay, the lower Columbia River estuar y, Willapa Bay, Grays Harbor and southeast186 Alaska. In general, green sturgeon are more common north of Point Conception, California. This DPS occupies a limited number of river system s, primarily the main stems and tributaries of the Sacramento, Feather, and San Joaquin rivers, all of which drain to San Francisco Bay (Israel

et al. 2009). The Sacramento Rive r is the only known spawning ha bitat for the Southern green sturgeon DPS (Moyle et al. 1992a, CDFG 2002). Spawning in the San Joaquin River does not presently occur, likely as a result of damming and irrigation development. Adult and juvenile green sturgeon occur throughout th e Sacramento River. Although Southern DPS individuals in freshwater occur primarily in the aforementioned waterways, some have been captured in rivers of the Northern DPS, including the Klamath, Rogu e, and Columbia rivers of northern California and Oregon. Individuals have also rarely been captured in Southe rn California waters, but their

abundance increases north of Point Conception.

Upon outmigration from fresh water, subadult green sturgeon disper se widely along through

continental shelf waters of the west coast within the 110 meter contour (NMFS and USFWS 2005).(Moyle et al. 1992a, Erickson and Hightower 2007). Captures provide the limited knowledge available on green sturge on distribution and habitat in es tuaries and the North Pacific

Ocean. Adults are known from coastal estuarie s and bays, including Willapa and Nehalem bays, Grays Harbor, and the Colu mbia River estuary (EPIC et al. 2001). It appears that Southern DPS green sturgeon preferentially distribute north of the river mouth from whence they emerge as juveniles during fall and move into bays a nd estuaries during summer and fall (Moser and Lindley 2007, Israel et al. 2009). Spawning individuals move south during spring to natal freshwater habitat (Lindley et al. 2008).

Reproduction . Adults infrequently occupy freshwater reaches primarily to spawn and spend the majority of their lives in es tuaries and coastal marine wate rs (Wilson and McKinley 2004). Sturgeon, like salmon, possess strong site fidelity a nd will return to their natal streams to spawn (Bemis and Canard 1997). This occurs between late February and July, with a peak in mid-April to mid-June, hold in deep pools and return to sa lt water in the fall early, often with the first increases in fall flows, although Northern DPS i ndividuals in the Rogue River seek out slowflow and off-channel coves (Moyle et al. 1992a, Moyle et al. 1995a, Erickson et al. 2001b, Rien

et al. 2001a, Erickson and Webb 2007a, Heublein et al. 2009). Fish then tend to aggregate in deep pools, where they will over-summer before outmigrating in the fall, although some fish have been observed outmigrating relatively soon after presumed spawning events (Heublein et
al. 2009). However, multiple spawning events may occur annually, unlike annual spawning of salmonids (Heublein et al. 2009). Adults may spend over 6 m onths in freshwater until water temperatures drop below 50o F in fall or winter or when a significant flow even occurs, although temperature or flow cues are not always needed to initiate do wnstream migration (Erickson and Webb 2007a, Heublein et al. 2009). In the Sacramento River a dult green sturgeon spawn in late spring and early summer above Hamilton City, ab ove Red Bluff Diversion Dam, and possibly as far upstream as Keswic k Dam (CDFG 2002, Heublein et al. 2009). It appears that specific habitat for spawning includes large cobblestone s (where eggs can se ttle between), although spawning is known to occur over clean sand or bedrock.

Adults likely return on a 2 to 5 year basis for sp awning starting at 15 to 30 years of age for males and 17 to 40 for females (Adams et al. 2002, Moyle 2002a, Van Eenennaam et al. 2006). Most male spawners are young (17 to 18 years) whil e females on the spawning grounds are often older187 (27 to 28 years). Females produce roughly 60,000 to 140,000 eggs per spawning event (Scott and Crossman 1973c, Moyle et al. 1992a). Each egg is large by sturgeon standards so as to contain greater nourishment for embryos (Cech Jr. et al. 2000). Temperature may trigger spawning behavior, with ranges of 48o to 62o F being influential (Moyle et al. 1995a). Water temperature is also critical fo r egg survival with temperatur es above 68o F being fatal to developing embryos (Cech Jr. et al. 2000).

Growth. Green sturgeon spend their first 1 to 4 years in their natal streams and rivers (Nakamoto et al. 1995b, Beamesderfer and Webb 2002), al though they are believed to be physiologically adapted to sea water survival at 6 months of age (A llen and Cech 2007, Allen and J.J. Cech 2007, Allen et al. 2009). Larvae are active at night, a behavior that likely reduces predation and avoids bei ng moved downstream more th an necessary (Cech Jr. et al. 2000). Green sturgeon larvae grow very rapidly, reaching about 300 mm by age one (Deng 2000). Growth of the larval stage is ideal at 59o F, reduced below 52o F and above 66o F, and highly reduced above 75oF (Cech Jr. et al. 2000). Juvenile green sturge on of the Northern DPS (Rogue River) have been captured in the Rogue River es tuary from April until the end of November (Rien et al. 2001a). Green sturgeon are a long-lived fish, and likely live for 60 to 70 years (Moyle 2002a).

Feeding. While in fresh water, juveniles feed on a variety of fishes and invertebrates (Moyle et
al. 1992a). One juvenile from the Sacramento-S an Joaquin estuary was found to have preyed most commonly upon opisthobranch mollusks ( Philline sp.), with bay shrimp ( Crangon sp.) and overbite clams ( Potamocorbula amurensis ) as secondary prey. Other juveniles in the Sacramento River delta feed on opossum shrimp ( Neomysis mercedis ) and Corophium amphipods (Radtke 1966).

The limited feeding data available for adult gr een sturgeon show that they consume benthic invertebrates including shrimp, clams, chironom ids, copepods, mollusks, amphipods, and small fish (Houston 1988, Moyle et al. 1992a)(Houston 1988; Moyle et al. 1992; Wilson and McKinley 2004; Dumbauld et al. 2008)(Moyle et al. 1992a, Wilson and McKinley 2004,

Dumbauld et al. 2008). Starting as larvae, sturgeon us e electroreception to identify prey. Olfaction and taste may also be important to foraging, while vision is thought play a minor role in prey capture (Miller 2004).

Status and trends. NMFS listed the southern population of the North American green sturgeon as threatened on April 7, 2006 (71 FR 17757). Trend data for green sturgeon is severely limited. Available information comes from two predominan t sources, fisheries and tagging. Only three data sets were considered useful for the popul ation tim e series analyses by NMFS’ biological review team: the Klamath Yurok Tribal fishery ca tch, a San Pablo sport fishery tag returns, and

Columbia River commercial la ndings (NMFS and USFWS 2005). Using San Pablo sport fishery tag recovery data, the California Department of Fish and Game produced a population time

series estimate for the southern DPS. San Pa blo data suggest that green sturgeon abundance may be increasing, but the data showed no significan t trend. The data set is not particularly convincing, however, as it suffers from inconsis tent effort and since it is unclear whether summer concentrations of green sturgeon provide a strong indicator of population performance (NMFS and USFWS 2005). Althoug h there is not sufficient info rmation available to estimate188 the current population size of southern green stur geon, catch of juveniles during state and federal salvage operations in the Sacramento delta are lo w in comparison to catch levels before the mid1980s.

Natural threats. Green sturgeon eggs and la rvae are likely preyed upon by a variety of larger fish and animals, while sub-adult and adult st urgeon may occasionally be preyed upon by shark sea lions, or other large body predators. P hysical barriers, change s in water flow and temperatures may also affect fresh water survival.

Anthropogenic threats. The principle threat to southern green sturgeon comes from a drastic reduction in available spawning area from impa ssible barriers (e.g., Oroville, Shasta and Keswick dams). Other threatens include poten tially lethal temperature limits, harvest, entrainment by water projects and to xins and invasive species (Adams et al. 2007, Erickson and Webb 2007b, Lackey 2009). Since this DPS is co mposed of a single spawning population within the Sacramento River, stochastic variati on in environmental conditions and significant fluctuations in demographic rates increases the risk of extinct ion for this DPS.

Climate change has the potential to affect stur geon in similar, if not more significant ways it affects salmonids. Elevated air temperatures coul d lead to precipitation fa lling as rain instead of snow. Additionally, snow would likely melt soone r and more rapidly, potentially leading to greater flooding during melting and lower water levels at other ti mes, as well as warmer river temperatures. Although sturgeon can spawn over va ried benthic habitat, they prefer localized depressions in riverbeds (Moyle et al. 1992b, Moyle et al. 1995b, Erickson et al. 2001a, Rien et
al. 2001b). Increased extremes in river flow (i.e., periods of flooding and low flow) can alternatively disrupt and fill in spawning habitat that sturgeon rely upon (ISAB 2007b). If water flow is low during migration events, it is likely that new obstacles can im pede or block sturgeon movement. As with other anadromous fishes, st urgeon are uniquely evolve d to the environments that they live in. Because of this specificity, broad scale changes in environment can be difficult to adapt to, including changes in water temperature (Cech et al. 2000). Sturgeon are also sensitive to elevated water temperatures. Te mperature triggers spaw ning behavior. Warmer water temperatures can initial spawning earlier in a season for salmon and the same can be true for sturgeon (ISAB 2007b). If river and lake temp eratures become anomalously warm, juvenile sturgeon may experience elevated mortality due to lack of cooler water refuges in freshwater habitats. Apart from direct changes to sturgeon survival, altered water te mperatures may disrupt habitat, including the availabil ity of prey (ISAB 2007b). Warmer temperatures may also have the effect of increasing water use in agriculture, both for existing fields and the establishment of new ones in once unprofitable area s (ISAB 2007b). This means that streams, rivers, and lakes will experience additional withdrawal of water for irrigation and increasing contaminant loads from returning effluent. Overall, it is likel y that global warming will increase pressures on sturgeon survival and recovery.

Studies from other sturgeon species indicate that sturgeon readily bioaccumulate contaminants. White sturgeon from the Kootenai River have been found to contain aluminum, arsenic, cadmium, chromium, cobalt, copper, iron, lead, manganese, mercury, nickel, selenium, zinc, DDE, DDT, PCBs, and other organochlorines (Kru se and Scarnecchia 2001). Mercury has also been identified from white sturgeon of the lower Columbia River (Webb et al. 2006). Numerous189 organochlorines, including DDT, DDD, DDE, chlordan e, and dieldrin have also been identified in these fish (Foster et al. 2001). Observed concentrations are likely sufficient to influence reproductive physiology.

Critical habitat. On October 9, 2009, NMFS designated cri tical habitat for southern green sturgeon (74 FR 52300). The geogra phical area identified as critic al habitat is based upon the overlapping distribution of the s outhern and northern DPS, and encompasses all areas where the presence of southern green stur geon have been confirmed or wher e their presence is likely. Therefore the geographical area defined as critical habitat is the entire range of the biological species, green sturgeon, from the Bering Sea, AK, to Ensenada, Mexico. Specific fresh water areas include the Sacramento River, Feather Rive r, Yuba River, and the Sacramento-San Joaquin Delta. Specific coastal bays and estuaries include estuaries from Elkhorn Slough, California, to Puget Sound, Washington. Coastal marine areas include waters along th e entire biological species range within a depth of 60 fathoms. Th e principle biological or physical constituent

elements essential for the conservation of sout hern green sturgeon in fresh water include: food resources; substrate of sufficien t type and size to support viable egg and larval development; water flow, water quality such that the chemical characteristics support normal behavior, growth and viability; migratory corrido rs; water depth; and sediment quality. Primary constituent

elements of estuarine habita t include food resources, water flow, water quality, migratory corridors, water depth, and sediment quality. Th e specific primary constituent elements of marine habitat include food resources, water quality, and migrator y corridors.

Critical habitat of the Southern DPS of green sturgeon is threatened by several anthropogenic

factors. Four dams and several other structur es currently are impassibl e for green sturgeon to pass on the Sacramento, Feather, and San Joaqui n rivers, preventing movement into spawning

habitat. Threats to these riverine habitats also include increasing temperature, insufficient flow that may impair recruitment, the introduction of striped bass that may eat young sturgeon and compete for prey, and the presence of heavy metals and contaminants in the river. The application of pesticides may adversely affect prey resources and water quality within the bays and estuaries, as well as the growth and repr oductive health of Southern DPS green sturgeon through bioaccumulation. Other activities of concern include those that may disturb bottom substrates, adversely affect pr ey resources, or degrade water quality through re-suspension of contaminated sediments. Of particular concern ar e activities that affect prey resources. Prey resources can be affected by: commercial sh ipping and activities ge nerating point source pollution and non-point source pol lution that can discharge c ontaminants and result in bioaccumulation of contaminants in green sturgeon ; disposal of dredged materials that can bury prey resources; and bottom trawl fisheries th at can disturb the bottom (but may result in

beneficial or adverse effects on prey resources for green sturge on). In addition, petroleum spills from commercial shipping activ ities and proposed alternative energy hydrokinetic projects may affect water quality or hinder the migrati on of green sturgeon al ong the coast (USDC 2009b).

Shortnose sturgeon Description of the species. Shortnose sturgeon occur along the Atlantic Coast of North America, from the St. John River in Canada, sout h to the St. John’s Rive r in Florida. NMFS’ recovery plan (1998b) recognized 19 wild populations based o n their strong fidelity to their natal streams (Table 29), and several captive populations (from a Savannah River broodstock) that are190 maintained for educational and research pur poses (NMFS 1998b). Alt hough these populations are geographically isolated, gene tic analyses suggest that indi vidual shortnose sturgeon move between some of these populations each generation (Quattro 2002, Wirgin et al. 2005a).

Table 29. Shortnose sturgeon populations and their estimated abundances.

Population (Location) a Data
Series Abundance Estimate (C.I.) b Population Segment Reference

Saint John River (Canada) 1973-1977 18,000 (+/-30%) Adults Dadswell 1979 Kennebecasis River (Canada) 1998-20 05 2,068 (801-11,277) COSEWIC 2005

Kennebecasis River 2005 4,836 (+/-69) Li et al. 2007, NMFS unpubl. Penobscot River (ME) 2006-2007 1,049 (673-6,939) UME 2008 2008 1739 (846-3653) Summer P. Dionne, pers. comm.. 667 (451-1013) Fall P. Dionne, pers. comm.. Kennebec River (ME) 1977-1981 7,222 (5,046-10,765) Adult Squiers et al. 1982 2003 9,488 (6,942-13,358) Adults Squiers 2003

Merrimack River (MA) 1987-1991 32 (20-79) Adults Kynard & Kieffer, unpubl.; NMFS unpubl. Connecticut River (MA, CT) 1989-2002 1,042-1,580 c Adults Savoy 2004 Upper Connecticut River d 1976-1977 516 (317–898) Total Taubert 1980; NMFS 1998a 1977-1978 370 (235–623) Total Taubert 1980; NMFS 1998a 1976-1978 714 (280-2,856) Total Taubert 1980; NMFS 1998a 1976-1978 297 (267–618) Total Taubert 1980; NMFS 1998a

1994 328 (188-1,264) Adults Kynard & Kieffer, unpubl.; NMFS unpubl.

1994-2001 143 (14-360) Spawning

Adults Kynard & Kieffer, unpubl.; NMFS unpubl.

Lower Connecticut River e 1988-1993 895 (799-1,018) Adult Savoy and Shake 1992;

NMFS 1998a Hudson River (NY) 1980 30,311 Total Dovel 1979; NMFS 1998b

1994-1997 61,057 (52,898- 72,191) Total Bain et al. 2007

Delaware River (NJ, DE, PA) 1981-1984 12,796 (10,288- 16,267) Partial Hastings et al. 1987

1981-1984 14,080 (10,079- 20,378) Partial Hastings et al. 1987

1999-2003 12,047 (10,757- 13,589) Brundage and O'Herron

2003 Chesapeake Bay (MD, VA) Cape Fear River (NC) Winyah Bay (NC, SC) Santee River (SC) Cooper River (SC) 1996-1998 300 Adults Cooke et al. 2004 ACE Basin (SC)

Savannah River (SC, GA) 1,000 - 3,000 Adults B Post, SCDNR 2003; NMFS unpubl. Ogeechee River (GA) 1993 266 (236 – 300) Weber 1996, 1998

1993 361 (326 – 400) Total Rogers and Weber 1994,

NMFS 1998b

1999-2004 147 (104-249) Fleming et al. 2003; NMFS unpubl. Altamaha River (GA) 1988 2,862 (1,069 - 4,226) Total NMFS 1998a 1990 798 (645 - 1,045) Total NMFS 1998a 1993 468 (316 – 903) Total NMFS 1998a

6,320 (4,387-9,249) Total DeVries 2006191 Population (Location) a Data

Series Abundance Estimate (C.I.) b Population Segment Reference

Satilla River (GA) Saint Mary's River (FL) Saint Johns River (FL) FFWCC 2007c

a The original 19 populations identified by NMFS in the 1998 recovery plan are left aligned in this column. Estimates for a tributary or river segment are indented.

b Population estimates are established using different techniques and should be viewed with caution. In some cases, sampling biases may have violated the assumptions of the procedures used or resulted in inadequate representation of a population segment. Some estimates (e.g., those without confidence intervals or are depicted by ranges only) are the “best professional judgment” of research ers based on their sampling effort and success.

c Range represents total population estimates using four di fferent techniques. All techniques suggest the population increased during the sampling period (see Savoy 2004 for more details).

d Above Holyoke Dam.

e Below Holyoke Dam.

Distribution. Shortnose sturgeon occur along the Atlantic Coast of North America, from the St.

John River in Canada to the St. John’s River in Florida. At the northern end of the species’ distribution, the highest rate of gene flow (which suggests mi gration) occurs between the Ponobscot and Androscoggin Rivers. At the southern end of the species’ distribution, populations south of the Pee Dee River appear to exchange between one and 10 individuals per generation, with the highest ra tes of exchange between the Ogeechee and Altamaha Rivers (Wirgin et al. 2005a). Wirgin et al. (2005) concluded that rivers separated by more than 250 miles were connected by very little migration wh ile rivers separated by no more than 12 miles (such as the rivers flowing into coastal South Carolina) would experience high migration rates. Coincidentally, at the geographic center of the shortnose sturgeon range, there is a 250 mile

stretch of river with no known populations occurring from the De laware River, New Jersey to Cape Fear River, North Carolina (Kynard 1997a). However, shortnose sturgeon are known to occur in the Chesapeake Bay, and may be transien ts from the Delaware River via the Chesapeake (Skjeveland et al. 2000, Welsh et al. 2002) or remnants of a population in the Potomac River (Kynard et al. 2009).

Rogers and Weber (1995a), Kahnle et al. (1998a), and Collins et al. (2000b) concluded that shortnose sturgeon are extinct from the St. Johns River in Florida and the St. Marys River along the Florida and Georgia border. In 2002, a shor tnose sturgeon was captured in the St. Johns River, Florida, suggesting either immigration or a small remnant population (FFWCC 2007c). Rogers and Weber (1995a) also concluded that shortnose sturgeon have become extinct in Georgia’s Satilla River.

Habitat. Habitat use in fresh water during summer and winter months overlaps between adult and age-1 shortnose sturgeon (O'Herron II et al. 1993, Rogers and Weber 1995b, Kynard et al. 2000). Kynard et al. (2000) found that both age classes pr eferred deep-water curves with sand and cobble to higher velocity r uns, particularly during winter months, and shifted to channel habitat as water temperatures rose in summer mo nths. Many fish also ex hibited diel movement patterns between deeper waters during the day and shallower waters at night (Kynard et al. 2000). During the summer, at the southern end of their range, shortnos e sturgeon congregate in cool, deep, areas of rivers where adult and juvenile sturgeon can ta ke refuge from high temperatures (Flournoy et al. 1992, Weber 1996a). In the Connect icut River and the Merrimack, Kynard et al. (2000) found shortnose generally used wa ter about 3 meters deep, ranging from192 less than a meter to about 15 meters deep.

Movements. The general migratory strategy of shortnose sturgeon is similar to many fresh water and diadromous fishes, which probably optimizes feeding opportunities, minimizes losses due to unfavorable conditions (winter refuge migra tions), and optimizes spawning success(Harden Jones 1968, Northcote 1978). Water temperatures, flow regimes, and barriers influence their movement patterns (Kynard 1997(Kynard et al. 2000). Adult shortnos e sturgeon will migrate upstream to spawning areas in the spring or in the fall. Fish th at migrate upstream in the fall generally overwinter in areas just downstr eam of spawning sites, while others including non-spawners will overwinter in estuarine wa ters. After spawning in the spring, spent (post-spawned) adults tend to migrate rapidly down stream to feeding areas in the estuary or to tidally influence fresh water during summer, wh en movement become limited, possibly due to stress asscoaited with high water temperatures and low oxygen concentrations (Dadswell et al. 1984a, Collins 2010).

In the Penobscot River, shortnose sturgeon o ccur year-round in estuarine waters, spending midOctober to mid-April in the upper estuarey and moving downstr eam to the middle estuary in spring and near the river mouth in summer (Fernandes et al. 2010). However, individuals move in and out of the Penobscot River, some traveling at least 150 km to other monitored rivers, such as the Kennebec (Fernandes et al. 2010).

Young-of-the year shortnose stur geon move downstream after hatching, remaining in fresh water for about 1 year (Kynard 1997b). Initially, young shortnose sturgeon will re side short distances from spawning areas, and as they grow will tend to move further downstream (Dadswell et al. 1984a). By age 3 or older juvenile sturgeon will spend a large portion of th eir year at the saltand fresh water interface of co astal rivers (NMFS 1998a).

Sturgeon are iteroparous, and based on limited data it appears that females sturgeon spawn every three to five years while males spawn every other year, although some may spawn in consecutive years (Kieffer and Kynard 1993, NMFS 1998a) (Dovel et al. 1992; Collins and Smith 1993). Spawning typically occurs during the spring, betw een mid-March and late May. Spawning areas are often located just below the fall line at the farthest accessible upstream reach of the river (NMFS 1998a). The onset of spawning may be cu ed to decreasing river discharge following the peak spring freshet, when water temperatures ra nge from 8 to 12 °C and bottom water velocities range between 25-130 cm/s, although photoperiod appears to control spawning readiness (Dadswell et al. 1984a, NMFS 1998a).

Length at maturity is about 45-55 cm fork length fo r shortnose sturgeon and age at first spawning appears to vary along a latitudinal cline. Acco rding to spawning checks, it appears that male shortnose sturgeon in southern rivers will first spawn between ages 2 and 5, while fish as far north as the St. Johns River, Canada first spawn at about 10 to 11 years of age (Dadswell et al. 1984a, NMFS 1998a). Age at first spawning for female shortnose sturgeon varies from about age 6 to 18 years, like males, vary ing on a latitudina l cline (Dadswell et al. 1984a, NMFS 1998a). In general, fish in the northern portion of the species’ range live longer than individuals in the southern portion of the species’ range (Gilbert 1989b). The m aximum age reported for a shortnose sturgeon in the St. J ohn River in New Brunswick is 67 years (for a female), 40 years193 for the Kennebec River, 37 years for the Hudson River, 34 years in the Connecticut River, 20 years in the Pee Dee River, and 10 years in the Altamaha River (Gilbert 1989b). Male shortnose sturgeon appear to have shorter life spans than females (Gilbert 1989b).

Feeding. Like all sturgeon, shortnose have ventrally located, sucker-like m ouths, structured for feeding on benthos. Foraging generally occurs in areas with abundant macrophytes, where juvenile and adult shortnose sturgeon f eed on amphipods, polychaetes, and gasteropods (Dadswell et al. 1984b, Moser and Ross 1995, NMFS 1998a). Starting as larvae sturgeon use electroreception to identify prey. Olfaction and taste are also like ly important to foraging, while vision is thought to play a mi nor role (Miller 2004). As adults, a significant portion of a shortnose sturgeon’s diet may consis t of freshwater mollusks (Dadswell et al. 1984b). Based on observations by Kynard et al. (2000), shortnose sturgeon will consume the entire mollusk, excreting the shell after ingestion.

Life span. Shortnose sturgeon in the northern porti on of the species’ range live longer than individuals in the southern portion of the spec ies’ range (Gilbert 1989a). The maximum age reported for a shortnose sturgeon in the St. John River in New Brunswick is 67 years (for a female), 40 years for the Kennebec River, 37 years for the Hudson River, 34 years in the Connecticut River, 20 years in th e Pee Dee River, and 10 years in the Altamaha River (Dadswell

et al. 1984b). Male shortnose sturge on appear to have shorter lif e spans than females (Gilbert 1989a).

Status and trends. Shortnose sturgeon were listed as endangered on March 11, 1967, under the Endangered Species Pr eservation Act (32 FR 4001) and remained on the endangered species list

with enactment of the ESA of 1973, as amended. Pollution and ove rfishing, includin g bycatch in the shad fishery, were listed as pr incipal reasons for the species' decline. Shortnose sturgeon are listed as an endangered species throughout all of their range.

Northern shortnose sturgeon population abundan ces are generally larger than southern populations (Kynard 1997b). Update d population estimates also suggest that three of the largest populations (Kennebec, Hudson, and Delaware Rive r) may be increasing or stable, although data is limited. The New York (Hudson River) shortnose sturgeon population is the largest extant population of this species and base d on available data exhibits a ppears to have increased (NMFS 1998b, Bain et al. 2000). The most recent population estim ate indicates this population consists of about 61,000-shortnose sturgeon (95% conf idence interval [CI] was between 52,898 and 72,191 fish (Bain et al. 2000)). A comparison of the Ba in estimate to the 1979/1980 population estimate of spawning adults by Dovel et al. (1992); about 13,0 00 fish) led Bain et al. (2000) to conclude that the population had made a drama tic increase (about 400 % increase) between 1979

and 1997. While still evidence of an increasi ng population, a comparison of total population estimates (30,000:60,000) would sugge st the population has only doubled in size during the study years. Similarly, the Kennebec River populat ion appears to be increasing. E arly estimates suggest that the Kennebec River contained an estimated 7,200 adult shortnose sturgeon in 1977- 81 (Squiers et al. 1982), while the most recent estimate for this population is about 9,500 fish (Squiers 2003), suggesting the popul ation has increased by about 30 % in about a twenty year period.194 Data from the Delaware River, suggests that the population may be stable. Brundage and O’Herron (2006) estimate that th e current population for the Delawa re River is 12,047 adult fish (1999-2003; 95% CI: 10,757-13,589), which is si milar to the 1981/84 estimate by Hastings et al. (1987) of 12,796 fish (95% CI: 10,288-16367). The recent capture of several fish that were tagged as adults by Hastings et al. (1987) suggests that older fish may comprise a substantial portion of the Delaware River population. Despite their longev ity, the viability of sturgeon populations is sensitive to variability in juvenile recruitment and survival (Anders et al. 2002,

Gross et al. 2002, Secor et al. 2002). Although intera nnual variation in ju venile recruitment would be expected as a result of stochastic f actors that influence spawning and egg/larval survival, if the mean population size does not change over the long- term it then it would appear there is sufficient juvenile surviv al to provide at least periodic recruitment into the adult age classes. Data on juvenile recruitment or age-1+ survival would, however, establish whether this population is at a st able equilibrium.

South of Chesapeake Bay, populations are rela tively small compared to their northern counterparts. The largest of the southern populations of shortnose sturgeon is the Altamaha River population. Population estimates have been calculated several times for sturgeon in the Altamaha since 1993. Total population estimates s hown pretty sizeable interannual variation is occurring; estimates have ranged from as low as 468 fish in 1993 to over 6,300 fish in 2006 (NMFS 1998a, DeVries 2006). The Ogeechee River is the next most studied river south of Chesapeake Bay, and abundance estimates indicate that the shortnose sturgeon population in this river is considerably sm aller than that in the Altamaha Rive r. The highest point estimate in 1993 using a modified Schnabel technique resulted in a total population es timate of 361 shortnose sturgeon (95% CI: 326-400). In contrast the most recent survey resulted in an estimate of 147 shortnose sturgeon (95% CI: 104- 249), suggesting that the popul ation may be declining.

Annual variation in population estimat es in many basins is due to ch anges in yearly capture rates, which are strongly correlated with weather conditi ons (river flow and water temperatures). In “dry years” fish move into deep holes upriver of the saltwater/freshwater interface, which can make them more susceptible to gillnet sampling. Consequently, rivers with limited data sets among years and limited sampling periods within a year may not offer a realistic representation

of the size or trend of the shor tnose sturgeon population in the ba sin. As a whole, the data on shortnose sturgeon populations is rather limited and some of the differences observed between years may be an artifact of the models and assu mptions used by the vari ous studies. Long-term data sets and an open population model woul d likely provide for more accurate population estimates across the specie s range, and could prov ide the opportunity to mo re closely link strongyear classes to ha bitat conditions.

Throughout the specie’s range there are other extant populations, or at least evidence that several other basins are used periodically. That is, shortnose sturgeon have been docum ented in the St. John’s River (FL), the St. Mary’s River, Chesapea ke Bay, Potomac River, Piscataqua River, the Housatonic River, and others. Some basins probably previously contained shortnose populations, but recent sampling has been la rgely unsuccessful. Despite the occasional observations of shortnose sturge on, populations may be extinct in several basins (e.g., St. John’s (FL), St. Mary’s, Potomac, Housatonic, and Neus e rivers). Those few fish that have been observed in these basins are generally presumed to be immigrants from neighboring basins. In195 some cases, (e.g. Chesapeake Bay) migratory inform ation collected from tagged fish and genetic evidence confirms that fish captured in Chesap eake Bay were part of the Delaware River population (Grunwald et al. 2002, Wirgin et al. 2005b).

Natural threats. Yellow perch, sharks, and seals are pred ators of shortnose sturgeon juveniles (NMFS 1998a). The effects of diseas e and parasites are generally unknown.

Anthropogenic threats. Shortnose sturgeon have declined fr om the combined effects from the construction of hydropower and water diversion projects, dredging and blasting, water pollution, fisheries, and hatcheries. The construction of dams has resulted in substantial loss of shortnose sturgeon habitat along the Atlantic seaboard. In many cases da ms divide shortnose sturgeon spawning habitat (e.g., Connecticut River, Penobs cot River) and impede passage or block it completely. Where it has occurred, remediation measures, such as obstruction removal or modification to allow for fish passage have improved shortnose sturgeon habitat and likely improved productivity and more such modifications are planned in certain basins. For instance, with the breaching of the Bangor Dam in the Pen obscot River in 1977 five river kilometers were opened to sturgeon and other anad romous fishes. With the recen t signing of the Penobscot River Restoration Trust, access may be restor ed to another 29 km of habitat.

Historic fishery harvests, as well as the incidental harvest in current fisheries, have had lasting effects on shortnose sturgeon. In the late nineteenth and early twentie th centuries shortnose sturgeon commonly were harvested incidental to Atlantic sturgeon, the larger and more commercially valuable of these two sympatric sturgeon species (NMFS 1998a). The effects of these harvests may have latent and long-lasting impacts on some populations. At present there is no legal directed fishing effort for shortnose sturgeon in the United States, although some illegal poaching is suspected. Additionally, shortnose sturgeon are often caught incidental to other fisheries. For instance, shortnose are caught in cidentally by bass anglers, and incidentally to alewife/gaspereau and shad fisheries in the St. John’s River in Canada, shad fisheries in the Altamaha River, Hudson River, and othe rs (COSEWIC 2005, Bahn and Peterson 2009).

Habitat alterations from discharg es, dredging or disposal of mate rial into waterways, and other developmental activities along riverine and es tuarine systems threaten shortnose sturgeon

habitat. Periodic maintenance of harbors and rivers likely results in the direct take of some sturgeon, but perhaps of greater impact is the manner in which dredging alters benthic topography and community structur e, and water quality (increas e in suspended sediments). Shoreline development of liquefied natural gas facilities and alternative power sources also alters coastal habitats through changes in benthic co mmunities by dredging, changes in water quality and water temperatures, and may increase the poten tial of ship strikes. In the Bay of Fundy, a

tidal turbine killed at least three Atlantic salmon in the 1980s , and may be a threat to shortnose sturgeon as well (Dadswell and Rulifson 1994). Although currently the only example of this type of turbine in North Ameri ca, increasing interests in findi ng alternative energy sources is expected to result in an increase in the num ber of marine turbines along the coast.

Fish kills have also been observed where es tuaries are affected by urban and agricultural discharges that cause vegetative blooms and eutrophic conditions. Extreme declines in dissolved oxygen levels have occurred period ically throughout the species ra nge. In the late 1960s and196 early 1970s, dissolved oxygen levels reached zero ppm in the Penobscot, Kennebec, and Androscoggin rivers and estuaries during the su mmer. Extreme low dissolved oxygen levels are have also plagued Chesapeake Bay. In most cas es, dissolved oxygen levels have improved since through improved treatment and control of waste discharges in the past twenty years, but degraded conditions of benthos are still common in many estuaries throughout the species range as a result of this historic loading of organic materials, waste, and legacy toxins such as dioxin. As a result, shortnose sturgeon a nd other benthic organisms are regul arly in direct contact with legacy pollutants, as well as a suite of common contaminants a dded from more current industrial and agricultural practices. Studies demonstrate that shortnose sturgeon ca rry a wide number of potentially hazardous contaminants. Individual s from the Delaware River contain numerous metals (mercury, aluminum, antimony, barium, cadmium, calcium, chromium, copper, iron, magnesium, manganese, nickel, potassium, s odium, vanadium, and zinc), PCDDs, PCDFs, PCBs, DDE, DDD, bis (2-ethylhexyl) phthalate, di -n-butylphthalate, and chlordane (ERC 2002). Most of these metals, PCDDs, PCDFs, and PCBs were also found in shortnose sturgeon in the Kennebec River (ERC 2003).

Climate change has the potential to affect sturgeon in similar, if not more significant, ways that it affects salmonids. Elevated air temperatures coul d lead to precipitation fa lling as rain instead of snow. Additionally, snow would likely melt soone r and more rapidly, potentially leading to greater flooding during melting and lower water levels at other ti mes, as well as warmer river temperatures (ISAB 2007c). Although sturgeon can spawn over varied benthic habitat, they prefer localized depressi ons in riverbeds (Moyle et al. 1992a, Moyle et al. 1995a, Erickson et al. 2001b, Rien et al. 2001a). Increased extremes in river flow (i.e., periods of flooding and low flow) can alternatively disrupt and fill in spaw ning habitat that sturge on rely upon (ISAB 2007c). If water flow is low during migration events, it is likely that new obstacles can impede or block sturgeon movement. As with other anadromous fishes, sturgeon are uniquely evolved to the environments that they live in. Because of this specificity, broad scale changes in environment can be difficult to adapt to, including changes in water temperature (Cech Jr. et al. 2000).

Sturgeon are also directly sens itive to elevated water temperatures. Temperature triggers spawning behavior. Warmer water temperatures can initial spawning earlier in a season for salmon and the same can be true for sturgeon (I SAB 2007c). If river and lake temperatures become anomalously warm, juvenile sturgeon may experience elevated mortality due to lack of cooler water refuges in freshwater habitats. If temperature rise beyond thermal limits for extended periods, habitat can be lost; this could be the case if southern habitats warm, resulting in range loss(Lassalle et al. 2010). Apart from direct changes to sturgeon survival, altered water temperatures may disrupt habita t, including the availability of prey (ISAB 2007c). Warmer temperatures may also have the effect of incr easing water use in agriculture, both for existing

fields and the establishment of new ones in once unprofitable areas (ISAB 2007c). This means that streams, rivers, and lakes will experience additional withdrawal of water for irrigation and increasing contaminant loads from returning efflue nt. Overall, it is likely that global warming will increase pressures on sturgeon survival and recovery.

Critical habitat. Critical habitat has not been es tablished for shortnose sturgeon.

Smalltooth sawfish197 Description of the species. The smalltooth sawfish is a tropical marine and estuarine elasmobranch fish (sharks and rays) that has been reported to have a circ umtropical distribution. Although they are rays, sawfish physically more resemble sharks, with only the trunk and especially the head ventrally flattened. Smallto oth sawfish are characterized by their “saw,” a long, narrow, flattened rostral bl ade with a series of transv erse teeth along either edge.

Distribution. In the western Atlantic, the smalltoot h sawfish has been reported from Brazil through the Caribbean and Central America, the Gulf of Mexico, and the Atlantic coast of the US. The smalltooth sawfish has also been recorded from Bermuda (Bigelow and Schroeder 1953). Forms of smalltooth sawfish have been reported from the eastern Atlantic in Europe and West Africa; the Mediterranean; South Africa; and the Indo-West Pacific, including the Red Sea, India, Burma, and the Philippines (Bigelow and Schroeder 1953, Van der Elst 1981, Compagno and Cook 1995). Whether populations outside of th e Atlantic are truly smalltooth sawfish or closely related species is unknown (Adams a nd Wilson 1995). Pacific coast records of

smalltooth sawfish off Central America need confirmation (Bigelow and Schroeder 1953, Compagno and Cook 1995).

The range of the smalltooth sawfish in the Atlantic has contracted markedly over the past century. The northwestern terminus of their Atlantic range is located in the waters of the eastern US. Historic capture records within the US range from Texas to New York. Water temperatures no lower than 61°F to 64.4°F and the availability of appropriate coastal hab itat serve as the major environmental constraints limiting the northern movements of smalltooth sawfish in the western North Atlantic (Simpfendorfer 2001). As a result, most records of this species from areas north of Florida occur during spring a nd summer periods (May to Augus t) when inshore waters reach appropriately high temperatures. The data also suggest that smalltooth sawfish may utilize warm water outflows of power stations as thermal re fuges during colder months to enhance their survival or become trapped by surrounding cold wa ter from which they would normally migrate.

Almost all occurrences of smalltooth sawfish in warm-water outflows were during the coldest part of the year, when water temperatures in these outfalls are typi cally well above ambient
temperatures.

Movement. Historic records of smalltooth sawfish indi cate that some large mature individuals migrated north along the US Atlantic coast as temperatures warmed in the summer and then south as temperatures cooled (Bigelow and Schroeder 1953). Recent Florida encounter data, however, do not suggest such migration. Only tw o smalltooth sawfish have been recorded north of Florida since 1963 (the first was captured o ff of North Carolina in 1999 and the other off Georgia 2002) but it is unknow n whether these individuals re sided in Georgia and North Carolina waters annually or if they had migr ated north from Florida (Schwartz 2003b, Burgess unpublished data). Given the very limited number of encounter reports from the east coast of Florida, Simpfendorfer and Wiley (2004) hypothesi ze the population previously undertaking the summer migration has declined to a point where th e migration is undetectable or does not occur.

Most specimens captured along the Atlantic coast nor th of Florida have also been large (>2.7 m) adults and likely represent seasonal migrators, wa nderers, or colonizers fro m a core population(s) to the south rather than being members of a continuous, even-density population (Bigelow and Schroeder 1953). It is likely th at these individuals migrated s outhward toward Florida as water198 temperatures declined in the fall, as there is only one winter record from the Atlantic coast north

of Florida. Based on smalltooth sawfish encounter data, the current core range for the smalltooth sawfish is from the Caloosahatchee River, Fl orida, to Florida Bay (NMFS 2000, Simpfendorfer and Wiley 2004). 3

Habitat. Smalltooth sawfish are euryhaline, occurring in waters with a broa d range of salinities

from freshwater to full seaw ater (Simpfendorfer 2001). Younger, smaller individuals tend to inhabit very shallow mud banks and tides are a major factor in their movement (Simpfendorfer et
al. 2010). At this size smalltooth sawfish spend the vast majority of their time on shallow mud or sand banks that are less than 1 foot (30 cm ) deep. As they grow, juveniles tend to occupy deeper habitat, but shallow areas (<1 m depth) remain preferre d habitat; juveniles also expand their ranges, whereas small individuals have very rest ricted ranges (Simpfendorfer et al. 2010). Acoustic tracking studies have shown that at this size sawfish will remain associated with the same mud bank over periods of several days. Th ese banks are often very small and daily home range sizes can be of the magnitude of 100–1,000 m 2 (Simpfendorfer 2003b). Acoustic monitoring studies have shown that juveniles have high levels of site fidelity for specific nursery areas for periods up to almost 3 months (Wile y and Simpfendorfer 2007). Their occurrence in freshwater is suspected to be only in estuarin e areas temporarily freshw ater from receiving high levels of freshwater input. Ma ny encounters are reported at the mout hs of rivers or other sources of freshwater inflows, suggesting estuarine area s may be an important factor in the species distribution (Simpfendorfer and Wiley 2004).

Information on juvenile smalltooth sawfish indicate s that they prefer shallow euryhaline habitats adjacent to red mangroves (NMFS 2006h). They do still have a preference for shallow water, remaining in depths mostly less than 90 cm. Several sawfish approximately 150 cm in length fitted with acoustic tags have b een relocated in the same general areas over periods of several months, suggesting a high level of site fideli ty (Simpfendorfer 2003). The daily home ranges of

these animals are consid erably larger (1–5 km 2 ) than for the very small sawfish and there is less overlap in home ranges between days.

Smalltooth sawfish are most common in shallow coastal waters less than 25 m (Bigelow and Schroeder 1953, Adams and Wilson 1995). Indeed, th e distribution of the smallest size classes of smalltooth sawfish indicate that nursery areas occur throughout Florida in areas of shallow water, close to shore and typically associated with mangroves (Simpfendorfer and Wiley 2004). However, encounter data indicate there is a tendency for smalltooth sawfish to move offshore and into deeper water as they grow. Larger anim als are more likely to be found in deeper waters. Poulakis and Seitz (2004b) reporte d that almost all of the sawfish <3 m in length were found in water less than 10 m deep and 46% of encounter s individuals >3 m in Florida Bay and the Florida Keys were reported at depths between 70 to 122 m. Since large animals are also observed in very shallow waters, it is believed that smaller (younger) animals are restricted to shallow waters, while large animals roam over a much larger depth range (Simpfendorfer 2001). Recent data from sawfish encounter reports and from satellite tagging indicate mature animals

3 See the 2006 Draft Recovery Plan for more detailed information on the historic and current distribution of smalltooth sawfish in four regions of the eastern U.S. This information is based on the Status Review Team’s analysis and the more recent encounter database resear ch {Poulakis, 2004 #1266;Seitz, 2002 #1426;Simpfendorfer, 2004 #1465;Simpfendorfer, 2004 #1465}.199 occur regularly in waters in excess of 164 f eet (Poulakis and Seitz 2004a, Simpfendorfer and Wiley 2004).

Growth and reproduction. As in all elasmobranchs, fertilization is internal. Bigelow and Schroeder (1953) report the litter size as 15 to 20. Simpfendorfer a nd Wiley (2004), however, caution this may be an overestimate, with recent anecdotal information su ggesting smaller litter sizes (about ten). Smalltoot h sawfish mating and pupping seas ons, gestation, and reproductive periodicity are all unknown. Gestation and repr oductive periodicity, however, may be inferred based on that of the largetooth sawfish, sharing th e same genus and having similarities in size and habitat. Thorson (1976) reported the ge station period for largetooth sawfish was approximately five months and concluded that females probably produce litters every second

year.

Bigelow and Schroeder (1953) describe smalltoot h sawfish as generally about 61 cm long at birth and growing to a length of 5.5 m or great er. Recent data from smalltooth sawfish caught off Florida, however, demonstrate young are bor n at 76 to 87 cm (Simpfendorfer and Wiley 2004), with males reaching maturity at approxim ately 2.7 m and females at approximately 3.6 m (Simpfendorfer 2002). A recent study by Simpfendorfe r suggests rapid juvenile growth occurs during the first two years after birth (Simpfendo rfer 2008). First year growth is 65-85 cm and second year growth is 48-68 cm. Growth ra tes beyond two years are uncertain; however, the average growth rate of captive smalltooth sa wfish has been reported between 13.9 and 19.6 cm per year. The maximum reported size of a sm alltooth sawfish is 24.9 feet (Last and Stevens 1994), but the maximum size normally observed is 7.6 m (Adams and Wilson 1995). No formal studies on the age and growth of the smalltooth sa wfish have been conducted to date, but growth studies of largetooth sawfish sugge st slow growth, late maturity (10 years) and long lifespan (25 to 30 years Thorson 1982, Simpfendorfer 2000b). These characteristics suggest a very low

intrinsic rate of incr ease (Simpfendorfer 2000b).

Simpfendorfer estimated intrinsic rates of natura l population increase at 0.08 to 0.13 per year and population doubling times from 5.4 to 8.5 years (Simpfendorfer 2000a).

Feeding. Smalltooth sawfish feed primarily on fish, w ith mullet, jacks, and ladyfish believed to be their primary food resources (Simpfendorfer 2001) . In addition to fish, smalltooth sawfish also prey on crustaceans (mostly shrimp and crabs), which are located by disturbing bottom sediment with their saw (Norman and Fraser 1937, Bigelow and Schroeder 1953).

Status and trends. The US smalltooth sawfish distinct population segment (DPS) was listed as endangered under the ESA on April 1, 2003 (68 FR 15674). The smalltooth sawfish is the first marine fish to be listed in the US. Despite being widely recognized as common throughout their historic range up until the middle of the 20 th century, the smalltooth sawfish population declined dramatically during the middle a nd later parts of the century. The decline in the population of smalltooth sawfish is attributed to fishing (both commercial and recreational), habitat modification, and sawfish life history. Large numbers of smalltooth sawfish were caught as bycatch in the early part of this century. Sma lltooth sawfish were histor ically caught as bycatch in various fishing gears throughout their historic range, including gillnet, otter trawl, trammel net, seine, and to a lesser degr ee, handline. Frequent accounts in earlier literature document200 smalltooth sawfish being entangled in fishing ne ts from areas where smalltooth sawfish were once common but are now rare (Evermann and Bean 1898). Loss and/or degradation of habitat

contributed to the decline of ma ny marine species and continue to impact the distribution and abundance of smalltooth sawfish. Simpfendorfer ( 2001) estimated that the US population size is currently less than 5% of its size at the time of European settlement. One dataset from shrimp

trawlers off Louisiana from the late 1940s thro ugh the 1970s suggests a rapid decline in the species from the period 1950-1964 (NMFS 2006h).

Seitz and Poulakis (2002) and Poulakis and Seitz (2004a) doc umented recent (1990 to 2002) occurrences of sawfish along the s outhwest coast of Florida, and in Florida Bay and the Florida Keys, respectively and includes a total of 2,969 sm alltooth sawfish encounters. Mote Marine Laboratory also maintains a smalltooth sawfish pub lic encounter database, established in 2000 to compile information on the distribution and a bundance of sawfish. A total of 434 sawfish

encounters have been validate d since 1998, most from recreationa l fishers (Simpfendorfer and Wiley 2004). Dr. Simpfendorfer reluctantly gives an estimate of 2,000 individuals based on his four years of field experience and data collect ed from the public, but cautions that actual numbers may be plus or minus at least 50%.

The majority of smalltooth sawfish encounters today are from the southwest coast of Florida between the Caloosahatchee River and Florida Bay. Outside of this core area, the smalltooth sawfish appears more common on the west coast of Florida and in the Florida Keys than on the east coast, and occurrences decr ease the greater the distance from the core area (Simpfendorfer and Wiley 2004). The capture of a smalltooth sa wfish off Georgia in 2002 is the first record north of Florida since 1963. New reports during 2004 extend the current ra nge of the species to Panama City, offshore Louisiana (south of Timbalie r Island in 100 feet of water), southern Texas (unconfirmed), and the northern coast of Cuba.

The abundance of juveniles enc ountered, including very small in dividuals, suggests that the population remains reproductively active and viab le (Seitz and Poulakis 2002, Simpfendorfer 2003a, Simpfendorfer and Wiley 2004). The declini ng numbers of individuals with increasing

size is consistent with the hi storic size composition data (G. Burgess, pers. comm. in Simpfendorfer and Wiley 2004). This informa tion and recent encounters in new areas beyond the core abundance area sugge st that the population may be increasing. From 1989-2004,

smalltooth sawfish relative abundance has incr eased by about 5 percen t per year (Carlson et al. 2007). However, recovery of the species expected to be slow on the basi s of the species’ life history and other threats to the species remain ing (see below), the popula tion’s future remains tenuous. Based on genetic sampling, the estimates of current effective population size are between 269.6 and 504.9 individuals (95% CI 139.3 – 1,515). (NMFS 2011l). This number is usually 25-50% of census population size (breeding adults) in elasmobranchs, so it is likely that the breeding population consists of high hundred s to low thousands of individuals (NMFS 2011l).

Natural threats. The primary natural threat to smalltooth sawfish survival is the species low reproductive rate. In the face of reduced populati on sizes, this biological parameter means that recovery, at best, will be slow, and that catastr ophic perturbations can have severer consequences
to recovery.201 Anthropogenic threats. Smalltooth sawfish decline has been largely due to fisheries interaction (see NMFS 2006g for a review). The distinctive “saw ” can easily become entangled in a variety of commercial and recreational fishing gear, re sulting in drowning or injury. Even when

individuals that have been enta ngled are retrieved alive, indivi duals may be killed for curio collection of the saw, fear of injury from fisherman, or injured from the gear or handling during gear removal. However, additional anthropogenic impacts result from habitat loss. Destruction of mangrove habitat, dredging, trawling and filling, and loss of reef habitat have negative impacts on all life stages of smalltooth sa wfish. Although a concern, pollution impacts on particularly reproductive biol ogy are unknown. However, habita t degradation due to runoff containing pesticides, eutrophying agents, and other contaminants can also have a negative impact on smalltooth sawfish habitat.

Critical habitat. On September 2, 2009, critical habitat was designated for smalltooth sawfish along the central and so uthwest coast of Florida (74 FR 45353). Although PCEs were not identified, the mangrove and adjacent shallow eu ryhaline habitat are impor tant nursery habitat

for smalltooth sawfish.

Bocaccio Description of the species. The bocaccio is a rockfish species that genetic analyses suggest is composed of two distinct populations (Wishard et al. 1980, Matala et al. 2004). A southern population exists along the Pacific coasts of Me xican and California and is separated from a northern population by a region of a pparent scarcity from northern California to southern Oregon (MacCall and He 2002b). It has been proposed that oceanographic features, such as current patterns restricting larval movement, are re sponsible for population discreteness (Matala et al. 2004, NMFS 2008d). The northern population is the entity that is proposed for listing. However, the presence of a third population has also been suggested (Queen Charlotte Island, Vancouver Island to Point Conception, Californi a, and south of Poin t Conception)(Matala et al. 2004). For stock management purposes, the NMFS and Pacific Fisheries Management Council recognize these populations as separate stocks.

Distribution. Bocaccio occur from the central Baja peninsula of Mexico north along the continental shelf and slope as far as Stepovac Bay, Alaska (Love et al. 2002).

Habitat and movement. Preferred bocaccio habitat is largel y dependent upon the life stage of an individual. Larvae and young juveniles tend to be found in deeper offshore regions (1-148 km offshore), but associated with the surface and occasionally with floating kelp mats (Hartmann 1987, Love et al. 2002, Emery et al. 2006). As individuals mature into older juveniles and adults, they transition into shallow waters and settle to the bottom, preferring algae-covered rocky, eelgrass, or sand habitats and aggregating into schools (Eschmeyer et al. 1983a, Love et
al. 1991). After a few weeks, fish move into slightly deeper waters of 18-30 m and occupy rocky reefs (Feder et al. 1974, Carr 1983, Eschmeyer et al. 1983a, Johnson 2006, Love and Yoklavich 2008). As adults, bocaccio may be found in depth s of 12-478 m, but tend to remain in

shallow waters on the continental shelf (20-250 m), still associati ng mostly with reefs or other hard substrate, but may move over mud flats (Feder et al. 1974, Kramer and O'Connell 1995,

Love et al. 2002, Love et al. 2005, Love and York 2005, Love et al. 2006). Artific ial habitats,202 such as platform structures, also appear to be suitable habitat for bocaccio (Love and York 2006). Adults may occupy territories of 200-400 hectares, but can venture outside of this territory (Hartmann 1987). Adults tend to occu py deeper waters in the southern population compared to the northern population (Love et al. 2002). Adults are not as benthic as juveniles and may occur as much as 30 m above the bottom and move 100 m vertically during the course of a day as they move between different areas (Love et al. 2002, Starr et al. 2002). Prior to severe population reductions, bocaccio appeared to frequent the Tacoma Narrows in Washington State (DeLacy et al. 1964, Haw and Buckley 1971, Miller and Borton 1980).

Reproduction. Bocaccio are live-bearers with internal fertilization. Once females become mature (at 54-61 cm total length), they produce 20,000-2.3 million eggs annually, with the number increasing as females age and gr ow larger (Hart 1973b, Echeverria 1987, Love et al. 2002). However, either sex has been known to at tain sexual maturity as small as 35 cm or 3 years of age and, in recent year s as populations have declined, average age at sexual maturity may have declined as well (Hart 1973b, Echeverria 1987, Love et al. 2002, MacCall 2002b). Mating occurs between August and November, with larvae bor n between January and April (Lyubimova 1965, Moser 1967, Westrheim 1975, Wyllie Echeverria 1987, Love et al. 2002, MacCall and He 2002b).

Growth. Upon birth, bocaccio larvae measure 4-5 mm in length. These larvae move into pelagic waters as juveniles when they are 1.5- 3 cm and remain in oceanic waters from 3.5-5.5 months after birth (usually until early June), where they grow at ~0.5-1 mm per day (Moser 1967, Matarese et al. 1989, Woodbury and Ralston 1991, Love et al. 2002, MacCall and He 2002b, MacCall 2003). However, growth can vary from year-to-year (Woodbury and Ralston 1991). Once individuals are 3-4 cm in length, they return to nearshore wa ters, where they settle into bottom habitats. Females tend to grow faster than males, but fish may take 5 years to reach sexual maturity (MacCall 2003). In dividuals continue to grow until they reach maximum sizes of 91 cm, or 9.6 kg, at an estimated maximum age of 50 years (Eschmeyer et al. 1983a, Halstead

et al. 1990, Ralston and Ianelli 1998, Love et al. 2002, Andrews et al. 2005, Piner et al. 2006). However, individuals tend to grow larg er in more northerly regions (Dark et al. 1983).

Foraging. Prey of bocaccio vary with fish age, with bocaccio larvae starting with larval krill, diatoms, and dinoflagellates (Love et al. 2002). Pelagic juvenile s consume fish larvae, copepods, and krill, while older, nearshore juve niles and adults prey upon rockfishes, hake, sablefish, anchovies, lanter nfish, and squid (Reilly et al. 1992, Love et al. 2002).

Acoustics and hearing. Data regarding bocaccio hearing ar e not availab le. However, field measurements have recorded bocaccio calls, whic h are more prevalent at night, as <900 Hz repetitive pulses of ~0.1 s duration (Sirovic et al. 2009).

Status and trends. Bocaccio were proposed for listing on April 23, 2009 (74 FR 18516). Bocaccio as a species has undergone severe decline in the past several decades, with the species currently estimated to be 3.6% of its abunda nce in 1970 (MacCall and He 2002b). Prior to World War II, commercial landings of rockfish species generally remained under 20,000 lbs, but sky-rocketed during the war to 375,000 lbs annually and fluctuated between 50,000 and 220,000 lbs until 1970, when landings increased linearly with fishing effort to a peak of 900,000 lbs by203 1980 (Palsson et al. 2008). Levels fluctuated after this between 48,000 and 300,000 lbs for the next decade and clearly crashed in the 1990’s, with landings below 30,000 lbs annually. At the cessation of commercial fishing in 2003, 2,600 lbs of rockfish were harvested. Similar trends are seen in recreational landings from Puget Sound (WDF 1975-1986).

Among rockfish of the Puget Sound, bocaccio appear to have undergone a particular decline (MacCall and He 2002b). This has likely because of the removal of the largest, most fecund individuals of the population due to overfishing and the frequent fa ilure of recrui tment classes, possibly because of unfavorable climactic/o ceanographic conditions (MacCall and He 2002b).

Bocaccio resistance to depletion and recovery is also hindered by demographic features (Love et
al. 1998a). Bocaccio are long-lived fishes, taking several years to reach sexual maturity and becoming more fecund with age (Dorn 2002). As harvesting targeted the largest individuals available, bocaccio have become less capab le of recovering population numbers (Love et al. 1998b). At present, in the complete absence of directed or bycatch fishing pressure, it is estimated that bocaccio populations would have to have frequent good recruitment to restrain their present decline (Tolimieri and Levin 2005) . In addition, bocaccio reproduction appears to be characterized by frequent recruitment failu res, punctuated by occasional high success years (Love et al. 1998b, MacCall and He 2002b). Over the past 30 years, 1977, 1984, and 1988 are the only years in which recruitment appears to have been significant successes (it should be noted that 1999 and 2002 also appear to have been strong, but survivorship into maturity is still pending). Recruitment success appears to be linked to oceanographic/climactic patterns and may be related to cyclic warm/cool ocean periods, w ith cool periods having greater success (Sakuma and Ralston 1995, MacCall 1996, Love et al. 1998b, Moser et al. 2000b). Harvey et al. (2006) suggested that bocaccio may have recently dive rted resources from reproduction, potentially resulting in additional impairment to recovery. Overall, bocaccio have the highest variability of recruitment of any rockfish studied to date, w ith recruitment exhibiting a random walk and high temporal variability (MacCall a nd He 2002b, Tolimieri and Levin 2005).

Although population estimates are not availabl e for the northern population, the southern population has been estimated to number 1.6 milli on fish of 1 year of age or older in 2002 (MacCall 2002a). Of these, 1.0 million were estim ated to occur south of Pt. Conception, where recruitment has been stronger. However, indivi duals north of Pt. Conception tend to be larger and, hence, more fecund. In 2002, the southern population was estimated to produce 720 billion eggs annually (243 billion south of Pt. Conception). North of Pt. Conception, bocaccio are most abundant in the Monterey Bay area, where prime habitat seems to be over the continental slope and, secondarily, over the shelf (Dark et al. 1983).

The rate of decline for rockfish in Puget S ound has been estimated at ~3% annually for the period 1965-2007. Various rebuilding estimates for bocaccio populations have predicted recovery, but require long pe riods (98-170 years) and assume no mortality from fishing (intentional harvests are closed, but bycatch still occurs)(MacCall and He 2002a, MacCall 2008, NMFS 2008d).

Natural threats. Interspecies competiti on, predators, and climactic regimes are the primary natural factors that depress bo caccio numbers. Copper and quillb ack rockfish may compete with204 bocaccio in Puget Sound for available resources (NMFS 2008d). King salmon, lingcod, terns and other seabirds, harbor seals, and Steller se a lions are known predators of bocaccio and other rockfish species (Love et al. 2002, Beaudreau and Essington 2007, Lance and Jeffries 2007). Bocaccio and other rockfish appear to be negati vely influenced by El Niño conditions, possibly reducing available prey supply (Moser et al. 2000a, Harvey 2005).

Anthropogenic threats. Although overfishing is the prim ary reason for bocaccio being proposed as a listed species, bycat ch and habitat loss are also hu man-related factors that have likely led to bocaccio decline. Although a frequent species captured in fisheries during the late 1970’s, bocaccio were not recorded from any recreational surveys from 1996-2007 (WDF 1975- 1986, Palsson et al. 2008). Apart from commercial fishi ng, recreational fishing (even catch-andrelease) appears to incur significant mortality on bocaccio and other rockfishes (Schroeder and Love 2002). The species is considered overfished by the Pacific Fisheries Management Council and is not presently harvested in tentionally. However, bycatch is still considered to be a high impact stressor to rockfish populations of Washington State waters (Palsson et al. 2008).

Habitat loss is also a factor in bocaccio decline, with rocky ha bitats (reportedly, there are only

217 km 2 in Puget Sound) being threatened by constr uction of bridges, sewer lines, cable and pipeline deployment, and dredge spoil (Palsson et al. 2008). Loss of kelp, which is valuable to juvenile fish recruitment, as well as anoxic conditions, exac erbate habitat loss (NMFS 2008d).

Critical habitat. Critical habitat has not been proposed or designated for the bocaccio.

Yelloweye rockfish Description of the species. Yelloweye rockfish are likely co mposed of at least two populations and possibly more. Yamanaka et al. (2006) found that those i ndividuals found within the Georgia Basin and Queen Charlott e Strait were genetically dis tinct from other samples from Oregon to Alaska. The Georgi a Basin/Queen Charlotte Sound population is the one which has been proposed for listing in US waters.

Distribution. Yelloweye rockfish occur from Baja Ca lifornia to the Aleutian Islands, but are

most common from central Ca lifornia to Alaska (Love et al. 2002).

Habitat. As with other rockfishes, yelloweye hab itat varies based upon life stage. Larvae maintain a pelagic existence but as juveniles, move into sha llow high relief rocky or sponge garden habitats (Eschmeyer et al. 1983a, Richards et al. 1985, Love et al. 1991). Juveniles may also associate with floating de bris or pilings (Lamb and Edge ll 1986). As adults, yelloweye rockfish move in to deeper habitats. Individual s have been found in waters as deep as 549 m, but are generally found in waters of less than 180 m (Eschmeyer et al. 1983a, Love et al. 2002). However, adults continue to associate with rocky, high relief habitats, par ticularly with caves and crevices, pinnacles, and boulder fields (C arlson and Straty 1981, Richards 1986, Love et al. 1991, O'Connell and Carlisle 1993, Yoklavich et al. 2000). Yelloweyes generally occur as individuals, with loose, re sidential aggregations infre quently found (Coombs 1979, DeMott 1983, Love et al. 2002). In the Puget Sound region, spor t catch records from the 1970’s indicate

that Sucia Island and other islands of the San Juans as well as Bellingham Bay had the highest concentrations of catches (Delacy et al. 1972, Miller and Borton 1980).205 Reproduction. Yelloweye rockfish are live bearers wi th internal fertilization. Copulation occurs between September and April, with ferti lization taking place later as latitude increases (Hitz 1962, DeLacy et al. 1964, Westrheim 1975, O'Connell 1987, Wyllie Echeverria 1987, Lea

et al. 1999). Puget Sound yelloweyes mate between winter and summer, giving birth from spring to late summer (Washington et al. 1978). Gestation lasts roughly 30 days (Eldridge et al. 2002). Although yelloweye rockfish were once be lieved to reproduce annually, evidence exists that indicate the potential for multiple births per year (MacGregor 1970, Washington et al. 1978). Females produce more eggs as they grow older and larger , with each individual producing roughly 300 eggs per year per gram of body weight (1.2-2.7 million eggs per year)(MacGregor 1970, Hart 1973b). In addition, older females of several rockfish species may be capable of provisioning their offspring better than their younger counterparts, meaning that they may be more a more influential component in a given year’s recruitment success (Sogard et
al. 2008).

Growth and development. Larvae are born at 4-5 mm in length and maintain a pelagic

existence for the first 2 months of life, before moving to nears hore habitats and settling into rocky reef habitat at ab out 25 mm in length (DeLacy et al. 1964, Matarese et al. 1989, Moser 1996a, Love et al. 2002). Yelloweye growth is thought to vary by la titudinal gradient, with individuals in more northerly regions growing faster a nd larger. Year cla ss strength appears to be most strongly linked to surviv al of the larval stage (Laidig et al. 2007). In general, sexual maturity appears to be reached by 50% of indi viduals by 15-20 years of age and 40-50 cm in length (Yamanaka and Kronlund 1997) . As with other rockfish, yelloweyes can be long-lived (reported oldest age is 118 y ears)(Munk 2001). Maximum size has been reported as 910 cm, but assymptotic size in Alaskan waters for both male s and females was estimated to be 690 cm and 659-676 mm along British Columbia (Clemens and Wilby 1961, Westrheim and Harling 1975,

Rosenthal et al. 1982, Love et al. 2005, Yamanaka et al. 2006).

Movement. Individuals shift to deeper habitats as they age. Juve niles tend to begin life in shallow rocky reefs and graduate to deeper rock y habitats as adults. Once adult habitat is established, individuals tend to remain at a particular site (Love 1978, Coombs 1979, DeMott 1983).

Foraging. As with other rockfish species, yelloweye rockfish prey upon different species and size classes throughout their development. Larval and juvenile rockfish prey upon phyto- and zooplankton (Lee and Sampson 2009). Adult yellow eyes eat other rockfi sh (including members of their own species), sand lance, gadids, fl atfishes, shrim p, crabs, and gastropods (Love et al. 2002, Yamanaka et al. 2006).

Status and trends. Yelloweye rockfish were proposed for listing on April 23, 2009 (73 FR 18516). Yelloweye rockfish abundance has been variable in the Puget Sound region over the past 60 years, ranging from less than 1% to grea ter than 3% of samples, although Wallace (2001) documented large historical population in the Strait of Georgia. The latest samples have been historic lows in abundance. Perhaps more importan tly, age classes appear to have been truncated to younger, smaller fish, severely hampering the abil ity of the species to recover from its primary cause of decline: overfishing (Berkeley et al. 2004).206 Prior to World War II, commercial landings of rockfish species generally remained under 20,000 lbs, but sky-rocketed during the war to 375,000 lbs annually and fluctuated between 50,000 and 220,000 lbs until 1970, when landings increased linearl y with fishing effort to a peak of 900,000 lbs by 1980 (Palsson et al. 2008). Levels fluctuated afte r this between 48,000 and 300,000 lbs for the next decade and clearly crashed in th e 1990’s, with landings below 30,000 lbs annually. At the cessation of commercial fishing in 2003, 2,600 lbs of rockfish were harvested. Over the period of 1965-2007, it is estimated that rockfi sh species has declined by 3% per year.

The most recent estimate of yelloweye rockfi sh abundance in the Puget Sound region was 3,000 individuals, with low abundan ce through spawning areas (Palsson et al. 2008).

Natural threats. Interspecies competiti on, predators, and climactic regimes are the primary natural factors that depress yelloweye rockfish numbers. Copper and quillback rockfish may compete with yelloweye rockfish in Puge t Sound for available resources (NMFS 2008d). Lingcod, killer whales, and Steller sea lions are likely predators of yelloweye and other rockfish species (Love et al. 2002, Beaudreau and Essington 2007, Lan ce and Jeffries 2007). Yelloweye and other rockfish appear to be negatively influenced by El Niño conditions, possibly reducing available prey supply (Moser et al. 2000a, Harvey 2005, Black 2009). Oceanographic conditions (such as sea level anomalies and nearshore temper ature conditions) appear to strongly influence the strength of each year ’s recruitment (Laidig et al. 2007). Rates of natural mortality have been reported to range from 2-4.6% annually (Yamanaka and Kronlund 1997, Wallace 2007).

Anthropogenic threats. Overfishing is considered the pr imary cause of yelloweye rockfish decline throughout their range, in cluding in Washington State and British Columbian waters (Wallace 2007, NMFS 2008d). Although commercial harvesting of the species has ended, bycatch is still considered to be a high impact stressor to rockfish populations of Washington State waters (Palsson et al. 2008). It has been estimated that yelloweye rockfish have fallen 30% in abundance within 1/3 of a gene ration in the past few decades, an astonishing rate of decline.

Habitat loss is also a factor in yelloweye decline, with rocky ha bitats (reportedly, there are only

217 km 2 in Puget Sound) being threatened by constr uction of bridges, se wer lines, cable and pipeline deployment, and dredge spoil (Palsson et al. 2008). Anoxic conditions and chemical

contamination are also considered threats to yelloweye rockfish recovery (NMFS 2008d).

Canary rockfish Description of the species. It is unclear how many populatio ns compose canary rockfish as a species. Genetic analysis have found that individuals south of Ca pe Blanco in southern Oregon

lack an allele that individuals north of this point have (Wishard et al. 1980). This has been used to support the proposal of a northern DPS. In addition, canary rockfish are managed as two stocks in Canadian waters (COSEWIC in press) . However, clear evidence of genetically or morphologically distinct populat ions is still lacking.

Distribution. Canary rockfish are found from the north ern Baja peninsula north to the western Gulf of Alaska, and with the greatest abundan ce along British Columbia to central California (Miller and Lea 1972, Hart 1973b, Cailliet et al. 2000, Love et al. 2002).207 Habitat. Canary rockfish occupy a variety of hab itats based upon their life stage. Larvae and younger juveniles tend to occupy sha llow waters at the beginning of their lives, but generally remain in the upper 100 m of the water column (Love et al. 2002). Juveniles in itially settle into tide pools and rocky reefs (M iller and Geibel 1973, Love et al. 1991, Cailliet et al. 2000, Love et
al. 2002). Juveniles have also been observed in diurnal movements, occurring near sand-rock interfaces in groups by day and moving over sandy areas at night (Love et al. 2002). After as much as 3 years, juveniles move into deeper rocky reefs, forming loose schools, rarely on but generally near the bottom (Phillips 1960, Bo ehlert 1980, Lamb and Edgell 1986, Rosenthal et al. 1998, Starr 1998, Cailliet et al. 2000, Johnson et al. 2003, Methot and Stewart 2005, Tissot et al. 2007). Adults may be found in waters of up to 400 m, but tend to be most common in the 80- 200 m range, or even shallower (Mos er 1996b, Methot and Stewart 2005, Tissot et al. 2007). Mid shelf locations seem to have the highest co ncentrations of canary rockfish off Washington and Oregon (Weinberg 1994). Adults tend to occur in shallow areas in highe r latitudes than their southern counterparts, although adul ts do appear to move into pr ogressively deeper waters as they age (Vetter and Lynn 1997, Methot and Stewar t 2005). It is believed that, within Puget Sound, canary rockfish were most common in the 1960’s and 1970’s in Tacoma Narrows, Hood Canal, San Juan Islands, Belli ngham, and Appletree Cove (Delacy et al. 1972, Miller and Borton 1980). A latitudinal gradient may be present by age class, with older and larger individuals preferably occupying more northerly habitat (Dark et al. 1983).

Movement. Individual canary rockfi sh can range widely (up to 700 km over several years), although patterns of residency have been observed (Gascon and Miller 1981, DeMott 1983,

Casillas et al. 1998, Lea et al. 1999, Love et al. 2002). In addition, seasonal movements have been found, with individuals moving from 160-210 m de pths in late winter to 100-170 m in late summer (COSEWIC in press).

Reproduction. Canary rockfish develop their young in ternally before giving birth to live young as larvae. During each annual spawning even t, a female can produce 260,000 to 1.9 million eggs, depending upon her size and age (Guillemot et al. 1985, NMFS 2008d). Unlike some other rockfish, there does not appear to be a latitudinal or geographic gr adient associated with number of eggs produced (Gunderson et al. 1980, Love et al. 2002). Birth takes pl ace in Oregonian and Washingtonian waters between September th roug h March, with a peak in December and January. The peak in British Columbian wate rs is slightly later (February)(Hart 1973b, Westrheim and Harling 1975, Wyllie Echeverria 1987, Barss 1989).

Growth and development. When born, larvae are 3.6-4.0 mm in length and take from 1-4 months to develop into juveniles (Waldr on 1968, Richardson and Laroche 1979, Stahl-Johnson 1985, Moser 1996a, Krigsman 2000, Love et al. 2002). As with other rockfish, females seem grow more quickly than do males, with females r eaching sexual maturity at 7-9 years of age (35- 45 cm in length) versus males at 7-12 years (~41 cm in length) off Oregon (Westrheim and Harling 1975, Boehlert and Kappenman 1980, Lenarz and Echeverria 1991, STAT 1999). Mean length at sexual maturity off Vancouver Isla nd is 41 cm for females and 48 cm for males (Westrheim and Harling 1975). Canary rockfish are known to frequently reach 60-75 years of age and have been found to be as old as 84 years (Cailliet et al. 2000, Cailliet et al. 2001,

Andrews et al. 2007). Maximum reported sizes are 76 cm and 4.5 kg (Boehlert 1980, IGFA208 1991, Williams et al. 1999, Love et al. 2002, Methot and Stewart 2005).

Foraging. Canary rockfish prey upon different speci es as they age. Larvae are planktivores, consuming invertebrate eggs, copepods, and nauplii (Moser and Boehlert 1991, Love et al. 2002). Juveniles feed upon zooplankton, including crustaceans, juvenile polychaetes barnacle cyprids, and euphasiid eggs and larvae (Gaines and Roughgarden 1987, Love et al. 1991). However, adults move into a carnivorous lif estyle as well as eati ng euphasiids and other

crustaceans. Adults consume othe r fishes such as shortbelly rockfish, mytophids and stomiatiods (Cailliet et al. 2000, Love et al. 2002). However, oceanographic and climactic shifts can alter foraging such that canary rockfish feed on other available species (Lee and Sampson 2009).

Status and trends. Canary rockfish were proposed fo r listing on April 23, 2009 (74 FR 18516). Canary rockfish were once considered common in Puget Sound, but has declined at a faster rate than any other rockfish species in the region (Holmberg et al. 1967, NMFS 2008d). Prior to World War II, commercial landings of rockfish species generally remained under 20,000 lbs, but sky-rocketed during the war to 375,000 lbs annually and fluctuated between 50,000 and 220,000 lbs until 1970, when landings increased linearly with fishing effort to a peak of 900,000 lbs by 1980 (Palsson et al. 2008). Levels fluctuated after this between 48,000 and 300,000 lbs for the next decade and clearly crashed in the 1990’s, with landings below 30,000 lbs annually. At the cessation of commercial fishing in 2003, 2,600 lbs of rockfish were harvested. Canary rockfish have been noted for being much less frequen tly caught in the Puget Sound and Georgia Basin region since 1965 (NMFS 2008d). The rate of dec line for rockfish in Puget Sound has been estimated at ~3% annually for the period 1965-2007.

Declines have been noted in both numbers as well as frequencies. This likely due to the targeted removal of larger, older, and more fecund i ndividuals by commercial fisheries, reducing the ability of canary rockfish to rebound from excessive mortality (NMFS 2008d). For example, recreational fishing data have not reported any individuals ca ught greater than 55 cm since 2000, whereas a variety of large size classes had former ly been caught. There ar e concerns that even now some populations have been lost entirely, primarily due to over harvesting, but also due to low dissolved oxygen levels in some areas of Puget Sound (NMFS 2008d).

Natural threats. Interspecies competiti on, predators, and climactic regimes are the primary natural factors that depress canary rockfish numbers. Coppe r and quillback rockfish may compete with canary rockfish in Puget Sound fo r available resources (NMFS 2008d). Predators of canary rockfish include other rockfishes, lin gcod (for which rockfish is a particularly important dietary component), cabezon, seabirds , salmon, sharks, dolphins, seals, Steller sea lions, and perhaps river otters (Mer kel 1957, Miller and Geibel 1973, Morejohn et al. 1978, Roberts 1979, Antonelis Jr. and Fiscus 1980, Ainley et al. 1981, Rosenthal et al. 1982, Stevens and Miller 1983, Love et al. 1991, Beaudreau and Essington 2007, Lance and Jeffries 2007). Canary and other rockfishes appear to be nega tively influenced by El Niño conditions, possibly reducing available prey supply (Moser et al. 2000a, Harvey 2005).

Anthropogenic threats. Overharves ting the primary cause of canary rockfish declines, but habitat loss is also important. Canary rockfish are considered overfishe d by the Pacific Fisheries Management Council and are not presently harveste d intentionally. However, bycatch is still209 considered to be a high impact stressor to ro ckfish populations of Washington State waters (Palsson et al. 2008). Habitat loss is also a factor in canary rockfish de cline, with rocky habitats (reportedly, there are only 217 km 2 in Puget Sound) being threaten ed by construction of bridges, sewer lines, cable and pipeline depl oyment, and dredge spoil (Palsson et al. 2008). Low oxygen levels as well as pollutant, chem ical, and nutrient loading are also considered significant threats to canary rockfish recovery (NMFS 2008d).

Critical habitat. Critical habitat has not been desi gnated or proposed for canary rockfish.

Pacific eulachon Description of the Species. Eulachon that spawn in rivers south of the Nass River of British

Columbia to the Mad River of California have been separated into the proposed listing of southern DPS eulachon. This is based upon timi ng of runs related to temperature, genetic distinctions, size at maturity, and ecological features of both oceanic and freshwater environments. Differences may also exist in mean number of vertebrae (Hart and McHugh 1944,

McLean et al. 1999, Hay and McCarter 2000, McLean and Taylor 2001, Beacham et al. 2005). Like salmon, eulachon likely imprint upon chemical cues in their natal system. However, unlike salmon, hatchlings spend far less time in freshwat er systems and likely retain homing only to the estuarine system that their natal river drains to. Based upon this, the smallest stock unit is likely the estuary that natal streams dr ain (Hay and McCarter 2000, Beacham et al. 2005). Specific spawning rivers within the nata l system are likely selected based upon environmental conditions at that time of return (Hay and Beacham 2005).

Distribution. Eulachon, or candlefish, are small smelt native to eastern North Pacific waters from the Bering Sea to Monterey Bay, Californi a, or from 61o N to 31o N (Hart and McHugh 1944, Eschmeyer et al. 1983b, Minckley et al. 1986, Hay and McCarter 2000). However, the

southern extent of their distribut ion has receded northward over th e past several decades. The lower Columbia River mainstem provides spawni ng and incubation sites, and a large migratory corridor to spawning areas in th e tributaries. Prior to the construction of Bonneville Dam,

eulachon ascended the Columbia River as far as Hood River, Oregon. Major tributaries that support spawning runs include the Grays, Sk amokawa, Elochoman, Kalama, Lewis and Sandy

rivers.

Growth and reproduction. Although primarily mari ne, eulachon return to freshwater to spawn.

Adult eulachon have been observed in several ri vers along the west coast, including the Umpqua and Rogue rivers in Oregon, California’s Humboldt Bay, Klamath, Mad, Russian, and

Sacramento rivers as well as Redwood Cree k, and Washington’s Puget Sound, Hood Canal, Bear, Naselle, Nemah, Wynoochee, Quinault, Queets, and Nooksack rivers (Odemar 1964, Moyle 1976b, Minckley et al. 1986, Emmett et al. 1991b, Jennings 1996, Wright 1999, Larson and Belchik 2000, Musick et al. 2000a, WDFW and ODFW 2001). Spawning has been

documented in the Elwha River on the Strait of Juan de Fuca, but sightings or spawning in these Oregonian and Washingtonian rivers is very limited or unknown (Wright 1999, Shaffer et al. 2007). For southern DPS eulachon, most spawning is believed to occur in the Columbia River and its tributaries (Grays, Skamokawa, Elochom an, Kalama, Lewis, and Sandy rivers), with secondary levels in the Mad and Klamath rive rs, as well as sporadic production in other Oregonian and Washingtonian rivers (Emmett et al. 1991b, Musick et al. 2000a, W DFW and210 ODFW 2001). Southern DPS fish likely take less time to mature than do fish from more northerly rivers and generally sp awn earlier in southern portions of their range than in northern rivers (Clarke et al. 2007). Most eulachon die following migration and spawning, but some have outmigrated and returned for a second spawning event.

North of Washington State, eulachon spawning is supported in two major river systems: the Fraser and Yukon (Hart and McHugh 1944, Richardson et al. 2000, NPS 2008). Although 33

rivers have been documented to support spawning in British Columbia alone, only half of these are considered to have sustained spawning (Hay and McCarter 2000, Willson et al. 2006).

Spawning takes place at differential time and temperatures, depending upon the river system involved (Willson et al. 2006). In the Columbia River and further south, spawning occurs from late January to March, although river entry occu rs as early as December (Hay and McCarter 2000). The peak of eulachon runs in Washington State is from February through March. Fraser River spawning is significantly later, in April and May (Hay and McCarter 2000). Rivers of

northern British Columbia host eulachon runs in late February through March (the same time as Washington State peak runs) and peaks from mid-March to mid-May (Wydoski and Whitney 1979a). Alaskan runs occur in May and river entr y may extend into June (Hay and McCarter 2000). Although a trend of earlier spawning in southerly rivers is apparent, euchalon spawning may occur as early as January in Alaska and as late as May in north ern California rivers. Additionally, smaller-scale resolution of Br itish Columbia spawning supports earlier runs (February to early March) in northern territori al rivers (i.e., Nass River) and later spawning (April and May) in more southerl y rivers (i.e., Fraser River).

The timing of euchalon entry into spawning rivers is likely tied to water temperature and tidal cycles (Ricker et al. 1954, Bishop et al. 1989, PRFR 1998, WDFW and ODFW 2001, Lewis et
al. 2002, Spangler 2002). Water velocity may be a factor in spawning timing, where water velocity greater than 0.9 miles per h our can limit upstream migration (Lewis et al. 2002). Spawning normally occurs when water temperatur e is between 39o and 50o F, although Stikine and Nass River spawning is known to occur at cooler temperatures (S mith and Saalfeld 1955, Langer et al. 1977, Franzel and Nelson 1981, WDFW a nd ODFW 2001). Similarly, deviations in spawning while temperatures are warmer have been noted for the Susitna River in Alaska (Barrett et al. 1984, Vincent-Lang and Queral 1984). A dults are known to be sensitive to temperature changes; in the Cowlitz River, an increase in temperature from 41o to 52o F over several days was followed by a 50% mortality in adults and failure to spawn (Blahm and McConnell 1971).

Adults may migrate up to 100 miles upstream to reach spawning grounds (Hart and McHugh 1944). Males tend to arrive on spawning grounds ear lier than females and tend to stay longer, making them more susceptible to commercial and recreationa l fisheries (Hart and McHugh 1944). However, males outnumber females by a r oughly 2:1 margin. Eulachon sperm is viable for only minutes and a key factor of eulachon spawning may be male grouping en mass to broadcast their sperm. Once milt reaches downstream females, they release 17,000 to 60,000 eggs (generally 25,000 on average) at which time fertilization occurs. Fem ales lay eggs over sand, course gravel, or detritial substrate. However, like abalone, this method requires high eulachon density to ensure fert ilization. Density may also imp act the ability of eulachon to211 undergo synchronization of spawning, mate choice, and gonadal sterol levels. Eggs attach to gravel or sand and incubate for 30 to 40 days af ter which larvae drift to estuaries and coastal marine waters (Wydoski and Whitney 1979a). Larv ae develop into juvenile s and, after being in

estuaries and marine waters fo r 3 to 5 years, migrate back to natal basins to spawn.

Eulachon generally die following spawning (Scott and Crossman 1973b). Maximum know lifespan is 9 years of age, but 20 to 30% of individuals live to 4 years and most individuals survive to 3 years of age, although spawning has been noted as early as 2 years of age (Barraclough 1964, Parente and Snyder 1970, Langer et al. 1977, Wydoski and Whitney 1979a,

Barrett et al. 1984, Hugg 1996, Hay and McCarter 2000, WDFW and ODFW 2001). However, the age distribution of spawners varies between river and from year-to-year (Willson et al. 2006).

Habitat. Adult eulachon are found in coastal and offs hore marine habitats possibly to 2,000 feet deep, but more frequently between 50 and 600 feet deep (Allen and Smith 1988, Hay and McCarter 2000, Willson et al. 2006). Following hatching in fre shwater, larvae and juveniles become thoroughly mixed in coastal waters genera lly less than 50 feet deep, but can be found as far down as 600 feet deep (Barracl ough 1964, Hay and McCarter 2000).

Foraging. Larval and post larval eulachon pr ey upon phytoplankton, copepods, copepod eggs, mysids, barnacle larvae, worm larvae, and othe r eulachon larvae until th ey reach adult size (WDFW and ODFW 2001) . At this time, the primary prey of eulachon are obtained by filer feeding on copepods and euphausiids, including Thysanoessa spp., unidentified malacostracans, and cumaceans (Smith and Saalfeld 1955, Barraclough 1964, Wydoski and Whitney 1979a, Drake and Wilson 1991, Sturdevant et al. 1999, Hay and McCarter 2000).

Status and trends. The southern DPS of eulachon was pr oposed for listing as a threatened on March 13, 2009 (74 FR 10857). It is considered to be at moderate risk of extinction throughout the DPSs range. This is likely due to a variet y of factors, including predation, commercial and recreational fishing pressure (directed and bycatch ), and loss of habitat. Population decline is anticipated to continue as a resu lt of climate change and bycatch in commercial shrimp fisheries. However, as highly fecund fish, eulachon have the ability to rebound quickly if given the opportunity, a feature that is likel y necessary to withstand significant predation pressure and high mortality likely experienced by pelagic larvae (Bailey and Houde 1989). The median minimum population doubling time for eulachon is estimated to be 1.4 to 4.4 years (Musick et al. 2000a).

Eulachon formerly experienced widespread, abunda nt runs and have been a staple of native American diets for centuries along the northwest coast. However, such runs that were formerly present in several California ri vers as late as the 1960s a nd 1970s (i.e., Klamath River, Mad River, and Redwood Creek) no longer occur (Lar son and Belchik 2000). This decline likely began in the 1970s and continued until, in 1988 and 1989, the last re ported sizeable run occurred

in the Klamath River and no fish were found in 1996, although a moderate run was noted in 1999 (Larson and Belchik 2000, Moyle 2002b). Eulachon have not been identified in the Mad River and Redwood Creek since the mid-1990s, although effort here may be low or non-existent (Moyle 2002b). Despite a brief period of improved returns in 2001–2003, the returns and associated commercial landings ha ve again declined to the very low levels observed in the m id-212 1990s (JCRMS 2009) , and since 2005, the fishery ha s operated at the most conservative level allowed in the management plan (JCRMS 2009).

As mentioned, the Columbia Ri ver and its tributaries form the backbone of southern DPS spawning grounds. Historically, th is system likely supported half of all spawning abundance, but has declined precipitously since the early and mid 1990s (ODFW and WDFW 2007). Although regulations on commercial and re creational catches have been implemented throughout southern DPS freshwater range , commercial catch record s suggest populations are a small fraction of their former abundance (998 metric tons from 1936 to 1992 on average versus 91 metric tons annually from 1993 to present; (ODFW and WDFW 2007)). The Fraser River also represents a key spawning area for the southern DPS of eulachon and also experienced extensive declines in abundance to a presently historic low (10 metr ic tons of commercial catch, or roughly 300,000 individuals)(Schweigert et al. 2007, DFO 2008). This is a declin e of over 90% since the early 2000s, or roughly two fish generations.

Natural threats. Eulachon are heavily predated upon by numerous marine and terrestrial species. These include: fishes (green, white , and mammoth sturgeon, salmon, piked and spiny dogfish, sablefish, salmon sharks, lingcod, arrowt ooth flounder, Dolly Varden, Pacific halibut, Pacific hake, and Pacific cod), marine mammals (baleen whales, porpoises, orcas, dolphins, fur seals, harbor seals, and Steller sea lions), birds (harlequin ducks, pigeon guillemots, common murres, mergansers, cormorants, gulls, and eagles), and terrestrial mammals (Clemens et al. 1936, Hart 1973a, Scott and Crossman 1973b, Je ffries 1984, Drake and Wilson 1991, Yang and Nelson 1999, Willson et al. 2006). The high fat content of eu lachon make them a particularly valuable prey for white sturgeon in the Columb ia and Fraser rivers during winter (Willson et al. 2006).

Anthropogenic threats. Eulachon face numerous anthropogenic e ffects that have likely lead to their decline. Fisheries harvests are likely a major contributor to eulachon decline. The best

available information for catches comes from the Columbia River, where catches have been as high as 5.7 million pounds per year, but averaged near 2 million pounds from 1938 to 1993 (Wydoski and Whitney 1979a). Since 1993, catches have not exceeded 1 million pounds annually and the median catch has been 43,000 pounds (97.7% reduction in catch), even when effort is accounted for (WDFW a nd ODFW 2001). Efforts to quantif y fish that will be returning to the Columbia River to spawn in subsequent years have suggested signif icant variation in run size (100 metric tons to greater than 4,000 metric tons), but average less than 1,000 metric tons. Catch from sport fisheries is also high (Wydoski and Whitney 1979a). Outside of the Columbia River, the next highest landing si ze has been in the Fraser River, but catch here is nearly ten times smaller than in the Columbia (Hay et al. 2003). Bycatch from shrimp trawling along US and Canadian coasts has also been high, com posing up to 28% of the total catch by weight, triggering periodic closures of the fishery in some years (Hay and McCarter 2000, DFO 2008).

Perhaps the largest threat to eulachon surviv al and recovery is changing ocean conditions associated with global warming. Increases in ocean temperatures have already occurred and will likely to continue to impact eu lachon and their habitats. For example, changes in glacial melt throughout the Alaskan, British Columbian, Washi ngtonian, and Oregonian coasts will likely change flow patterns in major rivers, such as the Fraser, that are important for eulachon213 spawning (Morrison et al. 2002). Alterations in the hydrolog y of the Columbia River include decreased snowpack, increased peak flows, decreased base flow, and increased water temperatures, which will pose similar challenges to eulachon spawning as in the Fraser River (ISAB 2007b). Reductions in glacier sizes may also shift peak flow throughout regional rivers (Meier et al. 2003, Mote 2003, Barry 2005). In the marine environment, eulachon rely upon cool or cold ocean regions and the invertebrate communities therein (Willson et al. 2006). As with El Niño and La Niña events, warming ocean temper atures will likely al ter these communities, making it more difficult for eulachon and their la rvae to locate or capture prey (Roemmich and McGowan 1995, Zamon and Welch 2005a). Warmer wa ters could also allow for the northward expansion of eulachon predator and competitor ranges, increasing an already high predation pressure on the species (Rexstad and Pikitch 1986, McFarlane et al. 2000, Phillips et al. 2007).

Construction projects have also had a negative impact on eulac hon stocks. Dams, such as the Bonneville Dam on the Columbia River, have blocked eulachon from moving into former spawning habitat (Smith and Saalfeld 1955). Such damming projects also alter sedimentation and flow dynamics that eulachon have developed around in their evolution. River substrate composition, likely critical to successful spawning, is also altered by dams. The impoundment of water tends to raise water temperatures; a factor that spawning eu lachon are particularly sensitive to (NMFS 2008c). Sediment retention st ructures constructed in response to the Mount St. Helens eruption to limit sediment transport downstream that could block navigation, have

been correlated with reduced eulachon runs in subsequent years (Lou Reebs, pers. comm. in 74 FR 10857). Dredging activities likely destroy eggs and remove th e benthic substrates they rely upon. Although poorly known, eulachon ecotoxicol ogical studies support contaminant burdens, particularly of arsenic and lead, which can be high (Futer and Nassichuk 1983, Rogers et al. 1990, EPA 2002).

Critical habitat. On October 20, 2011, the NMFS proposed to designate critical habitat for the southern DPS of eulachon, including roughly 539 km of riverine and estuarine habitat in Washington State (Grays, Skamokawa, Elochoman, Cowlitz, Kalama, Toutle, Lewis, Quinault, and Elwa rivers/creeks), Oregon (Columbia River), and California (Mad, Klamath, Redwood, Umpqua, and Sandy rivers as well as Tenmile Creek)(76 FR 65324). These areas contain physicalor biological features esse ntial to the conservation of th e DPS, including (1) freshwater spawning and incubation sites with water flow, quality and temperature conditions and substrate supporting spawning and incubation, (2) freshwater and estuarine migration corridors free of obstruction and with water flow, quality and temp erature conditions supporting larval and adult mobility, and with abundant prey items supporting larval feeding after the yolk sac is depleted, and (3) nearshore and offshore marine foraging habitat with water qualit y and available prey, supporting juveniles and adult survival.

Atlantic sturgeon Population structure. Atlantic sturgeon were once present in 38 river systems and, of these, spawned in 35 of them. Individual s are currently present in 36 ri vers, and spawning occurs in at least 20 of these. Modern genetic analyses suggest that Atlantic sturgeon exhibit high fidelity to their natal rivers (Harwo od 2010). B ecause of high natal river fidelit y, it appears that most rivers support independent populations (Waldman and Wirgin 1998, Wirgin et al. 2000, King et al. 2001, Wirgin et al. 2002, Grunwald et al. 2008).214 Distribution. Atlantic sturgeon once ranged from Hamilt on Inlet on the coast of Labrador to the Saint Johns River in Florida and extralimitally to Bermuda and Venezuela (Smith and Clugston 1997, ASSRT 2007, Read 2010).

Reproduction and growth. The general life history pa ttern of Atlantic stur geon is that of a long lived, late maturing, iteroparous, anadromous species.

Spawning intervals range from once every one to five years for males (Smith 1985, Bain 1997,

Collins et al. 2000a, Schueller and Peterson 2010) and thr ee to five years for females (Bain 1997, Stevenson and Secor 1999, Gales et al. 2010, Schueller and Peterson 20 10). Fecundity increases with age and body size (ranging from 400,000 – 8 million eggs)(Van Eenennaam and Doroshov 1998, Dadswell 2006, Hammond 2010). The average age at which 50% of maximum lifetime egg production is achieved estimated to be 29 ye ars, approximately 3-10 times longer than for other bony fish species examined (Boreman 1997).

Sturgeon eggs are highly adhesi ve and are deposited on the botto m substrate, usually on hard surfaces (e.g., cobble)(Gilbert 1989b, Smith and Clugston 1997). Hatching occurs approximately 94-140 hrs after egg deposition, an d larvae assume a bottom-dwelling existence (Anonmyous 2010a). The yolksac la rval stage is completed in about 8-12 days, during which time larvae move downstream to rearing grounds ove r a 6 – 12 day period (Kynard and Horgan 2002). During the daytime, larvae use benthic stru cture (e.g., gravel matrix) as refugia (Kynard and Horgan 2002). Juvenile sturgeon continue to move further downstream into brackish waters, and eventually become residents in estuarine waters for months or years.

Atlantic sturgeon may reach ages of 60 year s or more, but aging studies are limited by inaccuracy once individuals are older than 15 ye ars old (Rien and Beamesderfer 1994, Nakamoto

et al. 1995a, Rossiter et al. 1995, Van Eenennaam et al. 1996, Stevenson and Secor 1999,

Whiteman et al. 2004, Jackson et al. 2007). Individuals grow rapidly once they migrate out of natal streams, but experience slower growth once they reach sexual maturity and beyond (Harrison and Thurley 1974, Dovel and Berggren 1983). Individuals in southern waters may have shorter life spans.

Habitat. Estuaries along the coast that do not s upport Atlantic sturgeon spawning populations may still be important rearing habitats.

Movement. Atlantic sturgeon spawn in freshwater, but spend most of their sub-adult and adult life in the marine environment. While few speci fic spawning locations ha ve been identified in the United States, through genetic analysis, many rivers are known to support reproducing populations. Early life stage At lantic sturgeon coupled with upstream movements of adults suggest spawning adults generally migrate upriver in the spring and early summer; this includes February-March in southern systems, April-Ma y in mid-Atlantic systems, and May-July in Canadian systems (Smith 1985, Bain 1997, Smith and Clugston 1997, Kahnle et al. 1998b). Some rivers may also support a fall spawning migration.

Sub-adult and adult Atlantic st urgeon undertake long marine migr ations and utilize East Coast215 nearshore marine for rearing, feeding, and migrating (Harrison and Thurley 1974, Dovel and Berggren 1983, Bain 1997). Migratory sub-adults and adults normally occur in shallow (10- 50m) waters dominated by grav el and sand substrate (Stein et al. 2004). Tagging and genetic

data indicate that sub-adult and adult Atlantic sturgeon may travel widely after emigrating from rivers. Despite extensive mixing in coastal waters , Atlantic sturgeon display high site fidelity to their natal streams. Straying between rivers within a proposed DPS would sometimes exceed five migrants per generation, but between DPS exchanges usually less than one migrant per generation, with the exception of fish from the Delaware River straying more frequently to southern rivers (Grunwald et al. 2008).

Diet. Atlantic sturgeon feed primarily on polych aetes, isopods, and amphipods in the marine environment, while in fresh water, they feed on oligochaetes, gammarids, mollusks, insects, and chironomids (Moser and Ross 1995, Johnson et al. 1997, Haley 1998, Haley 1999, Brosse et al. 2002, Guilbard et al. 2007, Savoy 2007, Collins et al. 2008). There is some disagreement as to whether Atlantic sturgeon cease foraging during certain times or in certain places. Although there is some evidence to support a portion of indi viduals not foraging in sp ring or in freshwater, evidence also exists to support ha lf to almost all individuals foraging in these circumstances (Brosse et al. 2002, Collins et al. 2008).

Several authors have found that polychaetes cons titute a major portion of Atlantic sturgeon diets.

Brosse et al. (2002) reported that over 90% of Atlan tic sturgeon diet wa s polychaetes during spring, summer, and winter. Savoy (2007) f ound Atlantic sturgeon diets consisted of approximately 66% polychaetes and 27% decapods in Long Island Sound while at the mouth of the Connecticut River, individuals fed almost exclusively on polyc haetes. At the mouth of the Hudson River, Haley (1999) found that sturge on fed on 47% polychaetes, 27% amphipods, and 22% isopods. In North Carolina, Moser and Ro ss (1995) determined Atlantic sturgeon diets were different, feeding on 32% polychaetes, 28% isopods, 12% mollusks, and then other items. In South Carolina, Collins et al. (2008) identified th e proportion of the sampled Atlantic sturgeon with each species in their guts and most guts cont ained polychaetes (over 50% of the fish that had been feeding had polychaetes in their guts).

Status and trends. On October 6, 2010, the NMFS proposed to list five DPSs of Atlantic sturgeon: the New York Bight, Chesapeake Bay, Carolina, and South Atlantic DPSs as endangered and the Gulf of Maine DPS as thre atened (75 FR 61872 and 75 FR 61904). On June 6, 2011, the NMFS proposed protective measures for the Gulf of Maine DPS (76 FR 34023).

Prior to 1890, Atlantic sturgeon populations were at or near carrying cap acity. In the mid-1800s, incidental catches of Atlantic sturgeon in the sh ad and river herring haul seine fisheries indicated that the species was very abundant (Armstrong a nd Hightower 2002). A major, targeted fishery did not exist until 1870 when a caviar market was established (Smith and Clugston 1997). Record landings were reported in 1890, where over 3350 metric tons (mt) of Atlantic sturgeon were landed from coastal rivers along th e Atlantic Coast (Sm ith and Clugston 1997 , Matthiopoulos and Aarts 2010). Between 1890 and 1905, Atlantic sturgeon populations declined dramatically due to sale of meat and caviar. The majority of these landings (75%) were from the Delaware River fishery, which presumably supp orted the largest populat ion along the Atlantic Coast (Matthiopoulos and Aarts 2010). Ten years after peak landings, the fishery collapsed in216 1901, when less than 10% (295 mt) of its 1890 peak landings were reported. The landings

continued to decline to about 5% of the peak until 1920 and remained between 1-5% thereafter. Between 1920 and 1998, the harvest level remained very low due to depleted populations. Prompted by research on juvenile production between 1985 and 1995 (Peterson et al. 2000), the Atlantic sturgeon fishery was closed by the Atla ntic States Marine Fi sheries Commission in 1998, when a coastwide fishing moratorium was imposed for 20 to 40 years, or at least until 20 year classes of mature female Atlant ic sturgeon were present (ASMFC 1998).

Currently, the only populations that have been st udied well enough to provid e an estimate of size

are from the Hudson and Altamaha Rivers. Thes e two systems are considered the two largest spawning populations on the East Coast. Kahnle et al. (2007) reported that approximately 870 adults per year returned to the Hudson River between 1985 and 1995. Peterson et al. (2010) reported that approximately 324 and 386 adults per year returned to the Altamaha River in 2004 and 2005, respectively. Juvenile Atlantic sturgeon abundance may be a more precise way to measure the status of Atlantic sturgeon populations because it is believed that all age-1 and age-2 juveniles are restricted to their na tal rivers (Dovel and Berggren 1983, Bain et al. 1999).

Peterson et al. (2000) reported that ther e were approximately 4,300 age-1 and -2 Atlantic sturgeon in the Hudson River between 1985 a nd 1995. Schueller and Peterson (2010) reported that age-1 and -2 Atlantic sturgeon populati on densities ranged from 1,000 to 2,000 individuals over a 4 year period from 2004 to 2007.

The Hudson and Altamaha are presumed to be th e healthiest populations within the US. Thus, other spawning populations within the US are pred icted to have fewer than 300 adults spawning per year. However, evaluating the status of th e species depends on the status of the smaller extant populations because maintaining those populations maintains genetic heterogeneity and having a broad range prevents a single catastr ophic event from causing their extinction.

Natural threats. Naturally, these are small populations and in some rivers because of variable spawning returns, Allee affects c ould be an issue. During all stages of development, Atlantic sturgeon are sensitive to temperatures a bove 28°C (Niklitschek and Secor 2005, Anonmyous 2010b, Mcconnell et al. 2010) and dissolved oxygen levels below 4.3 to 4.7 parts per million (EPA 2003, Hindell et al. 2010, Taylor et al. 2010). Juvenile sturgeon are also stressed by high salinities until they mature and out migrate.

Anthropogenic threats. Factors likely play a larger role in this species’ current status. Water quality, ship strikes, bycatch, dams, and poachin g all contribute to the currently depressed populations of Atlantic sturgeon despite ha ving very few natura l predators.

The Atlantic Sturgeon Status Review Team (2007) determined Atlantic sturgeon in the Delaware River are at a moderately high risk of extincti on because of ship strikes and sturgeon in the James River are at a moderate risk from ship st rikes. Since that time , managers in the Hudson

River are concerned that ship strikes may also be threatening Atlantic sturgeon populations there. In these systems, large ships move upstream fr om the mouths of the river to ports upstream through narrow shipping channels. The channels are dredged to the approximate depth of the ships, usually leaving less than 6 feet of clearance between th e bottom of ships and the benthos

of the river. Because of the size of the propellers used on large ships, everything along the217 bottom is sucked through the prope llers. Large sturgeon are most often killed by ship strikes because smaller fish often pass through the prope llers without making contact but larger sturgeon get hit. As shipping increases in the future, as has been pred icted by the US Coast Guard, more Atlantic sturgeon are likely to be killed during encounters with ships. Besides the threats to Atlantic sturgeon from ships, th e act of dredging the channel can also kill sturgeon. Dredging projects in the Kennebec, Delaware, James, Cape Fear, and Savannah Rivers put Atlantic sturgeon at moderate risk (ASSR T 2007). Dredging primarily a ffects sturgeon by removing food resources and homogenizing habitat, eliminating holding areas and other high quality habitat. Also, sometimes Atlantic sturgeon are attracted to the sediment plume created during dredging operations and are killed by the dredge itself.

Atlantic sturgeon are caught as bycatch in seve ral fisheries both within river systems and along the coast. In the James River, bycatch in the st riped bass fishery poses a moderately high risk to the species, while it poses a moderate risk in nearly every other river system on the East Coast (ASSRT 2007). While these determinations were made for Atlantic sturgeon in each river system, the majority of the commercial fisherie s interactions occur in estuaries and along the coast, where sturgeon from all rive rs could be captured as bycatch.

On the East Coast, there is no good means of fish passage for Atlantic sturgeon in the systems with dams. Furthermore, as human populations grow along the Atlantic Coast and droughts were common over the past decade, it is likely that many more rivers on the East Coast could be dammed. Sturgeon in the Santee-Cooper River sy stem and the Cape Fear River are at a moderately high risk because of dams. Additionally, sturgeon in the Neuse River are at a

moderate risk from dams.

Atlantic sturgeon particularly were overfish ed during the late 1880s, peaking in 1890 and the fishery collapsed in 1901 (Jehl and Cooper 1980). While the fishery remained open following the initial peak harv est period, landings remained low through the 20 th century until 1996 when the fishery was closed due to concerns a bout the recovery of their populations.

Atlantic sturgeon have also been impacted by industrialization, poor wa ter quality, and loss of habitat (Van Eenennaam et al. 1996, Jager et al. 2001, Collins et al. 2002, Stein et al. 2004). Most Atlantic sturgeon managers and researchers consider water qu ality as a moderate risk to every DPS in the United States (ASSRT 2007). A tlantic sturgeon are sensitive to pesticides, heavy metals, and other toxins in the aquatic environment.

Critical habitat. Critical habitat has not been proposed for Atlantic sturgeon.

Marine invertebrates

Elkhorn coral Description of the species. Although they resemble plants, elkhorn coral is a colony of small shelled animals that collaboratively form frond-like branches radi ating from a central trunk that is firmly attached to the sea fl oor. The largest specie s of its genus, colonies can reach at least 6.6 feet high and 13 feet in diameter (Veron 2000). Corallites (branche s of radial arms of calcium carbonate) are tube-like and porous, 0.08 inch to 0.16 inch long, about 0.08 inch in diameter,218 white near the growing tip, and brown to tan away proximally.

Distribution. Elkhorn coral is found widely in the Ca ribbean, including in the Florida Keys, Abaco Island (The Bahamas), Alacran Reef, Mexi co, Belize, Colombia, Costa Rica, Guatemala, Honduras, Nicaragua, Panama, Venezuela, Bonaire, Cayman Islands, Jamai ca, Puerto Rico, US Virgin Islands, Navassa, and throughout the West Indies (Goreau 1959, Kornicker and Boyd 1962, Storr 1964, Scatterday 1974, Jaap 1984, Dustan and Halas 1987, NMFS 2006e). However,

abundance within the distri bution is reduced, largely due to wa ter temperature and quality issues.

Growth and reproduction. Elkhorn corals employ both sexu al and asexual reproduction. Sexual reproduction is accomplished by releas ing sperm and egg during spawning events.

Colonies are referred to as simu ltaneous hermaphrodites, meaning that a given colony contains both female and male reproductive sex organs (Szmant 1986). Spawning events are relatively short, with gametes released only a few nights dur ing July, August, and/or September. In some populations, spawning is synchronou s after a full moon. Annual e gg production in Puerto Rico was estimated to be 3,870 to 5,100 eggs per square inch of living coral tissue (Szmant 1986). Once fertilization occurs, planktonic larvae form before settling and metamorphosizing on appropriate substrates, pref erably coralline algae (Bak 1977, Sammarco 1980, Rylaarsdam 1983). Initial calcification ensues an d develop into da ughter corallites.

Studies indicate that larger colonies (as measured by surface area of the li ve colony) have higher fertility and fecundity rate s; over 80% of the colonies larger than 620 inches 2 were fertile. Estimated colony size at sexual maturity was 248 inches 2 and the smallest reproductive colony

observed was 6.3 inches 2 by 3.15 inches 2 (Soong and Lang 1992).

Biological and physical factors affect spatial and temporal patterns of recruitment. These include substrate availability and community structure, grazing pressure, fecundity, mode and timing of reproduction, behavior of larvae, hurricane dist urbance, physical oceanography, the structure of established coral assemblages, and chemi cal cues (Lewis 1974, Birkeland 1977, Goreau et al. 1981, Rogers et al. 1984, Baggett and Bright 1985, Ha rriott 1985, Hughes and Jackson 1985, Sammarco 1985, Morse et al. 1988, Fisk and Harriott 1990, Ri chmond and Hunter 1990).

Growth rates are relatively rapid, expressed as the linear extension of branches, ranging from 1.57 to 4.33 inches annually, and have enabled el khorn coral to construc t significant reefs in several locations throughout the Caribbean (Vaughan 1915, Jaap 1974, Adey 1978). Branching species, such as acroporid corals, grow differentia lly in response to light such that coral polyp growth maximizes exposure to available light (Kaniewska et al. 2009). Growth can also occur from fragmentation and disper sal (Tunnicliffe 1981, Bak and Criens 1982). A broken branch may be carried by waves and currents to another lo cation and, if favorable, branches grow into a new colony. Rapid growth and fragment dispersa l facilitate a competitive advantage for elkhorn

coral relative to othe r coral and benthic species (S hinn 1976, Neigel and Avise 1983, Jaap et al. 1989).

Elkhorn coral require relatively clear water and depend almost entirely upon symbiotic photosynthesizers (zoozanthelle) for nourishment (Porter 1976, Lewis 1977, Jaap et al. 1989, Mieog et al. 2009) and is much more susceptible to in creases in water turbidity than are some219 other corals. Different strains of symbiotic zoozanthelle ( Symbiodinium spp.) can confer different thermal and light tole rances to acroporiids (Abrego et al. 2009, Ainsworth and HoeghGuldberg 2009, Abrego et al. 2010). The type of Symbiodinium spp. may change during ontogeny or remain the same, depending upon acroporiid species, and may be the same as parent colonies or not (Baird et al. 2007, Gómez-Cabrera et al. 2008, Abrego et al. 2009).

Habitat. Colonies of elkhorn coral of ten grow in dense stands and form interlocking framework known as thickets in fringing and barrier reefs, ranging in depth from 3.3 to 49 feet (Jaap 1984, Dustan 1985, Dustan and Halas 1987, Tomascik and Sander 1987, Wheaton and Jaap 1988). However, optimal depth range is considered to be 3.3 to 16.4 feet in depth, with possible exposure at low tide (Goreau and Wells 1967). Co lonies generally do not form thickets below 16.4 feet, with maximum water depths of fram ework construction ranging from 10 to 39.4 feet (Lighty et al. 1982). Elkhorn coral thrive in shallo w reef zones where wave energy is a significant factor. In areas with strong wave energy conditions only isolated colonies occur, while denser thickets may develop in intermedia te wave energy conditions (Geister 1977). The preferred habitat of elkhorn co ral is the seaward face of a r eef (Shinn 1963, Cairns 1982, Rogers

et al. 1982).

Status and trends. Elkhorn coral was listed as threatened under the ESA on May 9, 2006 (71 FR 26852). Elkhorn coral underwent precipitous dec lines in the early 1980s throughout its range and this decline has continued. Although quant itative data on histor ical distribution and abundance are scarce, best available data indica te declines in abundance (coverage and colony numbers) by greater than 97%. Recovery from a bleaching event in 2005 is expected to take 10- 12 years; this is after a pr evious event in 1997. In a ll, roughly one-third of the Acropora palamata genotypes have been lost as a result of these events (Miller and Williamson 2010).

Natural threats. The overriding threats are disease, temperature-induced “bleaching” (loss of zoozanthelle), and physical damage from hurricanes (Carpenter et al. 2008, Mallela and Crabbe 2009, Baskett et al. 2010). Disease is widespread, episodi c, and unpredictable in its occurrence and results in high mortality. This is primarily du e to a disassociation of zoozanthelle from coral tissue. Just prior to this, coral epithelium and gastrodermis tissue begins to decay and die, likely

as a result of stress to th e individual coral (Ainsworth et al. 2008). Optimal water temperatures range from 77o to 84oF, with mortality obs erved at 61o and 96oF (Jaap 1979, Roberts et al. 1982). High light levels can also i nduce m ortality. Synergistic analyses have found that high temperature increases the risk of colony mortality under a variety of sediment loading conditions, but excessive sediment appears to reduce mort ality risk under high light and temperature regimes, possibly by reducing expos ure to these stressors (Anthony et al. 2007, Boyett et al. 2007). High sediment with otherwise good light and temperature conditions appears to increase colony mortality (Anthony et al. 2007). Elkhorn coral require near oceanic salinities (34 to 37 parts per thousand). High temperature or rapid heating can result in heat shock and alter cellular metabolism within the coral as well as possibl y hinder immune respons e or the ability of zoozanthelle to thrive (Rodriguez-Lanetty et al. 2009, Middlebrook et al. 2010). Bleaching can

occur due to adverse environmental conditions (Ghiold and Smith 1990, Williams and BunkleyWilliams 1990) and is currently a significant factor in deteriorating cora l reef health. In 2005, wide-scale bleaching occurred throughout the Cari bbean with wide-scale mortality, with some areas reaching 95% of coral colonies affected (Wilkinson and Souter 2008). The US Virgin220 Islands, a location of Acropora critical habitat, experienced great er than 50% mortality of corals, the greatest level ever recorded. Puerto Rico and Florida (additional areas of Acropora critical habitat) also experienced diseas e rates of 50% of co ral colonies or greater. Bleaching was associated with unusually warm waters in the re gion. Encouragingly, bleaching events can lead

to increased thermal tolerance in affected reefs, meaning that subsequent bleaching events are not as severe (Maynard et al. 2008). A record number of hu rricanes also caused extensive damage to coral reefs; the prevalence of hurrican es and subsequent coral reef damage has been linked to climate change (Wilkinson and Souter 2008) . Ocean acidification is also a threat due to the increased solubility of calcium carbonate in even slightly more acidic sea water (thereby eroding the shells which form coral hard parts)(Anthony et al. 2008, De’ath et al. 2009, Wei et
al. 2009, Crawley et al. 2010). Acidification also reduces the thermal tolerance of corals, meaning that bleaching can occur at lower temperatures (Anthony et al. 2008). Hurricanes can

cause wide-scale inhibition of r ecruitment in years following stor m passage as well as physical damage to coral colonies themselves (Mallela and Crabbe 2009).

Anthropogenic threats. Threats to elkhorn coral also incl ude eutrophication, sedimentation, anchoring, which degrade coral condition and increas e synergistic stress effects (e.g. bleaching).

Critical habitat. NMFS published a final rule to designa te critical habitat for elkhorn and staghorn corals on November 26, 2008 (73 FR 72210). This habitat serves as substrate of suitable quality and availability, in water depths from the mean high water line to 98 feet (except along some areas of Florida, where 6 foot cont our is the shoreward limit), to support successful larval settlement, recruitment, and reattachment of fragments. Four sp ecific areas are proposed for designation: the Florida unit, which compri ses approximately 1,329 square miles of marine habitat; the Puerto Rico unit, which comprise s approximately 1,383 square miles of marine habitat; the St. John/St. Thomas unit, which co mprises approximately 121 square miles of marine habitat; and the St. Croix unit, which comprises approximately 126 square miles of marine habitat. NMFS proposes to exclude one military site, comprising approximately 47 square miles, because of national security impact s. The lone PCE identified thus far is natural

consolidated hard substrate or dead coral skelet on that are free from fleshy or turf macroalgae

cover and sediment cover. This feature is esse ntial to the conservati on of these two species because of the extremely limited recruitment cu rrently being observed and the need for this species to have habitat to recruit into.

Staghorn coral Description of the species. Although they resemble plants, staghorn coral is a colony of small shelled animals that collaboratively form staghorn-antler-like coloni es, with cylindrical, straight or slightly curved diverging bran ches. Branches are 0.1 inch to 0.6 inch in diameter and rarely may grow back together. Colonies in turbulent wa ter are smaller than in calm water, with greater branch density. Branching is irregular and seco ndary branches form at 60 to 90 degree angles relative to a primary branch. Prominent axial co rallites (branches of radial arms of calcium carbonate) form at branch tips; bract-like corallites radiate symmetrically around branches. Individual colonies are up to 5 feet across and typically form m onospecific thickets. Tissue color ranges from golden yellow to medi um brown, with little or no color near the growing branch tips. The colony may or may not be firmly attached to the sea floor.221 Distribution. Staghorn coral is found widely in the Ca ribbean, including in the Florida Keys, Abaco Island (The Bahamas), Alacran Reef, Mexi co, Belize, Colombia, Costa Rica, Guatemala, Honduras, Nicaragua, Panama, Venezuela, Bonaire, Cayman Islands, Jamai ca, Puerto Rico, US Virgin Islands, Navassa, and throughout the West Indies (Goreau 1959, Kornicker and Boyd 1962, Storr 1964, Scatterday 1974, Jaap 1984, Dustan and Halas 1987, NMFS 2006e). However,

abundance within the distri bution is reduced, largely due to wa ter temperature and quality issues.

Growth and reproduction. Staghorn corals employ both se xual and asexual reproduction. Sexual reproduction is accomplished by releas ing sperm and egg during spawning events (Szmant 1986). Colonies are referred to as simult aneous hermaphrodites, meaning that a given colony contains both female and male reproductiv e sex organs. Spawning events are relatively short, with gametes released only a few nights dur ing July, August, and/or September. In some populations, spawning is synchronou s after a full moon. Annual e gg production in Puerto Rico was estimated to be 3,870 to 5,100 eggs per square inch of living coral tissue (Szmant 1986). Once fertilization occurs, planktonic larvae form before settling and metamorphosizing on appropriate substrates, pref erably coralline algae (Bak 1977, Sammarco 1980, Rylaarsdam 1983). Initial calcification ensues an d develop into da ughter corallites.

Studies indicate that larger colonies (as measured by surface area of the li ve colony) have higher fertility and fecundity rate s; colonies with a bran ch length longer than 3.5 inches were fertile and over 80% of colonies with branches longer than 6.7 inches were fertile (Soong and Lang 1992). Estimated size at sexual maturity is 6.7 inch es in branch length and the smallest known reproductive colony was 3.5 inches in branch length (Soong and Lang 1992).

Biological and physical factors affect spatial and temporal patterns of recruitment. These include substrate availability and community structure, grazing pressure, fecundity, mode and timing of reproduction, behavior of larvae, hurricane dist urbance, physical oceanography, the structure of established coral assemblages, and chemi cal cues (Lewis 1974, Birkeland 1977, Goreau et al. 1981, Rogers et al. 1984, Baggett and Bright 1985, Ha rriott 1985, Hughes and Jackson 1985, Sammarco 1985, Morse et al. 1988, Fisk and Harriott 1990, Ri chmond and Hunter 1990).

Growth rates are relatively rapid, expressed as th e linear extension of br anches, ranging from 1.2 to 4.3 inches annually, and have enabled staghorn co ral to construct signif icant reefs in several locations throughout the Caribbean (Vaughan 1915, Jaap 1974, Adey 1978). Branching species, such as acroporid corals, grow differentially in response to light such th at coral polyp growth maximizes exposure to available light (Kaniewska et al. 2009). During the 1970s there were vast fields, or thickets, of staghorn coral on many reefs. The nominal situation in 2004 was isolated branches and small thickets, 1.6 to 3.3 feet across. Growth can also occur from fragmentation and dispersal (Tunnicliffe 1981, Ba k and Criens 1982). A broken branch may be carried by waves and currents to another locati on and, if favorable, bran ches grow into a new colony. Rapid growth and fragme nt dispersal facilitate a comp etitive advantage for staghorn

coral relative to othe r coral and benthic species (S hinn 1976, Neigel and Avise 1983, Jaap et al. 1989). Larval recruitment is influenced by the t ype and availability of benthic substrate, with certain types of coral or rock s ubstrates resulting in greater or lesser recruitment success (RitsonWilliams et al. 2009).222 Habitat. Historically, staghorn coral so dominated r eef systems within the 23 to 49 feet depth that the area became known as the staghorn zone . In other reef systems (Jamaica, Cayman Islands, Belize, and eastern Yucat an), staghorn coral was a major mid-depth (33 to 82 feet) reefbuilder (Adey 1977, 1978). Historically, staghorn coral was reported from depths ranging from surface to 200 feet, although it is considered rare below 66 feet (Goreau and Goreau 1973). In southeastern Florida, this species historically occurred on the outer reef (52 to 66 feet), on spur, groove bank, and transitional reefs, and on oc tocoral-dominated hard-bottom (Goldberg 1973, Davis 1982, Jaap 1984, Wheaton and Jaap 1988). Co lonies were common in back- and patchreef habitats (Gilmore and Hall 1976, Cairns 1982). Although staghorn coral colonies are sometimes found interspersed among colonies of elkhorn coral, they are generally in deeper

water or seaward of the elkhorn zone and more protected from wave action.

Staghorn coral require relatively clear wate r and depend almost entirely upon symbiotic photosynthesizers (zoozanthelle) for nourishment (Porter 1976, Lewis 1977, Jaap et al. 1989, Mieog et al. 2009) and is much more susceptible to in creases in water turbidity than are some other corals. Different strains of symbiotic zoozanthelle ( Symbiodinium spp.) can confer different thermal and light tole rances to acroporiids (Abrego et al. 2009, Ainsworth and HoeghGuldberg 2009, Abrego et al. 2010). The type of Symbiodinium spp. may change during ontogeny or remain the same, depending upon acroporiid species, and may be the same as parent colonies or not (Baird et al. 2007, Gómez-Cabrera et al. 2008, Abrego et al. 2009).

Status and trends. Staghorn coral was listed as threatened under the ESA on May 9, 2006 (71 FR 26852). Staghorn coral underwen t precipitous declines in th e early 1980s throughout its range and this decline has con tinued. Although quantitative data on historical distribution and abundance are scarce, best available data indica te declines in abundance (coverage and colony numbers) by greater th an 97% (Figure 23).

Staghorn corals still occupy their historic range, but localized range redu ctions and extirpations have occurred with most populations experienci ng losses from 80-98% of their 1970s baseline (Bruckner 2002). Monitoring data from around the USVI indicates that staghorn corals have virtually disappeared from the north side of Buck Island, St. Croix, and only a few localized

areas off the southern reef cont ain staghorn corals, representing 2- 3% of the coral cover in these areas (Rogers et al. 2002). Surveys of fragments of staghor n from nearshore areas of St. Thomas and outlaying cays indicate that colonies of thes e corals were once much more abundant than the numbers recorded in the 2003 survey (Rogers et al. 2008).

In Puerto Rico, well-developed and dense thickets of staghorn coral were present through the late 1970s at many reefs surrounding the main island, and also the offshore islands of Mona, Vieques and Culebra (Almy and Carrión-Torres 1963, McKenzie and Benton 1972, Goenaga and Cintrón 1979, Boulon Jr. 1980). Later, in 1978-79 during an island-wide survey, staghorn coral was found on only 20% of those reefs (Bruckner 2002 ). Prior to Hurricane David in 1979, 20

random 0.6 m 2 photoquadrats were selected from each of 10, 40-m-long transects parallel to the depth contours across the reef (16.7 to 19.2 m dept h). Based on analysis of point count data, staghorn coral had a mean total cover of 31.1% (r ange of 9.9 to 56.9%); after the storm, total cover of staghorn coral dropped to a mean of 0.90% (range of 0.02 to 2.7%)(NMFS 2008a). In the summer of 2004, there was an epidemic outbr eak of white pox disease at Los Corchos Reef223 in Culebra, Puerto Rico. Prior to the outbreak, coral cover on th e reef reached values of 80%. However, three weeks after Tropical Storm Jeanne , 80 to 90% of the staghorn coral colonies at permanent monitoring sites at Los Corchos were already dead or dying; likely as a result of impacts from both disease and storm damage (NMFS 2008a). During the 2005 bleaching event, near Culebra Island, almost 100% of staghorn colonies suffered part ial to complete mortality due to bleaching (García-Saís et al. 2008). Similar to the situation in USVI, the bleaching event was followed by a white plague-like massive outbreak th at caused mass mortality and resulted in a net 20-60% decline in living coral cover at surveyed reefs of th e east coast within a period of approximately six months.

Figure 23. Percent loss of staghorn and elkhorn coral throughout the Caribbean Sea.

Following the 2005 bleaching event, monitoring data indicate that total coral cover is now less than 12% on many reefs (Rogers et al. 2008). Coral mortality due to the 2005 bleaching event was more severe than at any time in the last 40 years of monitoring in USVI (Woody et al. 2008). Staghorn corals suffered widespread mort ality associated with the 2005 bleaching event and current monitoring data does not indicate signifi cant recovery (Rothenberger et al. 2008, Woody et al. 2008). Overall, colonies of Atlantic Acropora have declined by up to 98% and live colonies were no longer present at many st udy sites in the USVI following the 2005-2006 bleaching event.

Natural threats. The overriding threats are disease, temperature-induced “bleaching” (loss of zoozanthelle), and physical damage from hurricanes (Carpenter et al. 2008, Mallela and Crabbe 2009, Baskett et al. 2010). Disease is widespread, episodi c, and unpredictable in its occurrence and results in high mortality. This is primarily du e to a disassociation of zoozanthelle from coral tissue. Just prior to this, coral epithelium and gastrodermis tissue begins to decay and die, likely224 as a result of stress to th e individual coral (Ainsworth et al. 2008). Optimal water temperatures range from 77o to 84oF, with mortality obs erved at 61o and 96oF (Jaap 1979, Roberts et al. 1982). High light levels can also i nduce mortality. Synergistic analyses have found that high temperature increases the risk of colony mortality under a variety of sediment loading conditions, but excessive sediment appears to reduce mort ality risk under high light and temperature regimes, possibly by reducing expos ure to these stressors (Anthony et al. 2007, Boyett et al. 2007). High sediment with otherwise good light a nd temperature conditions appears to increase colony mortality (Anthony et al. 2007). Elkhorn coral require near oceanic salinities (34 to 37 parts per thousand). High temperature or rapid heating can result in heat shock and alter cellular metabolism within the coral as well as possibl y hinder immune respons e or the ability of zoozanthelle to thrive (Rodriguez-Lanetty et al. 2009, Middlebrook et al. 2010). Bleaching can

occur due to adverse environmental conditions (Ghiold and Smith 1990, Williams and BunkleyWilliams 1990) and is currently a significant factor in deteriorating coral reef health. The major

El Niño/La Niña Southern Osci llation cycle in 1997- 1998 resulted in a large bleaching event in the Caribbean and the Atlantic, as well as massi ve losses of corals in the Indian Ocean and Western Pacific (Wilkinson and Souter 2008). Howe ver, the most significant bleaching event to

date in the USVI and other areas of the Caribbean occurred in 2005 when sea surface temperatures exceeded the 29.5°C coral bleaching threshold for twelve weeks, and maximum temperatures exceeded 30°C (Woody et al. 2008). Bleaching occurred in twenty-two species, including Acropora , over a wide range of depths and affect ed more than 90% of the coral cover, on average, between July and November in the USVI (Woody et al. 2008). Wide-scale mortality, with some areas reaching 95% of coral colonies affected, resulted from this event (Wilkinson and Souter 2008). The US Virgin Islands, a location of Acropora critical habitat, experienced greater than 50% mortality of corals, the greatest level ever recorded. Puerto Rico and Florida (additional areas of Acropora critical habitat) also expe rienced disease rates of 50% of coral colonies or greater. Bleaching was associ ated with unusually warm waters in the region. Encouragingly, bleaching events can lead to in creased thermal toleran ce in affected reefs, meaning that subsequent bleaching events are not as severe (Maynard et al. 2008). A record number of hurricanes also caused extensive damage to coral reefs; the prevalence of hurricanes and subsequent coral reef damage has been li nked to climate change (Wilkinson and Souter 2008). Ocean acidification is also a threat due to the increased solubility of calcium carbonate in even slightly more acidic sea water (thereby eroding the shells which form coral hard parts)(Anthony et al. 2008, De’ath et al. 2009, Wei et al. 2009, Crawley et al. 2010).

Acidif ication also reduces the thermal tolerance of corals, meaning that bleaching can occur at lower temperatures (Anthony et al. 2008). Hurricanes can caus e wide-scale inhibition of recruitment in years following storm passage as well as physical damage to coral colonies themselves (Mallela and Crabbe 2009).

White band disease is thought to be the major factor responsible for the rapid loss of Atlantic Acropora due to mass mortalities. White band disease is the only coral diseas e to date that has been documented to cause major changes in the composition and st ructure of reefs (Humann and Deloach 2003). In 2011, Sutherland et al. (2011) were able to definitively identify human waste as a cause for white pox di sease in elkhorn corals.

Reductions in long-term water clarity can also re duce the coral photosynthesi s to respiration ratio (P/R ratio). Telesc nicki and Goldberg (1995) and Yentsch et al. (2002) found that elevated turbidity225 levels did not affect gross photos ynthetic oxygen production, but di d lead to increased respiration that consumed the products of photosynthesis wi th little remaining for coral growth.

Unfortunately, since staghorn corals are broadcast spawners once colonies become rare, the distance between colonies may limit fertilizatio n success and there is substantial evidence to suggest that sexual recruitment of staghorn corals is currentl y compromised. Reduced colony density in some areas is compounded by low genot ypic diversity, indicati ng that fertilization success and consequently, larval availability, is like ly reduced. This can have long-term implications for genetic variability of remain ing colonies due to th e reduced potential for exchange of genetic material between populations that are spatially fu rther apart (Bruckner 2002).

Data on levels of genetic diversity and population structure sugge st that there is a population structure among islands, and even ov er spatial scales of no more than 20 km, as well as varying degrees of genetic diversity within local populations (L irman 2002, Vollmer 2002). For instance, one clone of staghorn coral may dominate areas up to 10 m 2 in size and the clones are generally spatially discrete with larval exchange between staghorn populations as close as 2 to 15 km being extremely limited, suggesting that larval sources need to be conserved on a very small spatial scale (Baums et al. 2005, Vollmer and Palumbi 2007).

Anthropogenic threats. Threats to staghorn coral are exace rbated further by eutrophication, sedimentation, and anchoring, which degrade co ral condition and increase synergistic stress effects (e.g. bleaching). Excessive sedimentatio n can smother corals and increased nutrient availability promotes algal growth on corals, leading to light bloc kage to zoozanthelle and death of corals ( Acropora Biological Review Team 2005). A lthough reefs in the Florida Keys currently experience about 10% macroalgal cover or less, much of the wider Caribbean Sea may exceed 20% cover (Bruno 2008), inhibiting and reduc ing coral survival. Global warming is also projected to have negative impacts on coral survival through coral bleaching, increased storm intensity, and reduced calcification ( Acropora Biological Review Team 2005).

Critical habitat. NMFS published a final rule to designa te critical habitat for elkhorn and staghorn corals on November 26, 2008 (73 FR 72210). This habitat serves as substrate of suitable quality and availability, in water depths from the mean high water line to 98 feet (except along some areas of Florida, where 6 foot cont our is the shoreward limit), to support successful larval settlement, recruitment, and reattachment of fragments. Four sp ecific areas are proposed for designation: the Florida unit, which compri ses approximately 1,329 square miles of marine habitat; the Puerto Rico unit, which comprise s approximately 1,383 square miles of marine habitat; the St. John/St. Thomas unit, which comprises approximately 121 square miles of marine habitat; and the St. Croix unit, whic h comprises approximately 126 square miles of marine habitat. NMFS proposes to exclude one military site, comprising approximately 47 square miles, because of national security impact s. The lone PCE identified thus far is natural

consolidated hard substrate or dead coral skelet on that are free from fleshy or turf macroalgae

cover and sediment cover. This feature is e ssential to the conservati on of these two species because of the extremely limited recruitment cu rrently being observed and the need for this species to have habitat to recruit into.226 White abalone

Distribution. White abalone occur along the US west coast among offshore islands and banks (particularly Santa Catalina and San Clemente islands) and mainland inshore waters from Point Conception, California south to Punta Abreojos, Baja California, Mexico (Bartsch 1940, Cox 1960, 1962). White abalone occur primarily alon g the mainland coast in their northern and southern range, but are more frequently at th e offshore islands (especially San Clemente and Santa Catalina islands) in the middle portion of the California range (Cox 1962, Leighton 1972). However, individuals have also been found around several Mexican islands including Isla Cedros and Isla Natividad (G uzmán Del Proó 1992). There are no recognized subspecies of white abalone although there is one possible subspecies of white abalone inhabiting Guadalupe Island, Mexico (Hobday and Tegner 2000).

Habitat. White abalone occupy kelp forests in rela tively exposed areas of low-relief rocky habitat surrounded by sand. Large juvenile and small adult abalone (3 to 4 inches) are cryptic

and seek shelter in crevices and under rocks befo re, as adults, moving to more open habitats on the tops and sides of rocks wher e food is more plentiful (Haaker et al. 1986, Hobday and Tegner 2000). Adult white abalone are f ound most abundantly at depths of 82 to 99 feet, but may occur

in waters from 66 to 197 f eet (Hobday and Tegner 2000).

Feeding. White abalone eat algae, feeding as postlarv ae and early juveniles on bacteria, sessile pennate diatoms, and other benthic microflora. As advanced juveniles and adults, drifting brown algae and microalgal films provide the pr imary source of nutriti on (Tutschulte 1976).

Growth and reproduction. Abalone aggregate for spawni ng, but low numbers and physical barriers can prevent large spaw ning aggregations from form ing (Babcock and Keesing 1999,

Leet et al. 2001). A brief annual spawning event occurs en mass generally between February and April (Tutschulte 1976). Although an av erage female is capable of producing over 20 million larvae over her lifetime, larval survival to adulthood is estimated at <1% (Leighton 2000).

Twenty four hours after fertiliz ation, a free-swimming larva emer ges from the fertilized egg and joins the plankton (Leighton 1989, 2000). After 2 to 3 weeks in the plankton the larvae settle to the bottom. One to 3 months after settlement j uveniles are fully formed and resemble adults. After 2 to 4 years, white abalone are mature and inhabit the tops and sides of rocky substrates. However, low food resource availability has been shown to stunt growth in blacklip abalone (Saunders et al. 2009a, Saunders et al. 2009b).

Movement. Abalone movements tend to vary by individu al, with some staying at a particular home site and others undertaking extensive movements (Momma and Sato 1969, Ault and Martini 1987). This has been further char acterized in green li p abalone by Cennie et al. (2009a), where individuals belonged to ei ther a sedentary or a wandering clade (which moved over larger areas and covered greater distan ces, occupied inner portions of hides more frequently, responded more rapidly to food odor, foraged for longer pe riods, and displaced conspecifics from food patches).

Life span. Researchers believe that ab alone are long-lived, but the av erage life span is unclear227 (Hobday and Tegner 2000). Models of growth pr edict maximum size is reached in 34 to 35 years and a maximum life span of 35 to 40 years has been estimated (Tutschulte 1976, Tutschulte and Connell 1988).

Status and trends. On May 29, 2001, the white abalone wa s listed as an endangered species throughout its range under the ESA (66 FR 29046). In the white abalone status review, Hobday and Tegner (2000) estimated pre-exploita tion abundance at 2,221,800 abalone, but the population in 1996 to 1997 was estimated at 1,613 indi viduals, representing a 99.9% decline, and was estimated to disappear by 2010. White ab alone recovery is hindered by low spawning densities, resulting in recruitment failure. Wh ite abalone along Mexico are believed to be depleted, but their status is genera lly unknown. Based upon survey data, Hobday et al. (2001) updated the 1996 to 1997 white abalone abundance estimate to 2,540.

White abalone recruitment is highly variable (Tutschulte 1976). However, estimates of population size have been difficult to calculate because estimates are onl y based upon adults, as juveniles are infrequently obser ved. White abalone observed duri ng surveys were of large size which corresponds to predicted ages near the en d of the predicted life span (Davis 1996, Davis et
al. 1998, Hobday and Tegner 2000, Hobday et al. 2001). Because no white abalone were observed in the smaller age/size classes during the surveys there appears to be a lack of

successful recruitment since th e 1960s (Hobday and Tegner 2000).

Natural threats. Natural pressures exist fr om sea otter predation (Johnson et al. 2009) and the low density of individuals during spawning even ts. Interspecific competition has not been studied in white abalone, but blacklip abalone fa ce significant competition for algal resources by sea urchins (Strain and Johnson 2009).

Anthropogenic threats. White abalone numbers were severely reduced due to excessive harvest. This has led to below-threshold spaw ning densities in many areas that are blamed for the inability of the species to recover. Although small-scale aqu aculture takes have occurred to attempt captive breeding and recovery, these take s are small in number. Otherwise, substantial human harvesting is not known. No commercial or recreational takes are permitted under ESA protection.

Although toxicology of abalone is poorly known, red abalone have demonstrated metabolic breakdown when exposed to the pesticide 3-trifluoromethyl-4-nitrophenol (Viant et al. 2001). Silver, cadmium, and mercury are also known to bi oaccumulate in abalone, likely from ingested algae (Huang et al. 2008). Update rates have been measured as 1.78 L g -1 d -1 for silver, 0.056 L

g -1 d -1 for cadmium and 0.32 L g -1 d -1 for mercury, of which 58 to 83%, 33 to 59%, and 65 to 78%, respectively, is assimilated. Abalone are known to bioaccumulate high levels of heavy metals in the presence of high e nvironmental concentrations (Wang et al. 2009).

Changes in sea surface temperatur es have been suggested as a driving force in altering red abalone distribution in the past (Braje et al. 2009); it is unknown what affect, if any climate change may have on white abalone.

Critical habitat. Critical habitat has not been designated for white abalone.228 Black abalone

Distribution. Black abalone historically occurred between Coos Bay, Oregon, and Cape San Lucas, Baja California (Cox 1962), but were rare north of San Francisco (Morris et al. 1980).

Distribution extended to the Channel Islands a nd from Cedros to Punta Asuncion along Baja, Mexico (Guzmán Del Proó 1992). Present occurr ence remains throughout much of this range, although greatly reduced.

Growth and reproduction. Spawning occurs during spring and summer (Cox 1960). Synchronicity of gamete release is vital, as likelihood of fertilization is reliant upon dense adult aggregation and subsequent high egg and sper m density (Davis 1996). Fecundity increases exponentially with size, with small mature fe males producing a few hundred thousand eggs each year, but older individuals producing 10 to 15 m illion eggs (Hahn 1989). Fertilized eggs sink and hatch into free-swimming larv ae within 72 hours. After one or two weeks, larvae settle (CDFG 1993). Larval mo rtality is assumed to be high (Lei ghton 1972). At 1.5 inches (roughly 3 years of age), black abalone are co nsidered sexually mature (Blecha et al. 1992). Size at sexual maturity appears to decrease with latitude, suggesting precocity in southern portions of the range (Munoz and Camacho 1976, Guzman del Proo et al. 1980).

Feeding. Black abalone eat algae, with larvae eati ng pelagic plankton and postlarvae feeding upon on bacterial films, benthic diatoms (Cox 1962, Ault 1985), and coralline algae. Juveniles and adults feed upon large-bodied algae, such as giant and feather-boa kelp (Cox 1962, Howorth 1978). However, low food resource availability has been shown to stunt growth in blacklip abalone (Saunders et al. 2009a, Saunders et al. 2009b).

Habitat. Black abalone occupy rock y intertidal and shallo w subtidal zones (Haaker et al. 1986). Depth distribution is usually from the shore to about 10 feet of water de pth. Mobility patterns are size dependent (Blecha et al. 1992). Individuals above 4.1 inches undergo limited movement in exposed rocky locations. Juveniles (smaller than 1.6 inches) remain within rocky crevices during daylight hours, but become more active at night (Cox 1960). At 3 to 4 inches in length, when they emerge into more open rocky habi tats where food may be more abundant (Haaker et
al. 1986). Movement patterns have been further characterized in green lip abalone by Cennie et
al. (2009a), where individuals belonged to either a sedentary or a wande ring clade (which moved over larger areas and covered greater distances, occupied in ner portions of hides more frequently, responded more rapidly to food odor , foraged for longer periods, and displaced conspecifics from food patches).

Life span. Black abalone lifespan is unknown, but abalon e are believed to survive for 30 years or more (Blecha et al. 1992).

Status and trends. On January 11, 2008, NMFS published a propos ed rule to list black abalone as endangered under the ESA (73 FR 1986). Speci es decline is reflected by the decrease in

commercial catches until 1993, when commercial harvests we re halted. H istoric levels approached 2,200 tons in California in 1879 and declined to around 1000 tons in the 1970's. Commercial landings then decrease d to 19.1 tons in the last year of harvests, when mortality from withering syndrome devastated remaini ng black abalone stocks throughout southern229 California (Haaker 1994). Over 20 years, densities of more than 100 individuals per cubic yard disappeared from most of their former range so uth of Point Conception (Davis 1993). A similar mass mortality was reported at Palos Verdes Peni nsula in the late 1950's, where average density decreased from more than 2.8 individuals per square yard from 1975 to 1979 down to about 0.03 individuals per square yard from 1987 to 1991 (Cox 1962). Island habitats experienced more severe trends; 99% of black abalone vanished from Anacapa, Santa Barbara, and Santa Rosa Islands in less than 5 years (Haaker et al. 1989, Richards and Davis 1993).

Black abalone have also experienced severe dec lines due to a temperature-related disease called withering syndrome. This bact eria-based disease prevents as similation of nutrients in the digestive system and results in abalone that “w ither” as individuals consume body tissues. The

disease was first identified west of Santa Cr uz and Anacapa islands in 1985 and 1986 before spreading to Santa Rosa Island and Santa Ba rbara Island by 1988. The disease made its appearance along the mainland in 1988 in San Luis Obispo county, where 85% of the resident

black abalone died in Diablo Cove. This die-o ff was attributed to the presence of warm-water effluent from a nuclear power facility. From 1988 to the early 1990’s, withering syndrome continued to spread th roughout the Channel islands to 2000, when it was estimated that only 1% of the original population remained (Richards 2000).

Natural threats. Along with depletion from commercial harvests, withering syndrome has also been significant in black abalone decline. Withering syndrome is a chronic, degenerative disease responsible for mass mortalities (Moore et al. 2000a). Warm temperature, although not associated with the initiation of withering syndrom e, is associated with increased mortality rates (Lafferty and Kuris 1993, Harvell et al. 2009). This has been observed from power plant

effluent and incursions of warm water into tr aditionally temperate regi ons. In red abalone, higher rates of infection and mo re prominent signs of infecti on are associated with El Niño events (Moore et al. 2009). Interspecific competition has not been studied in black abalone, but blacklip abalone face significant competition fo r algal resources by sea urchins (Strain and Johnson 2009).

Compounding these factors are reproductive factors that further hamper species recovery. At low densities, individuals aggregate for spawning are not close enough for fertilization to occur. As a result, annual recruitment of juvenile bl ack abalone has declined steeply since adult populations dropped below half of initial densities (Richards and Davis 1993).

Black abalone at various life stages experience predation from several species. Juvenile abalone hiding amongst rocks are food for crabs, lobsters, octopi, starfish, fish, and predatory snails (Haaker et al. 1986). Abalone of intermediate sizes are vulnerable to octopus and fish predation, particularly sheepshead and cabezon. As adults , black abalone are primarily preyed upon by sea otters, which can be major regulators of black abalone populations, but are not know to extirpate communities as other threats are known to (Braje et al. 2009, Johnson et al. 2009). Interactions with other species can hinder species recovery in other ways, namely competition for space and food resources. Purple and red sea urchins tend to feed on the same kelp and brown algae food as black abalone and, when in high abundance a nd food is plentiful, have the potential to outcompete abalone for food (Leighton 1968, Paine 1974, Tegner and Levin 1982, Tegner 1989, Miller and Lawrenz-Miller 1993). However, abalone tend to inhabit different habitats than these230 urchin species (CDFG 1993). Space competition may also occur between black abalone and sand castle worms. This species cements itself to the underside of rocks, the same habitat that black abalone seek for refuge (Connell et al. 1988). This could limit the habitat available for

black abalone to recruit into dur ing recovery. Other factors th at can threaten black abalone include storms (crushing abalone between rocks a nd sedimentation in roc ky habitat), fresh water input, sedimentation on gills leading to asphyxi ation, and temperature impacts on reproduction and growth (Cox 1960, 1962).

Anthropogenic threats. Although commercial harv ests historically l ead to black abalone depletion, current harvesting is a small fraction of those levels. However, small removals are still significant in small populations.

Although toxicology of abalone is poorly known, red abalone have demonstrated metabolic breakdown when exposed to the pesticide 3-trifluoromethyl-4-nitrophenol (Viant et al. 2001). Silver, cadmium, and mercury are also known to bi oaccumulate in abalone, likely from ingested algae (Huang et al. 2008). Update rates have been measured as 1.78 L g -1 d -1 for silver, 0.056 L

g -1 d -1 for cadmium and 0.32 L g -1 d -1 for mercury, of which 58 to 83%, 33 to 59%, and 65 to 78%, respectively, is assimilated. Abalone are known to bioaccumulate high levels of heavy metals in the presence of high e nvironmental concentrations (Wang et al. 2009).

Changes in sea surface temperatur es have been suggested as a driving force in altering red abalone distribution in the past (Braje et al. 2009); it is unknown what affect, if any climate change may have on black abalone.

Critical habitat. On October 27, 2011, the NMFS designated critical habitat for black abalone This includes rocky areas from mean high wate r to six meters water depth in the Farallon, Channel, and Año Nuevo islands, as well as th e California coastline from Del Mar Ecological Reserve south to Government Point (excluding so me stretches, such as in Monterey Bay and between Cayucos and Montaña de Oros State Pa rk) in northern and central California and

between the Palos Verdes and Torrance border south to Los Angeles Harbor. These areas include primary constituent elements required by black abalone, such as rocky substrates to cling to, food resources (bacteria l and diatom films, crustose coralli ne algae, and a source of detrital macroalgae), juvenile settlement habitat (rocky intertidal habitat containing crustose coralline algae and crevices or cryptic biogenic stru ctures (e.g., urchins, mussels, chiton holes, conspecifics, anemones)), suitable water qualit y (temperature, salinity, pH, and other chemical characteristics necessary for normal settlement , growth, behavior, and viability of black abalone), and suitable nearshore ci rculation patterns (where sperm, eggs, and larvae are retained in the nearshore environment).

Marine Plants

Johnson’s seagrass Description of the species. Johnson’s seagrass has only relatively recently been identified as a distinct species and ther efore no historical dist ribution information is available (Eiseman and McMillan 1980).231 Distribution. Current distribution includes lagoons along approximately 125 miles of

southeastern Florida between Sebastian Inlet and north Biscayne Bay which means that Johnson’s seagrass has the most limited geographi c distribution of any seagrass in the world (Kenworthy 1997). However, norther n range extentions ( likely temporary) have recently been observed (Virnstein and Hall 2009). The larges t known groups of patches are located near

Sebastian Inlet and Lake Worth.

Habitat. Patches of Johnson’s seagrass have been observed to grow from the intertidal zone down to 3.3 feet water depth and in waters with variable temperatures and salinities (15 to 43 parts per thousand) and temperatures (Dawes et al. 1989, Kenworthy 1993, Virnstein et al. 1997, Kahn and Durako 2009). Patches near freshwater di scharges have been observed (Gallegos and Kenworthy 1996), although Torquemada et al ( 2005) noted that highly hypo- or hypersaline conditions can negatively impact growth. Intert idal patches may be comp letely exposed at low tides, suggesting tolerance to desiccation and wide temperature ranges (Kahn and Durako 2009).

Growth and reproduction. Only female flowers have been obs erved; no fruit or seeds have been found to date (Eiseman and McMillan 1980, Heidelbaugh et al. 2000). Meiosis does occur however, meaning that if male pollen ware even rarely presen t, sexual reproduction could take place (York 2005). However, there is no evidence of male flowers, meaning Johnson’s seagrass probably reproduces by cloning or asexual branch ing and fragmentation (Jewitt-Smith et al. 1997, Hammerstom and Kenworthy 2003). Conseque ntly, genetic diversity is low (Freshwater and York 1999), putting Johnson’s seagrass at a po tential genetic disadvantage compared to other seagrasses, particularly if removed from an area. However, if male pollen are even rarely present, sexual reproduction could take place (York 2005).

Clonal reproduction occurs when plants form new l eaf-pair, root and rhizome segments that arise from terminal buds (Posluszny and Tomlinson 1990). On average, new buds are formed on rhizomes every two to four days and rhizomes can grow at 0.2 inch per day (Bolen 1997, Kenworthy 1997). However, these clones can expa nd rapidly (1 to 3 feet per month) during periods of prolific branching (Kenworthy 1997, Greening and Holland 2003, Kenworthy 2003). As clones expand, high density “patches” are formed ranging from three to 66 feet 2 in size (Kenworthy 1997, Virnstein et al. 1997, Kenworthy 2000, 2003, Virnstein and Morris 2007). Patches can expand rapidly (nine feet 2 per month)(Kenworthy 2003) leading to coalescence with adjacent patches and large meadows of up to 30 acres (Kenworthy 1997).

Fragments or entire plants can be uprooted a nd drift extensively, providing a mechanism for dispersal and colonization of new areas (Hall et al. 2006). Virnstein et al. (2009) recently proposed that Johnson’s seagrass occurs in “pul sating patch,” with two to three consecutive summers of growth followed by a rapid decline. Johnson’s s eagrass frequently undergo whole patch mortality followed by recolonization (Virnstein et al. 1997, Heidelbaugh et al. 2000, Greening and Holland 2003, Kenwort hy 2003, Virnstein and Morris 2007).

Johnson’s seagrass appears to be physiologically adapted to exploit unstable environments and unvegetated patches, with minimal resources a llocated to the holding of space (Dean and Durako 2007). This characteristic may allow for more rapid overall patch growth and the exploitation of areas in which Johnson’s seagrass could not ot herwise compete (Dean and Durako 2007). These232 growth characteristics also he lp explain its patchy distri bution (Kenworthy 1993, Virnstein et al. 1997). Although successful in unstable areas, Johnson’s seagrass may be out-competed by more stable-selected plants in areas not subject to regular distur bance (Durako 2003). Johnson’s seagrass thrive in unstable or newly-created unve getated environments, but have little capacity for holding occupied space. As a result, Johnson’s seagrass can be highly variable in its occurrence over relatively short time frames (Virnstein et al. 2009). Due to this species’ physiology, low capacity for storage, and shallo w root system, growth over large unsuitable patches may be unlikely, and its ability to rec over from widespread habitat loss may be limited.

Status and trends. On September 14, 1998, Johnson's seagrass was listed as threatened under the ESA (69 FR 49035).

Historical abundance estimates of Johnson’s seagra ss are not available due to the species having only recently been differentiated. Limited data indicate no large distributional gaps or changes in abundance over much of Johnson’s seagrass distribution from 1994 to 1999. However, recent increases in reported occurrence could be an ar tifact of recent increases in search efforts.

Natural threats. Storms pose the greatest natural threat to Johnson’s seagrass. Storms can easily uproot or rip apart indivi duals and scatter them widel y. Although this can serve to disperse individuals into new habitats, it can also catastrophically eliminate established meadows. Subsequent siltation fo llowing high turbidity events can also bury individuals or parts of plants.

Anthropogenic threats. Due to its delicate morphology, small range, lack of genetic diversity and a physiology ill equipped to hold space and compete with other seagrasses, Johnson’s seagrass is vulnerable to prolonged widespread human-induced disturbance and habitat loss and its potential for recovery may be limited. Johns on’s seagrass and its habitat are threatened by several natural and anthropogenic factors, including (1) dredging and filling, (2) construction and shading from in- and overwater structures, (3) prop scarring and anchor mooring, (4) trampling, (5) altered water quality (such as stormwater r unoff and turbidity), and (6 ) siltation, as well as climate change (Waycott et al. 2009).

Critical habitat. Critical habitat for Johnson’s seag rass was designated on April 5, 2000 (65 FR 17786) and includes (1) locations with popul ations that have pers isted for 10 years; (2) locations with persistent flower ing populations; (3) locations at th e northern and southern range limits of the species; (4) locations with unique genetic diversity; and (5) locations with a documented high abundance of Johnson’s seagrass compar ed to other areas in the species’ range.

These PCEs are critical to the conservation of the species because they protect persistently reproductive and genetically diverse populati ons, allow for protective buffers along the distribution limits (i.e., edges of survival), and protect region s of high density that without further knowledge of species bi ology, appear to serve the need s of Johnson’s seagrass. Ten regions of sheltered bay and inlet waters are designated, including north and south of Sebastian Inlet, near Fort Pierce Inlet, no rth of St. Lucie Inlet, a portio n of Hobe Sound, the southern side of Jupiter Inlet, Lake Worth Lagoon (north of Bingham Island and Boynton Inlet), waters of Lake Wyman, and wide areas of northern Biscay ne Bay. These regions occupy approximately 22,574 acres or 9,139 hectares. Simply the nature of Johnson’s seagrass critical habitat makes it233 variable and prone to change. The growth of boa ting in Florida and development of coastal areas has resulted in trampling, propeller scarring, dred ging, filling, shading, and altered water quality that has degraded these areas compared to historical conditions. Alt hough many of the factors that can negatively affect Johnson’s seagrass an d their habitat are generally well regulated and enforced, the species is still unde r threat from high development pressure and subsequent habitat degradation throughout its range.

Environmental Baseline

By regulation, environmental baselines for Opinions include the past and present impacts of all state, federal, or private acti ons and other human activities in the action area, the anticipated impacts of all proposed federal pr ojects in the action area that have already undergone formal or early section 7 consultation, and the impact of stat e or private actions that are contemporaneous with the consultation in pr ocess (50 CFR §402.02). The Environmental baseline for this Opinion

includes the effects of several act ivities affecting the survival a nd recovery of proposed or listed species as well as their proposed or designa ted critical habitats in the action area.

Because this is a programmatic consultation, however, with a broad geographic scope that encompasses waters of the United States, this Environmental baseline serves a slightly different purpose. The Environmental baseline for this consultation focuses on the status and trends of the aquatic ecosystems in the United States and the c onsequences of that status for listed resources that occur in a general region. Since our action area and the Environmental baseline encompass a very broad spatial scale with many distinct ecosystems, wherev er possible we have focused on common indicators of the biolog ical, chemical, and physical h ealth of the nation’s aquatic

environments. The Environmental baseline for this consultation provides the backdrop for evaluating the effects of the action on listed a nd proposed resources under NMFS’ jurisdiction.

We divided the Environmental baseline for this consultation into marine versus freshwater regions. The freshwater component includes estuaries as well as five broad geographic regions: the Northeast Atlantic Region, the Southeast Atlantic Region, the Gulf Coast Region, the Southwest Region, and the Pacific Northwest Region. In some instances regions were further subdivided according to ecoregions, importance to NMFS’ trust resources or other natural

features. In each freshwater s ection we described the biological and ecological characteristics of the region such as the climate, geology, and pred ominant vegetation to provide landscape context and highlight some of the dominant processes that influence the bi ological and ecological diversity of the region where proposed, threatene d, and endangered species reside. We then described the predominant land and water uses with in a region to illustrate how the physical and chemical health of regional waters and the impact of human activities have contributed to current status of listed and proposed resources.

Stressors within the marine envi ronment tend to be much more ubiquitous than in freshwater ecosystems and thus we have not generally divided st ressors in the marine environment into more specific components, although some areas are relatively unique in regards to som e stressors, such as oil and gas industrial activit ies or hurricane impacts, and are described in a more regional context.234 Climate change We primarily discuss climate change as a threat common to all species addressed in this Opinion, rather than in each of the species-specific na rratives. As we better understand responses to climate change, we will address these effects in relevant species-specific sections. In general, based on forecasts made by the Intergovernmental Panel on Climate Change (IPCC), climate change is projected to have substantia l direct and indirect effects on individuals, populations, species, and the structure and function of marine, coastal, an d terrestrial ecosystems in the near future (IPCC 2000, 2001b, a, 2002). From 1906-2006, global surface temperatures

have risen 0.74o C and continues at an accelera ting pace; 11 or the 12 warmest years on record since 1850 have occurred since 1995 (Poloczanska et al. 2009). Furthermore, the Northern Hemisphere (where a greater proportion of ESA-lis ted species occur) is wa rming faster than the Southern Hemisphere, although land temperatures are rising more rapidly than over the oceans (Poloczanska et al. 2009). The direct effects of climate change will result in increases in atmospheric temperatures, changes in sea surface temperatures, patterns of precipitation, and sea level. Oceanographic models project a weakenin g of the thermohaline ci rculation resulting in a reduction of heat transport into high latitudes of Europe, an increase in the mass of the Antarctic ice sheet, and a decrease in the Greenland ice sheet, although the magnit ude of these changes remain unknown. Species that are shorter-lived, of larger body size, or ge neralist in nature are liable to be better able to adapt to climate change over the long termversus those that are longerlived, smaller-sized, or rely upon specialized habitats (Purvis et al. 2000, Brashares 2003, Cardillo 2003, Cardillo et al. 2005, Issac 2009). Climate change is most likely to have its most pronounced affects on species whose populations ar e already in tenuous positions (Isaac 2008). As such, we expect the risk of extinction to listed species to rise with th e degree of climate shift associated with global warming.

Some indirect effects of climate change woul d result from changes in the distribution of temperatures suitable for whale calving and rear ing, the distribution and abundance of prey and abundance of competitors or predators. For species that undergo long migrations, individual movements are usually associated with prey availa bility or habitat suitability. If either is disrupted, the timing of migra tion can change or negatively impact population sustainability (Simmonds and Eliott. 2009). Climate change can influence reproductive success by altering prey availability, as evidenced by low-success of northern elephant seal s during El Niño periods, when cooler, more productive waters are associ ated with higher first year pup survival (McMahon and Burton. 2005). Reduced prey availa bility resulting from increased sea surface temperatures has also been suggested to explai n reductions in Antarctic fur seal pup and harbor porpoise survival (Forcada et al. 2005, Macleod et al. 2007). Polygamous marine mammal mating systems can also be perturbated by rainfall levels, with the most competitive grey seal males being more successful in wette r years than in drier ones (Twiss et al. 2007). For marine mammals considered in this Opinion, availabl e data suggest sperm whale females have lower rates of conception following periods of unusua lly warm sea surface temperature (Whitehead 1997). Marine mammals with restricted distri butions linked to water temperature may be particularly exposed to range restriction (Learmonth et al. 2006, Issac 2009). MacLeod (2009) estimated that, based upon expect ed shifts in water temperatur e, 88% of cetaceans would be affected by climate change, 47% would be negative ly affected, and 21% woul d be put at risk of extinction. Of greatest concern are cetaceans with ranges lim ited to non-tropical waters and preferences for shelf habitats (Macleod 2009). Variations in th e recruitm ent of krill and the235 reproductive success of krill predator s correlate to variations in s ea-surface temperatures and the extent of sea-ice coverage during winter months. Although the IPCC (2001b) did not detect significant changes in the extent of Antarcti c sea-ice using satellite measurements, Curran et al. (2003) analyzed ice-core samples from 1841-1995 and concluded Antarctic sea ice cover had declined by about 20% since the 1950s.

Roughly 50% of the Earth’s marine mammal bioma ss occurs in the Southern Ocean, with all baleen whales feeding largely on a single krill species, Euphausia superba , here and feeding virtually nowhere else (Boyd 2002). Atkinson et al. (2004) linked sea ice loss to severe decreases in krill populations over the past several decades in some areas of the Antarctic. Reid and Croxall (2001) analyzed a 23-year time series of the reproductive perf ormance of predators (Antarctic fur seals, gentoo penguins, macaroni penguins, and black-browed albatrosses) that depend on krill for prey and concluded that thes e populations experienced increases in the 1980s followed by significant declines in the 1990s; overall an increase in the frequency of years with reduced reproductive success occurred. These declines resulted, at least in part, from changes in the structure of the krill populati on, particularly reduced recruitmen t into older krill age classes, which lowered the number of predators krill c ould sustain. The author s concluded that the

biomass of krill within the largest size class was sufficien t to support predator demand in the 1980s but not in the 1990s. By 2055, severe reductions in fisherie s catch due to climate change have been suggested to occur in the Indo-Pacifi c, Red Sea, Mediterranean Sea, Antarctic, and tropical areas worldwide while increased catches ar e expected in the Arctic, North Pacific, North Atlantic, and northern portions of the Southern Ocean (Cheung et al. 2010).

Climate change has been linked to changing ocean currents as well. Rising carbon dioxide levels

have been identified as a reas on for a poleward shift in the Ea stern Australian Current, shifting

warm waters into the Tasman Sea and alteri ng biotic features of the area (Poloczanska et al. 2009). Similarly, the Kuroshio Current in the we stern North Pacific (an important foraging area for juvenile sea turtles) has shif ted southward as a result of al tered long-term wind patterns over the Pacific Ocean (Poloczanska et al. 2009).

Climate-mediated changes in the distribution a nd abundance of keystone prey species like krill and climate-mediated changes in the distributi on of cephalopod populations worldwide is likely to affect marine mammal populations as they re-distribute throughout th e world’s oceans in search of prey. If sea ice exte nt decreases, then larval krill ma y not be able to survive without access to underice algae to feed on. This may be a cause of decrease d krill abundance in the

northern western Antarctic Peni nsula during the last decade (Fraser and Hofmann 2003).

Meltwaters have also reduced surface water salinities, shifti ng primary production along the Antarctic Peninsula (Moline et al. 2004). Blue whales, as predator s that specialize in eating krill, are likely to change their dist ribution in response to changes in the distribution of krill (Payne et
al. 1986, Payne et al. 1990, Clapham et al. 1999). If they did not ch ange their distribution or

could not f ind the biomass of krill necessary to sustain their population numbers, their populations (and other large whales with similar life histories, su ch as humpback whales) would likely experience declines similar to those observe d in other krill predator s, including dramatic declines in population size and increased ye ar-to year variation in population size and demographics. These outcomes would dramatically increase the extinction probability of baleen

whales. Edwards et al. (2007) found a 70% decrease in one zooplankton species in the North236Sea and an overall reduction in plankton biomass as warm-water species invade formerly coldwater areas. Productivity may increase in other areas, though, providing mo re resources for local species (Brown et al. 2009). In addition, reductions in sea ice may alleviate “choke points” that allow some marine mammals to exploit additional habitats (Higdon and Ferguson 2009).

The indirect effects of climate change would result from changes in the distribution of temperatures suitable for reproduction, the dist ribution and abundance of prey and abundance of competitors or predators. For species that undergo long migrations, individual movements are usually associated with prey availability or habita t suitability. If either is disrupted by changing ocean temperature regimes, the timing of migrat ion can change or negatively impact population sustainability (Simmonds and Eliott. 2009). For sea turtles, warming ocean temperatures may extend poleward the habitat whic h they can utilize (Poloczanskaet al. 2009). Seagrass habitats have declined by 29% in the last 130 years and 19% of coral reefs have be en lost due to human degradation, reducing lower la titude habitat for some sea turtle species (Poloczanska et al. 2009). Primary production is estimated to have d eclined by 6% between the early 1980s and 2010, making foraging more difficult for marine species (Hoegh-Guldberg and Bruno 2010).

Foraging is not the only potenti al aspect that climate change could influence. AcevedoWhitehouse and Duffus (2009) proposed that the rapi dity of environmental changes, such as those resulting from global warming, can harm immunocompetence and reproductive parameters in wildlife to the detriment of population viability and persistence. An example of this is the

altered sex ratios observed in sea tu rtle populations worldwide (Mazaris et al. 2008, Reina et al. 2008, Robinson et al. 2008, Fuentes et al. 2009a). This does not yet appear to have affected population viabilities through reduced reproduc tive success, although average nesting and emergence dates have changed over the past severa l decades by days to weeks in some locations (Poloczanska et al. 2009). However, such a fundament al shift in population demographics causes a fundamental instability in population viab ility. Altered ranges can also result in the spread of novel diseases to new areas via shifts in host ranges (Simmonds and Eliott. 2009). It has also been suggested that increases in harmful algal blooms could be a result from increases in sea surface temperature (Simmonds and Eliott. 2009).

Changes in global climatic patte rns will likely have profound eff ects on the coastlines of every continent by increasing sea levels and the inte nsity, if not the frequency, of hurricanes and tropical storms (Wilkinson and Souter 2008). A half degree Celsius increase in temperatures during hurricane season from 1965 to 2005 correlated with a 40% increase in cyclone activity in

the Atlantic. Sea levels have rise n an average of 1.7 mm/year over the 20 th century and 3.3 mm/year between 1993 and 2006 due to glacial melting and thermal expansion of ocean water; this rate will likely increase, which is suppor ted by the latest data from 2009 (Wilkinson and Souter 2008, Arndt et al. 2010, Hoegh-Guldberg and Bruno 2010) . Based on computer models, these phenomena would inundate nesting beaches of sea turtles, change patterns of coastal erosion and sand accretion that are necessary to maintain those beaches, and would increase the

number of turtle nests destr oyed by tropical storms and hurricanes (Wilkinson and Souter 2008), although other areas might experience less frequent tropical activity and a subsequent reduction in tropical cyclone impacts to sea turtle nest s (Fuentes and Abbs 2010). The loss of nesting beaches, by itself, would have catastrophic effects on sea turtle populations globally if they are

unable to colonize new beaches that form or if the beaches do not provide the habitat attributes237 (sand depth, temperatures regimes, and refuge ) necessary for egg survival. In some areas, increases in sea level alone may be sufficient to inundate sea turtle nests and reduce hatching success (Caut et al. 2009a). Storms may also cause direct harm to sea turtles, causing “mass” strandings and mortality (Poloczanska et al. 2009). Increasing temperatures in sea turtle nests alters sex ratios, reduces incubation times (pr oducing smaller hatchlings), and reduces nesting

success due to exceeded ther mal tolerances (Fuentes et al. 2009b, Fuentes et al. 2009c, Fuentes

et al. 2010). Smaller individuals likely e xperience increased predation (Fuentes et al. 2009b). Taken together, the body of literature on climate change supports widespread and significant negative consequences to sea turtle species.

Climate change also poses significant hazards to the survival and recovery of salmonids. Ongoing global climate change has implications fo r the current and likely future status of salmon, but particularly so in the Pacific north west, where snow melt in to the Columbia River Basin has significant influence on regional hyd rology. Recent studies, particularly by the Independent Scientific Advisory Board (ISAB), de scribe the potential impacts of climate change in the Columbia River Basin. These effects may decrease snowfall, increase early-year runoff, decrease summer and fall flow, and generally increase water te mperatures. The ISAB (2007a) identified the following list of likely effects of projected climate changes on Columbia Basin

salmon:

1. Water temperature increase resulting in loss of cold-water habitat (temperatures exceed upper thermal limits for a species). Projected salmon habitat loss would be most severe in Oregon and Idaho, possibl y higher than 40% of 2007 by 2090. Habitat loss would be less extreme in Washington at 22% by 2090. However, this assumes a high rate of greenhouse gas emissions and used a climate model that projected a 5o C in global temperatures by 2090, a value that is higher than the scenarios considered most likely (ISAB 2007a). Although a liberal estimate of change, this does not account for changes to hydrology that c ould further imbalance salmon habitat.

2. Variations in rainfall intensity may alter seasonal hydrograp hy. With reduced snowpack and greater rainfall, the timing of stream flow will likely change, reducing spring and summer stream flow and increas ing peak river flows (ISAB 2007a). This reduction in stream flow may impact the quality and quantity of tributary rearing habitat, greatly affecting spring and summer salmon and steelhead runs. In addition, the Pacific northwest’s low late-summer and early-fall stream flows are likely to be further reduced, which will limit juvenile fall Chinook and chum salmon shallow mainstem rearing habitat.

3. Considering both the water temperature a nd hydrologic effects of climate change, abundance Snake River spring/summer Chi nook populations would be substantially decreased (20-50% decline from simulate d average abundance based on historical 1915-2002 climate; (Crozier et al. 2008). This significantly increases extinction risks in the long term.

4. Eggs of fall and winter spawning fish, including Chinook, Coho, chum, and sockeye salmon, may suffer higher levels of mortality when exposed to increased flood flows.

5. Increases in seasonal mainstem Snake and Columbia River water tem perature would accelerate the rate of egg development of fa ll Chinook that spawn in the mainstem of238 the Snake and Columbia rivers and lead to earlier (smaller size) hatching. Potential effects of increased water temperatures on adult salmon include delay in dam passage, failure to enter fish ladders, incr eased fallback, and loss of energy reserves

due to elevated metabolic demand. Thermal st ress may also lead to increased risk of parasitism and disease.

6. Earlier snowmelt and higher spring flows, warmer temperatures, more rain, and less snow may cause spring Chinook and steelhead yearlings to smolt and emigrate to the estuary and ocean earlier in spring. The early emigration coupled with a projected delay in the onset of coas tal upwelling could cause thes e fish to enter the ocean before foraging conditions are optimal. Th e first few weeks in the ocean are thought

to be critical to the survival of salm on off Oregon and Washington, so a growing mismatch between smolt migrations a nd coastal upwelling would likely have significant negative impacts on early ocean survival rates.

7. Within the Columbia estuary, increased sea le vels in conjunction with higher winter river flows could degrade estuary habitats. Numerous warm-adapted fish species, including several non-indige nous species, normally found in freshwater have been reported from the estuary and might expand th eir populations with the warmer water. Climate change also may affect the trophi c dynamics of the estuary due to upstream extension of the salt wedge in spring-early summer caused by reduced river flows. Changes in the upstream extension of the salt wedge will influence the location of fish prey, but it is difficult to forecast the effect this change will have on juvenile

salmon.

8. Physical changes in the ocean associat ed with warming include increases in temperature, increased water column stratification, and changes in the intensity and timing of coastal upwelling. These cha nges will alter primary and secondary productivity, the structure of marine communities, and, in turn, the growth, productivity, survival, and mi grations of salmonids.

9. Changing ocean temperatures may alte r salmon behavior, distribution, and migrations, increasing the distance from home streams to ocean feeding areas. Energetic demands increase at warmer temperatures, requiring increased feeding to maintain growth. This could lead to in tensified competition for food and reduction in growth rates, further exacerbating the prey/predator relationship.

10. Increasing concentrations of carbon dioxide in the oceans lowers pH, which reduces the availability of carbonat e for shell-forming marine animals. Pteropods are expected to be negatively affected, and they can comprise more than 40% of some salmon diets. If salmon migrate farther to the north and/or food is less available, longer times may be required to reach ma turity, delaying the usual times of adult migrations into coastal water and rivers.

Climactic shifts also occur due to natural phe nomenon. In the North Atlantic, this primarily

concerns fluctuations in the North Atlantic Os cillation (NAO), which results from changes in atmospheric pressure between a semi-permanent high pressure feature over the Azores and a subpolar low pressure area over Iceland (H urrell 1995, Curry and McCartney 2001, Stenseth et
al. 2002). This interaction affects sea surf ace temperatures, wind patterns, and oceanic circulation in the North Atlantic (Stenseth et al. 2002). The NAO shifts between positive and239 negative phases, with a positive phase havi ng persisted since 1970 (Hurrell 1995). North Atlantic conditions experience d during positive NAO phases include warmer than average winter

weather in central and eastern North America and Europe and colder than average temperatures Greenland and the Mediterranean Sea (Visbeck 2002). Effects are most pronounced during winter (Taylor et al. 1998). The NAO is significant for Nort h Atlantic right whales due to its influence on the species primar y prey, zooplankton of the genus Calanus , which are more

abundant in the Gulf of Main e during positive NAO years (Conversi et al. 2001, Drinkwater et
al. 2003, Greene et al. 2003a, Greene and Pershing 2004, Kiszka et al. 2010a). This subsequently impacts the nutritional state of North Atlantic right whales and the rate at which sexually mature females can produce calves (Greene et al. 2003a). Local distribution shifts of North Atlantic right whales may be tied to the NAO (Kenney 2007).

Periodic weather patterns such as El Niño, La Niña, and the P acific decadal oscillation (PDO) can fundamentally change oceanographic conditions in the northeastern Pacific and the biology that is based upon it (Stabeno et al. 2004, Mundy and Cooney 2005, Mundy and Olsson 2005). Roughly every 3-7 years, El Niño can influe nce the northeastern Pacific (JOI/USSSP 2003,

Stabeno et al. 2004). Typical changes include increased winter air temper ature, precipitation, sea level, and downwelling favorable c onditions (Royer and Weingartner 1999, Whitney et al. 1999). La Niña events tend to swing these c onditions in the negative direction (Stabeno et al. 2004). However, sea surface temperatures (SSTs) can take 1 year to change following an El Niño event or change to vary ing degrees (Freeland 1990, Bailey et al. 1995, Brodeur et al. 1996a, Royer 2005). Eddies occur with greater fre quency in the Gulf of Alaska during El Niño years; the corollary is true during La Niña years (Crawford et al. 1999, Melsom et al. 1999). The 1997/1998 El Niño event is attributed with reducing primary production in the Gulf of Alaska by at least 50% (likely due to nitrate depletion); zooplankton abundance decreases as their phytoplankton food supply diminishes. Phytoplankton in continental shelf waters tends not to be effected by these events (F reeland 2000, Coyle and Pinc huk 2003, Zamon and Welch 2005b). The 1982/1983 El Niño and other downwelling events are generally regarded to have reduced food supplies for marine mammals along the US west coast (Feldkamp et al. 1991, Hayward 2000, Le Boeuf and Crocker 2005). Marine mammal distribution and social organization (group size) is also believed to have shifted northward in response to persistent or extralimital prey occurrence in more northerly waters dur ing El Niño events (Shane 1994, 1995, Benson et al. 2002, Lusseau et al. 2004, Norman et al. 2004, Danil and Chivers 2005). Low reproductive success and body condition in hum pback whales have also been sugge sted to have resulted from the 1997/1998 El Niño (Cerchio et al. 2005). El Niño events in the winters of 1952-1953, 1957- 1958, 1965-1966, and 1982-1983 were associated with strong downwelling anomalies, which reduces nutrient availability for plankton (Bailey et al. 1995, Wheeler and Hill 1999, Thomas and Strub 2001). Plankton diversity also shifts, as smaller plankt on are better able to cope with reduced nutrient availability (Corwith and Wheeler 2002, Sherr et al. 2005).

The PDO is the leading mode of variability in the North Pacific and operates over longer periods than either El Niño or La Niña and is capab le of altering SST, surface winds, and sea level pressure (Mantua 2002, Mantua and Hare 2002, Stabeno et al. 2004). Unlike El Niño and La Niña events, PDO events can persist for 20-30 year s, are more prominent outside the tropics, and mechanisms controlling them are relativel y unknown (Minobe 1997, 1999, Hare and Mantua 2000, Mantua and Hare 2002). During positive PDOs, the northeastern Pacific experiences240 above-average SSTs while the central and wester n Pacific Ocean undergoes below-normal SSTs (Mundy and Olsson 2005, Royer 2005). Warm PDO regimes, as with El Niño events, tends to decrease productivity along the US west coast (Hare et al. 1999, Childers et al. 2005). However, during the 1977 warm phase of the PDO, eupha usiid biomass remained the same and copepod abundance actually increased in the Pacific nor thwest; zooplankton biomass doubled in offshore waters of the Gulf of Alaska (Brodeur and Ware 1992, McFarlane and Beamish 1992, Brodeur et
al. 1996b, Francis and Hare 1997, MacCall et al. 2005). Opposite SST regimes occur during negative PDOs (Mundy and Olsson 2005). Posi tive PDOs occurred from 1925-1946 and 1977- 1999. Negative PDOs occurred from 1890-1924, 1947-1976, and 1999-present (Mantua et al. 1997, Minobe 1997, Childers et al. 2005).

The potential for invasive species to spread unde r the influence of climac tic change is also a significant concern. If water temperatures warm in marine ecosystems, native species may shift poleward to cooler habitats, openi ng ecological niches that can be occupied by invasive species introduced via ships ballast wa ter or other sources (Ruiz et al. 1999, Philippart et al. 2011). A similar observation of “Caribbean creep” has been observed, with warmer waters facilitating the range expansion of warmer-water species into more northerly regions (Canning-Clode et al. 2011). Although these expansions may be tempor ary, they can include harmful algal bloom species whose presence even temporarily can cause major morbidity and mortality issues to a variety of endemic species (Hallegraeff 2010). Moore et al. (2011) estimated that the impacts of a dinoflagellate establishment would likely inte nsify with a warming climate, resulting in roughly 13 more days of potential bloom c onditions per year by the end of the 21 st century. Invasive species that are better adapted to warmer water temperatures can also outcompete native species that are physiologically geared towards lower water te mperatures; such a situation currently occurs along central a nd northern California, where th e Mediterranean blue mussel has established and is displaci ng a native mussel competitor (Lockwood and Somero 2011).



### Freshwater and Coastal Estuarine Systems

Ballast Water Bioregion NA-ET2 (Maine to Massechusetts)

This region encompasses Maine, New Hampsh ire, and Massachusetts. The region is ecologically diverse, encompassing several broad ecoregions—according to Bailey’s (1995) Description of the Ecoreg ions of the United States this region encompasses the warm continental, the hot continental and the hot con tinental mountains divisions —these ecoregions can be further subdivided into provinces base d on vegetation (Bailey 1995). This region encompasses the New England/Acadian mixed forest s and the Northeastern Coastal Forests.

In this section, we describe several basins a nd estuarine complexes to characterize the general ecology and natural history of the area, and past and current human activ ities and their impacts

on the area. In certain instances we described some river basins in fu rther detail to provide additional context for evaluating the influence of the environmental baseline on listed species under NMFS’ jurisdiction and the h ealth of the environment.

Natural History. This region encompasses drainages entering the Gulf of Maine, and encompasses all of Maine, parts of New Hampsh ire, Massachusetts the Canadian provinces of241 New Brunswick and Nova Scotia. Characterized by a temperate climate and a rocky coastline, the greater Gulf of Maine encompasses th e Bay of Fundy, Casco Bay, Massachusetts Bay, Merrymeeting Bay and Cape Cod Bay. Significant Rivers that drain into the Gulf of Maine include the St. John, St. Croix, Penobscot River Basin, Kennebec/ Androscoggin River Basin and

the Merrimack River Basin.

Estuaries within the Gulf of Maine were formed by glaciers and as a result have characteristically rocky shorelines, shallow soil s, and deeply carved channels. The Gulf of Maine is semi-enclosed—bounded to the south by Georges Banks and to the north by Brown’s Bank. The area is more strongly influenced by the Labrador Current, which makes the waters significantly colder and more nut rient rich than waters to th e south that are more strongly influenced by the Gulf Stream.

The cold waters of the Gulf of Maine make it one of the most productive marine ecosystems in the world. The Gulf is characterized by salt marshe s, kelp and seagrass be ds, tidal mudflats, and underwater rocky outcrops form the foundation of a complex ecosystem and provide habitat for Atlantic herring ( Clupea harengus ), American lobster ( Homarus americanus ), Atlantic salmon, several whale species including endangered Northe rn right whales—where they are regularly observed in the spring and summer at re gular nursery and feeding areas.

Penobscot River Basin. The Penobscot River flows 275 miles to the ocean, with the largest watershed in Maine of 8,592 square miles (mi 2 ) (Jackson et al. 2005). The river flows from the mountains of western Maine, incl uding Maine’s highest peak, Mt. Ka tahdin to the ocean near the town of Bucksport, Maine. The Penobscot basin was formed by glaciation during the last ice age and the river’s bed is composed of glacial deposits and granitic bedrock. The average precipitation is approximately 42 inches per year. At the mouth, the average discharge is 10.1 billion gallons each day, or 14,000 cubic feet per second, but the discharge fluctuates seasonally and with dam releases, with naturally higher flows in the spring (Hasbrouck 1995(MaineRivers 2007). The river and estuary are also important for many fish species, with 45 freshwater and 39 salt water species having been recorded in the ri ver or estuary. Despite being home to so many fish, there are only three nonna tive species (Baum 1983(Jackson et al. 2005). The Penobscot estuary extends from Bangor downstream to Penobsco t Bay in the Gulf of Maine, approximately 31 miles, making it the largest es tuary in Maine and one of the largest on the East Coast (PEARL 2007). Downstream of Bangor, the river is a tid ally influenced, salt-wedge estuary. The majority of the estuary is bedrock-based, and sedi ment deposits are limited to isolated coves and

near marshes.

Merrymeeting Bay Basin. Merrymeeting Bay is the larg est, freshwater tidal estuary, approximately 18.6 miles upstream of the mouth of the estuary that enters the Gulf of Maine (Kistner and Pettigrew 2001, Jackson et al. 2005). The Kennebec and Androscoggin Rivers, along with four smaller tributaries, converge to form the bay, although the two large rivers account for 98% of the inflow. Merrymeeting Bay t ypically has the largest freshwater outflow to the Gulf of Maine, usually exceeding 15,000 cubic f eet per second. These high flows thoroughly flush the bay and have prevented eutrophication. The bay substrate is mud, sand, and exposed

bedrock.242 In Merrymeeting Bay, sampling only sandy substrate, which doesn’t hold as much contaminant as muddy substrates due to less su rface area, some toxic substan ces were identified. Sediments associated with the Androscoggin River had higher levels of P AHs and mercury, while sediments from the Kennebec River had higher le vels of chromium, arsenic, and selenium (Hayden 1998). The bay has more moderate levels of these toxins than th e rivers themselves. Chilcote and Waterfield (1995) found that levels of arsenic are hi gher than levels identified by EPA as likely to have adverse effects. At one station, PAHs from the Androscoggin also

exceeded EPA identified levels of minimal effects. In this region of the Gulf of Maine, metal deposition is linked more to the Androscoggin an d Kennebec than the Sheepscot River. Based on benthic samples taken in 1980 and again in 1991, it appears that all meta ls are declining in Merrymeeting bay except for copper, which showed an increase (Hayden 1998). Commercially important fish also have elevated metal concentra tions in their livers, which is thought to be from their time spent in Merrymeeting Bay (Kistner and Pettigrew 2001).

The Kennebec River flows 230 miles from the headwaters to the ocean, with a watershed of 5,384 mi 2 (Jackson et al. 2005, Maine Rivers 2007). The Ke nnebec River basin is primarily medium to coarse sand with some glacial till overlaying bedrock. Average precipitation is 42.5 inches of rain per year (Jackson et al. 2005). The average discharge at the mouth of the Kennebec River is 5,893 million gallons per day, with natural and controlled discharges similar to those seen on other Maine rivers (Maine Rivers 2007). There are 48 species of freshwater fish that use the Kennebec, in cluding 10 nonnative species.

The Androscoggin River travels 164 miles, with a watershed of 3,263 mi 2 (Jackson et al. 2005)MaineRivers 2007c). The river flows from northwest Maine, into New Hampshire, and then back into Maine, where it meets th e Kennebec River in Merrymeeting Bay. The Androscoggin has been Maine’s principle industr ial river (MaineRivers 2007c). The average precipitation in the watershed is 43.7 inches per y ear, resulting in an av erage discharge at the mouth of the Androscoggin, entering Merrymeet ing Bay, of approximately 4,190 million gallons each day. The river is home to 33 freshwater fish and 7 estuarine fish, including 8 nonnative species (Jackson et al. 2005).

Merrimack River Basin . The Merrimack River is 180 miles long, with 16 sub-basins in a watershed of 5,014 mi 2 (Jackson et al. 2005, MRWCI 2007). Seventy five percent of the watershed is in New Hampshire, with the rest in northeast Massachusetts . The precipitation is approximately 36 inches per year, with an aver age discharge of 5,364 million gallons per day, or 8,299 cubic feet per second. The geology of the Merrimack is dominated by granitic bedrock. The river is home to 50 species of fi sh, including 5 nonnativ e species (Jackson et al. 2005). For the lowest nine miles of the Merrimack River, extending north into New Hampshire and south to Cape Ann, Massachusetts, there are 25,000 acres of estuarine habitat and 15,000 acres of salt marsh habitat, which is referred to as the Great Marsh (USGS 2003).

Human Activities and Their Impacts

Land Use. Most of the watersheds within this region are heavily forested with relatively small areas of highly urbanized lands. Land use in th e Penobscot watershed is 5% agriculture and 95% forest and wetland (90% forest and forested wetlands). There are approximately 21 people per square mile living in the Penobscot watershed, and the largest town is Bangor, consisting of243 33,000 people (Jackson et al. 2005). While there is not mu ch urban development in the watershed, Doggett and Sowles (1989) report ta nneries, metal finishing, pulp and paper mills, textile plants, chemical products, and municipa l sewage contribute ch romium, mercury, zinc, copper, lead, arsenic, hydrocar bons, dioxins, PAHs, pesticides, a nd other contaminants to the

river.

The Kennebec River watershed usage is 82% fore st, 10% water, 6% agri culture, 2% developed (Jackson et al. 2005). The only major town in the wate rshed is Augusta, Ma ine, but there are approximately 39 people per square m ile throughout the watershed (Jackson et al. 2005). Currently, the primary pollution source on the river is from two pulp and paper mills, but there were multiple historical polluters along the river. The river exceeds recommended levels of dioxins, arsenic, cadmium chromium, copper, lead, mercury, nickel, silver, zinc, and PAHs in the sediments and surface water (MDEP 1999) (Har ding Lawson Associates 1999, 2000). Since 1990, the levels of dioxins in other rivers in Ma ine have been decreasing, but the levels in the Kennebec have remained constant (Kahl 2001).

The Androscoggin River watershed usage is 5% agriculture, 86% foreste d, 7% water, and 2% developed (Jackson et al. 2005). Major towns in the Androscoggin watershed are Auburn, Lewiston, and Brunswick. The human population in the watershed is approximately 65 people per square mile (Jackson et al. 2005). Throughout the 20 th century, textile mills, paper and pulp mills, and municipalities contributed large quantiti es of pollutants to the river. At one time it was considered one of the 10 most polluted rivers in the country and was one of the reasons for the implementation of the Clean Water Act. The river has become much cleaner since the Clean Water Act (CWA) was passed, but pesticides, mercury, lead, se dimentation, total suspended solids, PCBs, and dioxins are sti ll considered too high (Chamberland et al. 2002).

The Merrimack River watershed is composed of 75% forest, 13% urba n, 6% agriculture, 5% surface water, and 1% other (Jackson et al. 2005). The Merrimack River flows through

industrial centers Manchester and Concord, New Hampshire, and Lowell and Lawrence, Massachusetts. There are approximately 404 people per square mile in the Merrimack watershed (Jackson et al. 2005). The biggest sources of pollution facing the river are combined sewage overflows, industrial discharge, urbanization an d its associated run-off (USACE 2003). The upper mainstem of the river has problems with bacteria, E. coli, and acidity, while the lower mainstem has problems with bacteria, metals, nut rients, dioxins, turbidity and suspended solids, and un-ionized ammonia. In all, over 125 miles of mostly lower watershed areas do not support their designated uses (USACE 2003).

Hydromodification Projects. There are five major hydroelectric dams along the mainstem of

the Penobscot River as well as 111 ot her licensed dams located along th e river and its tributaries. Atlantic salmon historically migrated as far as 143 miles upstream of the mouth, but due to development along the river, in the 1960s, Atlantic salmon were extirpated (Jackson et al. 2005). The population has since been re-established and runs of 2,000 to 4,000 occur with natural spawning as far upstream as 62 miles. Unfo rtunately, 6,000 to 10,000 salmon are required for a sustainable population, so the Penob scot run depends on fish from a local hatchery (Moore and Platt 1997).244 The Kennebec River has eight la rge hydroelectric dams on its ma instem, which restricts fish passage both up and downstream. In 1999, the Edwards Dam was removed, opening 17

additional miles of habitat for fish and macroinve rtebrates in the river. Removal of Edwards dam restored full access to historical spawni ng habitat for species like Atlantic sturgeon, shortnose sturgeon, and rainbow sm elt, but not for species like alewife, American shad, and Atlantic salmon that migrated much further up th e river. Since the re moval of Edwards Dam, DO levels and macroinvertebrate density have improved. Additionally, in 2007, the fish passage

facilities on the lowest dam on the Kennebec River as well as the second and third lowest dams on the Sebasticook River became operational. The lowest dam on the Sebasticook River has been decommissioned and may be breach ed in as early as 2007 (MDMR 2007).

The Androscoggin River has 14 hydroelectric dams on the mainstem of the river and 18 in the

watershed. Fish ladders have been installe d on the lower dams allowing anadromous fish passage to Lewiston Falls (Brown et al. 2006). The dams play a considerable role in the poor water quality of the river, causi ng reduced DO throughout the summer. During the 60s, most of the river had oxygen levels of 0ppm, resulting in massive fish kills. There is still a 14 mile stretch of river that requires aerators to provide dissolved oxygen to the river.

The Merrimack River watershed has over 500 dams , including three in Massachusetts and three in New Hampshire, that essentially make th e mainstem into a series of ponds (Dunn Jr 2002,

Jackson et al. 2005). Flow alteration is considered a pr oblem on the upper mainstem of the river and has resulted in the river not meeting EPA’s flow requirements (USACE 2003).

Mining. Mining in Northeast Atlantic watersheds first began prior to the Civil War. Since then, mining has been conducted for granite, peat, roofing slate, iron ore, sulfur, magnetite, manganese, copper, zinc, mica, and other material s. Currently, exploration for precious metals and basic metals is ongoing, but to a lesser extant than during the 1980s. Recent mining

activities were conducted in this region by The Penobscot Nation, Champion Paper Company, Oquossoc Minerals, Boliden Resources, Inc., Black Hawk Mining, and BHP-Utah. There are

several abandoned mines in the No rtheast Atlantic coast watersheds that have become superfund sites due to excessive pollutants being leached into groundwater, such as Elizabeth Mine, Pike Hill Mine, Calhoun Mines, and others. Common po llutants leaked by mining operations in this area are lead, mercury, arse nic, and selenium (Ayuso et al. 2006, Piatak et al. 2006). All mines that are not in use are supposed to be deco mmissioned and cleaned up, but the impacts could persist for years before the rivers return to their pristine state.

Commercial and Recreational Fishing

. The primary commercial fi sheries along the Northeast Atlantic coast by harvest weight exist for he rring (39%), lobster (26%), blue mussel (6%), hatchery-origin sea-run Atlant ic salmon (4%), groundfish (4%), quahog (4%), soft clam (3%), sea cucumber (3%), seaweed (3%), crabs (2%), and various other species (6 %). Directed harvest of shortnose sturgeon and wild Atlantic salm on is prohibited by the ESA; however, both are taken incidentally in other fisheries along the east coast and are probably targeted by poachers throughout their range (Dadswell 1979, Dovel et al. 1992, Collins et al. 1996). Since 2006, a 30 day recreational fishing season between mid Se ptember and mid October for hatchery-origin Atlantic salmon has been permitted on the Penobsco t River, the only river with listed Atlantic salmon that allows salmon fishing. On the Penobs cot, spring salm on fish ing has not taken place245 since 1999, but may be permitted again in 2008. Poach ing is likely another fishing threat, but its impacts to individual population segments is unknown. Entanglement of marine mammals in fishing gear is not uncommon and can lead to mortality or serious injury.

Ballast Water Bioregion NA-ET3 (Connecticut to Virginia)

This region consists of Connecticut, Rhode Island, New York, New Jersey, Delaware, Pennsylvania, Maryland and Virgin ia. The headwaters of the C onnecticut River originate in New England/Acadian forests, and as the river descends, it tran sitions from boreal forest to temperate deciduous forest. As the river flow s through the low gradient coastal region, the ecoregion transitions to Northeastern Coastal Fore st. The headwaters of the Hudson River flow through Eastern Forest/Boreal Transition ecoregions. As the river descends, it transitions to

Eastern Great Lakes Lowland Forest and then Nort heastern Coastal Forest. The headwaters of the Delaware River originate in the Allegheny Hi ghland Forest ecoregion, and then as the river descends, it transitions to Appa lachian/Blue Ridge Forest and then Northeastern Coastal Forest ecoregions.

Connecticut River/Long Island Sound Natural History. The Long Island Sound watershed incl udes portions of Connecticut, New York, Massachusetts, New Hampshire, Rhode Island and Vermont. Long Island Sound was designated a national estuary in 1987, due to its significance as an area where freshwater from the Connecticut, Thames, and Housatonic Rivers ( 90% of the freshwater input) mixes with the Atlantic Ocean. The sound ranges in salinity from 23 parts per thousand ( ppt) in the western end to 35ppt on the eastern side. The su rface area of Long Island Sound is 1,320 mi 2 , draining an area of over 16,000 mi 2 . Long Island Sound connects to the Atlantic Ocean on both the eastern and western side, called “The Race” and the East River, respectively. The sound substrate is primarily mud, sand, silt, and clay, with very small areas of exposed bedrock. The sound is home to more 120 species of fish and at le ast 50 species use the sound as spawning grounds.

The Connecticut River drains a watershed of 11,259 mi 2 and flows approximately 410 miles to Long Island Sound. The river flows from the highl ands of New Hampshire and Quebec, and is bordered by the Green and White Mountains. Th e Connecticut River’s bed is composed of glacial deposits and granitic bedr ock. The average precipitation is approximately 43 inches per year. At the mouth, the average discharge is 10.2 billion gallons each day, or 15,715 cubic feet per second, which accounts for approximately 70% of the freshwater inflow to Long Island Sound (Jackson et al. 2005). The final 56 miles of the river prior to Long Island Sound is a tidal estuary (Jackson et al. 2005). The river and estuary are also important for many fish species,

with 64 freshwater and 44 estuarin e species having been recorded in the river or estuary, but 20 of the fish are nonnative (Jackson et al. 2005).

Human Activities and Their Impacts

Land Use. More than eight million people live in the Long Island Sound watershed. With so many people in the watershed, both point and no n-point source pollution is a major concern.

Toxic substances often adsorb to the surface of sediments, which means sediments with high surface to volume ratios like sand, silt, and clay, can hold more pollutants than larger substrates. The sound has elevated levels of PCBs, PAHs, n itrogen, lead, mercury, cadmium, cesium, zinc,246 copper, and arsenic. Organic and metal contam inants in Long Island Sound are above national averages (Turgeon and O'Connor 1991). Lead, copper, and zinc are believed to be deposited via the atmosphere (Cochran et al. 1998). Cadmium, chlordane, a nd lead appear to be decreasing while copper is increasing (Turgeon and O'C onnor 1991). Studies on winter flounder showed PAHs and PCBs leading to alteration of DNA in the livers of those fish (Gronlund et al. 1991). One of the biggest problems facing the sound is DO depletion (Parker and O’Reilly 1991), resulting in dead zones. The governors of Connecticut and New Yo rk have signed agreements to reduce the total nitrogen input to Long Island Sound by 58.5% befo re 2015 in an effort to get the DO of surface water above 5ppm, of deeper water above 3.5ppm, and no water ever below 2ppm.

Within the Connecticut River watershed the domin ant land use is forest (80%), with 11% used for agriculture and the remaining 9% in mixed (other) uses (Jackson et al. 2005). Major towns in the Connecticut watershed are Holyoke and Springfield, Massachusetts and Hartford, Connecticut. The human population in the wate rshed is approximately 179 people per square mile (Jackson et al. 2005). Throughout the 20 th century, power plants, defense contractors, municipalities, and corporations such as General Electric, Union Carbide, and Pfizer contributed large quantities of pollutants to the river. Still to this day, ap proximately one billion gallons of raw sewage enters the river as a result of combined sewer overfl ow from Hartford, Connecticut alone (CRWC 2006). The river has become mu ch cleaner since th e CWA was passed, but chromium, copper, nickel, lead, mercury, and zi nc, chlordane, DDT, DDE, PCBs, and PAHs are found in quantities above the EPA recommended le vels in sediments and fish tissue throughout the watershed (Jackson et al. 2005). Acid rain also affects rivers in the no rtheast, as it reduces the pH of rivers and causes metals to leach from bedrock at a fast er rate (Usfws 2007).

Hydromodification Projects. The Connecticut River has 16 hydroelectric dams on the mainstem of the river and as many as 900 are estimated to have been built in the watershed. Fish ladders have been installed at Vernon, Turner Falls, and Holyoke Dams allowing fish passage to areas above Holyoke Dam in Massachusetts since 1981 (Usgs 2004). For some species, the ladders are not efficient, so fi sh passage continues to be comp romised. For instance, overall passage efficiency at Turner Fall s fish ladder is 17%, and has hi storically been inefficient at passing shad. Shortnose sturgeon are not able to migrate to spawning habitat above Holyoke Dam, which was recently re-licensed through 2039, so the only spawning shortnose sturgeon in

the river are the fish that reside above the dam. The dams also affect the river’s water quality, causing reduced DO and elevated water te mperatures throughout the summer.

Mining. Dating back thousands of years, ther e is evidence of native people mining and extracting natural resources from the headwaters of the Connecticut River. There are many mines along the Connecticut River, which curren tly degrade the river’s water quality, including the country’s first chartered copper mine. Town s such as Plymouth, Vermont were famous for mining gold, iron, talc, soapstone, marble, asbest os, and granite (Ewald 2003). Other towns through New Hampshire and Vermont also mined gold, silver, soapstone, talc, granite, slate, and copper (Ewald 2003). In many locations, far downstream of the mines, accumulated heavy metals are in concentrations high enough to threat en aquatic life. In other cases, the mines are abandoned or failing and need to be cleaned. Su ch is the case with Elizabeth Mine, an old copper mine perched above the Connecticut River that leaches heavy metals into the river. As a247 result, Elizabeth Mine has been declared a supe rfund site. There is little to no mining in Long Island Sound and the concept is generally fr owned upon in the region, although there has been

and continues to be discussions about mining for sand and gravel.

Commercial and Recreational Fishing. There are not many commercial fisheries in the Connecticut River. Shad is the primary comm ercial fishery here, al though shellfish, bluefish, striped bass, and flounder can be caught in the tidal estuary near the mouth. There are many recreationally angled fish, such as shad, stripe d bass, bluefish, northern pike, largemouth and smallmouth bass, perch, cat fish, and other fish.

Long Island Sound fisheries provide an estimat ed 5.5 million dollars to the Connecticut economy. The primary fisheries target oysters, lo bsters, scallops, blue crabs, flounder, striped bass, and bluefish. Recently, due to DO defici encies, the western portion of Long Island Sound has seen major declines in fish and shellfis h populations. Despite th ese recent declines, the sound houses the largest oyster fishery in the US , which provides 95% of the nation’s oysters. At this same time, lobsters have been suffe ring from an unknown diseas e and their population has been declining. Simultaneously, menhaden have made a dramatic recovery over the past 10 years, which has resulted in much better fishing fo r larger predatory fish such as striped bass.

Directed harvest of shortnose sturgeon is prohi bited by the ESA. However, shortnose sturgeon are likely taken incidentally in fisheries in the Connecticut Ri ver and Long Island Sound. Moser and Ross (1993) found that capture s of shortnose sturgeon in co mmercial shad nets disrupted spawning migrations in the Cape Fear River, North Carolina, and Weber (1996b) reported that these incidental captures caused abandonment of spawning migrations in the Ogeechee River, Georgia. Entanglement of marine mammals in fishing gear is not uncommon and can lead to mortality or serious injury.

Hudson River Basin Natural History. The Hudson River flows approximate ly 315 miles to the ocean, with a watershed of 13,365 mi 2 . The river flows from the Adirondack Mountains, draining most of eastern New York State, to the ocean wh ere the Hudson River canyon continues onto the continental shelf, marking where the original mouth of the Hudson was covered by rising sea levels after the last ice age. The Hudson Rive r’s bed is composed of metamorphosed plutonic rock in the Adirondack Mountains, then transiti ons to sedimentary rock, such as shale and limestone in the middle portion of the watershed, and the lower portion of the watershed is a mixture of sedimentary, metamorphic, and i gneous rocks. The average precipitation is approximately 36 inches per year. At the mout h, the average discharge is 13.5 billion gallons each day, or 20,906 cubic feet per second (Jackson et al. 2005). The Hudson is a freshwater tidal estuary between Troy, NY at river mile 154 to Ne wburgh Bay at river mile 62, and then it is a tidal brackish estuary for the lower 62 miles to the Atlantic Ocean (Jackson et al. 2005). The river and estuary are home to ove r 200 fish species, with approximately 70 native freshwater fish species and 95 estuarine species having been recorded (Jackson et al. 2005).

Human Activities and their Impacts

Land Use. The Hudson River watershed usage is 25% agriculture, 65% forested, 8% urban, and 5% other (Jackson et al. 2005). Major towns in the Hudson River watershed are New York City,248 Albany, Poughkeepsie, and Hudson, New York a nd Jersey City, New Jersey. The human population in the watershed is approximately 350 people per square mile, but there are no people living in the headwaters and the populati on density in Manhattan is over 25,907 people per square mile (Jackson et al. 2005).

Throughout the 20 th century, power plants, municipalities, pulp and paper mills, and corporations such as IBM, General Motors, and General Electr ic in particular, who the EPA estimates dumped between 209,000 and 1.3 million pounds of PCBs into the river, contributed large quantities of pollutants to the Hudson. The PCB levels in the Hudson River are amongst the highest nationwide. The upper basin is mostly unaffected by humans, with clear, soft water with low nutrients. The middle Hudson is more polluted, w ith 30 to 50% of the land in this region being used for agriculture and several cities such as Corinth, Glens Falls, Hudson Falls, and Fort Edward contributing industrial waste to the river. The tidal freshwater portion of the Hudson is nutrient rich with exceptionally lo w gradient. High tide in this st retch causes the river to flow backwards due to the low gradient and this prevents stratification. The br ackish tidal estuary portion of the Hudson is nutrient ri ch with hard water. Two hundr ed miles of the Hudson River, from Hudson Falls to New York City, were desi gnated as a superfund site due to the amount of pollution. There are still elevated amounts of cad mium, copper, nickel, chromium, lead, mercury, and zinc, DDT, PCBs, and PAHs are found in qua ntities above the EPA recommended levels in sediments and fish tissue throughout the watershed (Wall et al. 1998).

Hydromodification Projects. The mainstem Hudson River ha s 14 dams and there are dams near the mouths of many tribut aries, but the lower 154 miles of tidally influenced river is

undammed. Several flood control dams on tributarie s such as the Indian and Sacandaga Rivers have drastically altered the flow of the mainstem Hudson River. The Hudson is an important

river for anadromous fishes because it is unobstru cted for the lower 154 miles, resulting in the healthiest population of ESA-liste d endangered shortnose sturgeon in the United States. Prior to the Clean Water Act, the middle stretch of the Hudson and much of the lower reaches had low dissolved oxygen as a result of reduced flow be hind the dams, high nutrients, and the collection of waste with high biological oxygen demand.

Mining. The Hudson River has been periodically important as a source of metals and mined resources. The Adirondack Mountains, in the headwaters, have mined silver, iron, titanium, coal, talc, vanadium, graphite, garnet, and zinc at various times over the past 300 years. McIntyre Mine is an example of a mine that has produced different mi nerals during different generations. Initially bought as an iron mine, McIntyre sat dorman t for 75 years before titanium was discovered there, at whic h point National Lead purchased it and mined there until 1982

when NL Industries abandoned the mine.

Commercial and Recreational Fishing. The Hudson River commerci al fishery historically caught fish, blue crabs, and oysters. Now, th e only fish that is caught commercially in the

Hudson is American shad. Histor ically, Atlantic stur geon, striped bass, American eel, and white perch were productive co mmercial fisheries. The striped bass fishery closed in 1976 due to

PCBs in the river and fish tissue. Atlantic st urgeon were fished until the mid 1990s. Blue crabs

are still fished in the estu ary all the way to Troy, NY with recent catches over 88,185 pounds per249 year. There is no commercial fishery for oysters but they used to be taken commercially in the

brackish tidal section of the Hudson.

Delaware River Basin Natural History. The Delaware River flows approxima tely 329 miles to the ocean, with a watershed of 12,757 mi 2 . The river originates in the Cats kill Mountains with over half of the river flowing through Pennsylvania and the rest of the watershed occupying parts of New Jersey, New York, and Delaware. The De laware River’s geology is sandst one with shale conglomerate in the upper watershed transitioni ng to sandstone, shale, and lime stone in the middle watershed and igneous and metamorphic rock in the lo wer watershed. The average precipitation is approximately 43 inches per year. At the mout h, the average discharge is 9.6 billion gallons each day, or 14,903 cubic feet per se cond, and although it is only the 42 nd largest river by discharge, Philadelphia is home to the larg est freshwater port in the country (Jackson et al. 2005). The Delaware River estuary begins in Trenton, New Jersey and extends downstream for 144 miles (Jackson et al. 2005). The river and estuary are ho me to 105 species of fish, with approximately 8 nonnative fish (Jackson et al. 2005).

Human Activities and their Impacts

Land Use. The Delaware River watershed usage is 24% agriculture, 60% forested, 9% urban, and 7% surface water or other (Jackson et al. 2005). Major towns in the Delaware River watershed are Easton, Allentown, Reading, and Philadelphia, Pennsylvania; Trenton and Camden, New Jersey; and Wilmington, Delaware . The human population in the watershed is approximately 555 people per square mile (Jackson et al. 2005). The water quality was significantly degraded around Philadelphia by 1799. By the 1960s the average DO in the lower river was approximately 0.2ppm. A survey in the 1970s of organoc hlorine frequency in rivers

ranked the Delaware at Trenton a nd the Schuylkill, the la rgest tributary to the Delaware, as the

8 th and 1 st worst, respectively in the nation (Jackson et al. 2005). While there aren’t many point sources of pollution since the Clean Water Ac t was enacted, historic ally, power plants, municipalities, pulp and paper mills, and industries such as the Philadelphia Shipyard, Bethlehem Steel, New Jersey Zinc Company, cont ributed large quantities of pollutants to the Hudson. Approximately 95% of PCBs are intr oduced to the river th rough combined sewage overflows from treatment plants. Even 35 year s after the Clean Water Act, there are still elevated amounts of copper, chromium, lead, mercury, and zinc, DDT, PCBs, and PAHs are found in quantities above the EPA recommended le vels in sediments and fish tissue throughout the watershed (Wall et al. 1998). The heaviest concentrations of chemicals in the river occur in a 14 mile stretch between the Philadelphia na val yard and the Tacony-Palmyra Bridge.

Hydromodification Projects. The Delaware River has 16 dams in the headwaters but the middle and lower river is the longest undammed stre tch of river east of the Mississippi. This stretch of free-flowing river is beneficial to anadromous and catadromous species, such as American shad, striped bass, and American eels.

Mining. The Delaware River watershe d, particularly the eastern section was home to the majority of the nation’s anthracite coal. As a result, many mining towns were established in the watershed to exploit the abundant resources . By 1914, over 181,000 people were employed as miners in the region. Apart from the coal mini ng, other minerals such as sulfur, talc, mica,250 aluminum, titanium, and magnesium were mined. Mines were also established for sand and gravel. Eventually minerals from the watershed were used to produce steel.

Commercial and Recreational Fishing. In the Delaware River, commercial fisheries exist for American shad, weakfish, striped bass, Atlantic croaker, Atlantic silversides, bay anchovy, black drum, hogchoker, northern kingfish and American eel. Commercial fishermen use gillnets and trawls as the primary means of capturing fish. Bycatch is a concern for the recovery of endangered shortnose sturgeon, where the highest morta lity rates are recorded in gillnet fisheries. Recreational fishermen target weakfish, striped ba ss, croaker, drum, kingfish, and eel. No data exists on shortnose sturgeon poaching.

Chesapeake Bay Drainages Natural History. Chesapeake Bay, the largest estuary in the United States, was formed by glacial activity more than 18,000 years ago. The Bay stretches some 200 miles from Havre de Grace, Maryland to Norfolk, Virginia, with more than 11,000 miles of shoreline. At its widest point, Chesapeake Bay is about 35 miles wide (n ear the Potomac River). Despite its massive size, the Bay is relatively shallow—average depth is only 21 feet—making it susceptible to significant fluctuations in temperature.

The Bay lies totally within the Atlantic Coasta l Plain but the watershe d includes parts of the Piedmont Province and the Appal achian Province. The tributaries provide a mixture of waters with a broad geochemical range to the Bay with its own mixture of mi nerals, nutrients and sediments depending on the geology of the place wh ere the waters originate. In turn, the nature of the Bay itself depends on the characteristics and relative volumes of these contributing waters.

While more than 50 tributaries deliver freshwater to Chesapeake Bay, ma jor rivers include the Susquehanna, Potomac, and the James River, whic h we describe in greater detail below.

Susquehanna River. Ranked as the 18 th largest river in the United States based on discharge, drainage area, or length, the Susquehanna River flows approximately 448 miles to the ocean, with a watershed of 27,580 mi 2 (Kammerer 1990, Jackson et al. 2005). The river flows north to south from New York, through Pennsylvania, an d reaches the Chesapeake Bay in Havre de Grace, Maryland. The Susquehanna River’s bed is rocky throughout, being described as a mile wide and a foot deep, with distinct pool/riffl e formations even near the mouth. The average precipitation is approximately 39 inches per year. At the mouth, the average discharge is 26.3 billion gallons each day, or 40,718 cubic feet per second, and serves as the primary freshwater source of the Chesapeake Bay (Jackson et al. 2005). The Susquehanna isn’t tidally influenced and doesn’t have much estuary habitat (Jackson et al. 2005). The river is home to 103 fish species, but 27 of the fish are nonnative (Jackson et al. 2005).

Potomac River. The Potomac River is approximatel y 383 miles long and has a watershed of 14,670 mi 2 . The river’s headwaters begin in the A llegheny Mountains of West Virginia and the Potomac most famously flows through Washington, D.C., to the western side of the Chesapeake Bay. The substrate of the Potomac and its tribut aries is mostly schist, phyllite, and metavolcanic rock. The average precipitation is approximately 39 inches. At the mouth, the average discharge is 7.3 billion gallons each day, or 11, 301 cubic feet per second (Jackson et al. 2005). The Potomac River estuary begins two miles belo w the Washington, D.C. Maryland border, just251 below the Little Falls of the Potomac River. Ni nety-five fish species liv e in the Potomac, but only 65 of those are native to the area (Jackson et al. 2005).

James River. The James River is approximately 340 miles long and drains a watershed of 10,

432 mi 2 . The James River is one of the longest bodies of water in entirely in one state, beginning in the Allegheny Mountains of western Virginia and flowing across the st ate to the Chesapeake Bay. The upper James River’s geology is primarily schist and siliclastic rock. The middle James River is primarily course grained conglomerates and sandstone. The lower section of the James is almost entirely sedimentary rock. The aver age precipitation is approximately 40 inches. At the mouth, the average discharge is 6.5 billion gallons each day, or 10,030 cubic feet per second (Blue 1998). The James River estuary begins at th e fall-line in Richmond, Virginia. Ninety-five fish species live in the Potomac, but only 65 of those are native to the area (Jackson et al. 2005).

Human Activities and their Impacts

Land Use. The Susquehanna River watershed usag e is 20% agriculture , 63% forested, 9% urban, and 7% pasture (Jackson et al. 2005). Major towns in the Susquehanna River watershed are Scranton, State College, and Harrisburg, Penn sylvania and Havre de Grace, Maryland. The human population in the watershe d is approximately 145 people per square mile (Jackson et al. 2005). The water quality has not been well documented because the river wasn’t used as a primary source of drinking water for any major c ities. The three main events that had the greatest effect on the river we re logging, dam building, and mi ning. While most of these activities took place in the 1800s, the river is still responding to the disruption they caused (Jackson et al. 2005). Sediment transport in the early 1900s was nine times higher than it was 200 years earlier, due to logging and agriculture. Sediment transpor t and its associated nutrients remain a major concern for the Chesapeake Bay. Coal is abundant through the watershed, amounting to nearly 30 billion tons of coal mine d. Coal waste and acid mine drainage damaged

much of the river and its tributaries. There was so much coal silt in the Susquehanna at one point that a fleet of over 200 vessels began harvesting the silt from the rive r’s bed. From 1920 to 1950, over 3 million tons of coal were harvested from behind one dam. Later, between 1951 and 1973, over 10 million tons were harvested from behind another dam. Coal is no longer a primary industry in the watershed, but the impacts of the acid mine drainage are still prominent. Another major problem is untreated sewage and industrial wast e that is dumped directly into the river. In Binghampton, New York, there ar e 10 sewer outfalls, 70 in Scranton, Pennsylvania, 65 in Harrisburg, Pennsylvania, and the number of outfa lls totals over 400 in the watershed, generally with the number of outfalls being proportional to the size of the city. As a result, the Susquehanna contributes 44% of the nitrogen and 21% of the phosphorous to the Chesapeake Bay. This has led to large algal blooms in the bay and a resulting “dead zone” between Annapolis, Maryland and Newport News, Virgin ia. In 2005, the Susquehanna was named America’s most endangered river by American Rivers, who produce an annual list. Even 35 years after the Clean Water Act, there are still elevated amounts of copper, sulfur, selenium, arsenic, cobalt, chromium, lead, mercury, zi nc, and pesticides (Beyer and Day 2004).

The Potomac River watershed usage is 32% agricu lture, 58% forested, 5% developed, 4% water, 1% wetland, and 1% barren (Jackson et al. 2005). Major towns in th e Potomac River watershed are Washington, D.C.; Arlington and Alexandria, Virginia; and Hagerstown, Maryland. The human population in the watershe d is approximately 358 people per square mile (Jackson et al.252 2005). The water quality has significantly improved over the past 50 years. Even 35 years after the Clean Water Act, there ar e still elevated amounts of cadmi um, chromium, copper, lead, dioxin, PCBs, and chlordane, which may have resu lted in recent highly publicized reports of male fish producing eggs.

The James River watershed usage is 23% agri culture, 71% foreste d, and 6% urban (VDCR 2006). Major towns in the James River waters hed are Charlottesville , Richmond, Petersburg, and Hampton Roads, Virginia. The human popul ation in the watershed is approximately 2.5 million people, or approximately 240 people per square mile (VDCR 2006). The James River has 21 municipal dischargers permitted and 28 permitted industrial dischargers. There are also 18 EPA Superfund sites along the river, mostly found in the major cities along its corridor. In some cases, industries such as Allied Chemical we re fined and forced to clean up large areas of extreme toxicity. Even 35 years after the Clean Water Act, there are st ill elevated amounts of zinc, copper, cadmium, nickel, chromium, lead, arsenic, dioxin, PCBs, and pesticides.

Hydromodification Projects. There are many dams along the Poto mac River and its tributaries, but only three impoundments ar e larger than 1.5 square miles. One of the major tributaries, the Anacostia River, is having over 60 dams removed or altered to improve water quality and fish

passage.

The Susquehanna River has over 100 dams along the mainstem and the first major dam is located just 10 miles upstream of the mouth. In recent years modern fishways have been installed in some of these dams and migratory fish app ear to be responding positively. For instance, between 1928 and 1972, no shad passed Conowingo Da m, 10 miles upstream of the mouth of the Susquehanna River, but since fish began co ming back, their abundance has increased from approximately 100 to more than 100,000.

The James River has several large dams along its length. Many dams have been removed or improved to allow fish passage, and in 1999, a ladder was built over Bosc her Dam, which had prevented upstream fish runs since 1823. That ladder provided access to 137 additional miles of

the James and 168 miles of its tributaries.

Mining. In the Chesapeake Bay watershed, coal mining has likely had the most significant impact on water quality. Mining in this watershed was so extensive that while many mines have been reclaimed and others are currently being reclaimed, at the current level of funding, it will take decades or more to completely reclaim a ll of the old mines in the watershed. Abandoned

coal mines leach sulfuric acid as a result of na tural reactions with the chemicals found in coal mines. Many of these abandoned coal mines must be treated with doses of limestone to balance the pH of the water draining from the mines. Much of the Appalachian Mountain chain that was mined for coal is now leaching sulfuric acid into tributaries of the Chesapeake Bay and requires some sort of treatment to improve the water quality of the region.

Commercial and Recreational Fishing. The Chesapeake Bay supports fisheries for American eel, croaker, blue crab, black sea bass, bluefi sh, oyster, red drum, spot, striped bass, summer flounder, weakfish, menhaden, and white perch (C FEPTAP 2004). Stocks of striped bass got so

low in the mid 1980s that a moratorium started in 1985, but they recovered so well that well-253 regulated harvests are now permitted. Since th e mid 1990s, levels of blue crab and menhaden have dropped to the lowest levels in history. Species such as catfish and white perch are year round residents and managed by individual st ates around the bay. Species like Spanish mackerel, king mackerel, red drum, and summe r flounder have ranges that extend beyond the bay and are managed under multiple regional ma nagement plans. Some species such as American shad are allowed to be fished by so me states (Virginia a nd Maryland) within the Chesapeake Bay, but not by other states (Delaware and Pennsylvania).

Ballast Water Bioregion CAR-VII (North Carolina to Florida)

This region covers all the drainages that ultimatel y drain to the Atlantic Ocean between the states of North Carolina and Florida. This region includ es all of South Carolina and parts of Georgia, North Carolina, Florida, and Virginia. The region encompasses three ecoregions—the hot continental division, subtropical division, and savanna division (sout hern most tip of Florida’s panhandle). The hot continental division is ch aracterized by it’s wi nter deciduous forest dominated by tall broadleaf trees, soils rich in humus and moderately leached (Inceptisols, Ultisols, and Alfisols), and rainfall totals th at decrease with distance from the ocean (Bailey 1995).

Most of the Southeast Atlantic Coast Region is contained within the subtropical ecoregion and is characterized by a humid subtropical climate w ith particularly high humidity during summer months, and warm mild winters. Soils are str ongly leached and rich in oxides of iron and aluminum (Bailey 1995). The subtropical ecoregi on is forested, largely by second growth forests of longleaf, loblolly and slash pi nes, with inland areas dominated by deciduous trees. Rainfall is moderate to heavy with annual averages of about 40 inches in the north, decreasing slightly in the central portion of the region, and increasing to 64 inches in s outhern Florida. The savanna ecoregion has a tropical wet-dr y climate, controlled by moist warm topical air masses and supports flora and fauna that is adapted to fluctuating water le vels (Bailey 1995).

In the sections that follow we describe severa l basins and estuaries to characterize the general ecology and natural history of the area, and past and current human activ ities and their impacts on the area. The region contains more than 22 river systems that generally flow in a southeasterly direction to the Atlantic Coast. The diverse geology and clim ate ensures variability in biological productivity and hydrology. Ma jor basins include the Albemarle-Pamlico Watershed and its tributaries, the Cape Fear River, Winyah Bay and the Santee-Cooper Systems, the Savannah, Ogeechee, and the St. Johns River, to name a few. The more northerly river, the Roanoke which is part of the Albemarle-Pamlico Watershed, is cooler a nd has a higher gradient and a streambed largely characterized by cobble, gravel and bedrock.

The southern rivers are characterized by larger po rtions of low gradient reaches, and streambeds that are composed of greater amounts of sand a nd fine sediments—are often high in suspended solids, and have neutral to slightly acidic wate rs with high concentratio ns of dissolved organic carbon. Rivers emanating entirely within the Coastal Plain are acidic, low alkalinity, blackwater systems with dissolved organic carbon concentr ations often up to 50 mg/L (Smock and Benke 2005). We described several river basins in detail to provide additional context for evaluating254 the influence of the environmental baseline on listed species under NMFS ’ jurisdiction and the

health of the environment.

Albemarle-Pamlico Sound Complex Natural History. The Albemarle-Pamlico Sound Estu arine Complex, the largest lagoonal estuarine system in the United States, incl udes seven sounds incl uding Currituck Sound, Albemarle Sound, Pamlico Sound and others (EPA 2006b). The Estuarine Complex is separated from the Atlantic Ocean by the Outer Banks, a lo ng barrier peninsula, and is characterized by shallow waters, wind-driven tides that result in vari able patterns of water circulation and salinity. Estuarine habitats include salt marshes, hardwo od swamp forests, and bald cypress swamps.

The Albemarle-Pamlico watershed encompasse s four physiographic regions—the Valley and Ridge, Blue Ridge, Piedmont a nd Costal Plain Provinces. Th e geology of the basin strongly influences the water quality and quantity within the basin. The headwaters of the basin tributaries are generally steep and surface water flowing downstream has less opportunity to pick up dissolved minerals. However, as the surface water flows reaches the Piedmont and Coastal Plain, water velocity slows due to the low gradie nt and streams generally pick up two to three

times the mineral content of surface waters in the mountains (Spruill and Survey 1998). At the same time, much of the upper watershed is composed of fractured rock overlain by unconsolidated and partially consolidated sands. As a result, of the basin’s geology, as a general matter more than half of the water flowing in streams discharging to the Albemarle-Pamilico Estuarine Complex comes from ground water.

Primary freshwater inputs to the Estuary Comp lex include the Pasquotank, Chowan and Roanoke Rivers that flow into Albemarle Sound, and the Tar-Pamlico and Neuse Rivers that flow into Pamlico Sound. The Roanoke River is approximat ely 410 miles long and drains a watershed of 9,580 mi 2 . The Roanoke River begins in the mountains of western Virginia and flows across the North Carolina border before entering the Albemarle Sound. The upper Roanoke River’s geology is primarily a high gradient boulder-rubb le bedrock system. The middle Roanoke River is primarily course sand and gr avel. The lower section of th e Roanoke is almost entirely organic-rich mud. The average precipitation is approximately 43 inches. At the mouth, the average discharge is 5.3 billion gallons each da y, or 8,193 cubic feet per second (Smock and Benke 2005). The Roanoke River is home to 119 fish species, and only seven of those are not native to the area (Smock and Benke 2005). The Ro anoke is also home to nine endangered fish species, two amphibians, and seven mussels, in cluding several important anadromous fish species.

The Neuse River is 248 miles l ong and has a watershed of 6,235 mi 2 (Smock and Benke 2005).

The Neuse River watershed is also located entire ly within the state of North Carolina, flowing through the same habitat as the Cape Fear Rive r, but ultimately entering Pamlico Sound. The river originates in weathered crystalline rocks of the piedmont and crosse s sandstone, shale, and limestone before entering Pamlico Sound (Turekian et al. 1967). The average precipitation is approximately 48 inches. At the mouth, the averag e discharge is 3.4 billion gallons each day, or 5,297 cubic feet per se cond (USGS 2005).

Human Activities and their Impacts255 Land Use. Land use in the Roanoke River is domina ted by forest (68%) and the basin contains some of the largest intact, leas t disturbed bottomland forest floodpl ains along the eastern coast. Only 3% of the basin qualifies as urban land us es, and 25% is used for agriculture (Smock and Benke 2005). The only major town in the Roan oke watershed is Roanoke, Virginia. The population in the watershed is approximately 80 people per square mile (Smock and Benke 2005). In contrast, the Neuse River watershed is described as 35% agri culture, 34% forested, 20% wetlands, and 5% urban, and 6% other, with a basin wide density of approximately 186 people per square mile (Smock and Benke 2005). While the population increased in the Albemarle-Pamlico Complex more than 70% during the last 40 years, the rate of growth is relatively low for many coastal c ounties in the Southeast (EPA 2006b). Much of the estuarine complex is protected by large amounts of state an d federally protected lands, which may reduce development pressures.

Throughout the 20th century, mining, agriculture, paper and pulp mills, and municipalities contributed large quantities of pollutants to the Roanoke River and the Albemarle-Pamlico Estuarine Complex. Even so, today the Albema rle-Pamlico Estuarine Complex is rated in good to fair condition in the National Estuary Program Coastal Condition Report despite that over the past 40-year period data indica te some noticeable changes in the estuary, including increased dissolved oxygen levels, increased pH, decreased levels of suspended solids, and increased chlorophyll a levels (EPA 2006b).

Coal is mined from the mountainous headwaters of the Roanoke River in southwestern Virginia. Mining through the piedmont and coastal areas of North Carolina was conducted for limestone, lead, zinc, titanium, apatite, phosphate, crushe d stone, sand, and fossils. Many active mines in these watersheds are still in operation today. These mines are blamed for increased erosion, reduced pH, and leached heavy metals.

Agricultural activities are major source of nutri ents to the estuary and a contributor to the harmful algal blooms (HABs) in summer, a lthough according to McMahon and Woodside 1997 (EPA 2006a) nearly one-third of the total nitrogen i nputs and one-fourth of the total phosphorus input to the estuary are from at mospheric sources. Primary ag ricultural activities within the watershed include corn, soybean, cotton, peanut , tobacco, grain, potato, and the production of chicken, hog, turkey, and cattle.

In general, the Roanoke River is much cleaner since the passage of the CWA, although mercury, arsenic, cadmium, chromium, copper, lead, nickel, zinc, and PCBs are still considered high (NCDENR 1999). Fish tissues samp led within the estuary also s howed elevated concentrations of total PAHs and total PCBs—10% of the sampled stations exceeded risk-based EPA Advisory Guidance values (EPA 2006b). Water quality studi es in the mid-1990s showed the Neuse Basin contained the highest nitrogen a nd phosphorus yields, while the Chowan Basin had the lowest yields (Spruill and Survey 1998).

The Neuse River entered the national spotlig ht during the early 1990s due to massive and frequent fish kills within the basin. Over one billion American shad have died in the Neuse River since 1991. The problem is persistent but the cause of th e kills differs among events; in 2004 more than 700,000 estuarine fish died and more than 5,000 fresh fish died within the basin.256 Freshwater species most commonly identified during investigations included sunfishes, shad, and carp, while estuarine species most commonly repor ted included menhaden, perch, and croaker. Atlantic menhaden have historically been involved in a majority of estuarine kill events and have exhibited stress and disease in conjunction with fish kills. Fish kill events may often have different causative agents, and in many cases the pr ecise cause is not clea r, but high levels of nutrients, HABs, toxic spills, outbreaks of a marine organism, Pfiesteria pescicida , low DO concentrations and sudden wind changes that mix hypoxic waters, are some of contributing

factors or causes to th e basins persistent fish kills (NCDEQ 2004).

Both the Roanoke River and the Neuse Rivers are fragmented by dams. The reservoirs are used for flood control and recreation, but the amount of agricultural and urba n runoff that collects behind the dams has caused sanitation problems in the recent past. Three dams were removed recently in an effort to improve environmental conditions and fish passage. Widespread stream modification and bank erosion were rated high within the greater watershed relative to other sites in the Nation (Spruill and Survey 1998).

Commercial and Recreational Fishing. The Albemarle and Pamlico Sounds and associated rivers support a dockside commercial fisher y valued at over $54 million annually. The
commercial harvest includes blue crabs, southern flounder, striped bass, striped mullet, white perch, croaker, and spot, among others. Roughl y 100 species are fished commercially or recreationally in the region. The Neuse River su pports many of the same species as the Roanoke

River.

Commercial and recreational fisheries exist for oyster, crab, clam, American shad, American eel, shrimp, and many other species. Shellfish can be collected by dredging, which has adverse effects to benthic organisms, including shortnose sturgeon that use estuar ine areas for feeding. Commercial fisheries along the South Carolina coas t use channel nets, fyke nets, gillnets, seines,

and trawls. All of those methods must use some sort of turtle excluder device, but could still accidentally capture a shortnose sturgeon.

Major Southeast Coastal Plains Basins Natural History. More than five major river basins flow through the Coastal Plains of the Southeast and directly enter the Atlantic O cean including the Cape Fear, Great Pee-Dee, Altamaha, and the St. Johns Rivers (see Table 30 fo r a description of several basins within this region). Rainfall is abundant in the region a nd temperatures are gene rally warm throughout the year. Northern rivers originate in the Blue Ri dge Mountains or the Piedmont Plateau, but all the

rivers described in this section have sizeable reac hes of slack water as they flow through the flat Coastal Plain. Two rivers, The Satilla River in Georgia and the St. Johns River in Florida, are located entirely within the Coastal Plain. The hi ghest elevation of the St. Johns River is 26 feet above sea level, so the change in elevation is essentially one inch every mile, making it one of the most gradually flowing rivers in the country.

Smock et al. (2005) describe the mountains and plateau as areas of heavily dissected and primarily highly metamorphosed rock of Paleozoic age, with occasional areas of igneous and sedimentary rock. Underlying rock is varied with bands of limestone, dolomite, shale, sandstone, cherts, and marble, with a number of springs and caves scattered throughout the area.257 Where the Piedmont Plateau dips the sedimentary deposits of the coastal plain is termed the fall line. Here, steep changes in elevation result in ra pids or falls before the rivers level off in their

Coastal Plain reaches. In the Coastal Plain reaches of the areas rivers soils are acidic with a low cation exchange capacity and a sandy or loamy su rface horizon, and a loamy or clay subsurface. The acidic characteristics, slow flowing water with poor flushing and hi gh organic and mineral inputs gives these waters their characteristic “blackwater” (o r “brownwater” for those that originate in the Piemont Plateau) appearance. Th e Satilla River is a blackwater river that has a naturally low pH (between 4 and 6) and white sa ndbars--due to the low pH it also has naturally lower productivity than other ri vers that originate within th e mountains or the Plateau.

Table 30. General Information abou t Rivers of the Southeast United States (NCDENR 1999, Smock and Benke 2005).

Watershed Length (mi.) Basin

Size (mi 2 ) Physiographic

Provinces* Mean Annual Precipitation (in.) Mean Discharge

(cfs). No. Fish Species No. Endangered Species

Cape Fear River 320 9,324 PP, CP 47 7,663 95 8 fish, 1 mammal, 15 mussels

Great Pee Dee River 430 10, 641 BR, PP, CP 44 13,102 >100 6 fish, 1 reptile

Santee-Cooper River 440 15,251 BR, PP, CP 50 15,327 >100 5 fish, 2 reptiles

Savannah River 300 10,585 BR, PP, CP 45 11,265 >100 7 fish, 4 amphibians, 2 reptiles, 8 mussels, 3 crayfish

Ogeechee River 250 5,212 PP, CP 44 4,061 >80 6 fish, 2 amphibians, 2 reptiles, 1

mussel

Altamaha River 140 (>400) 14,517 PP, CP 51 13,879 93 1 mammal, 12 fish, 2 amphibians, 2 reptiles, 7

mussels, 1 crayfish

Satilla River 200 3,530 CP 50 2,295 52 2 fish, 1 amphibian, 2 reptiles, 1

mussel

St. Johns River 311 8,702 CP 52 7,840 >150 1 mammal, 4 fish, 2 reptiles, 2

birds
* Physiographic Provinces: BR = Blue Ridge, PP = Piedmont Plateau, CP = Coastal Plain

Human Activities and their Impacts

Land Use. Across this region, land use is dominated by agriculture and indus try, and to a lesser extent timber and paper production, although more than half of most basins remain forested. Basin population density is highly va riable throughout the region with the greate st density in the

St. Johns River watershed with about 200 people per square mile of catchment, most of whom are located near Jacksonville, Florida. In contra st, there are only 29 people per square mile in the

Saltilla River watershed in Ge orgia (Smock and Benke 2005). See Table 31 for a summary of land uses and population densities in several area basins across th e region (data from (Smock and Benke 2005).

The largest population centers in the region in clude Miami and Jacksonville, Florida, and Savannah, Georgia. Major towns include Greensboro, Chapel Hill, Fayetteville, South Carolina, and Wilmington, North Carolina in the Cape Fe ar River watershed; Winston-Salem, North Carolina and Georgetown, Florence, and Sumter , South Carolina in th e Great Pee-Dee River Watershed; Charlotte, Hickor y, and Gastonia, North Carolina and Greenville and Columbia, South Carolina in the Santee-Cooper River waters hed; Savannah and Augusta, Georgia, in the Savannah River watershed; Louisville, Statesbo ro, and Savannah, Georgia, in the Ogeechee258 River watershed; Athens, and Atlanta, Georgi a, in the Altamaha River watershed; and Jacksonville, Florida in the St . Johns River watershed.

Several of the rivers in the region have elevated levels of metals including mercury, fecal coliform, bacteria, ammonia, turbidity, and low DO. These impairments are caused by municipal sewage overflows, mining, and non-point source po llution, waterfowl, urban runoff, marinas, agriculture, and industries including textile manuf acturing, power plant operations, paper mills and chemical plants (Harned and Meyer 1983, Berndt et al. 1998, NCDWQ 1998, Smock and Benke 2005).

Several watersheds exhibit high nitrogen loads including the Cape Fear River, Winyah Bay, Charleston Harbor, St. Helena Sound, Savannah River, Ossabaw Sound, Altamaha River, and St. Mary’s River and Cumberland Sound (Brickeret al. 2007). Nitrate concentr ations (as nitrogen) tend to be higher in stream draining basins with agricultural and mixed land uses (Berndt et al. 1998). Based on studies in Georgi a, however, nitrate loads did not vary with growing season of crops (periods of heaviest fert ilizer application), but were infl uenced by high streamflow, which could be related to downstream transport by subsurface flows (Berndt et al. 1998).

Table 31. Land Uses and Population Density in Severa l Southeast Atlantic Basins (Smock and Benke 2005) Watershed Land Use Categories (Percent) Density (people/mi. 2 ) Agriculture Forested Urban Other

Cape Fear River 24 56 9 11 80

The Great Pee-Dee 28 58 8 6 127

Santee-Cooper River 26 64 6 4 168

Savannah River 22 65 4 9 91

Ogeechee River 18 54 1 17 (wetlands) 78

Altamaha River

-- 64 3 7 73

Satilla River 26 72 1 1 29

St. Johns River 25 45 6 24 (wetlands & water) 202

Sediment is the most serious pollutant in the Yadkin (Pee-Dee) River and has historically been blamed on agricultural runoff. In the mid 1990s, farmers in the region began using soil conservation techniques that ha ve reduced sediment inputs by 77% . Unfortunately, the reduction in sediment inputs from farms did not translate to a reduction in sediment in the river, as during this period there was a 25% reduction in ag ricultural land and a 38% increase in urban development.

Mining. Mining occurs throughout the regi on. South Carolina is ranked 25 th in the states in

terms of mineral value and 13 th among the eastern 26 states, and produces 1% of the total nonfuel mineral production value in the United St ates. There are curren tly 13 minerals being

extracted from 485 active mines in South Caroli na alone. Portland and masonry cement and crushed stone were the State’s leading nonfuel minerals in 2004 (NMA 2007). In contrast, Georgia accounts for 4%, Florida accounts for 5% , and North Carolina accounts for 1.76% of the259 total nonfuel mineral production va lue in the United States. No rth Carolina’s le ading nonfuel

minerals in 2004 were crushed st one, phosphate rock, and construc tion sand and gravel. Georgia produces 24% of the clay in the nation; other leading nonfuel mine rals include crushed stone and Portland cement. Florida is the top phosphate rock mining state in the United States and produces about six times more than any othe r state in the nation. Peat and zirconium concentrates are also produced in Florida.

The first gold mine discovered and operated in the United States is outside Charlotte, North

Carolina in the Pee Dee watershe d. Mines through Georgia are al so major producers of barite and crude mica, iron oxide, and feldspar. There is a proposed titanium mi ne near the mouth of the Satilla River. Unfortunately, mines release some toxic materials and negatively impact fish, as fish living around dredge ta ilings have elevated levels of mercury and selenium.

Hydromodification Projects. Several of the rivers within th e area have been modified by dams and impoundments. In contrast to rivers along the Pacific Coast, we found considerable less information on other types of hydromodification pr ojects in this area, such as levees and channelization projects. There are three locks and dams along th e mainstem Cape Fear River and a large impoundment on the Haw River. The lower river and its tributaries are relatively undisturbed. The lower reach is naturally a bl ackwater river with naturally low dissolved oxygen, which is compounded by the reduced flow and stratification caused by upstream

reservoirs and dams. The Yadkin (Pee Dee) River is heavily u tilized for hydroelectric power. There are many dams on Santee-Cooper River System. The Santee River Dam forms Lake Marion and diverts the Santee River to the Cooper River, where another dam, St. Stephen Dam, regulates the outflow of the Sant ee River. Lake Moultrie is fo rmed by both St. Stephen Dam and Pinopolis Dam, which regulates the flow of the Cooper River to the ocean. Below the fall line, the Savannah River is free-flowing with a meande ring course, but above th e fall line, there are three large dams that turn the piedmont secti on of the river into a 100-mile long stretch of reservoir. Although the Altamaha River is undammed, hydropower dams are located in its tributaries the Oconee and Ocmulgee Rivers abov e the fall lines. There are no dams, however, along the entire mainstem Satilla River. There are no major dams on the mainstem St. Johns River either, but one of the larg est tributaries has a dam on it. The St. Johns River’s flow is altered, however, by water diversions for drinking water and agriculture.

Commercial Fishing. The region is home to many commerc ial fisheries targeting species like shrimp, blue crab, clams, American and hickory shad, oysters, whelks, scallops, channel catfish, flathead catfish, snapper, and grouper. Shortnose sturgeon can be caught in gillnets, but gillnets and purse seines account for less than 2% of th e annual bycatch. Shrimpers are responsible for 50% of all bycatch in Georgia waters and of ten interact with sea turtles. There are approximately 1.15 million recreatio nal anglers in the state.

Ballast Water Bioregion CAR-I (Florida to Texas)

This region encompasses states of Alabama, Arkansas, Illinois, Iowa, Kansas, Kentucky, Louisiana, Mississippi, Missouri, Oklahoma, So uth Dakota, Tennessee, th e western portion of Florida including the Florida Keys, and parts of , Georgia, Texas, Minnesota, Montana, North Dakota, Nebraska, Colorado, Indiana, Ohio, New Mexico, North Carolina, Pennsylvania, Virginia, West Virginia, Wisc onsin, Wyoming, Mexico, and two Canadian provinces. Almost260 2/3 of the continental Un ited States drains to the Gulf of Mexico through the Mississippi River

Basin.

While the Mississippi River is the most notable basi n that drains to the Gu lf of Mexico in terms of overall size (and the largest river in the United States) more than ten major river basins flow through to the Gulf including the Atchafalaya, Mobile, Red, Brazos, Colorado, and Rio Grande Rivers several (see Table 34 for a description of several basins within this region). In the following sections, we describe several basins an d estuaries that enter the Gulf of Mexico to characterize the general ecology a nd natural history of the area, and past and current human activities and their impacts on the area.

Natural History. Due to the enormity of the drainage s in this region, several ecoregions are encompassed in this region including the subt ropical, the tropical/subtropical steppe, hot continental and mountain segments, temperate st eppe, and the prairie eco regions (Bailey 1995). Most of the region is within the subtropical ecoregion (d ivision) and is characterized by a humid subtropical climate with particularly high humidity during summer months and warm mild winters. Soils are strongly leac hed and rich in oxides of iron and aluminum (Bailey 1995). The region is forested, largely by sec ond growth forests of longleaf, loblolly and slash pines with inland areas dominated by deciduous trees. Rainfall is moderate to heavy with annual averages

of about 40 inches in the nort h, decreasing slightly in the ce ntral portion of the region, and increasing to 64 inches in southern Florida.

The geology of the eastern Gulf Coast is primar ily sedimentary rocks of both siliclastic (sand, silt, clay) and carbonate (limestone and dolomite) t ypes. Karst is a major mineral in Florida. The piedmont region of Georgia is composed of metamorphosed sedimentary rock and overlaid by decomposed rock called saprolite. Saprolite is rich in aluminum, silicon, and iron oxide. The metamorphosed sedimentary rock is also rich in minerals that intruded during earthquakes millions of years before.

Soils in the eastern Gulf are rich in oxides of iron and aluminum, moister and strongly leached (Bailey 1995), whereas soils in the western Gulf Coast highly varied, and reflect climate and geological differences. Arid pa rts of the region exhibit calcari ous and/or gypsum-rich soils, and tend to have a neutral pH, whereas prairie soils are commonly slightly acidic sandy to clay loams. There is a strong declin e in total rainfall moving east to west, which strongly affects vegetation patterns, river discharg e (see Table 32 – rivers are listed in their general east to west pattern).

Table 32. Select Rivers in the Gu lf Coast Region (Kammerer 1990, Brown et al. 2005, Dahm et al. 2005, Ward and Ward 2005)

Watershed Length (mi. [approx.) Basin Size (mi 2 ) Physiographic

Provinces* Mean Annual Precipitation (inches) Mean Discharge (cfs). No. Fish Species No. Endangered Species

Suwanee River 245 9,640 CP 53 10,804 81 1 fish

Apalachicola River System 106 (>530) 19,571 BR, PP, CP 50 26,804 104 1 fish, 1 reptile

Choctawhatchee River 170 4,646 CP 57 7,487 80 1 fish

Escambia-Conecuh River 231 4,233 CP 65 6,922 102 0261 Watershed Length (mi. [approx.) Basin Size (mi 2 ) Physiographic

Provinces* Mean Annual Precipitation (inches) Mean Discharge (cfs). No. Fish Species No. Endangered Species

Mobile River 774 43,000 CP, VR, AP,
PP, BR 50 67,592 236 12 fish, 3 reptiles, 19 mussels,

7 snails

Pascagoula River 140 (>400) 9,498 CP 61 15,256 119 1 fish, 2 reptiles

Pearl River 409 8, 494 CP 56 13,172 119 1 fish, 2 reptiles, 1 mussel

Mississippi River 2,320 1,151,000 39 450,000 375

Sabine River 555 9,756 CP 50 8, 405 >104 >4 fish, 2 crayfish

Neches River 416 10,011 CP 54 6,321 96 > 4 fish, 1 crayfish

Trinity River 550 17,969 CL, GP, CP 45 7,840 99 3 fish, 1 crayfish, 3 mussels

Brazos River 1,280 44,620 CL, GP, CP 32 8,793 93 >4 fish, 4 mussels

Colorado River 862 39,900 CL, GP, CP 32 2,649 98 >4 fish, 2 salamanders, 1 snake, 5 mussels

San Antonio/Guadalupe

Rivers 408 10,128 GP, CP 32 2,790 88 > 7 fish, several amphibians, 3 spring/cave pool-associated aquatic insects, 1 plant

Nueces River 315 16,800 GP, CP 24 706 > 66 > 3 fish

Rio Grande 1,759 335,908 SR, CO, B/R, GP, CP, SC,

SO 8 1,307 >160 > 16 fish, several mollusks, 6

birds

Physiographic Provinces: BR = Blue Ridge, PP = Piedmont Plateau, CP = Coastal Plain, VR=Valley Ridge, AP=Appalachian Plateau, SR=Southern Rock Mtns., CO=Col orado Plateau, B/R=Basin & Range, GP=Great Plains, SC=Sierra Madre Occidental, SO=Sierra Madre Oriental, CL=Central Lowlands

Human Activities and their Impacts

Land Use. Land use is dominated by forest in th e basins east of the Mississippi, whereas grass/shrub and rangeland uses domin ate in basins west of the Miss issippi. The Mississippi also appears to be a divide between the less developed easte rn basins, and the increasingly urbanized western basins. According to data presented in Table 33, the most developed watersheds are the Trinity River, the San Antonio and Guadalupe Rivers, the Brazos River, the Colorado River, and the Mississippi River. Most of the populati on within the San Antoni o River watershed is concentrated within the greater San Antonio area. Based on data from 2000, the population density of San Antonio is an estimated 1,122 people/mi 2 , and in other areas of the basin density is as little as 16 people/mi 2 (Dahm et al. 2005). The Trinity River Basin encompasses several urban areas including one of the most highly populated areas in the region--the City of Dallas. In stark contrast, overall there are only 29 people pe r square mile in the Neches River watershed (Dahm et al. 2005).

Major threats to the southwestern basins also include wastewatereffluent, water extraction, nonpoint source pollution, nonnative species, ex isting impoundments, and proposals for dams (Dahm et al. 2005), and new reservoirs are proposed fo r some basins (Lane-Miller and DeVries 2007). Municipal waste water discharge poses a serious problem in several rivers, including the Suwannee River basin, and the Chattahoochee and Flint Rivers. According to Dahm et al. (2005) the Rio Grande is one of the most impact ed rivers due to water quality and quantity

concerns. The basin suffers from elevated le vels of salinity, nutrients, bacteria, metals, pesticides, herbicides, organic solvents, and the basin is heavily hydromodified by dams and water diversions for irrigation. About 100 miles downstream of Atlanta the Chatahoochee is262 very polluted, with excessive amount s of nutrients, pesticide, fecal coliform bacteria, PAHs, and oils. The lower Mississippi River is degraded by excess fecal coliform bacteria, PCBs, chlordane, turbidity, siltation, nu trients, reduced DO, pesticides, and eutrophication. Most of the riparian habitat has been lost to agriculture and urban development (Brown et al. 2005).

In many basins agricultural practices associated with row crops (corn, soybeans, hay and cotton) confined animal feeding operations (poultry a nd lifestock—hog, cattle, sh eep, goats), and dairy production are significant source of nutrients, fecal coliform, and pesticides. Other basins are severely impacted by altered sediment regime s. The Choctawhatchee River watershed has highly erodable soils, heavy rains, and intermitte nt droughts that leads to excessive sediment loading. Erosion causes sediment and nutrient issues, while droughts cause low flow and low dissolved oxygen. In contrast, downcutting of reaches of the Brazo s River are a problem resulting from numerous dams interrupting sediment transport w ithin the basin.

Several rivers including the Pascagoula River and its tributaries, and the Sabine River are also impaired by sediment, pathogens, low DO, fecal coliform, nutrients, mercury, PCB, dioxin, ammonia, pesticides like atrazine, and BOD. Occasional fish kills occur within the Colorado

River as a result of storm runoff and low DO. The upper Colorado River has salinity problems and many reservoirs have problems with toxic golden algae (Dahm et al. 2005). The upper Brazos River basin has naturally high salinity, the middle basin has elevated nutrients from nearby dairy farms, several reservoirs have toxi c golden algae, and the lower basin has elevated atrazine, bacteria, phosphorous, and low DO (Dahm et al. 2005). Major polluters in the Mobile River include pulp and paper mills, textiles, ch emical plants, hydroelectric, iron and steel manufacturing, and coal plants.

Pollution of this nature can reduce productivity a nd health of the fish populations within the basin, and at times can lead to fish kills. Sin ce 1998, there have been at least 16 fish kills, at

least one of which was the result of elevated a mmonia levels, two were contributed to pesticides, 10 were from low DO, and 3 were from unknown causes (MSDEQ 2000). Large fish kills are the most severe and usually the most easil y observed response of aquatic ecosystems to pollution, but often the degradation is more elusive occurring at sublethal levels.

Table 33. Land Uses and Population Density in Several Gulf of Mexico Basins (Brown et al. 2005, Dahm et al. 2005, Ward and Ward 2005).

Watershed Land Use Categories (Percent) Density (people/mi. 2 ) Agriculture Forested Urban Other

Suwannee River 30 38 1 9 57

Apalachicola River System 25 55 2 18 (10% wetland) 133

Choctawhatchee River 25 57 1 17 (9% wetland) 46

Escambia-Conecuh River 15 72 <1 12 (7% wetland) 86

Mobile River 18 68 2 12 (7% wetlands) 114

Pascagoula River 17 66 1 16 (11% wetland) 75

Pearl River 24 58 2 15 (12% wetland) 109

Mississippi River 57 28 14

--- 26

Sabine River 10 67 8 15 grassland 47263 Watershed Land Use Categories (Percent) Density (people/mi. 2 ) Agriculture Forested Urban Other

Neches River 15 65 5 15 grassland 29

Trinity River 15 35 30 20 grassland 254

Brazos River 24 3 16 15 grassland 52

Colorado River 30

-- 15 55 range 91

San Antonio and Guadalupe Rive rs 15 -- 25 60 range 220

Nueces River 15

-- 5 55 shrubland 42

Rio Grande River 5 14 7 74 shrub & grass 42

Mining. Mining occurs throughout the region. Mining along the eastern Gulf of Mexico coast is primarily for clay, sand, limestone, phosphate, and peat. There are also some sulfide mines upstream on the Apalachicola River and gravel mines in the Escambia River.

Hydromodification Projects. Several of the rivers within th e area have been modified by dams, impoundments for navigation, levees, and drainage systems. Some rivers on both the eastern and western portion of the Gulf (including the Miss issippi River) have been heavily hydromodified— fragmented by hydroelectric power plants and navigational dams, ch annels have been deepened, straightened, and contained with in levees. For instance, ther e are 13 dams on the mainstem

Chattahoochee and three on the F lint River, but there are no ma jor dams on the Apalachicola River. There are 36 major dams in the Mobile River watershed, and the Trinity River watershed is also highly fragmented with 21 ma jor dams throughout the watershed.

There are more than a 132 dams on the Brazos Rive r—as a result of the dams there has been a reduction in sediment transport to reaches belo w the dams, consequently the river channel has deepened (downcut) resulting in the isolation of the mainstem from several of the oxbow lakes

and off channel habitat once available to the nati ve fishes and other an imals. According to

Dahm et al. (2005), although development is not prev alent in the lower river due to the frequency of flooding, the river is threatened by existing and proposed diversions to the neighboring cities of Houston and Fort Worth. Additionally, dredging activities have been documented to capture or kill 168 sea turtles fr om 1995 to 2009 in the Gulf of Mexico, including 97 loggerheads, 35 Kemp’s ridleys, 32 greens, and three unidentified sea turtles (USACOE 2010).

Commercial and Recreational Fishing. There is an extensive commercial fishery in the Gulf of Mexico. Fishermen fish with gillnets, traw ls, paired trawls, and cast nets. Recreational fishermen are allowed to use hand lines, rod and r eels, spears, and cast nets. This gear poses a risk to gulf sturgeon as a potenti al bycatch species. Gulf of Me xico fishing regulations require special gear to release turt les and smalltooth sawfish.

Ballast Water Bioregions NEP-VI (US/Mex ico Border to Los Angeles) and NEP-V (California north of Los Angeles)

The basins described in this section empty in to the coastal waters of California and are encompassed by the state of Califor nia and parts of Oregon. Select watersheds described herein264 characterize the general ecology and natural history of the area, and the past, present and future human activities and their impacts on the area. Essentially, this region encompasses all Pacific Coast Rivers south of Cape Blanco, California through southern California. The Cape Blanco area marks a major biogeographic boundary and has been identified by NMFS as a DPS/ESU boundary for Chinook and coho salmon, and steelhead on the basis of strong genetic, life history, ecological and habitat differences north and south of this landmark. Major rivers contained in this grouping of watersheds ar e the Sacramento, San Joaquin, Salinas, Klamath, Russian, Santa Ana and Santa Margarita Rivers see Table 34).

Natural History. The physiographic regions covered by the basins discusse d herein, include: (a) the Cascade-Sierra Nevada Mountains provin ce, which extends beyond this region as we

have defined it and continue north into British Co lumbia, (b) the Pacific Border province, and (c) the Lower California province (Carter and Resh 2005). The broader ecoregions division, as defined by Bailey (1995) is th e Mediterranean Division. Thr ee major vegetation types are encompassed by this region: the temperate c oniferous forest, the Mediterranean shrub and savannah, and the temperate grasslands/savanna h/shrub. The area, once dominated by native grasses, is naturally prone to fires set by lig htening during the dry season (Bailey 1995).

This region is the most geologically young and tect onically active region in North America. The Coast Range Mountains are folded and faulted formations, with a variety of soil types and nutrients that influence the hydr ology and biology of the indivi dual basins (Carter and Resh 2005). The region also covers the Klamat h Mountains and the Sierra Nevada.

The climate is defined by hot dry summers and we t, mild winters, with precipitation generally decreasing in southern latit udes although precipitation is strongly influences by topograph and generally increases w ith elevation. Annual precip itation varies from less than 10 inches to more than 50 inches in the region. In the Sierra Nevada about 50% of the precipitation occurs as snow (Carter and Resh 2005), as a re sult snowmelt strongly influences hydrological patterns in the area. Severe seasonal patterns of flooding and drought, and high interannual variation in total precipitation makes the general hydrological pattern highly predic table within a basin, but the constancy is low across years (Carter and Resh 2 005). According to Carter and Resh (2005) this likely increases the variability in the annual composition of the fish assemblies in the region (Table 34).

Table 34. Select Rivers in the Southw est Coast Region (Carter and Resh 2005)

Watershed Length (mi. [approx.) Basin Size (mi 2 ) Physiographic

Provinces* Mean Annual Precipitation (inches) Mean Discharge (cfs). No. Fish Species (native) No. Endangered Species

Rogue River 211 5,154 CS, PB 38 10,065 23 (14) 11

Klamath River 287 15,679 PB, B/R, CS 33 17,693 48 (30) 41

Eel River 200 3651 PB 52 7416 25 (15) 12

Russian River 110 1439 PB 41 2331 41 (20) 43

Sacramento River 400 27,850 PB, CS, B/R 35 23,202 69 (29) >50 T & E spp.

San Joaquin River 348 83,409 PB, CS 49 4,662 63 >50 T & E spp.

Salinas River 179 4241 PB 14 448 36 (16) 42 T & E spp.

Santa Ana River 110 2438 PB 13 60 45 (9) 54265 Watershed Length (mi. [approx.) Basin Size (mi 2 ) Physiographic

Provinces* Mean Annual Precipitation (inches) Mean Discharge (cfs). No. Fish Species (native) No. Endangered Species

Santa Margarita River 27 1896 LC, PB 49.5 42 17 (6) 52
* Physiographic Provinces: PB = Pacific Border, CS = Cascades-Sierra Nevada mountains, B/R=Basin & Range

The San Joaquin River, drains the largest basin in the region, origin ates within the Sierra Nevada

near the middle of California and flows in a nor thwesterly direction through the southern portion of the Central Valley. The alluvial fan of th e Kings River separates the San Joaquin from the

Tulare River basin.

Human Activities and their Impacts

Land Use. Land use is dominated by forest (and vacant land) in northern basins, and grass, shrubland, and urban uses dominate in southern basins (see Table 35). Overall, the most developed watersheds are the Santa Ana, Russia n, and Santa Margarita Rivers. The Santa Ana Watershed encompasses portions of San Bernar dino, Los Angeles, Riverside, and Orange

counties. About 50% of coastal subbasin of th e Santa Ana watershed is dominated by urban land uses and the population density is about 1,500 people per square mile. When steep and

unbuildable lands are excluded from this area, then the population density in the watershed is 3,000 people per square mile. However, the most de nsely populated portion of the basin is near the city of Santa Ana where density reach es 20,000 people per square mile (Burton et al. 1998,

Belitz et al. 2004). The basin is home to nearly 5 m illion people and the population is projected to increase two-fold in the next 50 years (Burton et al. 1998, Belitz et al. 2004).

Not only is the Santa Ana watershed the most h eavily developed watersheds in the region, the Santa Ana is the most heavily populated study site out of more than 50 assessment sites studied across the nation by the United States Geologi cal Survey under the Na tional Water-Quality Assessment (NAWQA) Program. Water quality and quantity in the basin reflects the influence of the high level of urbanization. For instance, the primary source of baseflow to the river is the treated wastewater effluent; secondary source s--sources that influence peak flows—include stormwater runoff from urban, agricult ural, and undeveloped lands (Belitz et al. 2004). Concentrations of nitrates and pe sticides are elevated within the basin, and were more frequently detected than in other national NAWQA sites (Belitz et al. 2004). Belitz et al. (2004)found that total nitrogen concentrations commonly exceeded 3 mg/L in the Santa Ana basin. In other NAWQA basins with elevated total nitrogen concentrations across th e country, the primary influencing factor was the level of agriculture and the applica tion of manure and pesticides

within the basin. In the Santa Ana basin the el evated nitrogen is attr ibuted largely to the wastewater treatment plants, where downstream reaches consistently exceeding 3 mg/L total nitrogen. Samples of total ni trogen taken upstream of the wast ewater treatment plants were commonly below 2 mg/L (Belitz et al. 2004). Other contaminants detected at high levels included volatile organic compounds (VOCs; in cluding chlorform, which sometimes exceeded water quality standards), pesticid es (including diuron, diazinon, carbaryl, chlophyrifos, lindane, malathion, and chlorothalonil), and trace elements (i ncluding lead, zinc, arseni c). As a result of the changes, the biological community in the basin is heavily altered (Belitz et al. 2004).266 Table 35. Land Uses and Population Density in Sev eral Southwest Coast Region (Carter and Resh 2005).

Watershed Land Use Categories (Percent) Density (people/mi. 2 ) Agriculture Forest Urban Other

Rogue River 6 83 <1 9 grass & shrub 32

Klamath River 6 66 <1 24 grass, shrub, wetland 5

Eel River 2 65 <1 31 grass & shrub 9

Russian River 14 50 3 31 (23 grassland) 162

Sacramento River 15 49 2 30 grass & shrub 61

San Joaquin River 30 27 2 36 grass & shrub 76

Salinas River 13 17 1 65 (49 grassland) 26

Santa Ana River 11 57 32

--- 865

Santa Margarita River 12 11 3 71 grass & shrub 135

In many basins, agriculture is the major water user and the major source of water pollution to surface waters. In 1990 nearly 95% of the water diverted fr om the San Joaquin River was diverted for agriculture, and 1.5% diverted for livestock (Car ter and Resh 2005). During the same period, Fresno, Kern, Tulare, and Kings C ounties ranked top in th e nation for nitrogen fertilizer use. Nitrogen fertili zer use increased 500% and phosphor us use increased 285% in the San Joaquin River basin in a 40 y ear period (Knatzer and Sheton 1998 in ) (Carter and Resh 2005). A study conducted by USGS in the mid-1990s on water quality within San Joaquin River basin detected 49 pesticides in the mainstem and three subbasins--22 pesticides were detected in 20% of the samples and concentrations of seven exceeded water quality standards (Dubrovsky et
al. 1998). Water chemistry in the Salinas River is strongly influence by intensive agriculture— water hardness, alkalinity, nutrients and conductivity are high in areas where agricultural uses predominate.

Mining. Famous for the gold rush of the mid 1800s, California has a long history of mining. In 2004, California ranked top in the nation for nonfue l mineral production with 8.23% of the total production (NMA 2007). Today, gold with silver and iron ore comprise only 1% of the production value. Primary minerals include c onstruction sand and gravel, cement, boron and

crushed stone. California is the only state to produce boron, ra re-earth metals and asbestos (NMA 2007).

The State contains some 1,500 abandoned mines a nd roughly 1% is suspected of discharging

metal-rich waters in the basins. The Iron Metal Mine in the Sacramento Basin releases more tha 500 kg of copper and more than 350 kg of zinc to the Keswick Reservoir below Shasta Dam, as well as elevated levels of lead (Cain et al. 2000) (Cain et al. 2000 in)(Carter and Resh 2005). Metal contamination seriously reduces the biolog ical productivity within a basin, can result in fish kills at high levels and at low levels co ntributes to sublethal effects including reduced feeding, overall activity levels, and growth. Th e Sacramento Basin and the San Francisco Bay watershed is one of the most heavily impacted basins within the st ate from mining activities, largely because the basin drains some of the most productive mineral deposits in the region. Methylmercury contamination within San Francisco Bay, the result of 19 th century mining practices using mercury to amalgamate gold in the Sierra Nevada Mountains, remains a persistent problem today. Based on sediment cores, we know that pre-mining concentrations267 were about 5 times lower than concentrati ons detected within the Bay today (Conaway et al. 2003, EPA 2006a).

Hydromodification Projects. Several of the rivers within th e area have been modified by dams, water diversions and drainage sy stems for agriculture and drinking water, and some of the most drastic channelization projects with in the nation. In all, there are about 1,400 da ms within the State of California, more than 5,000 miles of le vees, and more than 140 aqueducts (Mount 1995).

While about 75% of the runoff occurs in basins in the northern half of the State, 80% of the

water demand is in the southern half of the State. Two wate r diversion projects meet these demands—the Federal Central Valley Project an d the California State Water Project. The Central Valley Project, one of the world’s largest water storage and transport systems, has more than 20 reservoirs and delivers about 7 million acre-feet each year to southern California. The State Water Project has 20 major reservoirs a nd holds nearly 6 million acre-feet of water, delivering about 3 million acre feet. Together thes e diversions irrigate about 4 million acres of farmland and deliver drinking water to about 22 million residents.

Both the Sacramento River and the San Joaquin River are heavily modified, each with hundreds of dams. The Rogue, Russian, and Santa Ana Rivers each have more than 50 dams, and the Eel, Salinas, and the Klamath Rivers have between 14 and 24 dams. The Santa Margarita, considered one the last free flowing rivers in coastal southern California has 9 dams in its watershed. All major tributaries of the San Joaquin River ar e impounded at least once and most have multiple dams or diversions. The Stanislaus River, a trib utary of the San Joaquin River has over 40 dams. As a result, the hydrograph of the San Joaquin River is seriously altered from its natural state, the temperature regime and sediment transport re gime are altered, and su ch changes have had profound influences on the biological community within the basin—while the modifications generally result in a reduction of suitable habi tat for native species, these changes frequently

result in a concomitant increase of suitable hab itat for nonnative species. The Friant Dam on the San Joaquin River is attributed with the extirpation of spri ng-run Chinook salmon within the basin, a run once estimated as producing 300,000 to 500,000 fish (Carter and Resh 2005).

Commercial and Recreational Fishing. The region is home to many commercial fisheries. The largest in terms of total landings in 2006 were northern anchovy, Pacific sardine, Chinook salmon, sablefish, Dover sole, Pacific whiting, s quid, red sea urchin, and Dungeness crab (CDFG 2007). Red abalone are also harvested off of th e shores of California. Illegal poaching of abalone, including endangered white abalone continue s to be of concern in the state, with the demand for abalone in local restaurants, seafood markets and international businesses (Daniels and Floren 1998). The first salmon cannery esta blished along the west co ast was located in the

Sacramento River watershed in 1864 but it only operated for about two years because the sediment from hydraulic mining decimated the r uns in the basin (Hittell 1882, and Goode and others, 1884-1887, cited in NRC 1996).

Ballast Water Bioregion NEP-IV (Oregon and Washington Coasts)

This region encompasses Washington, Oregon, Id aho, and includes parts of Nevada, Montana, Wyoming, and British Columbia. The region is ecologically diverse, encompassing northern marine lowland forests, mountain fo rests, alpine meadows and Northe rn desert habitat. In this268 section we focus on three primary areas that ch aracterize the region, the Columbia River Basin and its tributaries, the Puget Sound Region, and the Coastal Drainages north of the Columbia River. The broader ecoregion divisions, as de fined by Bailey (1995), and encompassed within this region are the Marine and Marine Mountains Divisions, por tions of the Temperate Dessert, and Temperate Steppe and Temperate Steppe Mo untains. Puget Sound and the coastal drainages

are contained within the Mari ne Division, while the Columbia River watershed encompasses portions of all five ecoregions.

Columbia River Basin Natural History. The most notable of all basins within the region is the Columbia River. The largest river in the Pacific Northwest and the four th largest river in terms of average discharge the United States drains an area over 258,000 square miles (making it the sixth largest in terms of drainage area), the Columbia River Basin incl udes parts of Washington, Oregon, Nevada, Utah, Idaho, Wyoming, Montana and British Columbia and encompasses 13 terrestrial and three freshwater ecoregions, including arid shrub-steppes, high desert plateaus, temperate mountain forests, and deep gorges (Kammerer 1990, Hinck et al. 2004, Stanford and Synder 2005).

Major tributaries include the Sn ake, Willamette, Salmon, Flathead, and Yakima Rivers; smaller rivers include the Owyhee, Grande Ronde, Clearwater, Spokane, Methow, Cowlitz and the John Day Rivers (see Table 36 for a description of se lect Columbia River Tributaries). The Snake River is the largest tributary at more than 1,000 miles long; its headwaters originating in Yellowstone National Park, Wyoming. The second largest tributary is the Willamette River in Oregon (Kammerer 1990, Hinck et al. 2004). The Willamette River is the 19 th largest river in the nation in terms of average a nnual discharge (Kammerer 1990). The basins drain portions of the Rocky Mountains, the Bitteroot Range , and the Cascade Mountain Range.

The average annual runoff at the mouth of the Columbia River is 265,000 cubic feet per second (cfs)(Kammerer 1990). A saltwat er wedge extends 23 miles upstream of the mouth with tidal influences extending up to 146 miles up river (Hinck et al. 2004). The climate within the basin is a mix of arid, dry summers, cold winters, and maritime air masses entering from the west. It is not uncommon for air temperatures in the Rocky Mountains to dip below zero in mid-winter, but summer air temperatures can reach more than 100 °F in the middle basin.

Table 36. Select Tributaries of the Columbia River (Carter and Resh 2005)

Watershed Length (mi. [approx.) Basin Size (mi 2 ) Physiographic

Provinces* Mean Annual Precipitation (inches) Mean Discharge (cfs). No. Fish Species (native) No. Endangered Species

Snake/Salmon River 870 108,495 CU, NR, MR,

B/R 14 55,267 39 (19) 5 fish (4 T, 1 E), 6 (1 T, 5 E) snails, 1 plant (T)

Yakima River 214 6,139 CS, CU 7 3,602 50 2 (T)

Willamette River 143 11,478 CS, PB 60 32,384 61 (~31) 5 fish (4 T, 1 E),
* Physiographic Provinces: CU = Columbia-Snake River Plateaus, NR = Northern Rocky Mountains, MR = Middle Rocky Mountains, B/R=Basin & Range, CS = Cas cade-Sierra Mountains, PB = Pacific Border

The river and estuary were once home to more th an 200 distinct runs of Pacific salmon and steelhead, and represented adaptatio n to the local environment within a tributary or segment of a river (Stanford and Synder 2005). Salmonids within the basin include Chinook, chum, coho,269 sockeye salmon, steelhead and redband trout, bull trout, and cutthroat trout. Other fish species

within the basin include st urgeon, eulachon, lamprey, and sculpin (Wydoski and Whitney 1979a). According to a review by Stanford et al. (2005), the basin contained 65 native fish species and at least 53 nonnative fishes. The mo st abundant non-native fi sh is the American shad, which was introduced to the basin in the late 1800s (W ydoski and Whitney 1979a).

Human Activities and their Impacts

Land Use. More than 50% of the United State’s portion of the Columbia River Basin is in Federal ownership (most of which occurs in high desert and mountain area s), 39% is in private land ownership (most of which occurs in river va lleys and plateaus), and th e remainder is divided among tribes, state, and local governments (Hinck et al. 2004). See Table 37 for a summary of land uses and population densities in several subbasins within the Columbia River watershed (Stanford and Synder 2005).

Table 37. Land Uses and Population Density in Select Tributaries of the Columbia River (Stanford and Synder 2005)

Watershed Land Use Categories (Percent) Density (people/mi. 2 ) Agriculture Forest Urban Other

Snake/Salmon River 30 10-15 1 54 scrub/rangeland/barren 39

Yakima River 16 36 1 47 shrub 80

Willamette River 19 68 5 -- 171

The interior Columbia Basin has been altered su bstantially by humans causing dramatic changes and declines in many native fish populations. In general the basin suppor ts a variety of mixed uses. Predominant human uses include loggi ng, agriculture, ranchi ng, hydroelectric power generation, mining, fishing and a variety of recreational activities, and urban uses.

The decline of salmon runs in the Columbia is at tributed to loss of ha bitat, blocked migratory corridors, altered river flows and pollution, over ha rvest, and competition from hatchery fish. Critical ecological connectivity (m ainstem to tributaries and ri parian floodplains) has been disconnected by dams and associated activities su ch as floodplain deforest ation and urbanization. The most productive floodplains of the waters hed are either flooded by hydropower dams or dewatered by irrigation diversions. Portions of this basin are also subject to impacts from cattle grazing and irrigation withdrawals. In the Yakima River 72 stream and river segments are listed as impaired by the Washington Department of Ecology and 83% exceed temperature standards. In the Willamette River riparian vegetation was greatly reduced by land conversion. By 1990 only 37% of the riparian area wi thin 120 m was forested, 30% was agricultural fields and 16%

was urban or suburban lands. In the Flathead Ri ver aquatic invasive pl ants such as pondweed, hornwort, watermilfoil, waterweed, cattail and du ckweed grow in the floodplain wetlands and

shallow lakes and in the Yakima River non-nativ e grasses and other plant are commonly found along the lower reaches of the ri ver (Stanford and Synder 2005).

Agriculture and Ranching. Roughly 6% of the annual flow from the Columbia River is diverted for the irrigation of 7.3 m illion acres of croplands within th e basin. The vast majority of these agricultural lands are located along the lower Columbia River, the Willamette, Yakima, Hood, and Snake Rivers, and the Columbia Plateau (Hinck et al. 2004). The Yakima River Basin is one of the most ag riculturally productive areas in the United States (Fuhrer et al. 2004).270 Croplands within the Yakima Basin account for a bout 16% of the total basin area of which 77% is irrigated.

Agriculture and ranching increase d steadily but slowly within the Columbia River basin from the mid to late 1800. By the early 1900s, agricultural opportunities be gan increasing at a much more rapid pace with creation of more irrigation cana ls and the passage of the Reclamation Act of 1902 (NRC 2004). Today, agriculture represents the larg est water use within the basin. More than 105,000 acre feet per day (more than 90 per cent) is used for agri cultural purposes. Agriculture, ranching, and the rela ted services employ more than nine times the national average (19% of the households within the basin; (NRC 2004)).

Ranching practices have led to increased soil erosion and sediment loads within adjacent tributaries, the worst of these effects may have occurred in the late 1800 s and early 1900s with deliberate burning to increase grass production (NRC 2004). Several measures are in use to reduce the impacts of grazing including restrict ing grazing in degraded areas, reduced grazing allotments, and lower stocking rates. Today ag ricultural impacts to water quality within the basin are second to large scal e influences of hydromodifica tion projects for both power generation and irrigation. Water quality impacts from agricultural activities include alteration of the natural temperature regime, and insecticid e and herbicide contamination, and increased suspended sediments.

The US Geological Survey has a number of fixed water quality sampling sites throughout

various tributaries of the Columb ia River, many of which have been in place for decades. Water volumes, crop rotation patterns, crop-type, and lo cation of within the basin are some of the

variables that influence the distri bution and frequency of pesticides within a tributary. Detection frequencies for a particular pesticide can vary widely. One study conducted by the US Geological Survey between May 1999 and Ja nuary 2000, detected 25 pesticide compounds (Ebbert and Embry 2001). Anothe r study detected at least two pesticides or their breakdown products in 91% of the samples collected, with the median number of chemicals being eight, and the maximum was 26. The herbicide 2,4-D occurre d most often in the mixtures, along with azinphos-methyl, the most heavily applied pesticide, and atrazi ne, one of the most mobile pesticides in water (Fuhrer et al. 2004). However, the most freque ntly detected pesticides in the Yakima River Basin are total DDT, as well as its breakdown products DDE and DDD, and dieldrin (Johnson and Newman 1983, Joy 2002, Joy and Madrone 2002, Fuhrer et al. 2004). In addition to current use-chemicals these legacy ch emicals continue to pose a serious problem to water quality and fish communities despite th eir cancellation in the 1970s and 1980s (Hinck et
al. 2004).

Fish and macroinvertebrate commun ities exhibit an almost linear decline in condition as the level of agriculture intensity increases within a basin (Cuffney et al. 1997, Fuhrer et al. 2004). A study conducted in the late 1990s examining 11 species of fish, including anadromous and resident fish collected throughout the basin for a suite of 132 contaminants, which included 26 pesticides revealed organochlorines, specifica lly hexachlorobenzene, chlordane and related compounds, and DDT and its metabolites, were the most frequently detected pesticides within fish tissues (Hinck et al. 2004).271 Urban and Industrial Development. The largest urban area in the basin is the greater Portland metropolitan area, located at the mouth of th e river. Portland’s population exceeds 500,000 people, whereas the next larges t cities, Spokane, Salem, Eugene, and Boise, have more than 100,000 people (Hinck et al. 2004). Overall, however the popul ation within the basin is onethird the average, and while the basin covers about 8% of United States’ land, only about 1.2% of the United States population lives within the basin (Hinck et al. 2004).

Discharges from sewage treatment plants, paper manufacturing, a nd chemical and metal production represent the top three permitted sources of contaminants within the lower basin according to discharge volumes and concentrations (Rosetta and Borys 1996). According to Rosetta and Borys (1996) based on their review of 1993 data, 52% of the point source waste water discharge volume is from sewage treatment plants, 39% from paper and allied products, 5% from chemical and allied products, and 3% from primary metals. However, suspended sediment loading is predominantly from point sources from the paper and allied products industry (71%), while 26% comes from sewage tr eatment plants and 1% is from the chemical and allied products industry. Non-point source di scharges (urban stormwater runoff) account for more of the total pollutant loading to the lowe r basin for most organics and over half of the metals. Although rural non-point sources contribu tions were not calculated, Rosetta and Borys (1996) surmised that in some areas and for some contaminants rural areas may contribute a large portion of the load; this is particularly the cas e for pesticide contamination in the upper river basin where agriculture is the predominant land use.

A study conducted in the late 1990s examining 11 species of fish, including anadromous and resident fish collected throughout the basin for a suite of 132 contaminants, which included 51 semi-volatile chemicals, 26 pesticides, 18 me tals, seven PCBs, 20 dioxins, and 10 furans revealed PCBs, metals, chlorinated dioxins and furans (products of wood pulp bleaching operations) and other contaminants within fish tissues—white sturgeon tissues contained the greatest concentrations of chlori nated dioxins and furans (Hinck et al. 2004).

Hydromodification Projects. More than 400 dams exist in th e basin ranging from mega dams that store large amounts of water to small diversi on dams for irrigation. Every major tributary of the Columbia except the Salmon River is totally or partially regulated by dams and diversions. More than 150 dams are major hydroelectric proj ects of which 18 dams are located on mainstem Columbia River and its major tributary, the Snak e River. The Federal Columbia River Power System encompasses the operations of 14 major da ms and reservoirs on the Columbia and Snake Rivers, operated as a coordinated system. The Army Corps of Engineers operates nine of 10 major Federal projects on the Columbia and Snake Rivers, and Dworshak, Libby and Albeni Falls dams. The Bureau of Reclamation opera tes Grand Coulee and Hungry Horse dams. These Federal projects are a major source of power in the region, and provide flood control, navigation, recreation, fish and wildlife, municipal and i ndustrial water supply, a nd irrigation benefits.

The Bureau of Reclamation has operated irrigati on projects within the basin since the 1904. The irrigation system delivers water to about 2.9 million acres of agricultural lands; 1.1 million acres of land are irrigated using wate r delivered by two structures, the Columbia River Project (Grand Coulee Dam) and the Yakima Project. Grand C oulee Dam delivers water for the irrigation of272 over 670,000 acres of crop lands and the Yakima Project delivers water to nearly 500,000 acres of crop lands (BOR 2007).

The Bonneville Power Administration, an agency of the US Department of Energy, wholesales electric power produced at 31 Federal dams (67% of its production) and non-hydropower facilities in the Columbia-Snake Basin, selling a bout half the electric power consumed in the

Pacific Northwest. The Federal dams were developed over a 37-year period starting in 1938 with Bonneville Dam and Grand Coulee in 1941, a nd ending with construction of Libby Dam in

1973 and Lower Granite Dam in 1975.

Development of the Pacific Northwest regional h ydroelectric power system , dating to the early twentieth century, has had prof ound effects on the ecosystems of the Columbia River Basin (ISG 1996). These effects have been especially advers e to the survival of anadromous salmonids. The construction of the Federal power system modi fied migratory habitat of adult and juvenile salmonids, and in many cases presented a complete barrier to habitat acce ss. Both upstream and downstream migrating fish are impeded by the da ms, and a substantial number of juvenile salmonids are killed and injured during downstrea m migrations. Physical injury and direct mortality occurs as juveniles pa ss through turbines, bypasses, and spil lways. Indirect effects of passage through all routes may incl ude disorientation, stress, delays in passage, and exposure to high concentrations of dissolved gases, warm water, and increase d predation. Dams have also flooded historical spawning and rearing habitat with the creation of massive water storage

reservoirs. More than 55% of the Columbia River Basin that was accessible to salmon and steelhead before 1939 has been blocked by larg e dams (NWPPC 1986). C onstruction of Grand Coulee Dam blocked 1,000 miles of habitat from migrating salm on and steelhead (Wydoski and Whitney 1979a). The mainstem habitats of th e lower Columbia and Willamette Rivers have been reduced primarily to a single channel. As a result, floodplain area is reduced, off-channel

habitat features have been eliminated or disconne cted from the main channel, and the amount of large woody debris in the mainst em has been reduced. Remaini ng areas are affected by flow

fluctuations associated with reservoir mana gement for power generation, flood control and irrigation. Overbank flow events, im portant to habitat diversity, have become rare as a result of controlling peak flows and associated revetments . Consequently, the dynamics of estuary has changed substantially.

Artificial Propagation. There are several artificial pr opagation programs for salmon production within the Columbia River Basin, many of which were instituted under Fede ral law to ameliorate the effects of lost natural produc tion of salmon within the basin from the dams on fishing. The hatcheries are operated by Federal, state, a nd tribal managers. For more than 100 years,

hatcheries in the Pacific Northwest have been used to produce fish for harvest and replace natural production lost to dam construction, and have only minima lly been used to protect and rebuild naturally produced salmonid population (e.g., Redfish Lake sockeye salmon). In 1987, 95% of the coho salmon, 70% of the spri ng Chinook salmon, 80% of the summer Chinook salmon, 50% of the fall Chinook salmon, and 70% of the steelhead returning to the Columbia River Basin originated in hatche ries (CBFWA 1990). More recent estimates suggest that almost half of the total number of smolts produced in the basin come from hatcheries (Mann et al. 2005).273 The impact of artificial propaga tion on the total production of P acific salmon and steelhead has been extensive (Hard et al. 1992). Hatchery practices, among other factors, are a contributing

factor to the 90% reduction in na tural coho salmon runs in the lowe r Columbia River of the past 30 years (Flagg et al. 1995). Past hatchery and stocki ng practices have resulted in the transplantation of salmon and st eelhead from nonnative basins, a nd the impacts of these practices are largely unknown. Adverse effects of these pr actices likely included: the loss of genetic variability within and among populations (Busack 1990 and Riggs 1990 cited in Hard et al. 1992) (Reisenbichler 1997), disease transfer; incr eased competition for food, habitat, or mates; increased predation; altered migration; and di splacement of natural fish (Steward and Bjornn

1990 cited in Hard et al. 1992) (Hard et al. 1992, Fresh 1997); and species with extended freshwater residence are likely to face higher risk of domestication, predation, or altered migration than are species that spend only a brief time in fresh water (Hard et al. 1992) to name a few. Nonetheless, artificial propagation also may contribute to the conservation of listed salmon and steelhead although it is unclear whet her or how much artificial propagation during the recovery process will compromise the distinctiveness of na tural population (Hard et al. 1992).

Currently, NMFS is working on hatchery reform project in the Columbia River Basin, which will

include a collaborative review of how harvest and hatcheri es—particularly federally-funded hatcheries— are affecting the rec overy of listed salmon and steelh ead in the Basin. Eventually, the project team would create a management appr oach that allows tribal, state and Federal managers to effectively manage Columbia Rive r Basin hatcheries to meet conservation and harvest goals consistent with their respective lega l responsibilities. This effort was mandated by Congress in 2005, and is currently in its early stages.

Mining. Most of the mining in the basin is focu sed on minerals such as phosphate, limestone, dolomite, perlite, or metals such as gold, silver , copper, iron and zinc. Mining in the region is conducted in a variety of methods and places within the basin. Alluvial or glacial deposits are often mined for gold or aggregate, and ores are often excavated fr om the hard bedrocks of the Idaho batholiths. Eleven percent of the nation’ s output of gold has come from mining operations in Washington, Montana, and Idaho, and more than half of the nation’s silver output has come from a few select silver deposits with 30% co ming from two deposits located in the Columbia River Basin (the Clark Fork River and Coeur d’Alene deposits; (Hinck et al. 2004, Butterman and Hilliard 2005). According to Wydoski and Wh itney (1979a) one of th e largest mines in the region, located near Lake Chelan, once produced up to 2,000 tons of copper-zinc ore with gold and silver on a daily basis. Most of the phos phate mining within the basin occurs within the headwaters of the Snake River, but the overall output from these deposits accounts for 12% of the United States production of phosphate (Hinck et al. 2004).

Many of the streams and river reaches in the basin are impaired from mining and several abandoned and former mining site s are designated as superfun d cleanup areas (Stanford and Synder 2005, EPA 2007). According to the US Bureau of Mines, there are about 14,000

inactive or abandoned mines within the Columb ia River Basin of which nearly 200 pose a potential hazard to the environmen t (Quigley and Graham 1997). Contaminants that have been

detected in the water include lead and other trace metals. Mining of copper, cadmium, lead, manganese, and zinc in the upper Clark Fork River have contributed wastes to this basin since 1880 (Woodward et al. 1994). Benthic macroinvertebrates and fish within the basin have274 bioaccumulated metals—the exposure and bioaccumulation of these metals in native fishes in the basin are suspected of reducing th eir survival and growth (Farag et al. 1994, Woodward et al. 1994). In the Clark River, several fish kills have occurred since 1984 and are attributed to

contamination from trace metals such as cadmium, copper, lead and zinc (Hinck et al. 2004).

Commercial, Recreational, and Subsistence Fishing. Archeological records indicate that indigenous people caught salmon in the Columbia River more than 7,000 years ago. One of the most well known tribal fishing sites within the basin was located near Celilo Falls, an area in the

lower river that has been occ upied by Dalles Dam since 1957. Salmon fishing increased with better fishing methods and pres ervation techniques, such as drying and smoking, such that harvest substantially increased in the mid-1800s with canning techniques. Harvest techniques also changed over time, from early use of hand-held spears and dip nets, to river boats that used seines and gill-nets, eventually, transitioning to large ocean-going vessels with trolling gear and

nets and the harvest of Columbia River salmon a nd steelhead off the waters of the entire west coast, from California to Alaska (Mann et al. 2005).

During the mid 1800s, an estimated 10 to 16 milli on adult salmon of all species entered the Columbia River each year. Large harvests of returning adult salmon during the late 1800s ranging from 20 million to 40 million pounds of sa lmon and steelhead annually significantly reduced population productivity (Mann et al. 2005). The largest harv est of Chinook salmon ever

recorded occurred in 1883 when Columbia River canneries processed 43 million pounds of salmon (Lichatowich 1999). Commercial landings declined steadily from the 1920s to a low in 1993, when just over one million pounds were harvested (Mann et al. 2005).

Harvested and spawning adults reached 2.8 million in the early 2000s, of which almost half are hatchery produced (Mann et al. 2005). Most of the fish caught in the river are steelhead and spring/summer Chinook salmon, while ocean harv est consists largely of coho and fall Chinook

salmon. Most ocean catches are made north of Ca pe Falcon, Oregon. Over the past five years, the number of spring and fall salmon commercially harvested in tribal fi sheries has averaged between 25,000 and 110,000 fish (Mann 2004). Recreat ional catch in both ocean and in-river fisheries varies around 140,000 to 150,000 fish (Mann et al. 2005).

Puget Sound Region Natural History. The Puget Sound watershed defined by the crest lines of the Olympia Mountain Range (and the Olympic Peninsula) to the west and the Cascade Mountain Range to the east. The Olympic Mountains reach height s of about 8,000 feet above sea level, and are extremely rugged and steeply peaked with abrupt descents into the Puget Lowland. The Cascade Mountains on the east range in heights of 4-8,000 feet above sea level w ith the highest peak, Mount Rainer towering over the re gion at 14,410 feet above sea le vel. As the second largest estuary in the United States, Puget Sound has ab out 1330 miles of shoreline, extends from the mouth of the Strait of Juan de Fuca east, includi ng the San Juan Islands and south to Olympia, and is fed by more than 10,000 rivers and streams.

Puget Sound is generally divided into four majo r geographic marine basins: Hood Canal, South Sound, Whidbey Basin, and the Main Basin. The Ma in Basin has been further subdivided into two sub-basins: Admiralty Inlet a nd Central Basin. Each of the above basins forms a depression275 on the sea floor in which a shallower ledge or si ll separates the relatively deep water from the adjacent basin. The waters of Puget Sound functi on as a partially mixed, two-layer system, with relatively fresh water flowing seaward at the surf ace and salty oceanic water entering at depth. The main ledge of Puget Sound is located at th e north end of Admiralty Inlet where the water shoals to a depth of about 200 f eet at its shallowest point (Kin g County 2001). The deepest point in Puget Sound is found in the Central Basin a nd is over 920 feet. Appr oximately 43% of the Puget Sound’s tideland is lo cated in the Whidbey Island Basin. This reflects the large influence of the Skagit River, which is the largest river in the Puget Sound system and whose sediments are responsible for the extensive mudflats and tidelands of Skagit Bay.

Habitat types that occur within the nearshore envi ronment include eelgrass meadows, kelp forest, mud flats, tidal marshes, subestua ries (tidally influenced portions of river and stream mouths), sand spits, beaches and backshore, banks and bl uffs, and marine riparian vegetation. These habitats provide critical f unctions such as primary food production, support habitat for invertebrates and juvenile and adult fishes, a nd provide foraging and refuge opportunities for

birds and other wildlife.

The Puget Sound ecoregion is a glaciated area cons isting of glacial till, glacial outwash and lacustrine deposits with high quality limestone is found in the San Juan Islands (Wydoski and Whitney 1979a). Relief in the valley is moderate with elevation ranging from sea level to about 1300 feet. Geology in the region consists of mos tly Tertiary sedimentary bedrock formations.

The land and vegetation surrounding Puget Sound wate rs is classified as Puget Lowland Forest and occupies the depression or valley between the Olympic Peninsula on the west and the Cascade Mountains on the east (F ranklin and Dyrness 1973). The alpine zone is expressly devoid of trees. Vegetation changes abruptly along the mountain slopes and across minimal

horizontal distances as a resu lt of steep topography, soil, and microclimate (sun exposure, temperature, and precipitation). Dominant ve getation types include from the Puget lowland region – the lowland forest, the mid-mont ane forest of Pacific silver fir ( Abies amabilis ) with Alaska yellow cedar ( Chamaecyparis nootkatensis ); the subalpine forest of mountain hemlock ( Tsuga mertensiana )with subalpine fir ( Abies lasiocarpa ) and Alaska yellow cedar; and the alpine tundra or meadow above th e treeline (Kruckeberg 1991).

The Puget Sound region has a Medi terranean-like climate, with warm, dry summers, and mild wet winters (Franklin and Dyrne ss 1973). Annual precipitation varies from 28-35 inches, and falls predominantly as rain in lowland areas. Annual snowpack in the mountain ranges is often high—although the elevation of the Olympia Mountains is not as high as that of the Cascade Mountain Range, abundant accumulation occurs, su ch that it will sometimes persist throughout much of the summer months. Average annual ra infall in the north Cascades at Mount Baker Lodge is about 110 inches, and at Paradise Stat ion at Mount Rainer is about 105 inches, while average annual snowfall is 550 inches and 582 inches respectively--sometimes reaching more than 1,000 inches on Mount Rainer (Wydoski and Whitney 1979a, Kruckeberg 1991).

Major rivers draining to Puget Sound from the Cascade Mountains include the Skagit River, the Snohomish River, the Nooksack River, th e Puyallup/Green River, and the Lake Washington/Cedar River watershed. Major rive rs from the Olympic Mountains include the276 Hamma Hamma, the Duckabush, the Quilcene, a nd the Skokomish Rivers. Numerous other smaller rivers drain to the Sound, many of which are significant producers of salmonids despite

their small size.

The Puget Sound basin is home to: more than 200 fish species, representing more than 50 families; more than 140 mammals, of which less than a third are marine mammals. Salmonids within the region include coho salmon, Chinook salmon, sockeye salmon and kokanee, chum salmon, pink salmon, steelhead and rainbow trout, co astal cutthroat trout, bull trout, and Dolly Varden (Wydoski and Whitney 1979a, Kruckeberg 1991). Important commercial fishes include the five Pacific salmon species and several rock fish species. A number of introduced species occur within the region including brown trout, br ook trout, Atlantic salmon, bass, tunicates (sea squirts), and a saltmarsh grass (spa rtina). Estimates suggest that more than 90 species have been intentionally or acciden tally introduced in the region (Ruc kelshaus and McClure 2007). At present over 40 species in the region are listed as threatened and endangered under the ESA.

Human Activities and the Impacts

Land Use. Land use in the Puget Sound lowland is composed of agricultural areas (including forests for timber production), ur ban areas (industrial and residen tial use), and rural areas (low density residential with some agricultural activ ity). In the 1930s, all of Western Washington

contained about 15.5 million acres of “harvestabl e” forest land and by 2004 the total acreage was nearly half that surveyed more than 70 years earlier (PSAT 2007a). Forest cover in Puget Sound alone was about 5.4 million acres in the early 1990s and about a decade later the region had lost another 200,000 acres of forest cover with some watersheds losing more than half the total forested acreage. The most intensive loss of forest cover has occurred in the State’s Urban Growth Boundary, which encompasses specific parts of the Puget Lowl and; in this area forest cover declined by 11.1% between 1991 and 1999 (R uckelshaus and McClure 2007). Projected land cover changes (Ruckelshaus an d McClure 2007) indicate that tr ends are likely to continue over the next several decades with population changes—coniferous forests are projected to decline at an alarming rate as urban uses increase.

The Puget Sound Lowland contains the most densely populated area of Washington. The regional population in 2003 was an estimated 3.8 million people, with 86% residing in King, Pierce and Snohomish Counties (Snohomish, Ce dar-Sammamish Basin, Green-Duwamish, and Puyallup River watersheds), and the area is expe cted to attract four to six million new human residents in the next 20 years (Ruckelshaus and McClure 2007).

According to the State of the Sound report ( PSAT 2007a) in 2001, impervious surfaces covered 3.3% of the region, with 7.3% of lowland areas (below 1,000 feet elevation) covered by impervious surfaces. In one decade, 1991 – 2001 impervious surfaces increased 10.4% region wide. The Snohomish River watershed, one of the fastest growing in the region, increased 15.7% in the same period.

Much of the region’s estuarine wetland losses ha ve been heavily modified, primarily from agricultural land convers ion and urban development (NRC 1996b). Although most estuarine277 wetland losses result from convers ions to agricultural land by ditching, draining, or diking, these wetlands are also experiencing increasing eff ects from industrial and urban causes.

The most extreme case of river delta conversi on is observed in the Duwamish Waterway in Seattle. As early as the mid- 1800s, settlers in the region began discussing the need for a ship canal that linked Lake Washingt on directly with Puget Sound. Af ter several private and smaller attempts, by the early 1900s locks were built achi eving this engineering feat. The resultant outcome was that the Black River, which former ly drained Lake Washington to the Green and White Rivers (at their confluence, these rivers formed the Duwamish River), dried up. The lower White River, which historically migrat ed sporadically between the Puyallup and the Green/Duwamish basins, was permanently diverted into the Puyallup River basin in 1914 with

the construction of concrete dive rsion at river mile 8.5, resulting in a permanent increase of the Puyallup River flows by about 50% and a doubli ng of the drainage area (Kerwin 1999). The Cedar River, on the other hand was permanently diverted to Lake Washington. The oxbow in

the lower Duwamish River was lost with the lo wer river dredging in th e early 1900s reducing the

lower nine miles of the river to 5 miles in le ngth. Overtime the Waterway has been heavily armored and diked, result in the loss of all tidal swamps, 98% of the tidal forests, marshes,

shallows and flats and 80% of th e riparian shoreline (Blomberg et al. 1988).

By 1980, an estimated 27,180 acres of intertidal or shore wetlands had been lost at eleven deltas in Puget Sound (Bortleson et al. 1980). Tidal wetlands in Puget Sound amount to about 17-19% of their historical extent (Collins and Sheikh 2005). Coastal marshes close to seaports and population centers have been espe cially vulnerable to conversion with losses of 50-90% common

for individual estuaries.

More than 100 years of industria l pollution and urban developmen t have affected water quality and sediments in Puget Sound. Many different kinds of activities a nd substances release contamination into Puget Sound and the contributi ng waters. Positive cha nges in water quality in the region, however, are also evident. One of the most notable improvements was the elimination of sewage effluent to Lake Wa shington in the mid 1960s, which significantly reduced problems within the lake from phosphorus pollution and triggered a concomitant reduction in the cyanobacteria (Ruc kelshaus and McClure 2007).

Even so, as the population and industry has rise n in the region a number of new and legacy pollutants are of concern. According to the State of the Sound Report (PSAT 2007a) in 2004, more than 1,400 fresh and marine waters in the region were listed as “impaired.” Almost twothirds of these water bodies were listed as im paired due to contaminants, such as toxics, pathogens, and low dissolved oxygen or high te mperatures, and less than one-third had established cleanup plans; more than 5,000 acres of submerged lands (primarily in urban areas; 1% of the study area) are contaminated with high levels of toxic substances, including polybrominated diphenyl ethers (PBDEs—flame retardants), and roughly one-third (180,000 acres ) the submerged lands within Puget Sound are considered moderately contaminated. PBDEs biomagnified in the food chain, and in th e past 20 years the body burden in harbor seals has increased dramatically from 50 ppb to mo re than 1,000 ppb. Primary pollutants of concern in Puget Sound include heavy metals, organi c compounds, PAHs, PCBs, dioxins, furans, DDT, phthalates, and PBDEs.278 Areas of highest concern in Puget Sound are Southern Hood Canal, Budd Inlet, Penn Cove, Commencement Bay, Elliott Bay, Possession Sound, Sa ratoga Passage, and Sinclair Inlet (DOE 2002). Hypoxic dissolved oxygen concentration (<3 mg/L) were found at se veral (11 out of 54) stations. Dissolved oxygen concentrations less th an 3 mg/L were measured in Hood Canal, Penn Cove, Saratoga Passage, Bellingham Bay, Discovery Bay, Elliott Bay, Strait of Georgia and

West Point. Conditions in South Hood Canal were especially severe, wi th low DO concentration (<5 mg/L) evident year-round. Penn Cove also exhibited re-occurring hypoxia. Low DO was found at 18 other stations, including Saratoga Passage, Discovery Bay, Bellingham Bay, Elliott Bay, Budd Inlet, and Commencement Bay.

In 1989 the Washington State Depa rtment of Ecology (DOE) began a program to monitor marine sediment conditions called the Puget Sound A ssessment and Monitoring Program (PSAMP). The PSAMP is a multi-agency partnership admini stered by the Puget Sound Action Team. From 1989-1995 the Marine Sediment Monitoring Progra m was implemented to characterize baseline sediment quality conditions and trends throughout the Greater Puget Sound area. This was the first large scale evaluation of Puget Sound sedime nt quality at ambient (i.e. away from point sources of contamination) sta tions through the Sound. Eighty-si x stations were established throughout Puget Sound, Hood Canal, the Strait of Ge orgia, and the Strait of Juan de Fuca. Stations were grouped in two cat egories: core stations sample d annually, and rotating stations sampled once every three years alternating between North, Central and South Puget Sound regions. At each station, replicat e sediment samples were collected for the analysis of chemical contaminants, sediment variables, and benthic commun ity structure.

Overall, contaminant concentrations at monitori ng stations were genera lly low and below state sediment quality standards. Metals and semi-volatile organic compounds were most frequently detected. The highest metal a nd organic contamination was found in locations associated with

urban and industrial centers. Low metal concentrat ions were also detected in some rural areas and in deep depositional environments. Contamin ant concentrations occasionally exceeded state regulatory sediment quality standards. However, there was not a consistent pattern across years. An exception was mercury in Sinclair Inlet and Dy es Inlet, with concentr ations above standards for each of the seven years monitored.

By 2000, annual monitoring of sediments at ten hist orical PSAMP stations showed mixed trends in recent years for some chemicals found in se diments (DOE 2005). Less than one third (32 percent) of almost 13,000 chemical measurements made were detected during testing. Those detected most often exceeded sediment quality guidelines in urban embayments: Sinclair Inlet (mercury), Thea Foss Waterway (PAHs).

In general, metals concentrations in 2000 were lower than in 1989 thru 1996 more often than they were higher, while the opposite was true of PAHs (DOE 2005). At the Port Gardner and Inner Budd Inlet station, concentr ations of a number of priori ty pollutant and metals also decreased significantly. Individual PAH levels decreased at the Po int Pully station, but increased significantly at the Belli ngham Bay, Port Gardner, and East Anderson Island stations. Total HPAH and total PAH levels increased significantly at the Strait of Georgia, Bellingham Bay, East Anderson Island, and Budd Inlet stations . These changes may reflect changes in279 anthropogenic input of contaminants to the estuarine system over this 12-year study period. Also, changes in grain size and benthic infauna l community composition seen at the Strait of Georgia station were probably linked to increas ed precipitation and subs equent increased flow and sediment loading from the Fraser River in 1996 and 1997.

From 1997 to 1999, sediments were collected throughout Puget Sound as part of a joint monitoring program conducted by the DOE and NOAA (DOE 2003). Analyses were performed to quantify concentrations of potentially toxic chem icals, responses in laboratory toxicity tests,

and the structure of benthic infauna communities in sediments.

Degraded conditions, as indicated by a combination of relative high chemical concentrations, statistically significant responses in one or more tests of toxic ity, and adversely altered benthos, occurred in samples that repres ented about 1% of the total area (5,700 acres) (DOE 2003). These conditions occurred in samples collected within urbanized ba ys and industrial waterways, especially near the urban centers of Everett, Seattle, Tacoma, and Bremerton, where degraded conditions had been reported in previous studies. Sediments w ith high quality (as indicated by no elevated chemical concentrations, no significa nt responses in the t oxicity tests, and the presence of abundant and diverse infauna and or pollution sensitive taxa) occurred in samples that represented a majority, 68% of the total study area (400,000 acres). Sediments in which

results of the three kinds of anal yses were not in agreement were classified as intermediate in quality and represented about 31% of the total area (179,000 acres).

Although the highly degraded sediments comprise a small percentage of Puget Sound’s area these hot spots upload pollution into the food we b, and the resulting damage to the ecological health and function of the Puget Sound ecosystem may be much greater than the small area
suggest.

Researchers detected arsenic, copper, lead, and mercury throughout the Sound. They found cadmium at 59% of the stations and tributulin, an antifouling chemical f ound in ship hull paint,

at 50% of the stations. PAHs were co mmon while phthlalate esters, PCBs, DDTs and dibenzofurans appeared at fe wer stations (PSAT 2004). Degraded sediments were most prevalent in the Whidbey Basin and Central Sound regions (Everett Harbor, Elliott Bay, Commencement Bay). A higher degree of degr adation in critical ne arshore habitat may disproportionately affect important fish, sh ellfish and aquatic pl ant species (DOE 1997-2003 posters).

The USGS assessed water quality of streams, ri vers and groundwater in the Puget Sound Basin as part of the National Wa ter-Quality Assessment (NAWQA) Program between 1996 and 1998. This assessment focused on the quality of surfac e and ground waters and biological indicators such as fish status, algal status and invertebrate status in rela tion to land use. A widespread detection of pesticide compounds was observed in surface waters of the Puget Sound Basin (Bortleson and Ebbert 2000). Slightly more th an half of the pesticide compounds (26 of 47 analyzed) were detected. The study found that la rge rivers in the Puget Sound Basin were more likely to meet Federal and state guidelines than were small streams (Ebbert et al. 2000). A total of 74 manmade organic chemicals we re detected in streams and rive rs, with different mixtures of chemicals linked to agricultural and urban settings including at razine, prometon, simazine and280 tebuthiuron, carbaryl, diazinon, and malathi on (Bortleson and Ebbe rt 2000). Commonly detected volatile organic compound in the agricult ural land-use study area was associated with the application of fumigants to soils prior to planting (Ebbert et al. 2000). The average concentration of total nitrogen in small streams draining agricultural lands was twice the concentration in streams draining urban areas and over 40 times the concentration in streams draining undeveloped areas (Ebbert et al. 2000). The study concluded that contaminants in runoff from urban and agricultural land surfaces we re major influences on the water quality of streams and rivers (Ebbert et al. 2000), and according to the State of the Sound report water quality impacts from stormwater and wastewater runoff is a major limiting factor in the recovery of salmon and bull trout (Psat 2007b).

Hydromodification Projects. More than 20 dams occur within the region’s rivers and overlap with the distribution of salmonids, and a number of basins contain water w ithdrawal projects or small impoundments that can impede migrating sa lmon. The resultant impact of these and land use changes (forest cover loss and impervious surface increases) has been a significant modification in the seasonal flow patterns of ar ea rivers and streams, and the volume and quality of water delivered to Puget Sound waters. Se veral rivers have been hydromodified by other means including levees and revetments, and bank hardening for erosion control, and agriculture uses. The first dike built in the Skagit Rive r delta was built in 1863 for agricultural development (Ruckelshaus and McClure 2007), other basins like the Snohomish River are diked and have active drainage systems to drain water after high flows that top the dikes. Dams were also built on the Cedar, Nisqually, White, Elwha, Skokomish, Skagit and several other rivers in the early 1900s to supply urban areas with water, prev ent downstream flooding and allow for floodplain activities (like agriculture or development), and to power local timber mills (Ruckelshaus and McClure 2007).

In the next couple of years, however a highly publicized and long discu ssed dam removal project is expected to begin in the Elwha River. The removal of two dams in the Elwha River, a short but formerly very productive salmon river, is ex pected to open up more than 70 miles of high quality salmon habitat (Wunderlich et al. 1994). Estimates sugge stion that nearly 400,000 salmon could begin using the basin within 30 ye ars after the dams are removed (Psat 2007b).

About 800 miles of Puget Sound’s shorelines are hardened or dredged (Psat 2007b). The area most intensely modified is the urban corridor (eastern shores of Puget Sound0 from Mukilteo to Tacoma); here nearly 80% has been altered, mos tly from shoreline armoring associated with the Burlington Northern Railroad tracks (Ruckels haus and McClure 2007). Levee development

within the rivers and their deltas has isolated signifi cant portions of former floodplain habitat that was historically used by salmon and trout during rising flood waters.

Mining. Mining has a long history in the State of Washington, a nd in 2004 the state was ranked

13 th nationally in total nonfuel mineral production value and 17 th in coal production (Palmisano

et al. 1993a, NMA 2007). Metal mining for all metals (e.g., zinc, copper, lead, silver, and gold) peaked in the State be tween 1940 and 1970 (Palmisano et al. 1993a). Today, construction sand and gravel, Portland cement a nd crushed stone are the predominant materials mined. Where sand and gravel is mined from riverbeds (gravel ba rs and floodplains) it may result in changes in channel elevations and patterns, in stream sediment loads, and serious ly alter instream habitat. In281 some cases, instream or floodplain mining has resulte d in large scale river avulsions. The effect of mining in a stream or reach depends upon he rate of harvest and the natural rate of replenishment, as well as flood and precip itation conditions during or after the mining operations.

Commercial and Recreational Fishing. Most of the commercial landings in the region are groundfish, Dungeness crab, shrimp, and salmon. Ma ny of the same species are sought by Tribal fisheries, and by charter, and re creational anglers. Nets and tr olling are used in commercial and Tribal fisheries, whereas recreational anglers typically use hook and line, and may fish from boat, river bank, and docks. Entanglement of mari ne mammals in fishing gear is not uncommon and can lead to mortality or serious injury.

Oregon/Washington Coastal Drainages Natural History. This region encompasses drainages or iginating in the Klamath Mountains, the Oregon Coast Mountains and the Olympic M ountains--the Coast Range ecoregion where elevations range from sea level to about 4,000 feet. More than 15 watersheds drain the region’s steep slopes including the Umpqua, Alsea, Yaquina , Nehalem, Chehalis, Quillayute, Queets, and

Hoh Rivers. Numerous other small to moderate ly sized streams dot the coastline. Many of the basins in this region are relatively small—th e Umpqua River drains a basin of 4,685 sq. miles and is a little over 110 miles long and the Nehalem River drains a basin of 855 sq. miles and is almost 120 miles long—yet represent some of the mo st biologically diverse basins in the Pacific Northwest (Johnson 1999(Kagan et al. 1999, Carter and Resh 2005).

The region is part of a coastal, temperate rainforest system, a nd is characterized by moderate maritime climate marked by long wet seasons with short dry seasons and mild to cool year-round temperatures. Average annual precipitation rang es from about 60 inches to more than 180 inches, much of which falls as rain, and suppor ts a rich temperate forest. Vegetation is characterized by giant coniferous forests of Sitk a spruce, western hemlock, Douglas fir, western red cedar, and red alder and black cottonwood

The Oregon Coast supports a unique coastal sand dune system. The sand dunes were largely created by the sand deposited from the coastal ri vers, in particular the Umpqua and Columbia Rivers. North, steep headlands and cliffs are separated by stretches of flat coastal plain and large estuaries. Significant estuaries in the region (o utside of the Columbia River estuary) include Coos Bay, Tillamook Bay and the Nehalem Rive r Estuary in Oregon, and Grays Harbor, and Willapa Bay in Washington.

Human Activities and their Impacts

Land Use. The rugged topography of the western Ol ympic Peninsula and the Oregon Coastal Range has limited the development of dense population centers. For instance, the Nehalem River and the Umpqua River basins consist of less than 1% urban land uses. Most basins in this region have long been exploited for timber production, and are still dominated by forestlands. In Washington State, roughly 90% of the co astal region is forested (Palmisano et al. 1993a). Approximately 92% of the Nehalem River basi n is forested, with only 4% considered agricultural (Maser and Johnson 1999). Similarl y, in the Umpqua River basin about 86% is282 forested land, 5% agriculture and 0.5% are considered urban la nds—with about half the basin under Federal management (Carter and Resh 2005).

Tillamook County boasts about its dairy farming and cheese production—having a higher density of cows than people but even so, T illamook County like many others in the region is dominated by forested lands (EPA 2006a). Roughly 90% of Tillamook County is forestland, held by Federal and state governments and privat e entities. In the Ne halem Basin, state and private landowners own more than 90% of the forestlands, and about 80% of the private land holdings are large timber companies (Maser and Johnson 1999).

Hydromodification Projects. Compared to other areas in the greater Northwest Region, the coastal region has fewer dams and several rive rs remain free flowing (e.g., Clearwater River). The Umpqua River is fragmented by 64 dams, the fewest number of dams on any large river basin in Oregon (Carter and Resh 2005). According to Palmisano et al. (1993b) only about 30 miles of salmon habitat are permanently blocked by dams in the coastal st reams of Washington.

In the past, temporary splash dams were construc ted throughout the region to transport logs out of mountainous reaches. The ge neral practice involved building a temporary dam in the creek adjacent to the area being logged, the pond was fill ed with logs and when the dam broke the floodwater would carry the logs to downstream r eaches where they could be rafted and moved to market or downstream mills. Thousands of spla sh dams were constructed across the Northwest in the late 1800s and early 1900s. While the da ms typically only temporarily blocked salmon habitat, in some cases they remained long enoug h to wipe out entire r uns, the effects of the channel scouring and loss of channe l complexity resulted in the long term loss of salmon habitat (NRC 1996a).

Mining. Oregon is ranked 35 th nationally in total nonfuel mineral production value in 2004, while Washington was ranked 13 th nationally in tota l nonfuel mineral pro duction value 2004 and

17 th in coal production (Palmisano et al. 1993a, NMA 2007). Metal mining for all metals (e.g., zinc, copper, lead, silver, and gold) peaked in Washington between 1940 and 1970 (Palmisano et
al. 1993a). Today, construction sand and gravel , Portland cement and crushed stone are the predominant materials mined in both Washington and Oregon. Where sand and gravel is mined from riverbeds (gravel bars and floodplains) it ma y result in changes in channel elevations and patterns, instream sediment loads, and seriously al ter instream habitat. In some cases, instream or floodplain mining has resulted in large scale river avulsions. Th e effect of mining in a stream or reach depends upon he rate of harvest and the natural rate of replenishment, as well as flood and precipitation conditions during or after the mining operations.

Commercial and Recreational Fishing. Most of the commercial landings in the region are groundfish, Dungeness crab, shrimp, and salmon. Ma ny of the same species are sought by Tribal fisheries, and by charter, and re creational anglers. Nets and tr olling are used in commercial and Tribal fisheries, whereas recreational anglers typically use hook and line, and may fish from boat, river bank, and docks. Entanglement of mari ne mammals in fishing gear is not uncommon and can lead to mortality or serious injury.



### Marine Systems (Bioregions NA-S1, SP -IX, SP-XII, SP-XXI, NEP-I, NEP-II)283 Hurricanes The Gulf of Mexico, Caribbean Sea, and southern US Atlantic seaboard is prone to major tropical weather systems, including tropical stor ms and hurricanes. The impacts of these storms

on sea turtles in the marine environment is not known, but storms can cause major impacts to sea turtle eggs on land, as nesting frequently overlaps with hurricane seas on, particularly Kemp’s ridley sea turtles (NRC 1990a). Mortality can result both from drowning of individuals while still in the egg or emerging from the nest as well as causing major topographic alteration to beaches, preventing hatchling entry to marine waters . Kemp’s ridley sea turtles are likely highly sensitive to hurricane impacts, as their only ne sting locations are in a limited geographic area along southern Texas and northern Mexico (Milton et al. 1994). In 2010, Hurricane Alex made landfall in this area; su rprisingly, few nests were lost (Jaime Pena, Gl adys Porter Zoo, pers. comm.). Tropical storm Hermine a rrived too late in 2010 to impact eggs or hatchlings at Rancho Nuevo (Donna Shaver, NPS, pers. comm.).

Habitat degradation A number of factors may be dire ctly or indirectly affecting listed species in the action area by degrading habitat; perhaps most significant among them is anthropogenic noise in the ocean. Natural sources of ambient noise include: wind, waves, surf noise, precipitation, thunder, and biological noise from marine mammals, fishes, and crustaceans. Anth ropogenic sources of ambient noise include: transportation and shi pping traffic, dredging, c onstruction activities, geophysical surveys, and sonars. In general, it has been asserted that ocean background noise levels have doubled every decade for the last si x decades in some areas, primarily due to shipping traffic (IWC 2004). The acoustic noise that commercial traffic contributes to the

marine environment is a concern for listed species because it may impair communication between individuals (Hatch et al. 2008).

Vessel noise could affect marine animals in the action area. Shipping and seismic noise generally dominates ambient noise at fr equencies from 20 to 300 Hz (Richardson et al. 1995b,

Andrew et al. 2002, Hildebrand 2009). Background noise ha s increased significantly in the past 50 years as a result of increasing vessel traffic, and particularly shippin g, with increases of as much as 12 dB in low frequency ranges a nd 20 dB versus preindustrial periods (NRC 1994,

Richardson et al. 1995b, NRC 2003, Jasny et al. 2005, NRC 2005, McDonald et al. 2006, Hildebrand 2009). Over the past 50 years, the number of co mmercial vessels has tripled, carrying an estimated six times as much cargo (requiring larger, more powerful, and consequently louder vessels) (Hild ebrand 2009). Seismic signals also contribute significantly to the low frequency ambient sound field (Hildebrand 2009). Baleen whales may be more sensitive to sound at those low frequencies than are toothed whales. Dunlop et al. (2010) found that humpback whales shifted from using vocal communication (which carries relatively large amounts of information) to su rface-active communicat ion (splashes; carry relatively little information) when low-frequency background noise in creased due to increased sea state. Sonars and small vessels also contribute significan tly to mid-frequency ranges (Hildebrand 2009).

The northeastern US hosts som e of the busiest co mmercial shipping lanes in the world, including those leading into Boston, Providence, Newar k, and New York (MARAD 2011). In addition to vessel traffic, marine construction activities occu r in the Cape Cod area (liquefied natural gas284 terminal construction, pile driving, offshore wind farm construction, dredging, cable laying, drilling, and others) that c ontributes to local and regi onal background sound levels.

Commercial shipping in the Gulf of Alaska is dominated by cargo transports, container freight, crude oil tankers, and barges. Military vessels, ferries, and ot her commercial and recreational fishing vessels also converge in the Gulf of Alaska. Two primary shipping lanes radiate from the

Gulf of Alaska to Honolulu and San Francisc o. Important Alaskan ports include Kodiak, Alaska’s largest commercial fi shing port, and Valdez, the sout hern terminus of the 1,300 km trans-Alaska pipeline. Additional minor ports are located throughout the region and include: Anchorage, Cordova, Homer, Kodiak, Ni kiski, Seward, Whittier, and Yakutat.

The Puget Sound and nearby waters experience ve ry high levels of ve ssel traffic from both

commercial and recreational source s, producing the potential for sh ip strike, high ambient noise levels, and behavioral harassment of southern re sident killer whales. Commercially, a quarter million vessels move within the Puget Sound re gion annually, with the Ports of Seattle and Tacoma combining to be the third largest po rt in the US (www.wash ingtonports.org). These vessels include tankers, tugs, cargo containers, fe rries, and a variety of other vessel types. Several cruise ships are also based out of S eattle. Recreationally, 244 marinas, nearly 40,000 moorage slips, and 331 boat launches are located within the Pacific No rthwest, servicing 180,000 registered recreational vessels and countless vessels not requ iring registration (WSDE 2006). Haro Strait, one of the regions primary shipping lanes, is freque ntly used by southern resident

killer whales.

Several major ports occur further south along the US west coast, including Portland, San Francisco, Los Angeles, Long Beach, and San Di ego (DoT 2005). These ports service a wide variety of vessels, including car go, tug and barges, small ships, liquid bulk, dry bulk, break bulk, intermodal (container, roll-on/roll-off, lighter aboard ship), ferry, tourist passenger vessels (sailboats, ferry, party-boat fishing, whale watching) and cruise ships. Long Beach is among the largest ports in the US, accoun ting for 6% of the total cargo entering the US, and increasing rapidly (growing 122% between 2003 and 2006) (DoT 2007b, a). Los Angeles is also the fifth largest cruise ship terminal in the US. A shi pping lane runs along the US west coast south to southern California and additional shipping lanes extend westward from San Francisco and near Santa Barbara Island. While there are many major ports along the Pacific Coast, the states of Washington, Oregon, and Californi a have proposed more stringe nt ballast water discharge standards than those proposed by the USCG’s ba llast water management program. Upon initial review, the USCG determined these standards were not practicable and to date they have not been implemented. As part of the USCG’s ballast water management program, they will assess the performance of all technologies every three years in their prac ticability and periodic reviews to determine whether it is practicable to impl ement a more stringent ballast water standard.

In-water construction activities (e.g., pile driving associated w ith shoreline projects) in both

inland waters as well as coastal waters in the action area can produce sou nd levels sufficient to disturb proposed and listed species under some conditions. Pressure levels from 190-220 dB re 

### 1

μ Pa were reported for piles of different si zes in a number of studies (NMFS 2006b). The majority of the sound energy associated with p ile driving is in the low frequency range (<1,000 Hz) (Illingworth and Rodkin Inc. 2001, Reyff 2003, Illingworth and Rodkin Inc. 2004).285 Dredging operations also have the potential to emit sounds at levels that could disturb individuals of many taxa. Depending on the type of dredge, peak sound pressure levels from 100 to 140 dB

re 1 μ Pa were reported in one study (Clarke et al. 2003). As with pile driving, most of the sound energy associated with dredging is in the low-frequency range, < 1000 Hz (Clarke et al. 2003). Three large, in-water construction projects are kno wn to be underway or underdevelopment that have the potential to impact No rth Atlantic right whales. A 130-turbine wind farm is proposed

or reviewed for construction o ff Long Island, New York and a nother in Massachusetts Bay. Both projects would involve pile driving that produces large amounts of sound in the frequency range used by North Atlantic right whales. As this sound would likel y persist for extended periods, there is the potential for North Atlantic right whales to abandon local areas in favor of areas where they can better us ed their primary mechanism for gaining information about their environment. Although neither ar ea co-occurs locally with high No rth Atlantic right whale use, individuals do forage in waters near the Massachusetts site and may migrate past the New York

site to and/or from their south eastern/northeastern US breeding and foraging areas. In addition, a liquefied natural gas terminal is planned for c onstruction off Gloucester, Massachusetts. This project would involve similar stressors, but locat ed in prime North Atlantic right whale foraging

habitat.

Several measures have been adopt ed to reduce the sound pressure le vels associated with in-water construction activities or preven t exposure of marine mammals to sound. For example, a sixinch block of wood placed between the pile and the impact hammer used in combination with a bubble curtain can reduce sound pressure levels by about 20 dB (NMFS 2008e). Alternatively, pile driving with vibratory hammers produces pe ak pressures that are about 17 dB lower than those generated by impact hammers (Nedwell an d Edwards 2002). Other measures used in the

action area to reduce the risk of disturbance from these activities include avoidance of in-water construction activities during times of year when marine mammals may be present; monitoring

for marine mammals and sea turt les during construction activities ; and maintenance of a buffer zone around the project area, within which sound-producing activ ities would be halted when marine mammals enter the zone (NMFS 2008e).

The state of marine and estuarin e wetlands are a particular concer n because of the reliance that listed anadromous fish have on them (Qui nn 2005). Wetlands are the most diverse and productive systems in the marine environment, supporting 80% of Earth’s fish and shellfish through feeding, nesting, shelter, high tide refuge, spawning grounds, and nursery habitat (Bertness 1999, RAE/ERF 1999, Airamé et al. 2003, Quinn 2005). Washington’s wetlands serve as designated critical habitat for one-third of the state’s listed species, including ESA-listed salmonids (Lane and Taylor 1997).

Ingestion of marine debris can have fatal conseq uences even for large whales as well as sea turtles. In 1989, a stranded sperm whale along the Mediterranean was found to have died from ingesting plastic that blocked its’ digestive tract (Viale et al. 1992). A sperm whale examined in

Iceland had a lethal disease t hought to have been caused by the complete obstruction of the gut with plastic marine debris (Lambertsen 1990). Th e stomach contents of two sperm whales that stranded separately in California included extens ive amounts of discarded fishing netting (NMFS 2009). A fifth individual from the Pacific was found to contain nylon netting in its stom ach when it washed ashore in 2004 (NMFS 2009). Further incidents may occur but remain286 undocumented when carcasses do not strand.

For sea turtles, marine debris is a problem due primarily to individuals ingesting debris and blocking the digestive tract, causing d eath or serious injury (Lutcavage et al. 1997a, Laist et al. 1999). Gulko and Eckert (2003) estimated that betw een one-third and one-half of all sea turtles ingest plastic at some point in their lives; this figure is support ed by data from Lazar and Gracan (Lazar and Gra č an 2010), who found 35% of loggerheads ha d plastic in their gut. A Brazilian study found that 60% of stranded gree n sea turtles had ingested mari ne debris (primarily plastic and oil; (Bugoni et al. 2001)). Loggerhead sea turtles had a lesser frequency of marine debris ingestion. Plastic is possibly ingested out of curi osity or due to confusio n with prey items; for example, plastic bags can resemble jellyfis h (Milton and Lutz 2003). Marine debris consumption has been shown to depress growth rates in post-hatchling loggerhead sea turtles, elongating the time required to reach sexual matur ity and increasing predation risk (McCauley and Bjorndal 1999). Studies of shore cleanups ha ve found that marine debris washing up along

the northern Gulf of Mexi co shoreline amounts to about 100 kg/km (ACC 2010, LADEQ 2010, MASGC 2010, TGLO 2010). Sea turtles can also become entangled and die in marine debris,

such as discarded nets a nd monofilament line (O'Hara et al. 1988, NRC 1990a, Lutcavage et al. 1997a, Laist et al. 1999). This fundamentally reduces the reproductive potential of affected populations, many of which are alre ady declining (such as loggerh ead and leatherback sea turtle populations in the action area).

Invasive species have had majo r impacts on aquatic, estuarine, and marine ecosystems. As information presented here would be larg ely duplicative of that presented in the Effects analysis

, we detail the impacts of invasive species to listed resources in the Effects analysis

.

Oil and Gas Development

The Arctic Ocean off Alaska and Gulf of Mexi co are the epicenters for marine oil and gas development and extraction within the action area. The Arctic’s major sources of industrial underwater noise appear to be offshore oil, gas or mineral exploration a nd exploitation. These activities increase vessel traffic, produce loud so unds for seismic profiling, place structures in areas used by whales, and introduce noises from drilling and production into the environment (NMFS 1999, 2006f). Malme et al. (1985) exposed feeding humpb ack whales in southeastern Alaska to noise from a single air gun or to pla yback of recorded sounds of oil drilling, production platforms and aircraft. Whales showed no overa ll pattern of avoidance during 13 experiments,

each of which included between 10 and 40 different animals. Whales startled as soon as the airgun was turned on in three experiments. Thes e startle responses, which occurred at received sound levels between 150 to 169 dB (re 1 mPa), were thought to be caused more by the novelty of the air gun sound than by its intensity. There are presently two offshore production facilities

within state waters in the Beaufort Sea: Nort hstar and Endicott. Multiple marine geophysical (seismic) projects are planned for th e Beaufort and Chukchi Seas.

The northern Gulf of Mexico is the location of ma ssive industrial activity a ssociated with oil and gas extraction and processing. Ov er 4,000 oil and gas structures are located outside of state waters in the northern Gulf of Mexico; 90% of these occur off Louisiana and Texas (USN 2009).

This is bo th detrimental and beneficial for sea tu rtles. These structures appreciably increase the

amount of hard substrate in the marine envir onment, providing shelter and foraging opportunities for species like loggerh ead sea turtles (Parker et al. 1983, Stanley and Wilson 2003). However,287 the Bureau of Ocean Energy Management requires that structures must be removed within one year of lease termination. Many of these structures are removed by explosively severing the underwater supportive elements, which produces a s hock wave that kills, injures, or disrupts marine life in the blast radius (Gitschlag et al. 1997). For sea turtles, this means death or serious injury for individuals within a few hundred me ters of the structure and overt behavioral (potentially physiological) impacts fo r individuals further out (Duronslet et al. 1986, Klima et al. 1988). Although observers and procedures are in pl ace to mitigate impacts to sea turtles (i.e., not blasting when sea turtles are present), not all s ea turtles are observed all the time and low-level sea turtle injury and mortality still oc curs (Gitschlag and Herczeg 1994, Gitschlag et al. 1997); two loggerheads were killed in August 2010 (G . Gitschlag, NOAA, pers. comm.). Current

annual authorized takes due to Bureau of Ocean Energy Management OCS oil and gas exploration, development, production, and abando nment activities are 30 sea turtles, including no more than one each of Kemp's ridley, green, or hawksbill turtles and no more than ten loggerhead turtles (NMFS 1988). These levels were far surpassed by the Deepwater Horizon incident (see oil spills and rele ases below). Overall, these ac tivities provide both positive and negative effects at the individua l level and have no clear impact at the population and species

levels.

Oil pollution has been a significant concern in the Gulf of Mexico for several decades due to the large amount of extraction and re fining activity in the region. Routine discharges into the northern Gulf of Mexico (not including oil spills) include ro ughly 88,200 barrels of petroleum per year from municipal and industrial wastewater treatment plants and roughly 19,250 barrels from produced water discharged overboard during oil and gas operations (MMS 2007b, USN 2008). These sources amount to over 100,000 barrels of petroleum discharged into the northern Gulf of Mexico annually. Alt hough this is only 10% of the am ount discharged in a major oil spill, such as the Exxon Valdez spill (roughly 1 million barrels), th is represents a significant, continual, and “unseen” threat to Gulf of Mexico wildlife and habitats. Generally, accidental oil spills may amount to less than 24,000 barrels of o il discharged annually in the northern Gulf of Mexico, making non-spilled oil normally one of the le ading sources of oil discharge into the Gulf of Mexico, although incidents such as the 2010 Deepwater Horizon incident are exceptional (MMS 2007a). The other major source from year to year is oil natura lly seeping into the northern Gulf of Mexico. Although exact figures are unknown, natural seapage is estimated at between 120,000 and 980,000 barrels of oil annually (MacDonald et al. 1993, MMS 2007b).

Although non-spilled oil is the primary contributor to oil introduced into the Gulf of Mexico, concern over accidental oil spills is well-founded. Over five million barrels of oil and one million barrels of refined petroleum products are transported in the nort hern Gulf of Mexico daily (MMS 2007b); worldwide, it is estimated th at 900,000 barrels of oil ar e released into the environment as a result of oil and gas activitie s (Epstein and (Eds.). 2002). Even if a small fraction of the annual oil and gas extraction is re leased into the marine environment, major, concentrated releases can result in significant environmental impacts. Due to the density of oil extraction, transport, and refining facilities in the Houston/Galveston a nd Mississippi Delta areas (and the extensive activ ities taking place at these facilities), these locations have the greatest probability of experiencing oil spills. Oil rel eased into the marine environment contains aromatic organic chemicals known to be toxic to a variety of mari ne life; these chemicals tend to dissolve into the air to a gr eater or lesser extent, dependi ng upon oil type and composition288 (Yender et al. 2002). Solubility of toxic components is generally low, but does vary and can be relatively high (0.5-167 part s per billion; (Yender et al. 2002)). Use of dispersants can increase oil dispersion, raising the levels of toxic constituents in the wa ter column, but speeding chemical degradation overall (Yender et al. 2002). The remaining oil becomes tar, which forms floating balls that can be transported t housands of kilometers into the North Atlantic. The most toxic

chemicals associated with oil can enter marine food chains and bioaccumulate in invertebrates such as crabs and shrimp to a small degree (prey of some sea turtles (Marsh et al. 1992, Law and Hellou 1999)), but generally do not bioaccumulate or biomagnify in finfish (Varanasi et al. 1989,

Meador et al. 1995, Baussant et al. 2001, Yender et al. 2002). The loss of invertebrate communities due to oiling or oil toxicity would also decrease pr ey availability for hawksbill, Kemp’s ridley, and loggerhead sea turtles (NOAA 2003). Furthermore, Kemp’s ridley and loggerhead sea turtles, which commonly forage on crustaceans and mollusks, may ingest large amounts of oil due oil adhering to the shells of these prey and th e tendency for these organisms

to bioaccumulate toxins found in oil (NOAA 2003). It is suspected that oil adversely impacted the symbiotic bacteria in the gut of herbivor ous marine iguanas when the Galapagos Islands experienced an oil spill, contributing to a >60% decline in local populati ons the following year. The potential exists for green sea turtles to e xperience similar impacts, as they also harbor symbiotic bacteria to aid in their digestion of plant material (NOAA 2003). Seagrass beds may be particularly susceptible to oi ling as oil contacts grass blades and sticks to them, hampering photosynthesis and gas exchange (Wolfe et al. 1988). If spill cleanup is attempted, mechanical damage to seagrass can result in further injury and long-term scarring. Loss of seagrass due to oiling would be important to green sea turtles, as this is a significant component of their diets (NOAA 2003). Sea turtles are known to ingest and attempt to ingest tar balls, which can block their digestive systems, impairing foraging or digestion and potential ly causing death (NOAA 2003). Dispersants reduce the formation of tar balls. Although the effects of dispersant chemicals on sea turtles is unknown, testing on ot her organisms have found currently used dispersants to be less toxic th an those used in the past ( NOAA 2003). It is possible that dispersants can interfere with surf actants in the lungs (surfactants prevent the small spaces in the lungs from adhering together due to surface te nsion, facilitating large surface areas for gas exchange), as well as interfere with digesti on, excretion, and salt gla nd function (NOAA 2003). Oil exposure can also cause acute damage upon direct exposure to oil, including skin, eye, and respiratory irritation, reduced respiration, burns to mucous membranes such as the mouth and eyes, diarrhea, gastrointestinal ulcers and bleeding, poor digestion, anemia, reduced immune response, damage to kidneys or liver, cessati on of salt gland function, reproductive failure, and death (Vargo et al. 1986c, b, a, NOAA 2003, 2010b). Nearshore spills or large offshore spills

can oil b eaches on which sea turtles lay their eggs , causing birth defects or mortality in the nests (NOAA 2003, 2010b).

Several oil spills have impacted the northern Gulf of Mexico over the past few years, largely due to hurricanes. The impacts of Hurricane Iv an in 2004 on the Gulf Coast included pipeline damage causing 16,000 barrels of oil to be rel eased and roughly 4,500 barrels of petroleum products from other sources (USN 2008, BOEMRE 2010). The next year, Hurricane Katrina caused widespread damage to onshore oil storag e facilities, releasi ng 191,000 barrels of oil (LHR 2010). Another 4,530 barrels of oil were released from 70 other smaller spills associated with hurricane damage. Shortly thereafter, Hurricane Rita damaged offshore facilities resulting in 8,429 barrels of oil to be released (USN 2008).289 Major oil spills have impacted the Gulf of Mexico for decades (NMFS 2010). Until 2010, the largest oil spill in North America occurred in the Bay of Campeche (1979), when a well “blew out”, allowing oil to flow into the marine e nvironment for nine months, releasing 2.8-7.5 million

barrels of oil. Oil from this release eventual ly reached the Texas coas t, including the Kemp’s ridley sea turtle nesting beach at Rancho Nue vo, from where 9,000 hatchlings were airlifted and released offshore (NOAA 2003). Over 7,600 m 3 of oiled sand was eventually removed from Texas beaches and 200 gallons of oil were removed from the area around Rancho Nuevo (NOAA 2003). Eight dead and five live sea turtles were recove red during the oil spill event; although cause of deaths were not determined, o iling was suspected to play a part (NOAA 2003). Also in 1979, the oil tanker Burmah Agate collided with another ve ssel near Galveston, Texas, causing an oil spill and fire th at ultimately released 65,000 barr els of oil into estuaries, beachfronts, and marshland along the northern and central Texas coastline (NMFS 2010). Clean-up of these areas was not attempted due to the environmental damage such efforts would have caused. Another 195,000 barrels of oil are estimated to have been burned in a multi-monthlong fire aboard the Burmah Agate (NMFS 2010). The tanker Alvenus grounded in 1984 near Cameron, Louisiana, spilling 65,500 barrels of oil which spread west along the shoreline to Galveston (NMFS 2010). One oiled sea turtle was recovered and released (NOAA 2003). In 1990, the oil tanker Megaborg experienced an accident near Galveston during the lightering process and released 127,500 barrels of oil, most of which burned off in the ensuing fire (NMFS 2010).

On April 20, 2010, a fire and explosion occurred aboard the semisubmersi ble drilling platform Deepwater Horizon roughly 80 km southeast of the Mississippi Delta (NOAA 2010a). The platform had 17,500 barrels of fuel aboard, wh ich likely burned, escaped, or sank with the platform (NOAA 2010a). However, once the pl atform sank, the riser pipe connecting the platform to the wellhead on the seafloor broke in multiple locations, initiating an uncontrolled release of oil from the exploratory well. Over the next three months, oil was released into the Gulf of Mexico, resulting in oi led regions of Texas, Louisian a, Mississippi, Alabama, and Florida and widespread oil slicks throughout the nor thern Gulf of Mexico that closed more than

one-third of the Gulf of Mexico Exclusive Ec onomic Zone to fishing due to contamination concerns. Apart from the widespread surf ace slick, massive undersea oil plumes formed, possibly through the widespread us e of dispersants, and reports of tarballs washing ashore throughout the region were common. Although estimates vary, NOAA has estimated that 4.9 million barrels of oil were released (Lubchenco et al. 2010). A total of 720 sea turtles have been verified in the spill zone of which 172 were verified as having been exposed to oil (NOAA 2010c). However, specific causes of injury or death have not yet been established for many individuals as investigations into the role of oil in these animal s’ health status continue.

Perhaps the most famous shipwreck of all time o ccurred in the Gulf of Alaska when, in 1989, the
Exxon Valdez released at least 11 million gallons of Alaskan crude oil into one of the largest and most productive estuaries in North America. Th e spill was the worst in US history until the Deepwater Horizon event in 2010. The Alaska Departme nt of Environmental Conservation estimated that 149 km of shoreline was heavily oi led and 459 km were at least lightly oiled. Oil spills, both small and large, occur widely along US shores at refining and transfer facilities and

extraction sites.290 Sea turtles may also be harassed by the high leve l of helicopter activity over Gulf of Mexico

waters. It is estimated that between r oughly 900,000 and 1.5 million helicopter take-offs and landings are undertaken in associa tion with oil and gas activities in the Gulf of Mexico annually (NRC 1990a, USN 2008). This likely includes numer ous overflights of sea turtles, an activity

which has been observed to startle and at le ast temporarily displace sea turtles (USN 2009).

Pollution Chemical pollution of the freshwater, estuarine, and marine environment is a pervasive problem throughout the US, although the si gnificance of specific pollutant s varies between regions or

watersheds. The Gulf of Mexico is a sink for massive levels of pollution from a variety of marine and terrestrial sources, which ultimately can interfere with ecosystem health and particularly that of sea turtles (see Status of listed resources section). Sources include the petrochemical industry in and along the Gulf of Mexico, wastewater treatment plants, septic systems, industrial facilities, agriculture, animal feeding operations, and improper refuse disposal. The Mississippi River drains 80% of United States cropland (i ncluding the fertilizers, pesticides, herbicides, and other contaminants that are applied to it) and discharges into the Gulf

of Mexico near the actio n area (MMS 1998). Agricultural discha rges, as well as discharges from large urban centers (ex.: Houston and New Orleans) contribute contaminants as well as coliform

bacteria to Gulf of Me xico habitats (Garbarino et al. 1995). These contaminants can be carried long distances from terrestrial or nearshore source s and ultimately accumulate in offshore pelagic environments (USCOP 2004). The ultimate im pacts of this pollution are poorly understood.

Significant attention has been paid to nutrient enrichment of Gulf of Mexico waters, which leads to algal blooms (including harmful algal blooms), oxygen depletion, loss of seagrass and coral reef habitat, and the formation of a hypoxi c “dead zone” (USCOP 2004). This hypoxic event occurs annually from as early as February to as late as October, spanning roughly 12,700 km 2
(although in 2005 the “dead zone” grew to a record size of 22,000 km 2 ) from the Mississippi River Delta to Galveston, Texas (MMS 1998, Rabalais et al. 2002, LUMCON 2005, USGS 2010). Although sea turtles do not ex tract oxygen from sea water, numerous staple prey items of sea turtles, such as fish, shrimp, and crabs, do and are killed by the hypoxic conditions (Craig et
al. 2001). More generally, the “dea d zone” decreases biodiversity, alters marine food webs, and destroys habitat (Craig et al. 2001, Rabalais et al. 2002). High nitrogen lo ads entering the Gulf of Mexico from the Mississippi River is the likely culprit; nitrogen concentrations entering the

Gulf of Mexico have increased three fold over the past 60 years (Rabalais et al. 2002). Through these indirect effects, sea turtles are unable to utilize this region during this time for foraging and can only utilize it to a limited extent when the “dead zone” does not occur while the underlying

food web recovers.

Due to rapid human population growth and urban development in the Puget Sound region (Grant and Ross. 2002), significant pollutant loads have led to generally poor water quality in the region, including Puget Sound itself (Long et al. 2001, Grant and Ross. 2002, NMFS 2008e). Some of these pollutants have apparently bioaccumulated in southern resident killer whales, with persistent organic pollutants now being found in high levels in killer whales along British Columbia and Washington State (Koski 2007b) and considered to pose the greatest pollution

threat to southern reside nt killer whales (Ross et al. 2000c, CBD 2001a, Krahn et al. 2002b,291 Koski 2007b, NMFS 2008e). These pollutants have the potential to caus e immune, endocrine, and reproductive system effects in killer whales; concern has been expres sed by researchers that these pollutants may be a factor in the decline of southern resi dent killer whales (Calambokidis

et al. 1984, Hayteas and Duffield. 2000, Ross et al. 2000b, Ross et al. 2000c, Baird 2001b,

Ylitalo et al. 2001, de Wit 2002, Krahn et al. 2002b, Darnerud 2003, Hall et al. 2003, Krahn et
al. 2004b, Krahn et al. 2004c, Waring et al. 2004a, Krahn et al. 2007, NMFS 2008e). Exposure is believed to be prim arily through diet (Hickie et al. 2001); salmon preyed upon by southern resident killer whales may also be at risk. In addition, several hundred new chemicals enter the global marine environment annually; many of these chemicals have unknown effects to any lifeform (Grant and Ross. 2002, NMFS 2008e). Encouragingly, many persistent organic pollutants have been banned in the US and, over the past few decades, regulatory actions, Superfund clean-up, improved waste handling, and ongoing cleanup efforts have led to improvements in regional water quality (NMFS 2008e ). This has led to decreasing levels of many organochlorine residues in the envi ronment (Gray and Tuominen 2001, Mearns 2001, Grant and Ross. 2002, EVS Environmental Consultants 2003), although it may take up to 60 years for some chemicals to fall below levels known to cause health effects in marine mammals (Hickie et al. 2001).

Another significant form of pollution is from the petrochemical industry utilizing the Puget Sound region for transport and refining. Puget Sound is one of the leading petroleum refining

centers in the US with about 333 million barrels of crude oil and refined petroleum products transported through it annually (Puget Sound Action Team 2005). Inbound oil tankers carry crude oil to five major refineries in the sound, while outbound tankers m ove refined oil products to destinations along th e US west coast (Neel et al. 1997). In 2005, a total of 716 oil tankers passed through Washington’s waters bound for ports in Puget Sound, Canada, and along the Columbia River (WSDE 2006). In general, th e Strait of Juan de Fuca and areas near Washington’s major refineries (located in Anacortes, Ferndale, Blaine, and Tacoma) are

considered the locations most at risk of major spills in the action area (Neel et al. 1997, NMFS 2008e). Since the 1960s, there have been at least nine major oil sp ills of at least 2,222 barrels; the largest was ove r 50,000 barrels (Neel et al. 1997, Puget Sound Water Quality Action Team 2002).

In addition to the substantial volume of shipping traffic and the prevalence of petroleum refining centers in the action area, of equal or perhaps greater concer n are the chronic small-scale discharges of oil, which together can be substa ntial (Clark 1997). Thes e discharges originate from numerous sources, including the dumping of tank washings and ballast water by tankers, the release of bilge and fuel oil from general shipping, and th e disposal of municipal and industrial waters. Chronic oil pollution kills large numbers of seabirds (e.g., (Wiese and Robertson 2004)); however, its impact on southern resident killer whal es and other marine mammals is poorly documented (NMFS 2008e). In addition, the long-term effects of repeated ingestion of sub-lethal quanti ties of petroleum hydrocarbons on marine mammals are not well understood. As a result, the magnitude of the ri sks posed by oil discharg es in the proposed action area is difficult to preci sely quantify or estimate.

Commercial Fisheries Commercial fisheries are a major f actor in the decline of several li sted species in this Opinion.292 Directed harvesting of whales as been the greate st single reason several wh ale species are listed as threatened or endangered. However, harvests of listed whales have been largely eliminated and the species dealt with in this Opinion sp ecies appear to be ge nerally recovering. The significance of whale reductions compared to historical leve ls is provided under Status of Listed

Resources . Salmonids have also experienced signif icant fishing pressure, which is discussed below. Sea turtle species also continue to su ffer from commercial harv esting in their nesting

areas. While there is an international moratori um on harvesting whales, Japan harvests up to 101 sei whales annually (IWC 2008).

Although protected under the ESA, li sted salmonids are killed in commercial harvests along the

US west coast. As salmonid DPSs/ESUs ar e readily differentiated genetically, but not morphologically, the extent of impact that ha rvesting has is unknown, although documentation of species-level mortality is. Chinook and Coho salmon have historically been the primary salmonids caught along the Oregonian coast, although in 2002, Chinook salmon comprised only 2.4% (2,247 metric tons) of th e all landings and Coho salmon was only 0.5% (526 metric tons) due to crashes in both species (State 2004). In Washington, salmon accounted for ~15% of commercial fishery landings in 2007. This in cluded 1,377 metric tons of Chinook, 5,352 metric tons of chum, 1,139 metric tons of Coho, 36 metric tons of sock eye, and 147 metric tons of steelhead (WDFW 2007). Sockeye and steelhead are not subject to directed fisheries in the area (PFMC 2005). Along both Oregon and Washingt on, harvests generally occur over the

continental shelf within 65 km of shore and generally operate from March-November (NMFSNWR 2003). Across the entire US west coast, chum and Chinook salmon dominate commercial harvesting, with Coho composing a smaller component (Figure 24).

Figure 24. Annual landings of Chinook, Coho, pink, chum, and sockeye salmon for California, Oregon, and Washington State between 1995 and 2004 (from USN (2006)).

Apart from direct and bycatch mo rtality, there are additional consequences of fishing. Trawls can disturb the ocean floor, causing physical dama ge to exposed structures, increase turbidity, alter sediments, and remove prey of listed species. Fishing gear can also be lost, but subsequently entangle or trap fish and marine mammals years later. Recreational fishing can also reduce populations by focusing on certain species or locations.293 The Gulf of Alaska supports one of the most productive ecosystems in the US and is host to a very large fishing industry. The Alaskan groundfish fishery has received considerable attention for its exploitation of some fish species that are al so prey to Steller sea li ons. Significant concern exists that the level of fishing results in signi ficant prey reductions in the region, limiting prey availability for Steller sea lions. In 2005, 2006, and 2007, 38,000-48,000 metric tons of pollack, 20,000-28,000 metric tons of Pacific cod, 25,000 me tric tons of arrowtooth flounder, 8,000- 8,500 metric tons of rockfish, a nd ~5,500 metric tons of skates were allowed to be harvested annually (NMFS 2006a).

The waters surrounding Washington St ate host extensive commercial and recreational fishing; in 2007, 180 million pounds of fish and invertebrates were harvested from state waters; of this, 15% was salmon (Commander U.S. Pacific Fleet 2010). Although closely re gulated, there is the potential for overfishing, especia lly considering the selective natu re of southern resident killer whale diets (Fraser River populations of Chinook salmon). Lost gill nets, purse seines, and longlines may foul and disrupt bo ttom habitats as well as create the potential for “ghost fishing,” where marine animals are continually trapped and die in derelict gear . Recent reports of entanglement are unknown, but Sheffer and Slipp ( 1948a) documented several deaths of animals caught in gillnets between 1929 and 1943. Typically, killer whales are able to avoid nets by swimming around or underneath them (Jacobsen 1986, Matkin 1994). Recreational fishing also has the potential to affect fish habitats because of the large number of participants and the intense, concentrated use of specific habitats. Historically, killer whales have commonly been subject to shooting (some likely fatal) by fisherman due to perceived competition for target fish resources (Scheffer and Slipp. 1948a, Pike and Macaskie. 1969, Haley 1970, Olesiuk et al. 1990d, Baird 2001b). This practice has largely abated in the past fe w decades and is unlikely to occur today (Young et al. 1993, Carretta et al. 2001).

Three of the biggest threats to sea turtles result from harvest for commercial and subsistence use. These include egg harvest, the harvest of female s on nesting beaches, and directed hunting of sea turtles in foraging areas. These factors have le d to the precipitous declines in worldwide sea turtle populations. In the Atlantic , green sea turtles are captured and killed in turtle fisheries in Colombia, Grenada, the Lesser Antilles, Nicaragua , St. Vincent and the Grenadines; the turtle fishery along the Caribbean coast of Nicaragua, by itself, has captured more than 11,000 green sea turtles annually over the past decade (La gueux 1998, Bräutigam and Eckert 2006a). While these threats have been largely eliminated in Florida due to successful conservation measures, the hunting of juvenile and adu lt turtles continues bot h legally and illegally in many foraging areas where green sea turtles originating from Florida ar e known to occur (Fleming 2001, Chacon 2002). The killing of nesting hawksbill fema les continues to threat en the stability of hawksbill subpopulations in many areas. The centuries- old historic trade in tortoise shell greatly impacted hawksbill populations in the Insular Caribbean. Increases in nesting hawksbills in the region coincide with the decline of international trade in hawk sbill shell (Milliken and Tokunaga 1987), and in particular with th e 90% reduction in the annual ta ke of large hawksbills from Cuban waters (Carrillo et al. 1999).

Entrapment and Entanglement in Fishing Gear

Fisheries interactions are a significant problem for several marine mammal species and particularly so for humpback whales. Aside from the potential of entrapment and entanglement,294 there is also concern that many marine mammals that die from entanglement in commercial fishing gear tend to sink rather than strand ashore, thus ma king it difficult to accurately determine the frequency of morta lities. Entanglement may also make whales more vulnerable to additional dangers, such as predation and ship strikes, by restricting agil ity and swimming speed. Between 1998 and 2005, observers identified 12 humpb ack whales injured or killed by fisheries off the US west coast (NMFS, unpublished data). An estimated 78 rorquals were killed annually in the offshore southern California drift gilln et fishery during the 1980s (Heyning and Lewis. 1990). From 1996-2000, 22 humpback whales of the Central No rth Pacific population were found entangled in fishing gear (Angliss and Lodge. 2004). In 1996, a vessel from the Pacific Missile Range Facility in Hawaii rescued an entangled humpback, removing two crabpot floats from the whale. A photography study of humpback whales in southeastern Alaska in 2003 and

2004 found at least 53% of indivi duals showed some kind of scarring from fishing gear entanglement (Neilson et al. 2005).

Fisheries interactions are a part icularly problem for North Atlan tic right whales (Figure 25). Aside from the potential of entrapment and entang lement, there is also concern that many marine mammals that die from entanglement in commercial fishing gear tend to sink rather than strand ashore, thus making it difficult to accurately determine the frequency of such mortalities. Entanglement may also make whales more vulnera ble to additional dangers, such as predation and ship strikes, by restricting agility and swim ming speed. Along the Atla ntic coast of the US and the Maritime Provinces of Canada, there were 46 confirmed reports of North Atlantic right whales entangled in fishing gear between 1990 and 2007 (Cole et al. 2005a, Nelson et al. 2007a, Waring et al. 2009). Of the 39 reports that the NMFS coul d confirm, North Atla ntic right whales were injured in five of the entanglements and killed in four entanglem ents. Three of the 24 entangled whales between 2004 and 2008 died and one other resulted in serious injury (Glass et
al. 2009). Recent efforts to disentangle right whales have met with success (Anonmyous. 2009b). However, over 60% of the North Atlantic right whale population show some evidence of entanglement (Hamilton et al. 1998).295 Figure 25. A North Atlantic right whale entangled in fisheries gear off Florida, with Georgia Department of Natural Resources and Coastwise Consulting staff attempting to cut rope off (Credit: EcoHealth Alliance and Georgia Department of Natural Resources, ESA permit number 932-1905).

Recent reports of entanglement are unknown, but Sheffer and Slipp (1948b) documented several deaths of killer whales caught in gillnets between 1929 and 1943 in Washington State waters. Typically, killer whales are able to avoi d nets by swimming around or underneath them (Jacobsen 1986, Matkin 1994). Recreational fishing al so has the potential to affect fish habitats because of the large number of participants and th e intense, concentrated us e of specific habitats. Historically, killer whales have commonly b een subject to shooting (some likely fatal) by fisherman due to perceived competition for target fish resources (Scheffer and Slipp. 1948b, Pike and Macaskie. 1969, Haley 1970, Olesiuk et al. 1990c, Baird 2001c). This practice has largely abated in the past few decades a nd unlikely to continue today (Young et al. 1993, Carretta et al. 2001).

Steller sea lions routinely become entangled in fishing gear, but are most prone to entanglement in packing and rubber bands around their necks (Raum-Suryan et al. 2009b)(Figure 26). As an individual grows, these items ev entually cut into tissue, lead ing to gruesome injuries and, frequently, death. Several dozen individuals may become entangled and drown in commercial fishing gear annually (Atkinson et al. 2008, NMFS 2008g). Marine debris is also concerning for the health of Steller sea lion populations. It is estimated that 0.26% of Steller sea lions have marine debris around their necks or are hooked by fishing gear (0.07%); this equates to 100-200 individuals annually (FOC 2008, Raum-Suryan et al. 2009a). Individuals also commonly ingest296 flashers and hooks used in co mmercial fisheries (Raum-Suryan et al. 2009b). These can cause several internal injuries potentially leading to death.

Figure 26. Steller sea lions having ingested a commercial fish ing hook (A) and entangled by a rubber packing band. Photos taken and provided by Alaska Department of Fish and Game during research conducted under NOAA MMPA permits 14325 (A) and 358-1888 (B).

Fishery interaction remains a major factor in sea turtle recovery a nd, frequently, the lack there of. NMFS (2002) estimated that 62,000 loggerhead sea turtles have been ki lled as a result of incidental capture and drowning in shrimp trawl gear. Although turtle excluder devices and other bycatch reduction devices have significantly reduced the level of bycatch to sea turtles and other marine species in US wa ters, mortality still occurs in Gulf of Mexico waters.

In addition to commercial bycatch, recreationa l hook-and-line interactio n also occurs. Cannon and Flanagan (1996) reported th at from 1993 to 1995, at least 170 Kemp ’s ridley sea turtles were hooked or tangled by recreational hook- and-line gear in the northern Gulf of Mexico. Of these, 18 were dead stranded turtles, 51 were rehabilitat ed turtles, five died during rehabilitation, and 96 were reported as released by fishermen.

Naval Activities Naval activity, notable sonar use during traini ng exercises, has gained notoriety for its coincidence with marine mammal strandings. Ho wever, other activitie s (also during training exercises in designated naval ope rating areas and training ranges ) also have the potential to adversely impact marine mammals and sea tur tles. The action area overlaps several naval training ranges or facilities listed below. Listed whales travel widely in the North Atlantic and could be exposed to naval act ivities in several ranges.

 Marianas Island Range Complex, where humpback whales breed and po ssibly give birth,

 The Southern Califor nia Range Complex,

 The Northwest Training Range Complex, where humpback whales and Steller sea lions forage,

 The Gulf of Alaska Operating Area, where several listed whale species are known to forage and Steller sea lions reside, and

 The Hawaiian Islands Operating Area, where humpback whales and Hawaiian monk seals regularly breed and give birth.297  Northeast Training Range Complex, where No rth Atlantic right and humpback whales forage,

 The Virginia Capes, Cherry Point, and J acksonville-Charleston Operating Areas, which are situated consecutively along the migrator y corridor for North Atlantic right whales (the later operating area overlap s with North Atlantic right whale critical habitat, where calving and breeding take place), and

 The Key West, Gulf of Mexico, Bermuda, and Puerto Rican Complexes have the potential to overlap the dist ribution of North Atlantic ri ght and humpback whales, but presence within these areas is rare or undocumented.

Naval activities to which individu als could be exposed include, am ong others, vessel and aircraft transects, munition detonations, and sonar use. Responses by marine mammals could include no response, short-term and long-term behavioral responses and changes (altered vocal activity, changes in swimming speed and direction, respiration rates, dive times, and social interactions), temporary or permanent hearing loss, debris ingestion, ship-strike injury, and death.

Although naval

' vessels represent a small fraction of th e total sound level and are designed to operate quietly, these ships are large and equi pped with high-output sonar equipment such as ANISQS-53C tactical sonar, which produces signals at sour ce levels of 235 dB re 1 μPa

rms at 1 m. The signals emitted from these devices have the potential to affect marine mammals in the action area; however, empirical data are limited. No stranding or mortality events have been documented in or around other operating areas or training ranges within the action area that appear linked to naval sonar. Ho wever, an event that occurred in the Strait of Juan de Fuca and Haro Strait on May 5, 2003 demonstrates the potent ial for naval activities to impact southern

resident killer whales. The US Na vy guided missile destroyer U.S.S. 

### Shoup passed through the strait operating its mid-frequency sonar dur ing a training exercise. Members of 

### J pod (a family group of southern resident killer whales) were in the strait at the same time and exhibited unusual behaviors coincident with exposure to the sona r, as reported by local researchers (Commander U.S. Pacific Fleet 2003, NMFS 2005a, 2006b). Based on the duration of exposure, the received levels experienced by the whales, and informa tion on sound levels known to cause behavioral reactions in other cetaceans, NMFS concluded J pod was exposed to levels likely to cause behavioral disturbance, but not temporary or permanent hearing loss (NMFS 2005a, 2006b). Underwater detonations are sometimes performed at this site and there was an occasion when J pod was less than 1.5 km away when a blast occurred, which caused the whales to suddenly change their direction of travel (NMFS 2006b). In another incident, five beaked whales were discovered stranded or floating de ad coincident in time with the Alaska Shield/Northern Edge 2004 exercise between June 17-19, 2004 in the Gulf of Alaska Operating Area. However, no mid-frequency sonar or explosiv es were used during this exer cise and evidence linking the

exercise to mortalities is circumstantial at best.

Acoustic Harassment Devices Acoustic harassment devices (AHDs) are another source of underwater sound that may occur in the action area and may be disrup tive to southern resident kille r whales. AHDs used at salmon aquaculture farms emit "loud" signal s intended to displace harbor seals and sea lions and thereby reduce depredation (Petras 2003, NMFS 2008e). Howe ver, these signals can also cause strong avoidance responses in cetaceans (Olesiuk et al. 2002). Morton and Symonds (2002) describe298 one AHD model that broadcasts a 10 kHz signal at 194 dB re 1 μ Pa at 1 m. A large majority of these occur in Arctic waters during explorati on for petroleum products, although other detectable above ambient levels in open water for up to 50 km. Activation of AHDs at an aquaculture farm

near northeastern Vancouver Island corresponded with drastic declines in the presence and use of nearby passages and inlets by both resident and transient killer whales (Morton and Symonds 2002). The only AHD still in use in Washington State operates at th e Ballard locks in Seattle,

where NMFS uses it to dete r sea lions (NMFS 2008e).

Vessel Approaches – Commercial and Private Marine Mammal Watching Although considered by many to be a non-consump tive use of marine mammals with economic, recreational, educational and sc ientific benefits, marine ma mmal watching is not without potential negative impacts. Whale watching ha s the potential to harass whales by altering feeding, breeding, and social behavior or even injure them if the vessel gets too close or strikes

the whale. Another concern is th at preferred habitats may be ab andoned if disturbance levels are too high. In the Notice of Availability of Revised Whale Watch Gu idelines for Vessel Operations in the Northeastern United Stat es (64 FR 29270; June 1, 1999), NMFS noted that whale watch vessel operators seek out areas where whales concentrate, which has led to numbers of vessels congregating around groups of whales, in creasing the potential for harassment, injury,

or even the death of these animals. In additi on to whale watching vessels , large cruise vessels also operate in waters off the coast of Alas ka, and may pose a threat to humpback whales. Whale watching, particularly of humpback whales, is extensive in Hawaiian waters during

winter. The interactions that individuals experience in these wa ters likely influence how they react to approaches by vessels in the future (Herman 1979).

Several studies have specifically examined th e effects of whale watching on marine mammals, and investigators have observed a variety of s hort-term responses from animals, ranging from no apparent response to changes in vocalizations, duration of time spent at the surface, swimming speed, swimming angle or directi on, respiration rate, dive time, feeding behavior, and social behavior (NMFS 2006b). Responses appear to be dependent on fact ors such as vessel proximity, speed, and direction, as well as the number of vessels in the vicinity (Watkins 1986, Corkeron 1995, Au and Green. 2000, Erbe 2002b, Magalhaes et al. 2002, Williams et al. 2002b, Williams

et al. 2002d, Richter et al. 2003, Scheidat et al. 2004). Foote et al. (2004b) reported that

southern resident killer whale call duration in the presence of whale watching boats increased by 10-15% between 1989-1992 and 2001-2003 and suggested this indicated compensation for a noisier environment. Disturbance by whale watc h vessels has also been noted to cause newborn calves to separate briefly from their mothers' sides, which leads to gr eater energy expenditures by the calves (NMFS 2006b). Although numerous short-term behavioral responses to whale watching vessels are docum ented, little information is available on whether long-term negative

effects result from whal e watching (NMFS 2006b).

It is difficult to precisely quantify or estimat e the magnitude of the risks posed to marine mammals in general and southern resident kill er whales specifically (who possibly have the greatest exposure to whale watching activities of any listed marine mammal) by whale watching and recreational vessels (NMFS 2008e). Commer cial whale watching in Washington State has increased dramatically from small scale operati ons during the late 1970s to early 1980s, to 13 vessels by 1988, and a total of 76 vessels (and over 500,000 people) in 2006 (Osborne 1991,299 Koski 2006a, Koski 2007b, NMFS 2008e). Most companies belong to the Whale Watch Operators Association Northwest, which has established whale viewing guidelines for commercial operators (WWOANW 2007). Currently, over 50% of vessels involved with whale watching are commercially owned, with the San Ju an Islands and adjacent area also attracting large numbers of private boaters for recreational activities such as opportunistic viewing of killer whales (Koski 2007b, NMFS 2008e). In addition, private floatplanes, helicopters, and small aircraft regularly take adva ntage of whale watching opportu nities (MMMP 2002b). Weather

conditions in the Pacific Ocean in winter limit wh ale watching during winter months and activity is greatest during summer (NMFS 2008e). From May to September 2005, an average of over 19 boats (up to 94) surrounded southe rn resident killer whales on a daily basis (Koski 2006a). In Washington State, southern resident killer whales are the primary target species, pa rticularly in Haro Strait (Hoyt 2001, 2002, NMFS 2008e).

The increase in whale watching traffic over the past two decades has resulted in increased exposure of southern resident killer whales to vessel traffi c and sound emitted by it. Whale watching activities have the potenti al to affect southern resident killer whales in the action area, resulting in possible disturban ce or displacement. Increasing anthropogenic sound levels in the Puget Sound region have been associ ated with increased call duration by southern resident killer whales when vessels are present (Erbe 2002b, Foote et al. 2004b). Vessels also appear to cause whales to alter their direction of travel (Williams et al. 2002b, Williams et al. 2002d). Furthermore, vessel presence has been linked to reduced foraging su ccess and/or inhibiting foraging all together (Bain et al. 2006a, Williams et al. 2006). Based on a study in Johnstone Strait, British Columbia, north ern resident killer whales decreased feeding behaviors significantly and increased time e ngaging in behaviors which require d less energy such as resting and socializing (Williams et al. 2006).

Live-Captures for Aquaria Killer whales have been displayed in aquaria worldwide since the early 1960s. For 15 years,

killer whales were collected from the wild to populate display facilities; all but one individual came from Washington State or British Columbia until 1976, when local laws banned captures (Hoyt 1990, NMFS 2006b). During this time, from 275-307 killer whales were captured, of which 55 were sent to aquaria, 12-13 died, and 208-240 were re leased or escaped. Of the individuals captured and displayed or killed, 70% (47 or 48 individu als) were southern resident killer whales, including 17 immature males, 10 immature females, nine mature females, and seven or eight mature males; 15 individuals we re from K pod, five from L, and one from J (Olesiuk et al. 1990c, Baird 2001c, NMFS 2006b). The se lective removal of younger animals and males produced a skewed age and sex composition in the southern reside nt killer whale DPS, which probably affected its ability to recover (Olesiuk et al. 1990c).

Ship-Strikes Ship-strike is a significant concern for the recove ry of listed baleen whal es and sea turtles to a lesser degree. We believe the vast majority of ship-strike mortalit ies go unnoticed, and that

actua l mortality is higher than currently documented. More humpback whales are killed in collisions with ships than any other whale spec ies except fin whales (Jensen and Silber 2003). Along the Pacific US coast, a humpback whale is known to be killed about every other year by ship-strikes (Barlow et al. 1997). Two whales have been struck offshore of Japan (Jensen and300 Silber 2003).

Despite these reports, the magnitude of the risks commercial ship traffic poses to large whales in the proposed action areas has been difficult to quan tify or estimate. We struggle to estimate the

number of whales that are killed or seriously injured in ship strikes within the US Exclusive Economic Zone and have virtually no informati on on interactions between ships and commercial

vessels outside of US waters. With the informa tion available, we know those interactions occur but we cannot estimate their significance to whale species. Several humpback whales are also known to have become entangled in the North Pacific (Hill et al. 1997, Angliss and Outlaw 2007a). Along the Pacific coast, a humpback whale is known to be killed about every other year by ship strikes (Barlow et al. 1997).

Ship-strike is a significant concern for the recovery of North Atlantic righ t whales in the region. We believe the vast majority of ship-strike mort alities go unnoticed, and that actual mortality is higher than currently documented; Kraus et al. (2005) estimated that 17 % of ship strikes are actually detected. The magnitude of the risks commercial ship tra ffic pose to large whales in the proposed action areas has been difficult to quantify or estimate. We struggle to estimate the number of whales that are killed or seriously in jured in ship strikes within the US EEZ and have virtually no information on inter actions between ships and comme rcial vessels outside of US waters. With the information available, we know those interactions occur but we cannot estimate their significance to whale species.

Ship strikes are the largest single contributor to North Atlantic right wh ale deaths, accounting for approximately 35% of all known mort alities, even though right whales should be able to hear the sound produced by vessels (Richardson et al. 1995a, Ketten 1998, Knowlton and Kraus 2001,

Laist et al. 2001). Some information suggests righ t whales respond only within very close proximity to ships (Nowacek et al. 2004). Injury is generally caused by the rotating propeller blades, but blunt injury from direct impact with the hull also occu rs. There have been 18 reports of North Atlantic right whales being struck by vessels between 1999 and 2005 (Cole et al. 2005b, Nelson et al. 2007b). Of the 17 reports that NMFS could confirm, right whales were injured in two of the ship strikes and killed in nine. Recent records show that from 2004-2008, there were 17 confirmed reports of North Atlantic right whales being struck with eight whales dying of their wounds and two additional right whales sustaining serious injuries (Glass et al. 2009). Deaths of females are esp ecially deleterious to the ability of the North Atlantic right whale population to recover. For instance, in 200 5, mortalities included si x adult females, three of which were carrying near-term fetuses and four of which were just starting to bear calves, thereby representing a lost reproductive potential of as many as 21 indi viduals over the short term (Kraus et al. 2005). Voluntary recommendations fo r slower vessel speeds in the Bay of Fundy appear to be largely ignored (Vanderlaan et al. 2008a). Rules for seasonal (varies by location, but January through July) slowing of vessel traffic to 10 knots or changing shipping lanes by less than one nautical mile to avoid the greatest con centrations of right whales are predicted to be capable of reducing ship strike mortality by 62% in the Bay of Fundy region; the same rule applies from November through Apr il from Brunswick, Georgia to Jacksonville, Florida, where North Atlantic right whales go for calving and breeding. Speed rules also apply to medium and large ports along the eastern seaboard during this time frame when right whales migrate to and from northern feeding and sout hern breeding areas. Nearly a dozen shipping301 lanes transect through coastal wate rs of the southeastern US from the North-South Carolina to Cape Canaveral, Florida. The vast majority of ship strike mortalities are never identified, and that actual mortality is highe r than currently documented.

A total of six instances have been documented of northern and southern resident killer whales being struck by vessels since th e 1990s, including le thal interactions (Visser 1999, Visser and Fertl. 2000, Baird 2001c, Carretta et al. 2001, Carretta et al. 2004).

Sea turtle ship strikes are a poorly -studied threat to sea turtles, but has the potenti al to be highlysignificant (Work et al. 2010). All sea turtles must surface to breath an d several species are known to bask at the surface for long periods, in cluding loggerhead sea turtles. Although sea turtles can move rapidly, sea turtles apparently are not well able to m ove out of the way of vessels moving at more than 4 km/hr; most vessels move far faster than this in open water (Hazel and Gyuris 2006, Hazel et al. 2007, Work et al. 2010). This, combined with the massive level of

vessel traffic in the Gulf of Mexic o, has the potential to result in frequent injury and mortality to sea turtles in the region (MMS 2007b). Hazel et al. (2007) suggested that green sea turtles may use auditory cues to react to approaching vesse ls rather than visual cues, making them more susceptible to strike as vessel speed increases. Each state along the Gulf of Mexico has several hundred thousand recreational vessels registered, including Florida with nearly one million-the highest number of registered boats in the United States-and Texas with over 600,000- ranked sixth nationally (USCG 2003, 2005, NMMA 2007). Co mmercial vessel operations are also

extensive. Vessels servici ng the offshore oil and gas indus try are estimated to make 115,675- 147,175 trips annually, apart from commercial vessels travelling to and from some of the largest ports in the United States (such as New Or leans and Houston)(MMS 2007a, USN 2008). Extensive shrimping and other fisher y effort is also expended in the area. Overall, ship strike is likely highly underestimated as a source of injury or mortality to sea turt les in the action area.

Scientific Research and Permits Scientific research permits issued by the NMFS currently authorize studies of listed species in

the Pacific and Atlantic Oceans which occur primarily in the ac tion area. Tables 38 identifies the

cumulative number of takes for each listed mari ne mammal and sea turtle species, as well as smalltooth sawfish, rockfishes, sturgeon, and eulac hon in the action area aut horized in scientific research permits. Due to the large numbers of research permits authorizing salmonid takes, we did not quantitatively assess the numbers of act ions on listed salmon a nd steelhead. We did, however, address these in a qualitative manner. Cetacean takes include approach, biopsy, suction cup and implantable ta gging, breath sampling, acoustic pl aybacks, and/or ultrasound. Pinniped takes consist of approach, capture, ha ndling, restraint, measurements, anaesthesia or drugging, temporary captivity, exte rnal tagging, translocation, mortality, swab sampling, blood and tissue sampling, lavage, hot branding, tooth extraction, and/or ultrasound. Sea turtle researcher involves approach, capture , handling, restraint, PIT, flipper, satellite, or sonic tagging, lavage, mortality, ultrasound, blood or tissue sam pling, captive experiments, laproscopy, imaging, and/or antibiotic injections. Smallt ooth sawfish may be captured via a variety of means, measured, tagged, tissue sampled, and/ or ultrasounded. Research actions on sturgeon species include capture, handli ng, restraint, anaest hesia, laproscopy, lavage, boroscopy, fin, operculum, or barbel clipping, PIT, floy, sonic, or satellite tagging, gonad sampling, prophylactic, and/or mortality. Rockfishes may u ndergo capture, handling, re straint, anaestheia,302 floy or sonic taggi ng, and mortality.

Behavioral disruption of Steller sea lions occurs as a resu lt of human disturbance (FOC 2008). Research efforts to collect scats, count and we igh pups, and other human activities on or near rookeries can lead to stampedes into the wate r. Mortality can occu r directly due to pup trampling, separating from mothers, or drowning. If disturbance is too frequent, haulouts may be completely abandoned. Although habituation to so me activities, such as boating, can occur, unusual activities and sounds, such as blasting or demolition, can remotely trigger stampedes.

Table 38. Authorized takes of non-salmonid listed species in the action area. Species 2009-2013 lethal take 2009-2013 sub lethal take Humpback whale-North Atlantic 0 42,196 Humpback whale-North Pacific 0 340,137 North Atlantic right whale 0 31,805 Cook Inlet beluga whale 0 41,769

Southern resident killer whale 0 23,458

Hawaiian monk seal 59 41,590

Steller sea lion-eastern DPS 908 1,363,064

Green sea turtles-North Atlantic 71 69,045

Green sea turtles-North Pacific 0 12,490

Hawksbill sea turtle-North Atlantic 18 25,334

Hawksbill sea turtle-North Pacific 0 1,500 Kemp’s ridley sea turtle 34 31,865

Leatherback sea turtle-North Atlantic 12 18,759

Leatherback sea turtle-N orth Pacific 0 2,957 Loggerhead sea turtle-North Atlantic 402 92,385 Loggerhead sea turtle-North Pacific 0 1,886 Olive ridley sea turtle- North Atlantic 5 3,535 Olive ridley sea turtle -North Pacific 0 7,686

Smalltooth sawfish 0 3,940 Shortnose sturgeon 34,794 101,280 Green sturgeon-southern DPS 927 15,755

Bocaccio 155 191 Canary rockfish 2,919 3,182 Yelloweye rockfish 397 476
Permit numbers: 1420, 1440, 1447, 1449, 1450, 1462, 1475, 1486, 1501, 1505, 1506, 1507, 1514, 1516, 1518, 1522, 1526, 1527, 1537, 1538, 1540, 1542, 1544, 1547, 1549, 1551, 1552, 1556, 1570, 1571, 1575, 1576, 1578, 1580, 1581, 1591, 1596, 1599, 10014, 10018, 10022, 10027, 10037, 10042, 10045, 10115, 10137, 13306, 13307, 13430, 13543, 13544, 13573, 13675, 13815, 13846, 13908, 13927, 13958, 14073, 14097, 14122, 14176, 14204, 14210, 14233, 14245, 14249, 14266, 14272, 14274, 14296, 14325, 14326, 14336, 14337, 14353, 14381, 14390, 14394, 14438, 14442, 14451, 14457, 14462, 14477, 14506, 14508, 14510, 14534, 14585, 14586, 14591, 14599, 14603, 14604, 14616, 14622, 14655, 14682, 14726, 14759, 14780, 14791, 14824, 14949, 14971, 15112, 15119, 15135, 15225, 15237, 15249, 15269, 15271, 15274, 15275, 15330, 15372, 15408, 15415, 15431, 15449, 15476, 15483, 15488, 15545, 15552, 15566, 15595, 15606, 15614, 15677, 15682, 15741, 15750, 15783, 15791, 15801, 15847, 15849, 15883, 15994, 16017, 16039, 16073, 16146, 16166, 16174, 16194, 16208, 16220, 16262, 16269, 16271, 16292, 16327, 16328, 16347, 16408, 16414, 16429, 16439, 16458, 0642-1536, 0662-1661, 1036-1744, 1049-1718, 1058-1733, 1071-1770, 1120-1898, 1121-1900, 1127-1921, 1128-1922, 1410-4M, 14457-2M, 14457-3A, 15119- 2A, 1586-3R, 473-1700, 532-1822, 540-1811, 545-1761, 587-1767, 594-1759, 605-1904, 633-1763, 633-1778, 715- 1885, 716-1705, 727-1915, 731-1774, 753-1599, 774-1714, 775-1875, 781-1824, 782-1719, 808-1753, 945-1776, 948-1692, 965-1821, and 981-1707.303 EFFECTS OF THE PROPOSED ACTION:

EFFECTS OF THE NUMERICAL STANDADRD

Exposure Analysis

In this section of the Opinion, we assess the probable direct and i ndirect effects to listed species

and critical habitat that are r easonably likely to occur as a resu lt of activities au thorized, funded, or carried out by the ballast water management program including the recently proposed

numerical standard. As discusse d in the approach to the assessment, in order to determine the probability and extent of exposure, we rely on the spatial and temporal occurrence of the species

identified in the Status of the Species section of this Opinion and the extent of the impacts anticipated in the Action Area section. The specific stressors a ssociated with ballast water intake and discharge are capture of liste d species during uptake of ballast water and the introduction of non-native species through ballast water discha rge at the proposed standard. NMFS does not expect any effects from three methods of ballast water maintenance identifi ed in the FR (33 CFR 151.1510 and 33 CFR 151.2025): (1) retaining ballast water on board, (2) di scharging ballast water to on-shore or vessel/barge based holding facilities, or (3) using public drinking water, meeting the requirements of the Safe Drinking Wa ter Act (42 U.S.C. §§ 3 00f-300j), as ballast

water.

Ballast Intake Global shipping moves 80% of the world’s commodities (Tamburri et al. 2002, Bax 2003). When ships arrive in waters of the US to unloa d cargo, the ships take on ba llast water to offset the cargo weight being removed. When these ships take on or disc harge ballast, wa ter enters or leaves the ship through two to si x sea chests (depending on the size of the vessel) along the sides or bottom of the ship. When ships take on ballast, they are docked at ports and stationary. In the past, during ballast water exchange, they woul d take on ballast water while underway at offshore sites, but under the USCG rule, ballast water exchange is no longer required. Ballast intake occurring at the same time and place as threatened or endangered species poses the potential for listed species to be harmed by impinge ment, entrainment, or capture in ballast tanks or pipes.

Impingement occurs when organisms that are suffi ciently large to avoid passing through screens are trapped against the screen by the force of flowing water (for example, see Hanson et al

. 1977). Entrainment occurs when organisms are draw n into an intake system with the water flow (for example, see Beck and Miller 1974, Hanson et al . 1977). Impingement typically involves

adult fish and invertebrates and often results in death, physical damage to scales and dermal tissues, or physiological stress responses with st ress pathologies (for exam ple, see Brett 1958). In contrast, entrainment typically involves small organisms (such as invasive species), eggs, and

fish larvae.

In most cases, large marine mammals such as whales will not be present in ports and will not be subjected to ballast water intake operations. Even in cases when smaller marine mammals such as steller sea lions may be present near ships du ring ballast intake, these animals are too large to304 get impinged or entrained on pipes or screens le ading into the sea chest and strong enough to swim out of the area. Therefore, marine mamma ls under NMFS jurisdiction will not be affected by ballast water intake operations.

Sea turtles, like marine mammals, attain large sizes; however they comparatively may be more vulnerable to ballast water intake pipes. Sea turt les, especially recently hatched turtles, could be sucked into sea chests during ballast water intake activities. Recently ha tched sea turtles leave their nesting beaches and swim to offshore rearing areas near de bris collecting at convergence zones. It is unlikely that a ha tchling sea turtle would be in a port because ports tend to be in harbors or upstream in estuaries away from beach environments. Adult sea turtles would also not be expected to enter ports, but if they did, they are powerful enough swimmers to avoid being impinged or entrained in sea chests or pi pes during ballast water in take. Therefore, sea turtles may be affected but are not likely to be adversely affected by ballast water intake.

In the event that juvenile Atlant ic or Pacific salmonids or shor tnose, Atlantic, Gulf, or green sturgeon are in close proximity to ships taking up ballast water, these fish are not likely to get impinged or taken up. Based on a report on ballast water exchange prepared by the American Bureau of Shipping (2010), a moderately-sized ba llast intake system woul d intake less than 250 cubic meters per hour (or about 0.069 m 3 /s). Flow rates through sea ch ests are calculated to be less than 0.6 meters per second (m/s) (NMFS 2004), which would be the equivalent of a pipe that

is 10 centimeters in diameter in the scenario presented above, but normal intake pipes are 1.1m 2

. Fish must be able to swim faster than 0.6 m/s or not occur in the same time or place as the ballast intake to avoid being impinged, entrained, or captured. Salmon can swim at speeds ranging from two to eight m/s (Taylor and McPhail 1985, Colavecchia et al. 1998) (Booth et al. 1997) and sturgeon have sustained swimming speeds above 1m/s and burst speeds over 4m/s (Adams et al. 1999, Allen et al. 2006, Webber et al. 2007). Therefore, salmon and sturgeon would be able to swim away from intake structures. It is possibl e that these species wouldn’t avoid the intake current, however, adults of both sp ecies would be too large to be affected. Juvenile sturgeon, while very small, move to the deepest parts of main channels where they feed and grow. Juvenile salmonids rear in headwater systems as small juveniles and move to larger estuaries for parts of their lives as large juve niles or smolts. In both cases, in teractions with vessels would be limited if they ever occurred and the most vulnerable sizes would not be exposed to ballast water intake systems.

Smalltooth sawfish occupy shallow nursery areas such as inshore bars, mangrove swamps, and sea grass beds as juveniles and then move offshore as adults. Ships would not be expected to uptake ballast water in these shallow areas, so in teraction with juvenile smalltooth sawfish would be rare. If ships attempted to intake ballast water in a location with juvenile smalltooth sawfish, their uptake would be unlikely due to their large size at birth (ver y small juveniles are considered to be less than one meter, Wiley and Simpfendor fer 2010). Adults can grow to six meters in length (Simpfendorfer et al. 2008) and would not be vulnerable to uptake in ballast water tanks.

Rockfish are unlikely to be affected by balla st water intake because their habitat does not overlap with shipping ports. Rockfish adu lts live at depths between 50 and 250 meters (Gunderson and Sample 1980). Larval and juvenile rockfish are found near the thermocline in open ocean waters (Lenarz and Echeverria 1991).305 Pacific eulachon, like rockfish, spen d most of their lives in deep waters off the Pacific coast. However, Pacific eulachon are anadromous and return to freshwater to spawn before dying (Emmett et al. 1991a, Musick et al. 2000b, WDFW and ODFW 2001). During this brief time

when adults enter freshwater during late winter and early spring and larvae are carried downstream to the ocean in spring and early summer, there is the potential for some interactions with vessels uptaking ballast water. Adults, because of their strong swimming ability, are expected to avoid ballast water intake pipes. However, larv ae moving downstream would be vulnerable to ballast water intake for approxima tely three weeks until they are large enough to move to deep water areas in th e river (Howell 2001). There are seven large ports in the lower Columbia River, where interactions between out -migrating larval Pacific eulachon and ballast intake pipes are possible. The currents in th e Columbia River are unlikely to carry larval eulachon into them to make them vulnerable to ballast uptake, but there is a negligible chance during approximately a five week time in spri ng/early summer that some eulachon larvae could be removed from the system by ballast uptake.

There are several invertebrate s under NMFS jurisdiction that , like Pacific eulachon, produce larval offspring which are passively carried by ocean currents. The black and white abalones live offshore in the Pacific Ocean and their offspr ing that will successfully establish in offshore rocky habitats would be unlikely to have any in teraction with vessels intaking ballast water. Elkhorn and staghorn coral typica lly live in shallow water reef systems ranging from 1 to 10 meters deep and 5 to 30 meters deep, respectively (Wapnick et al. 2004, Mayor et al. 2006) (Gladfelter et al. 1977). Both species produce planktonic la rvae. Neither would be expected in ports, which are generally dredged environments w ith sandy bottoms. However, these corals live near port environments and their offspring coul d be taken up during balla st water intake if currents have carried their offspr ing to a port location. Like with abalone species, any offspring that has drifted to a port environment would not successfully settle and establish a new coral body because that is unacceptable hab itat. It is possible that th e planktonic coral could have continued drifting to suitable habitat but the ch ances of a coral plankton being removed from a port environment that would have otherwise successfully settle d is discountable.

Johnson’s sea grass grows in bays, inlets, and lago ons. The grass typically occupies areas with turbid water and high tidal currents. In some cas es, it may grow near ports or areas where ships could take up ballast water; howev er, as a rooted plant, there is essentially no chance it would be uprooted and removed from the environment. Fu rthermore, Johnson’s sea grass is known to have female flowers, but no male flowers have been identified. Because sexual reproduction and potential offspring are unknown, balla st water intake would not be expected to remove any seeds

from the environment.

Ballast Water Discharge An often overlooked ballast water discharge issu e is the presence of chemical pollutants in

ballast water taken from locati ons which are relatively more contaminated than the water receiving the ballast discharge. The natural presence of chemical pollutants in ballast water discharge is not under the USCG’s legal jurisdiction. The potential for exposure to ballast water chemical pollutants and the direct and indirect ef fects of such exposures will be analyzed in EPA’s biological evaluation for the vessel discharge permit. However, the effects of any306 pollutant generated in association with ballas t water treatment will be analyzed in tiered consultations to this programmatic because the types of treatments and their affects are unknown

at this time.

In this Opinion, the only concern with ballast wate r discharge is the establ ishment of non-native species and how those non-native species may affect listed species and their critical habitat by altering the basic functional compon ents of ecosystems. Several au thors have argued that ballast water discharge in the global shipping industry represents a planetary scale homogenization of aquatic life due to the establishment of formerly isolated biota that are presently becoming mixed and effecting environmental change throughout marine and estuarine ecosystems (Lövei 1997,

Ruiz et al. 1997, Lodge et al. 1998, Rahel 2000, Leppaekoski and Olenin 2001) (Carlton 1995). The ability of invasive species to alter popul ation, community, and ecosystem structure and function is well documented (Elton 1958, Mooney and Drake 1986, Vitousek et al. 1987, Drake

et al. 1989) and their direct and indirect effects to listed species and their critical habitat will be the focus of the effects analysis. Non-native sp ecies most often have indirect impacts to listed species through habitat alteration, which can result in changes in prey availability, changes in accessible habitat or cover, changes in predation risk due to effects on water clarity, and changes in water quality. Non-native species can also affect listed species or their critical habitat directly through competition, predation, or disease.

There are five groups of organisms that cause co ncern in ballast water: (1) planktonic life in suspension in the ballast tanks a nd ballasted cargo holds, (2) fouling organisms that settle on the

lower walls of ballast tanks a nd ballasted cargo holds, (3) infa unal organisms living in benthic sediments of ballast tanks and ballasted cargo holds, (4) encysted stages and resting eggs of the above organismal groups in ballast tanks and ball asted cargo holds, and (5) fish and fish eggs in ballast tanks and ballasted cargo holds. The prop osed ballast water standards focus on reducing planktonic life but may not fully address fouling organisms, infaunal organisms, or fish (USCG 2011). Because the BWDS only addresses discharg e water, which would contain the suspended organisms, all potential introducti ons associated with ballast wa ter are not addressed; however, treatment technologies and methods used to reduce the concentrations of planktonic organisms in

ballast water should reduce members of other organismal groups.

In ballast water, microorganisms may constitute a numeric majority of the species found. (Galil and Hulsmann 1997, (Carlton 1999, McCart hy and Crowder 2000, Carlton 2001). Microorganisms, such as phytoplankton, particul arly diatoms and dinoflagellates, can be especially abundant and may cause the harmful algal bloom known as “red tides” Hallegraeff 1993, NRC 1996) (Carlton and Geller 1993) or pr oduce toxic compounds. While there was previously not a scientific, enforceable, or pract ical discharge standard for microorganisms, the

USCG is now able to test ballast water and dete rmine whether the ship being tested is meeting the proposed numerical standard.

Fish may not be as numerically abundant in ballast water as other phyla (Wonham et al. 2001), but once established in a new environment, fish can spread over broad geographic areas. A classic example is the Eurasian ruffe ( Gymnocephalus cernuus ), which was introduced via

ballast water to the St. Lawrence River in the mid-1980s. Since its intr oduction, it has spread307 rapidly through rivers, bays, and lakes in the Grea t Lakes area. Furthermore, the ecological and geographical expansion of the ruffe has injured native fish populations (Anttila et al. 1998).

Susceptibility to Invasion In areas with high species richne ss, non-native species are unlikely to be able to establish in a

new environment because the native flora and fauna are better able to comp ete for and utilize the resources they have evolved to use (Stachowicz et al. 1999, Seabloom 2003) (Barko and Smart 1981). Increasing the number of reproductive members of an i nvasive population increases the probability that a successful es tablishment will occur (Kolar a nd Lodge 2001). The structure and function of aquatic and coasta l ecosystems are adversely aff ected by increased stress from human activities, which have altered these syst ems for a long period of time. The scale, intensity, and rate of human activ ities and associated impacts have significantly increased in the past century from factors su ch as population growth, highe r levels of consumption, and technological advances, making these system s more vulnerable to invasions by non-native species (Barko and Smart 1981, Cohen et al. 1984, Bogan 1993, Cohen and Carlton 1998, Maggs and Stegenga 1999, Ruiz et al. 1999, Cozzi et al. 2010, Strayer 2010, Norkko et al. 2011).

In addition to directly altered ha bitat, indirect habitat alterations such as climate change can

contribute to the increased risks of invasions. Increasing winter water temperatures in the mid and high latitudes (Christensen et al. 2007) can provide more favorable conditions for invasive species to become established. Stachowitz et al. (2002) found that increased winter water temperatures were correlated with increased recruitment of non-native sea squirts and concluded that continued climatic warming could facilitate invasions of species from less extreme climates.

Due to their tolerance for a variety of salinity conditions and port locations in estuarine areas, estuarine or brackish water specie s are more likely to establish than species from other habitats (Nehring 2006a). Estuaries and co astlines are particularly suscep tible to invasives because of their proximity to sources of invasives, such as ballast discharge (Carlton and Geller 1993, Ruiz

et al. 2000, Williams and Grosholz 2008). Species capab le of asexual reproduction may also be

better able to establish in new habitats due to the lack of allee affect s, bottlenecks, and other genetic effects associated with sexual reproducers (Roman and Darling 2007).

Progression of an Invasion When assessing when a listed species or its critical habitat may be affected by a non-native species, the most important question is when in the invasion process can listed species or their

critical habitat be affected ? Carlton (1985), Baldwin (1992) , Hallegraff (1998), and Carlton (2005) identified seven steps in the ballast mediat ed invasion process. Th e first step requires the potentially invasive species to be present at that time of year or under those water quality conditions creating the potential for an interactio n between that species and the vessel. Second, if the potential invasive species is present at the port with th e ship, it must then be taken up during ballast water intake, but al so survive being taken up into the ba llast tanks. Now in the ballast tanks, that individu al or those individuals of the potentially invasive species must survive during transport from that foreign port to a US port. Fourth, upon arrival in the US, the potentially invasive species must be in the di scharged ballast water and survive the discharge308 process. Fifth, upon release, the potentially invasive spec ies must survive in the new environment. Sixth, the newly introduced sp ecies must establish a reproductive population, producing viable offspring. And finally, after that species is able to re produce, it must survive and reproduce effectively enough to estab lish a long-term population in the area.

While the likelihood that any give n individual of a part icular species might go through all the steps previously described and b ecome successfully established in the long-term is quite low, the

likelihood that a vessel will releas e live organisms from ballast wa ter (stage five) is very high, especially given the large volumes of ballast wa ter discharged annually in US ports. It is possible that listed species or cr itical habitat exposed to rece ntly released organisms could exhibit a response, but the potential affects to listed species and their critical hab itat increase with the duration of exposure (stage six and seve n) and the wider the exposed area. Therefore, exposure to listed species is cons idered to have occurred when reproduction occurs in US waters and viable offspring ar e created (stage six).

The contention in the final PEIS is non-indige nous species “management is best accomplished by preventing introduction and establishment, whic h translates into minimizing release. Setting a discharge standard may reduce release and thus decrease the probability of introduction and establishment” (p. 4-6, 4-7, Davis 2005). While, as previously discussed, if no organisms are present in the ballast water discha rge or if no ballast water is di scharged into the environment, there is no invasive species effect to the receivi ng environment. However, most vessels will have initial organism densities greater than the organism densiti es proposed by the USCG so that vessels meeting any of the alternatives will achi eve a general reduction in the non-native species densities. Therefore, if ballast water discharge is going to have non-native species present, and indeed, globally between 3,000 and 7,000 individual species are likely being transported in ballast tanks (Carlton and Gelle r 1993) (Dunston and Bax 2003), th e level of inoculation is

critical to the risks of non- native species establishing viab le reproductive populations (Bax 2003).

For transoceanic introductions, vessel-mediated transport is virtually the only vector, but organisms may arrive in the US via several vector s, including the hull, anchor system, sea chest, internal seawater piping system, and the bilg es (Keller and Lodge 2007) (Padilla and Williams 2004, Coutts and Dodgshun 2007). Furthermore, non-nati ve species establish in US waters from

other vectors besides vessels such as, birds (Johnson and Carlton 1996, Green and Figuerola 2005), aquaculture (Padilla and Williams 2004) , Keller and Lodge 2007), aquarium trade (Padilla and Williams 2004), Keller and Lodge 2007), and live bait (Johnson and Carlton 1996). Nonetheless, the comparatively large volume of ballast water and its protection from open-ocean conditions during transit implies th at ballast water discha rge is likely to be a primary source of non-native species associated wi th shipping. The reductions of organisms released via ballast water discharge is dependent on the complex inte rplay of a number of factors: the types of organisms; their reproductive mode; their life stag e; their nutritional st atus; the physical match between source and receiving waters; the time of the year; suitable substrate (if the species has a benthic phase); the presence of other invaders that might facilitate establishment; presence of native or non-native competitors, predators, parasites, or diseases that might inhibit establishment; pollutants in the receiving waters; and frequency of release of a given organism.309 Invasion rates vary considerably with region and propagule delivery. One of the main factors to be considered is the large variability in the de nsities (concentrations) of organisms transported via ballast water. Sampling studies show that or ganism densities in a ballast tank or ballasted cargo hold can range from as little as one individual per cubic meter (very low densities) to thousands per cubic meter (very high densit ies) (Cohen 1998, (Carlton 2001). These studies rarely focus on the bacterial and viral component s of ballast, but these organisms may exist in even greater densities. The gr eater the volume discharged the more non-native species that can

be introduced to the local environment. Additio nally, multiple ballast discharges in the same port from similar source waters likely contain many of the same species, increasing the invasion

risk. As Minton et al. (2005) noted, “Even with the more stringent IMO standards, however, per-ship discharges in excess of 10 6 total zooplankton remain possibl e, as a result of the large ballast discharge volumes of some ships. Moreover , the size and frequency of ship discharges to individual port systems are additive in terms of propagule delivery, and may operate in concert to affect the risk of invasions.” Non-native species from ballast water discharg e are also additive to non-native species already established (stage 6 and 7) or recen tly discharged from another ballast tank or ballasted cargo hold (stage 5).

The amount of shipping traffic and the number of incidents of invasions per year are both increasing. In this way, viable populations can be artificially increased or small populations could become viable. Additive impacts from multiple ballast discharges can increase small, but non-invasive viable populations (s tage 6) into problematic populations (stage 7) or increase small numbers of individuals (stage 5) into viable populations (stage 6) (Ricciardi 2001). Furthermore, in some cases, benign viable populations can also be allowed to grow to problematic sizes when a second introduced species competes with a native species that ha d been dominant to the initial non-native species (Ricciardi 2001). This was demonstrated in a study of native mussels, nonnative clams, and non-native crabs, where the ne wly introduced crabs selectively preyed on the native mussels, thereby freeing the non-native clam s (which had been living in the area for at least 50 years without having a population explosi on) from competition and facilitating their spread in the ecosystem (Grosholz 2005).

When ballast water is discharged, there are ge nerally many non-native species poured into the environment, but due to biotic and abiotic factors, those species rarely live beyond their introduction. While it is gene rally accepted that the most im portant vector responsible for transporting and introducing non-native species is ballast water discha rge (Carlton and Geller 1993, Tamburri et al. 2002) USCG 2011), NMFS was only able to find one study that discussed the relative importance of various vectors. Of the 329 aquatic invasi ve species analyzed by

Molnar et al. (2008), potential sources of those organisms were up to 69% via shipping, 41% by aquaculture, 17% by construction, 6% by aquari um trade, and 2% by live seafood trade. According to Molner et al. (2008), of those 69% (227 species) introduced by shipping, 39% were exclusively introduced by hull fouling, 31% excl usively by ballast water, and the other 30% by

both.

Invasive Species History

Pre-Ballast Water Regulations310 Invasion corridors have been open since ships first sailed from one port to another; however, non-native species generally did not esta blish in the US until early in the 19 th century (Grosholz 2002) USCG 2011). As ships have increased in size, their ballast tanks are larger and capable of discharging a greater volume . Untreated ballast water has between 3,000 and 50,000 large organisms (over 50 microns) per cubic meter, resulting in a discharge of up to 100,000,000 large organisms and a national average of 10,000,000 large organisms per discharge (Minton et al. 2005). These discharges and resu lting invasions have occurred at an increasing rate through the 1980s, 1990s, and 2000s.

Identifying the precise time of an invasion is of ten difficult because there is a lag between when the species are discharged (stage five) and when they are discovered (generally stage six or seven). To illustrate how complicated predicting the invasion process can be, Carlton (1996) noted that in some cases, invasion pathways ha ve been open for decades with vessels moving between ports and discharging ballast water wit hout a non-native species es tablishing, then for no explainable reason, a species lik e the zebra mussel becomes a pr oblematic invade r. While it may be impossible to know for sure why the zeb ra mussels established in the 1980s instead of the 1960s, increased habitat altera tion has made many areas more susceptible to colonization by non-native species (Brothers and Spi ngarn 1992, Carlton 1996, Lake 2004, Didham et al. 2007). Furthermore, some non-native species that are curre ntly in stage six of the invasion process may have been present for decades without being noticed because of a lack of environmental monitoring or their harmless nature.

Annually, reproductively viable populations of non-native species were being introduced at the rate of one species per year in the early 1800s, increasing to approximately 3 new species a year in San Francisco Bay alone between the 1960s and 1990s (Cohen and Carlton 1998). By the early 1990s, non-native species comprised 40 to 100% of the common species in the San Francisco Bay, 97% of the total organisms, a nd 99% of the biomass there (Cohen and Carleton 1995, Cohen and Carlton 1998). This story was sim ilar throughout the US as habitat alteration and economic globalization led to increasing reports of ecologi cal and economic impacts from invasive species. As a result of the social problems resulting from invasive species introductions, Congress passed NANPCA in 1990 and amended the law in 1996 when they passed NISA in an effort to cont rol the rate of invasions.

Ballast Water Exchange Ballast water exchange, when ocean water from outside the EEZ is pum ped through the ballast tanks to exchange the ballast wa ter taken up from nearshore port environments and replaced with less species-rich water from the mid-ocean, was ma de mandatory for the great lakes in 1993. It was then made voluntary for ships entering al l US waters in 2001. In 2004, BWE was made mandatory as a ballast water treatment method in tended to protect US waters from invasive species and their impacts. The effectivene ss of BWE varies depe ndent upon vessel type (design), exchange method, ballasting system c onfiguration, exchange location, and method of study. One group of studies suggests that the efficacy of ballast water exchange is 80 to 99% per event (Rigby and Hallegraeff 1993, Smith et al. 1996, Dickman and Zhang 1999, Zhang and Dickman 1999, Hines and Ruiz 2000, Taylor and Bruce 2000). Othe r studies demonstrate that the volumetric efficiency of BWE ranges fr om 50 to 90% (Zhang and Dickman 1999, USCG 2001, Battelle 2003). Recent work by Ruiz et al. (in NOAA 2007) has shown that BWE can311 range from 88% efficacy of BW removal in container ships to between 97% and 99% in tanker ships, container ships, and bulk carriers. Locke et al. (1993) found a 67 to 87% exchange efficacy in removal of brackish-water-tolerant organisms from ballast water after mid-ocean exchange. Furthermore, studies using methlylene blue dye in ballast tanks have shown a 75 to 95% (for the phytoplankton community) efficacy was demonstrated in a test where ballast water was flushed through tanks for a continuous nine hours (equivalent to three tank volumes) (Wonham et al. 1996). Smith et al. (1996) showed a reduction in concentrations of larvae and plankton by 50 to 90% in ballast tanks after mid-ocean exchange. Ruiz et al. (in NOAA 2007) found that empty-refill BWE removed between 80% and 95% of zooplankton and 74% and 93% of phytoplankton (mean values from multiple samples). Dunston and Bax (2008) showed that exchanging ballast water three times, removed 95% of the organisms, but reduced the risk of invasion by only 21%.

Under BWE management, the reductions in organism s in ballast tanks resu lted in reductions in the number of invasions from the highs seen in the 1990s. Mid-ocean BWE reduces the organism density by a factor of zero (upper boun d) to one-tenth (lower bound) (USCG 2003). As noted above, without any treatment, the mean ballast water di scharge contained 10,000,000 large organisms (range of 100,000 to over 100,000,000) and under BWE the number of introduced large organisms fell to a mean of 1,000,000 (Minton et al. 2005). The reduction in organisms corresponded to an annual reduction in the number of established, reproductively viable, non-native populations by 53.7 to 62.7%. However, due to the variable success of BWE and some vessels not being able to conduct BWE b ecause of safety issues, the USCG decided to

establish a numerical standard to further reduce the potential for establishing invasive species. Under the proposed management program, BWE w ill no longer be an authorized treatment method and would only be used by ve ssels entering the Great Lakes.

Despite the dramatic reductions in the number of organisms released, by each ballast discharge, the number of organisms being released in small ge ographic locations is very high. In the Port of Hampton Roads in the Chesapeake Bay, Drake et al. (2007) estimated that 10 20 microorganisms,

6.8x10 19 viruses, and 3.9x10 18 bacteria cells are discharged via ballast water annually. Lawrence (2008) estimated 3 x 10 22 viruses to be transported annua lly. In 2005, a total of roughly 22 million tons of ballast water was discharged at the four largest west coast ports; 27% of this ballast water originated from one of the othe r three major US west coast ports (Simkanin et al. 2009). Of the taxa assessed in these discharges, Drake et al. (2007) estimated that 56% could survive in the discharge habitat. Egg, larvae, an d juvenile life stages appear to be the most common life stages of animals and plants establishing invasive populations (Mahtab et al. 2005, Von Halle and Simberloff 2005)).

Proposed USCG Numerical Standard The goal of the numerical standard is to further reduce the number of organisms released in

ballast water from BWE and to also increase the total number of vessels that treat ballast water.

While BWE results in the di scharge of 1,000,000 large organi sms, under the worst possible conditions, the numerical standard would release that many large organisms, but on average, the number of large organisms discharg ed would be reduced to 43,500 (Minton et al. 2005). And

while research on the number of smaller or ganisms discharged under BWE hasn’t been conducted, the largest expected discharge of small organisms is 100,000,000,000 and the mean312 number of small organisms that would be discharged is 4,350,000,000 under the numerical

standard.

BWE is more effective than numerical standards fo r freshwater when ballast water is discharged due to the salinity differences (Gray et al. 2007). Despite the loss of th e lone potential benefit of BWE, this numerical standard will reduce the nu mber of organisms per volume from the current BWE levels. To assess the impacts of establishing this new standard, it is necessary to calculate

the likelihood of a new non-nati ve population establishing. Furt hermore, NMFS must consider whether more stringent standards would provide additional protection for listed species and their critical habitat. NMFS uses analyzed the pr obable exposure associated with a standard 1000 times more stringent th an the currently proposed standard because that is the alternative identified in the final PEIS, the Notice of Proposed Rule Making, and IFR.

The final PEIS uses a population viability anal ysis (PVA) to relate initial population size and extinction probability to determin e the probability of successful establishment of new non-native populations. A PVA is normally used to dete rmine the size of a na tive, viable population necessary to ensure a high probabi lity of survival. In the context of this final PEIS, the concern is the initial non-native species population size mu st be small enough to ensure a high probability

of extinction. While PVAs are intended for native species relying on native resources, NMFS is unaware of a better model and furthermore, theo retical ecology suggests that stochastic events and allee effects are more likely to cause the extinction of small popula tions, regardless of whether they are native or non-native. Theref ore, NMFS is unaware of a better model for predicting the likely changes in invasion ri sk from BWE to a numerical standard.

The PVA analysis in this final PEIS is based on the diffusion model proposed for the analysis of population viability by Dennis et al. (1991). The model involves two main parameters, one relating to population grow th rate and the other to population variability around mean growth. In conjunction with an estimate of the proporti onal reduction in initial population size resulting from the alternatives, this model allows for a co mparison of the alternatives in terms of their effects on the probability of reduced introductions of organisms.

The diffusion model provides an assessment of the relative, proportional effectiveness in increasing extinction probability of a particular ballast water sta ndard as compared to untreated ballast water or BWE. The USCG did not perform a comparison in absolute terms because it would be very difficult and the results would be highly uncertain given th e scientific knowledge of invasive species ecology when this anal ysis first began in 2001 (Ludwig 1999).

The main premise of the model is that a decr ease in the number of living organisms initially introduced reduces the probability that a population becomes successfully established. This is because there is a relationship between initial population size of the introduced organisms and their extinction probability (Colautti et al. 2006) (Hopper and Roush 1993, Memmot et al. 1998, Grevstad 1999)(Colautti et al. 2006). While this model takes into account allee effects and reproductive events (stage six) of the newly introduced species, it fails to account for supplemental ballast mediated introductions of the same species to the area from discharges from other vessels (Dunstan and Bax 2008). Introducti on of individuals in separate events can

increase th e genetic diversity avai lable for an invasive species, allowing for more rapid evolution313 and exploitation of the new environment; the same can occur from high levels of variation within an establishing population (Roman 2006, Williams and Grosholz 2008). In the late 1990s, 21 billion gallons of ballast were discharged into ports of the US (Bright 1999, EPA 2001). By 2010, over 64 billion gallons of ballast water were being discharged by ships discharging between 1,500 and 103,000 cubic meters at a time (Minton et al. 2005). Despite increased precautions associated with NISA, new introducti ons are occurring throughout the United States at an increasing rate as a resu lt of increased ship traffic (G rosholz 2005). Therefore the PVA model used by the USCG likely over-estimates the extinction likelihood of non-native species and therefore, the effectiveness of the numerical standard.

The PVA modeled by the USCG provided a range of possible effectiveness using models that assume either a single species discharged from ba llast tanks at the numerical standrad levels or multiple species discharged from ballast tanks at numerical standard levels. The single species model reported up to an 8% reduc tion in the introduction rate of small organisms, but between 0 and 50% reduction in the likely introduction of large organisms. The long-term objective was also modeled, which would result in a 0 to 33% reduction of small organisms and a 0 to 59% reduction of large organisms. However, this mo del is not ecologically realistic because it is unlikely that every organism discharged by ballast water is the same species. The multiple species scenarios presented in the final PEIS ar e more ecologically real istic than evaluating a single species scenario. Under the multiple spec ies scenarios, the numerical standard could reduce the chance of invasion by between 37- and 63% from the invasion risk present in BWE. However, because the PVA model likely overest imates the effectiveness of the numerical standard by not considering the volume of ballast water from multiple ships, NMFS chose to use

the 37% reduction in risk to be conservative an d protective of listed species. Following the same rationale, the long-term objective of the ballast water program could be conservatively 85% more

effective than BWE.

Estimating Future Invasion Rates Many studies have been conducted to determine a valid model for predicting invasion rates, including recent assessments assessing the prop osed numerical standard performed by the National Academy of Sciences (2012) and EPA (2011) . The goal of both of those studies was to predict the invasion probability of individual species given propa gule pressure. The two studies

determined there are no valid models to address th e questions they were tasked with answering. Both of these reviews are complete and addre ss the limitations of our knowledge of invasive species modeling to date. NMFS agrees no models are infallible and that the invasive species science is not well developed enough to predict sp ecific invasive species probability of invasion; however, for the purposes of this national level programmatic, a more general model is used.

Drake and Lodge (2004) and Drake et al. (2005) provide a global estimate for the probability of an invasive species establishing given the volume of ballast water discharged. The goal of their model is "not to predict establishment probabilit ies of any particular sp ecies, but to predict invasion rates over a range of species" (NRC 2012, p. 105). The National Academy of Sciences goes on to say that, "the chance of a single invasion by a size class of organisms can be predicted as a function of the number and volume of releases, independent of the number of individual organisms" (NRC 2012, p. 105). The shortcoming of this model, as noted by the National Academy of Science is that it has not been validated.314 With the intention of attempting to validate th e global probability analysis provided by Drake and Lodge (2004) and Drake et al. 2005, NMFS estim ated the probability of an invasion using

their model as well as a valida tion technique using hi storic information from a well studied location. Applying the best scie ntific information available, the validation method requires knowledge of historic invasions for specific por ts and a conversion from no ballast water

treatment to the current ballast water standard to estimate the cu rrent invasion rate. This method presents two shortcomings: (1) there is limited in formation on historic invasion rates, with only sufficient knowledge from San Francisco Bay, and (2) like the USCG PVA analysis, it does not consider changes in ballast wa ter discharge volumes or vess el traffic through time.

In our analysis, we rely on the PVA presented in the USCG final PEIS to estimate the anticipated invasion rate because it is currently the best av ailable estimate of the impacts of implementing

the numerical standard. Because the ballast wa ter standard establishes a numerical limit of organisms of a particular size a llowed in a specific volume of ballast water, ideally a predictive model would assess the probability of an invasion given the volume of ballast water discharged in a bioregion or the entire US in order to establish the anticipa ted annual invasion rate.

The PVA designed by the USCG provides informati on on the generally anticip ated results of the

numerical standard relative to th e results of BWE. It indicates that initially at the numerical standard, assuming multiple species are discharged by ballast water, that there is conservatively a 37% proportional expected reduction in the numbe r of invasive species introduced. Also, Dunstan and Bax (2008) show BWE reduces the likelihood of an inva sive species establishing by between 43 and 48%. However, this doesn’t predict the absolute number of new non-native species that will be introduced each year, which is the number actually needed to estimate the

likelihood that threatened or enda ngered species and th eir critical habitat may be exposed to nonnative species.

The method of predicting invasive species based on their historic rate of introductions can only be done using San Francisco Bay; however becau se San Francisco Bay is the most invaded body of water in the US and possibly the world (Cohen and Carleton 1995, Connor et al. 2002,

Costello et al. 2007), that invasion rate can be used as a conservative estimate for all ports in the US. Cohen and Carlton (1998) ar gue that while it is difficult to determine whether a newly discovered invasive species was recently establ ished or had been pres ent for a long time, all evidence suggests the rate of invasions was increasing in exponential fashion between 1850 and 1995. By the mid 1990s, invasions by non-native species were occurring at th eir greatest rate in over 200 years of introductions (Cohen and Ca rlton 1998). The NMFS analysis uses the estimated number of invasive species introdu ced in 1995, and the rate of introductions under BWE to estimate the expected rate of introductions using the PVA provided in the final PEIS. In all cases, the best available numbers used in th e analysis were highly variable. Therefore, the numbers used for the estimates below represent the most conservative calculations by always using the upper error bounds of the most pr obable estimate or upper end of a range. NMFS acknowledges there is insufficient knowledge to ge t accurate numbers, but believes the estimated rates of invasion presented below repr esent the most conservative estimate.

In San Francisco Bay research first began on invasive species in 1851. At that time315 approximately one new non-native species would establish in the bay each year (Cohen and Carlton 1998). As shipping traffic increased, the number of non-native sp ecies that established in San Francisco Bay increased to 2.3 each ye ar between 1961 and 1995 (Callaway and Josselyn 1992, Cohen and Carlton 1998). However, there are several sources of n on-native species, and vessels account for approximately 69% of the time (Molner et al. 2008) (2.3*0.69=1.587 species introduced by vessels), of which as many as 61 % of the organisms that establish from vessel

mediated sources come from ballast wate r (1.587*0.61=0.968 introduced by ballast water) (Molner et al. 2008). Using this information, we estimate that by 1995, approximately 0.97

individuals introduced to the San Francisco Bay area each year were from ballast water. Dunstan and Bax (2008) showed that BWE reduced the pr obability of an esta blished population of nonnative species by 43 to 48%. This correlates with annual ballast-media ted non-native species introductions being reduced from 0.968 pe r year to approximately 0.504 (0.968*0.52) to 0.553 (0.968*0.57) per year. Then, using the multispeci es analysis in the final PEIS (0.504*0.63 and 0.553*0.63), the proposed numeric standard would be expected to reduce the number of invasive species to approximately 0.318 to 0.348, or approxim ately one new species every three years.

The calculated rate of 0.318 to 0.348 invasions per ye ar is comparable to the global multi species model used by NMFS to estimate invasion risk at the continental scale (D rake and Lodge (2004). Drake and Lodge (2004)(, p.578) used a generaliz ed linear regression with Poisson errors developed at a global scale to show “that shippi ng traffic was significantly related to the number of nonindigenous species discovered ( y = 8.47 × 10 -8 x where y is the number of non-indigenous species discovered and x is the shipping volume, in metric tonnes; p ≤ 0.0001).” Ballast water discharge volumes were obtained from the Natio nal Ballast Information Clearinghouse, which is a record of all reported ballast water discharges. However, re porting rates vary between 80 and 90% of the actual value. NMFS applied a co rrection factor assuming the conservative 80% reporting rate. This rate is expected to incr ease with the ballast water standard because

numerical values are easier to test and therefore there is greater risk in not complying with the mandatory reporting requirements. In San Fran cisco Bay the corrected average from 2005 to

2010 was 6.469x10 6 tonnes of ballast water were disc harged under BWE management. Using Drake and Lodge’s (2004) equa tion would result in approximat ely 0.548 species per year being introduced. Then, using the multispecies analys is in the final PEIS, which conservatively

estimates a 37% reduction in new invasions, the expected number of invasions in SF Bay would be approximately 0.345 individuals per year (with out the correction fact or, the estimate is 0.32 species introduced per year).

While this test is not an e xhaustive validation, both models in San Francisco Bay determine nearly identical estimates. Given the similarity of their estimates and the flexibility of the second method using the ballast water discharge volume at various ports, we believe this method provides the best available estimate for future in vasions at individual por ts as well as throughout

the US.

Despite setting a ballast water standard, listed sp ecies and habitat will likely still be exposed to ballast water discharges with non- native organisms at any standard other than zero discharge of non-natives. If the volume of ballast discharg e remains at the same level and there are noadditional habitat alterations making the environment more susceptible to invasions, ballastmediated invasions are likely to be reduced under the proposed numerical standard. However,316 changes in the volume of shipping and in trade routes, changes in habitat health and habitat accessibility, as well as changes in ballast water systems and their operation, would either

decrease or increase the level of non-native species introductions. Thus, depending on factors outside the scope of the ballast water management program, the total number of introductions

could still increase due to increases in global trade and habitat accessibility.

Case Studies As was just discussed, San Francisco Bay is a lo gical location to use as a case study because of the vast research history as well as invasion history there. Furthe rmore, the San Francisco Bay is home to six listed species and five critical habi tats in two different functional groups under NMFS jurisdiction. NMFS also decided to anal yze Puget Sound and Chesapeake Bay to assess the impacts of invasive species on listed species and th eir critical habitat. By choosing those locations, several listed species, their critical habitats, or reasonable surrogates were not considered, so we analy zed those individually.

San Francisco Bay. Among the best studied port in the US is San Francisco Bay, with major ports in both San Francisco and Oakland as well as many smaller ports in what may be the most heavily developed large es tuary in the US (Herborg et al. 2007). San Francisco Bay is also home to central valley spring-run Chinook salmon, S acramento River winter-run Chinook salmon, green sturgeon, and California central valley st eelhead (Table 39). Additionally, while they do not reside within the bay, black and white ab alone live along the California Coast within the same marine bioregion.

Carlton and Geller (1993), Kennedy (2001), and the final PEIS estimate approximately 234 invasive species are now established in the es tuary. Many of these species were established before any ballast water treatment was utilized. And as noted, there is c onsiderable uncertainty as to when a species establishes due to the delay between it being discharged and it being

discovered.

Table 39: Listed species pres ent and critical habitat designated in San Francisco Bay. Species Listing Status Designated Critical Habitat Central valley spring-run Chinook salmon Threatened Yes

Sacramento River winter-run Chinook salmon Endangered Yes Green sturgeon Threatened Yes California central valley steelhead Threatened Yes

Black abalone Endangered Yes

White abalone Endangered No

The reported ballast water discha rge into the San Francisco Bay bioregion has increased between

2005 and 2010 with 4.43x10 6 , 4.6x10 6 , 4.49x10 6 , 5.9x10 6 , 5.7x10 6 , and 5.9x10 6 tonnes discharged annually from around the globe, respectively. These num bers were corrected for the unreported discharges. If we accept the USCG’s P VA results, future invasions in San Francisco Bay at the discharge rate averaged over 2005 to 2010, would be approximately one new species establishing every three years unde r the proposed numeric ballast discharge standard. Following this trend in discharge volume s; however, an increase of approximately 372,000 tonnes per year in San Francisco Bay would be expected each year beyond 2010, increasing the risk of an invasion by approximately 0.032 new non-native species ea ch year. In other wo rds, if the rate of 0.345 new species per year in 2010 is correct, gi ven the upward trend in ballast discharge317 provided by the National Ballast Information Clear inghouse, the potential for an invasive species to establish in 2012 would be approximately 0.407 new non-native species each year. On the other hand, if the USCG ballast water management program increas es the standard in the event better technology is practic able, their goal is a sta ndard that is 1000 times more stringent than the proposed numeric standard. In that case, the an ticipated rate of invasi ons would fall to about 0.09 individuals per year or about one new non-na tive species every 11 year s. Although we can

estimate the rate of invasion w ithin a given system, we cannot predict the species that would

constitute future invasions.

Puget Sound. Puget Sound has a number of ports, includ ing Seattle, one of the largest ports in the US (Herborg et al. 2007). Puget Sound is home to bocaccio, canary rockfish, Puget Sound Chinook salmon, Hood Canal summer-run chum salmon, Pacific eulachon, Puget Sound steelhead, and yellow-eye rockfish (Table 40).

Table 40: Listed species pr esent and critical habitat designated in Puget Sound. Species Listing Status Designated Critical Habitat Bocaccio Endangered No Yellow-eye rockfish Threatened No Canary rockfish Threatened No

Pacific eulachon Threatened Yes Puget Sound steelhead Threatened No Puget Sound Chinook salmon Threatened Yes

Hood Canal summer-run chum salmon Threatened Yes

Pacific eulachon Threatened Yes

Resident killer whale Endangered Yes Humpback whale Threatened No

Eastern Steller sea lion Threatened Yes

The number of non-native species in Puget Sound has increased dramatically over the past 30 years. The first analysis identified 18 invasive invertebrates (Carlton 1979). However, research nearly 20 years later by Elst on (1997), Ruiz and Hines ( 1997), WDFW (1998), and Cohen et al. (1998) determined there were between 31 and 67 non-native species repro ducing in Puget Sound. Cohen (2004) later re-sampl ed the sound and identified 76 non-native species.

Ballast water discharge into Puget Sound has increased between 2005 and 2010 with corrected

volumes of 4.99x10 6 , 5.16x10 6 , 4.92x10 6 , 6.12x10 6 , 5.25x10 6 , and 5.8x10 6 tonnes discharged annually from around the globe, resp ectively. The mean corrected di scharge rate over this time

was 5.37x10 6 . Again, if NMFS adopts the USCG’s PVA fr om the final PEIS, we would expect a

37% reduction in the number of new non-native species under the numerical standard and an 85% reduction under the ballast management program’s published goal. Therefore, over the past six years, an invasion rate of approxima tely 0.455 non-native species per year (8.47x10 -8 x 5.37x10 6 x 0.63) would be expected or approximately one new species in Puget Sound every two years. And as we did for San Francisco Ba y, observing the increasi ng trend in discharge volumes, there is an increase of approximate ly 161,000 tonnes per year each year beyond 2010, increasing the risk of an inva sion by approximately 0.014 new non-na tive species each year. In other words, if the rate of 0.455 ne w species per year over the past six years is correct, given the upward trend in ballast discha rge provided by the National Ballast Information Clearinghouse, the potential for an invasive species to esta blish in 2012 would be approximately 0.483 new non-318 native species each year. On the other hand, if the USCG ballast water management program increases the standard to 1000 times more stringent than the proposed numerical standard, the anticipated rate of invasions would fall to about 0.068 individuals per year or about one new non-native species every 15 years. Although we can estimate the rate of invasion within a given system, we cannot predict the species th at would constitute future invasions.

Chesapeake Bay. The Chesapeake Bay is home to two of the largest ports in the US (Herborg

et al. 2007) as well as a number of smaller ports that also receive ballast water discharge. The Chesapeake Bay is the focus of national attention in an effort to restore the largest estuary and one of the most productive fisheries in the US. At various times throughout the year, shortnose sturgeon, five distinct populations of Atlant ic sturgeon, and green, hawksbill, leatherback, Kemp’s Ridley, and Northwest Atlantic DPS of loggerhead sea turtles use the Chesapeake Bay as their home. (Table 41).

Table 41: Listed species pres ent and critical habitat designated in Chesapeake Bay. Species Listing Status Designated Critical Habitat Shortnose sturgeon Endangered No Atlantic sturgeon Endangered and Threatened No

Green sea turtle Threatened Yes Hawksbill sea turtle Endangered Yes Leatherback sea turtle Endangered Yes Loggerhead sea turtle Threatened Yes Kemp’s ridley sea turtle Endangered No

Carlton (1996) assessed the non- native species in the Chesapea ke Bay and found only 20 aquatic invasive species. Following this assessment, Smith et al. (1999a) noted there were relatively few invasive species in the Chesapeake Bay because of poor water quality conditions. Simultaneously, another study accounting for known invasive species as well as species of unknown origin was conducted and determined ther e were 196 exotic or cryptogenic species in the Chesapeake (Ruiz et al. 1999). It is possible the second study uncovered more non-native species by taking more samples or because e fforts to improve water quality are allowing nonnative species to capitalize on the nutrient rich wa ters and greatly distur bed habitat in the bay.

Ballast water discharge into Chesapeake Bay has increased between 2005 and 2010 with

corrected volumes of 6.72x10 6 , 8.0x10 6 , 10.76x10 6 , 16.63x10 6 , 14.45x10 6 , and 21.06x10 6 tonnes discharged annually from around the globe, resp ectively. This is th e greatest proportional

increase in ballast water discha rge in the US during this time period. Again, if NMFS adopts the USCG’s PVA from the final PEIS, we would e xpect a 37% reduction in the number of new nonnative species under the numerical standard and an 85% reducti on under the ballast management program’s published goal. Therefore, using the six year average from 2005 to 2010, an invasion rate of approximately 1.1 non-na tive species per year (8.47x10 -8 x 12.94x10 6 x 0.63) would be expected or approximately three new species in Chesapeake Bay every three years. As we did for San Francisco Bay and Puget Sound, observing the increasing trend in discharge volumes, there is an increase of approximately 2,869,000 t onnes per year each year beyond 2010. This increase in ballast water discharge is likely to increase the risk of an invasion by approximately 0.24 new non-native species each year. In other word s, if the rate of 1.1 new species per year is correct, given the upward trend in ballast discharge provided by th e National Ballast Information Clearinghouse, the potential for an invasive species to esta blish in 2012 would be approximately319 1.34 new non-native species each year. On the other hand, if the USCG ballast water management program increases the standard to 1000 times more stringent than the proposed numerical standard, the anticipat ed rate of invasions would fa ll to about 0.164 individuals per year or about one new non-native species every six years. Although we can estimate the rate of invasion within a given system, we cannot predic t the species that woul d constitute future

invasions.

Smalltooth Sawfish, Johnson’s Sea Grass, Elkhorn and Staghorn Coral Smalltooth sawfish, Johnson’s sea grass, and el khorn and staghorn coral occupy similar ranges around the southern tip of Florida. The range of these corals extends to Puerto Rico, while smalltooth sawfish may range betw een Florida and some Caribbean Islands. Johnson’s sea grass is found in isolated bays, inlets, and lagoons alon g Florida’s Atlantic Coast. While smalltooth sawfish, Johnson’s sea grass, and the corals exist in the same marine bioregions as ports in Houston and New Orleans (ballast water bioregion CAR-I), they only use the far eastern areas in this marine bioregion and ther efore discharges from Mobile Ba y and west were considered unlikely to affect these species. Despite excluding some major ports, this ar ea is still exposed to discharges in Tampa, Miami, and the Florida Keys.

In the waters of Florida alone, 23 non-indigenous aquatic plant sp ecies have become established (McCann et al. 1996). Approximately 50 non-indigenous aqua tic species have established in the state, but not all from ballast wa ter discharge (Courtenay 1997).

Ballast water discharge into coastal Florida and Caribbean waters has remained steady between

2006 and 2010 with corrected volumes of 14.43x10 6 , 15.19x10 6 , 15.52x10 6 , 14.26x10 6 , and

14.08x10 6 tonnes discharged annually from around the globe, respectively. If NMFS accepts the USCG’s PVA from the final PEIS, we would e xpect a 37% reduction in the number of new nonnative species under the numerical standard and an 85% reducti on under the ballast management program’s published goal. Therefore, over the last 5 years, an aver age invasion rate of approximately 0.784 non-native species per year (8.47x10 -8 x 14.698x10 6 x 0.63) would be expected or approximately four new species along the southern Florida coast every three years. And as we did for the three case studies, observi ng the trend in discharge volumes, there is a decrease of approximately 70,000 tonnes per year each year beyond 2010, decreasing the risk of an invasion by approximately 0.006 new non-native species each year. In other words, if the rate of 0.784 new species per year is correct, given th e downward trend in balla st discharge provided by the National Ballast Information Clearinghouse, the potential for an invasive species to establish in 2012 would be approxi mately 0.772 new non-native species each year. On the other hand, if the USCG ballast water management progr am increases the standard to 1000 times more stringent than the proposed numerical standard, th e anticipated rate of invasions would fall to about 0.187 individuals per year or about one new non-native species every five years.

United States estimate Shipping is one of the primary m eans that the US imports and expor ts goods to the rest of the world. There are three general types of aquatic ecosystems in th e US: freshwater, estuarine, and saltwater, and all of those systems receive balla st water discharges and can be affected by introduced species. EPA estimates that approximately 75,000 large ballasting vessels use US waters, but not all of those sh ips are regulated by this FR. Additionally, approximately 6,000320 new ships are built each year around the world and th e size of the new ships are much larger than the ships currently used for international trade (Bax et al. 2003).

The USCG’s final PEIS estimates approxi mately 150 non-native species have become established in US waters since 1970. Many of these likely became established prior to 1993

when BWE was first introduced. The rate of invasion may have slowed, but Grosholz (2005) estimates there are over 500 different species of non-native organisms now in US coastal waters and furthermore, new introductions are occurring at an increasing rate due to increased shipping and ballast discharge.

To estimate the relevant ballast water discharg es to NMFS resources, NMFS analyzed the total

coastal and estuarine ball ast water discharges in the US over the past six years (excluded the Mississippi River Basin and Great Lakes ports). Ballast water di scharge into US waters has

increased between 2005 and 2010 w ith corrected volumes of 123.6x10 6 , 155.5x10 6 , 173.0x10 6

, 192.0x10 6 , 215.1x10 6 , and 236.5x10 6 tonnes discharged annually from around the globe, respectively. Again, if NMFS adopts the USCG’s PVA from the final PEIS, we would expect a

37% reduction in the number of new non-native species under the numerical standard and an 85% reduction under the ballast management program ’s published goal. Therefore, using the six year average, an invasion rate of approxi mately 9.7 non-native species per year (8.47x10 -8 x 182.6x10 6 x 0.63) would be expected in US waters ever y year. And as we did for local port case studies, observing the increasi ng trend in discharge volumes, there is an increase of approximately 22,582,000 tonnes per year each y ear beyond 2010, increasing the risk of an invasion by approximately 1.91 new non- native species each year. In other words, if the rate of 9.7 new species per year is correct, given the up ward trend in ballast discharge provided by the

National Ballast Information Cleari nghouse, the potential for an inva sive species to establish in 2012 would be approximately 13.52 new non-native spec ies each year. On the other hand, if the USCG ballast water management program increa ses the standard to 1000 times more stringent than the proposed numerical standard, the antici pated rate of invasions would fall to about 3.005 individuals per year nationwide.

Response Analysis

Consequences of exposing listed speci es and designated critical habitat The preceding section of this Opinion presented th e evidence that leads us to conclude that endangered or threatened speci es and proposed or designate d critical habitat under NMFS jurisdiction are likely to co-occur with discharges of ballast wa ter on, over, or near U. S. waters. In this section, we summarize information on the probable physic al, physiological, behavioral, social, and ecological responses of endangered or threatened species or constituent elements of critical habitat given exposure to potentially invasive organisms. Our purpose is not to provide a comprehensive review of the probable responses of proposed, endangered, or threatened species to these stressors; instead, our in tention is to identify the range of representative responses we would expect proposed or listed species and their cr itical habitats to exhibit given exposure to these stressors. We do provide probability estim ates of harmful invasions at the end of the Integration and Synthesis section.321 Direct effects to listed species Invasive species are a major threat to many ESA-listed species. For species listed by the USFWS, 26% were listed partially because of the impacts of invasive species and 7% were listed because invasive species were th e major cause of listing (Anttila et al. 1998). Pimentel et al. (2004) found that roughly 40% of US listed species are at risk of becoming endangered or extinct completely or in part due to invasive species, while Wilcove et al. (1998) found this to be 49%, with 27% of invertebrates, 37% of reptiles, 53% of fishes, and 57% of plants imperiled partly or wholly due to non-native invasions. In some re gions of the world, up to 80% of species facing

extinction are threatened by invasive species (Yan et al. 2002, Pimentel et al. 2004). Clavero and Garcia-Bertro (2005) found that invasive speci es were a contributing cause to over half of the extinct species in the IUCN database; invasive species were the only cited cause in 20% of

those cases.

Introduction of non-native species was found to be a leading cause of endangerment in some western US areas (Flather et al. 1998). Richter et al. (1997) also identified invasive species as one of three top threats to threat ened and endangered freshwater species in the US as a whole, but particularly so in the wester n US, where recovery of 60% of listed fishes is impaired due to non-native species.

In Hawaii, 44 species of native freshwater fishes are endangered partly due to the 33 non-native species introduced there, while another 27 native species have been negatively affected by these introductions(Maciolek 1984, Pimentel et al. 2004). Once invasive species have established, they can potentially migrate widely into su rrounding regions. Mitt en crabs have been documented to occur up to 1,000 km upstream of their likely establishment origin and are currently an invasive species in the Chesap eake and San Francisco Bays (Leichsenring and Lawrence 2011). Similarly, the Asian kelp expande d its range into San Francisco Bay following

its initial establishment in southern California and Mexico (Whitman Miller et al. 2004).

Pathogens, Parasites, and Toxins The impacts of introduced pathogens in the aqua tic environment has been poorly explored and we likely know very little about the true fr equency and significance of pathogen invasions (Drake et al. 2001). Although most available literature addresses pathogens as an invader, Dobbs and Rogerson (2005) contend that free-living microbial organisms lack a true range due to their potentially high reproductive rates, broad to lerances to physical conditions, and ability to form resting stages and, as such, can be cons idered cosmopolitan and not invaders. A similar cosmopolitan nature has been suggested for bryozoans, barnacles, polychaete worms, hydroids,

and molluscs that reside on marine de bris (Barnes 2002). Despite these hypotheses, NMFS would expect that if this asse rtion were true, the occurrence of novel diseases would not occur on a geographic basis or, in the presen t situation, in association with re latively high levels of ballast water discharge, which a majority of the availabl e literature contradicts. Furthermore, although a pathogen may already exist in waters receiving ba llast water discharges, some pathogens such as

viruses and bacteria occur in different strains which may be novel to potential hosts, thereby exacerbating the potential for morbidity and/or mortality from a “native” pathogen (Cohen 2010).322 Diseases. There are several pathogens that have adve rse effects to native fish populations. Viral hemorrhagic septicemia virus was initially introd uced to North America from western Europe via ballast water discharge in th e Great Lakes. Following its introduction, it caused extensive morbidity and mortality to both farmed and wild fish populations including salmonids, smelts, and rockfish, but not sturgeon (McKenna 2007, Bain et al. 2010). Since its establishment during or before 2005, the virus has been found wi dely in the Northern Hemisphere.

Pathogens also have adverse eff ects to invertebrate communities . Molluscs such as black and white abalone seem to be particularly sensitiv e to pathogens. Various species of the genus

Vibrio , known to cause cholera in humans, white pox a nd white plague type II diseases in corals, and mortality in abalone of the same genus as black and white abalone, have been identified in ports and ballast water of vessels (Anguiano-Beltrán et al. 1998, Ben-Haim and Rosenberg 2002, Aguirremacedo et al. 2008). Oyster species have sustai ned several outbreaks from invasive pathogens, including Haplosporidium nelsoni (the cause of MSX disease, which Chesapeake Bay eastern oysters have shown 75-92% mortality to) and Perkinsus marinus (the cause of Dermo disease) in California, eastern Nort h America, and Europe (Ford and Haskin 1982, Andrews 1984, Burreson et al. 2000, Renault et al. 2000, Burreson and Ford 2004), Bonamia

ostreae in Europe (Van Banning 1987, Ciguarria and Elston 1997), and in the northeastern US, respectively (Ford 1996).

As described in the Status of Listed Resources , elkhorn and staghorn coral f ace severe declines in part due to morbidity and mortal ity stemming from disease. The introduction of novel pathogens into a habitat can cause overwhelming infections, particularly when coupled with stresses of a

disturbed environment which corals must already cope with (Lesser et al. 2007).

Parasites. Parasites can directly affect fish specie s, including listed fis h. The Asian parasitic copepod, Neoergasilus japonicas

, began parasitizing endangered ci chlid species in Mexico upon its introduction there (Suárez-Morales et al. 2010). When salmon fry were imported to Sweden in the 1950’s, a parasitic fluke worm accompanied the fry and infected native Norwegian wild salmon stocks, resulting in 95% mortality am ong the naïve natives (Josefsson and Andersson 2011).

Like fish, invertebrates are high ly susceptible to introduced para sites and there are a number or examples of adverse effects. A parasitic worm, Terebrasabella heterouncinata , attacked red abalone in California aquacultu re subsequent to the paras ites’ introducti on, significantly reducing growth rate. The pathogen escaped into the marine environment from waste water and shells discharged from these facilities (C ohen and Webb 2002, Culver and Kuris 2002, Bower 2006) and although ballast water was not the transfer vector in this case, it is possible for vessels to transport this pathogen within California wate rs or re-introduce it to other parts of the US. The aquaculture impacts likely highlight the pote ntia l for morbidity and mortality in the wild

because T. heterouncinata is likely capable of infecting most or all California abalone species, including black and white abalone (Kuris & Culver 1999). A signi ficant risk to black abalone’s survival and recovery is a disease called w ithering syndrome, caused by a Rickettsia-like bacterium, Xenohaliotis californiensis that may have been introduced from Asia to southern California via ballast water discharge (Friedman et al. 2000)(Smith et al. 2003, Bower 2009, Cohen 2010).323 Competitors There are several good examples of introduced species competing with listed species. In the case of listed fish, Waknitz et al. (2003) examined what impacts introducing Atlantic salmon along

the US west coast would have on Pacific salm on species, many of whom are listed under the ESA. Atlantic salmon escapees from aquaculture could potentially dilute or alter the gene pool of Pacific salmon species, direc tly compete with native salmon, in troduce new diseases to naïve populations, and compete for limited resources. Th e introduction of brook trout into freshwater systems of the western US, particularly the Colu mbia River Basin, has directly resulted in the displacement of native trout in these locations through competitive interaction and hybridization (Dunham et al. 2002, Feyreret al. 2003). Mills et al. (2004) used laboratory experiments to show competitive interactions be tween least chub and mosquitofish led to reduced survival of the

native least chub. Other studies have investigated the impacts of mosquito fish with native cyprinids (Alcaraz et al. 2008).

Like with fish, there are examples of invertebra tes being introduced that directly compete with native invertebrates, and sometimes listed spec ies. Stat and Gates (2008) found a non-endemic symbiotic dinoflagellate of the same genus used by elkhorn and st aghorn coral were present in a Hawaiian Acroporid coral; genetic analyses sugges t the symbiotic dinoflagellate was transported

to and established in the isolated Northweste rn Hawaiian Islands National Monument via ballast

water.

Prey Ballast water potentially in troduces species that become prey for native biota in the region. For instance, Hawaiian green sea turtle s historically fed upon native al gal species, but have recently included the non-native algal species Acanthophora spicifera

, Hypnea musciformis , and

Gracilaria salicornia in their diets. The contribution of th ese three alga to green sea turtle diets in the area are now so significant that they com pose 40% of the diet am ong individuals studied (Russell and Balazs 2009). Anot her two invasive algal species comprised an additional 18% of the diet (Russell and Balazs 2009).

Predators Ballast water discharges sometime s contain invasive species that are generalist predators. The comb jellyfish, Mnemiopsis leidyi , was introduced into the Black Sea in the late 1980’s as a result of ballast water discharge and was estimat ed to constitute 100 million tons of biomass in the Black Sea in 1994. This rapid increase in pop ulation caused a decline of native species due to predation (mostly copepods and cladoceran s), the extinction of some native pontellid copepods, declines in benthic zoop lankton, 70% declines in inshore fish larvae (also prey to the invasive species), benthic anoxia, mucus within the water column, and re ductions in biodiversity as a whole (Vinogradov et al. 1989, Tsikhon-Lukanina and Re znichenko 1991, Zeitsev and Ozturk 2001, Moncheva and Kamburska 2002). This species preyed on native larval fish, while also effecting several areas with in the native food web. If a sp ecies having sim ilar ecological impacts was introduced to the US, there is the ch ance larval stages of listed fish could be consumed by an invasive predator.324 Direct effects to designated critical habitat an d indirect effects to lis ted species (ecological responses)

In general, species located higher within a food web (including most ESA-listed species under NMFS’ jurisdiction) are more likely to become extinct as a result of an invasion; conversely, species that are more centrally or bottom-oriented within a food web are more likely to establish (Harvey and May 1997, Byrnes et al. 2007). Propagule pressure is generally the reason for this trend, as individuals lower in the food web tend to have higher fecundity a nd lower survival rates (r-selection). This unbalancing of food webs ma kes subsequent introductions more likely as resource utilization shifts, increasing resour ce availability, and e xploitation success by nonnative species (Barko and Smart 1981, Byrnes et al. 2007). Such shifts in the base of food webs fundamentally alters predator-prey dynami cs up and across food chains (Moncheva and Kamburska 2002). The number of extinction even ts seems to be roughly correlated with the

number of invasive establishments w ithin an area (Harvey and May 1997).

Marins et al. (2010) found that port environments, compared to natural areas, had greater species diversity owing to the presence of invasive species. Many invasive species, such as Chinese mitten crabs, zebra mussels, the ascidian Pyura praeputialis , and the polychaete Ficopomatus enigmaticus , act as ecosystem engineers that become in strumental in the environments in which they invade, changing metabolic and trophic interac tions of predator-prey re lationships as well as altering physical and chemical processes within an envi ronment (Colwell 1996, Castilla et al. 2004, Hecky et al. 2004, Wallentinus and Nyberg 2007). Howe ver, while a large majority of

available information indicates th at invasive species alter the eco systems into which they invade, this is not always the apparent result (Hewitt and Inglis 2004). For instance, changes in native species abundance appears to facilitate the expa nsion of other invasive species (Daehler and Strong 1996). Grosholz and Ruiz (1996) studied the imp acts of green crabs at three different sites, finding that ecological impacts were similar at all thre e sites, although habitat use and expansion rate exhibited c onsiderable variation.

Pathogens, Parasites, and Toxins Primarily, pathogens and species with toxic effects have direct effects to listed species, however there are instances when these spec ies affect PCEs of listed species critical habitat or indirectly affect the species in other ways . There are a number of non-native species that have the potential to either expel toxins at low levels, only becoming problematic for other members of the ecosystem if their population grows to very large sizes, resulting in very large amounts of toxins being released. In other cases, pathogens are introduced to an environment affecting organisms in the environment that would directly affect cr itical habitat PCEs or i ndirectly affect listed species. Pathogens are in some cas es very specific to hosts, but when a species similar to a listed species is introduced, eventually that parasite th at specific to the non-nati ve species can shift to also affecting similar native populations. In thes e cases, the effects m ay be directly adverse to listed species or indirect to food re sources as identified in a species’ critical habitat. And in other cases, parasites can have direct effects to PCEs of designated critical habita t or indirectly affect listed species.

Toxins. Red tide dinoflagellates have been introduced via ballast water disc harges and have the potential to undergo extreme seasonal population fluctuations. During bloom conditions, high325 levels of neurotoxins are released into local and regional surface water and air that can cause illness and death in fishes, sea turtles, marine mammals, and invertebrates (as well as their larvae) (Hallegraeff and Bolch 1992, McMinn et al. 1997, Hallegraeff 1998, Hamer et al. 2000,

Hamer et al. 2001, Lilly et al. 2002).

The dinoflagellate, Alexandrium catenella , was likely introduced to Puget Sound via ballast water discharge and is now a major source of pa ralytic shellfish poisoning and harmful algal blooms in the region (Moore et al. 2011). Another species pote ntially introduced by ballast

water, Pseudonitzschia australis , resulted in bird mortalit y in Monterey Bay (Fritz et al. 1992)

and Thalassiosira spp. have been introduced via ballast water, damaging the gills of fishes (including Atlantic salmon) and mollusks when they bloom (Takano 1956, Kent et al. 1995, Miyahara et al. 1996, Macleod 2002). In the Black Sea, invasive dinoflagellates such as Gymnodinium uberrimum and Alexandrium monilatum have become dominant phytoplankton, peaking at up to 70% of the biomass compared to native dinoflagellates, clogging fish gills and increasing turbidity (Moncheva et al. 1995)(Moncheva and Kamburska 2002). The brown alga, Aureococcus anophagefferens , causes brown tide when it blooms and has been found to be transported via vessels along the US east coast, causing diebacks of eelgrass habitat due to blooms decreasing light availabil ity and failure of scallops and mussels to recruit (Doblin et al. 2004).

Diseases. Parasites and diseases in some cases are not able to invade without a specific host also invading. However, once established with their invasive hosts, they can infect native, naïve hosts. This was demonstrated along the US Pacifi c Coast where abalone of the same genera as listed white and black abalone resulting in 50% mortality or highe r (Chew 1990, Ciguarria and Elston 1997)(Simon et al. 2006)(Nell 2002, Lleonart et al. 2003, Nehring 2006b, Haupt et al. 2010). In another case, an invasive oyster was blamed as a vector for carrying the protozoan,

Bonamia ostreae , which infected native oysters and resulted in 50-80% mortality in those infected, severely reducing th e abundance of native oysters (Chew 1990, Ciguarria and Elston 1997, Minchin and Rosenthal 2002).

The introduction of American signa l crayfish into Swedish lakes resulted in an 80% reduction in the native noble crayfish’s abundance, due primarily to the crayfish plague ( Aphanomyces

astaci ) that accompanied the invasive crustacean (Minchin and Rosenthal 2002, Josefsson and Andersson 2011). In Britain, the inadvertent introduction of the same species may have been the vector carrying a fungal diseas e that caused the extirpation of native crayfish(Minchin and Rosenthal 2002).

Parasites. Parasites have the ability to reduce healthy populations of species that may benefit listed species or their critical habitat in some way. A parasite of eels in Japan recently appeared in eels along South Africa, likely as an initial result of ballast water di scharge (Christison and Baker 2007). A commonly-occurring nematode para site of Asian eels infe cted and caused high mortality in the native eel Anguilla anguilla when the former was introduced into European rivers (Minchin and Rosenthal 2002).

Predators. There are a few examples of indirect pred atory effects caused by invasive species. European green crabs have invaded both the east and west Coasts of the US, having trophic scale326 effects to both environments. In Massachuse tts Bay, green crabs prey upon native mussels and oysters, altering community structure (La fferty and Kuris 1996, Grosholz 2002, Pimentel et al. 2004). Along the US west coast, green crabs can quickly become the dominant predator of native clams, exerting strong top- down control of at le ast 20 invertebrate sp ecies (Grosholz and Ruiz 1995), with some species showing five to te n fold declines in abundance within the first three years of the invade r’s detection (Grosholz et al. 2000, Grosholz 2002). In both cases, the suppression of these native invertebrates led to increases in their natural prey; however, organisms at higher trophic levels did not in crease in response to the green crabs.

In the early 1980s, the ctenophore Mnemiopsis leidyi invaded the Black Sea via ballast water discharge, where only three other endemi c gelatinous zooplankt on existed (Kideys et al. 2005). By the end of the decade, the invader was in such high abundance that it had become the dominant zooplankton predator, causing abnormal changes in plankton dynamics and subsequent collapses in pelagic fish stocks (Kideys et al. 2005).

The amphipod, Gammarus tigrinus , introduced via ballast water, interacts with the native opossum shrimp along Northern Ireland through direct predation on th e native and also by facilitating predation by fishes, while the sh rimp itself preyed upon the amphipod heavily in areas where it had become dense (Bailey et al. 2006). Decapods as a w hole reduce the presence of macrophytes, snails, and other bent hic animals (Strayer 2010). Also, P. macrodacylus may act as a new prey for bethivore genera lists, such as shortnose sturgeon.

Competitors. The most commonly reported impact of non- native species in the freshwater and coastal environment is competition for limited resources (Nyberg 2007). Molluscs, decapods, and aquatic plants as taxonomic groups tend to be especially ca pable invaders and have proven to be disruptive to food webs. The most comm on impacts are alteration of habitat and nutrient availability as well as altering species composition and diversity within an ecosystem (Strayer 2010). Crabs, polychaetes, and mussels can incr ease bioturbation and aerate the sediment (Nyberg 2007). Gastropods can alter the biogeochemical cycle through excretion of biogenic

silicate in the faeces an d pseudofaeces (Ragueneau et al. 2005). Molluscan invasions can also provide substrate for epibionts, shelter for bent hic species, remove nutrients from the water, decrease turbidity and increase light penetration, remove sediments, and promote phytoplankton blooms by releasing nutrients from sediments (Bertness 1984, Gutierrez et al. 2003, Hecky et al. 2004).

Invertebrates. There are many examples of invertebrate competition either indirectly affecting similar species to listed species under NMFS juri sdiction or directly affecting the habitat they rely on. The compound tunicate, Botrylloides sandiegensis , was released near Woods Hole, Massachussetts and has outcompeted other encrus ting organism in the coastal environment of southern New England (Lafferty and Kuris 1996). The invasive snail Batillaria a ttramentaria outcompetes a native species, Cerithidea californica , along northern California via a more efficient ability to convert food into body tissue and possibly greater disp ersal potential (Savino and Kolar 1996, Carlton 1999, Pothaven et al. 2001). The slipper limpet ( Crepidula fornicata ) is a voracious feeder, outcompeting native oysters and scallops in Europe, where it is invasive (Wallentinus and Nyberg 2007). The altered nutrient availability that this species produces may contribute to harmful algal blooms (Wallentinus and Nyberg 2007).327 Intentional introductions also pr ovide an opportunity to study th e impacts of non-native species on native communities. Aquaculture is often responsible for intentional introductions, such as farmed short-necked clams ( Tapes philippinarum ), which can occur in such high densities that

anoxic conditions in the benthos and water column (which can cause extensive mortality of many marine organisms) can occur as bacteria bloom in response the large am ount of pseudofeces the clams produce (Bartoli et al. 2001, Wallentinus and Nyberg 2007). Additionally, in these farmed areas, other macrobenthic organisms also occur in relatively low abundance. Pacific oysters, also intentionally introduced in many locations , can fundamentally alte r the nutrient composition and phytoplankton availability in the habitats th ey occupy and monitoring of their aquaculture introductions has shown they establish under a variety of environmental conditions (Nehring 2006b, Molnar et al. 2008).

Filter feeding benthic species such as clams and oysters convert nutrients from the water column and deposit them in the benthos (Nehring 2006b). Pacific oysters are converting mud flats along Germany and the Netherlands into oyster r eefs (Nehring 2006b). A mollusk of the genus Rapanus has been blamed for the decreased abundan ce of native oysters in some areas of the Black Sea (Vinogradov et al. 1989). The Asian mussel, Musculista senhousia , established a mat over mud flats in Mission Bay, San Diego form ing a unique habitat that increased species richness. Tanaids, gastropods, amphipods, insect larvae, and polychaetes and those species that prey upon them appear to have most benefitte d from this altered habitat (Crooks 1998, Grosholz 2002). However, larger bivalves and eelgrass appear to be inhi bited by the mat covering the mud flats (Crooks 1998). In additi on to altering habitat and shif ting species distributions and abundances, this Asian mussel also caused incr eased sediment and organic accumulations converting sand flats into mud flats (Crooks 1998).

Maybe the most well known exampl e of a ballast-mediated invasi ve species and its subsequent environmental impacts is the zebra mussel, Dreissena polymorpha , which established in Lake St. Clair from Europe and have spread to most fr eshwater systems in the eastern US (Benson 1995, Strayer 1999, Pimentel et al. 2004). Shortly after zebra muss els were introduced, their close relative the quagga mussel was in troduced via ballast water disc harge into Lake Michigan and have become very abundant and pose significant impacts to habitat there, even outcompeting zebra mussels in some areas (Ciborowski 2007). The quagga mussel occupies deeper regions of the Great Lakes and has not spread to ot her freshwater systems in the US.

Zebra and quagga mussels displace native species in freshwater habitat to the point of local and regional extinction. A baseline study of the De troit River found that 97 % of shells collected were from live, native unionid species, but this declined to only 10% a decade later after these invasive dreissenid mussels invaded the river (Schloesser et al. 1998). A similar decline from 84% to 3% occurred along nearby Lake St. Clair over the same time frame, virtually extirpating at least eight native mussel species (Schloesser et al. 1998). Nalepa et al. (1996) had sim ilar, although not as dramatic findings in reduced di versity of native mussel species. Michigan Technological University (2010) re ported that zebra mussel number s are so large that they can remove significant amounts of phytoplankton from the lake that fish once fed upon. Both

Dreissena species compete with native taxa for space and food.328 Zebra mussels compete with native mussels, clams, and even snails for space, reaching densities as high as 700,000 per square meter (Bens on 1995, Keniry 1995, Drake and Bossenbroek 2004,

Pimentel et al. 2004). Overall, habitats into which zebra mussels invade experience a decline in biodiversity (McNickle et al. 2001) and an increase of waste products and bacteria (Strayer 1999). Furthermore, when zebra mussels excrete, they are transferring nutrients from being pelagically available in the water column to the benthos and generally unavailable to fish species. Native Great Lake unionid species, af ter having zebra mussels grow on them, fail to open and close their shells and thus are unabl e to burrow, have impaired valve movement, deformed valve margins, and depleted prey (Strayer 1999).

Zebra mussels are also highly effective filter feeders and between that and their abundance compete with native species for available nutr ition in the water column (Strayer 1999, Martel et
al. 2001)(Roberts 1990, Ciborowski 2007, Falkner et al. 2009). As a consequence, nearly all native unionid mollusks and the amphipod, Diporeia hoyi , have disappeared from the four lower Great Lakes, Lake St. Clair, as well as the St. Lawrence River (Strayer 1999). Diporeia was a major component of lake whitefish diet, whic h is a primary component of many predatory salmonid diets including the economica lly important lake trout. But as Diporeia declined, the condition of lake whitefish and sculpin declined (McNickle et al. 2001, Pothaven et al. 2001) leading many researchers to suggest the four lo wer Great Lakes may be on the verge of trophic collapse (Nalepa et al. 2006, Riley et al. 2008, Pothoven et al. 2011).

Related to the zebra mussel’s ability to effectivel y filter feed is increased water clarity, altered benthic and water chemistry, as well as blue-gre en and benthic algae blooms and increases in gastropod populations (Strayer 1999, Janssen and Jude 2001, McNickle et al. 2001). Their feces, because there is so much of it, could lead to cyanobacterial blooms (producing toxins) in sediments, causing waters to become anaerobic leading to mass mortalities of resident fish populations and the upper trophic levels, with other tertiary effects possibl e within the ecosystem (Vanderploeg et al. 2001, Ciborowski 2007). These impact s have generally been observed throughout the Great Lakes region (Schloesser et al. 1998).

The Asiatic clam, Corbicula fluminea , significantly reduces native phytoplankton abundance and can be found in excess of 10,000 indivi duals per square meter, which likely reduces the fitness of

native filter feeders in the area due to competition for space and food resources, causing severe declines or local extirpations of native mussels in some areas Kennedy and Huekelem 1985,

Gardner et al. 1976, Ilarri et al. 2011, Fuller and Imlay, 1976; Clarke, 1986(Cohen et al. 1984,

Leff et al. 1990, Strayer 1999, Cordeiro et al. 2007). In addition to f iltering food from the water

colum n, young Asiatic clams also feed on nutrients in the sediment, potentially competing with benthic sphaeriids (Strayer 1999). When C. fluminea experiences die-offs, the amount of ammonia produced by decomposition may a dversely affect surrounding native taxa Cooper et al. (2005a, 2005b). Overall, Asiatic clams can fundamentally alter the ecosystems into which they invade (Ruiz et al. 1997).

The invasive green mussel Perna viridis may competitively displace the native scorched mussel

Brachidontes exustus through its greater growth rate a nd maximum size in Tampa Bay (Ranwell 1964). Mya arenaria , an invasive soft-shelled clam, may be the most common large clam in many embayments along the Pacific Northwest coast (Carlton 1999). Mytilus galtoprovincialis

,329 an invasive mussel, has displa ced the native mussel to become the dominant mussel throughout the Western Cape region of South Africa (Carlton 1999)

. The invasive Japanese mussel,

Musculista senhousia , has outcompeted numerous native spec ies of coralline algae, serpulid polychaetes, bryozoans, and hydroi ds to exclusively occupy larg e patches of intertidal and subtidal rocky habitat along New Zealand ((Carlton 1999).

Vegetation. Invasive plants can ca use widespread habitat alte ration, including native plant displacement, changes in benthi c and pelagic animal communities , altered sediment deposition, altered sediment characteristics, and shifts in ch emical processes such as nutrient cycling (Grout

et al. 1997, Wigand et al. 1997, Ruiz et al. 1999). Introduced s eaweeds alter habitat by colonizing previously unvegetate d areas, while algae form extensive mats that exclude most native taxa, dramatically reducing habitat complex ity and the ecosystem se rvices provided by it (Wallentinus and Nyberg 2007). Invasive algae can alter native habitats through a variety of impacts, including trapping sedime nt, reducing the number of suspe nded particles that reach the benthos for benthic suspension and deposit feeder s, reduce light availability, and adversely impact foraging for a variety of animals (Gribsholt and Kristensen 2002, Britton-Simmons 2004, Levi and Francour 2004, Sanchez et al. 2005). The green alga Caulerpa taxifolia can become so

dominant in Mediterranean areas as to form dense mats and ex clude other algal species during some seasons and alter the diversity of other species present, includ ing fishes such as red mullet, which could not forage in certain areas due to physical barriers created by the alga (Boudouresque 2002, Levi and Francour 2004, Wallentinus and Nyberg 2007). The dense aggregations of this alga impa ir suspended particles reaching the bottom, reducing available food for benthic feeders (Wallentinus and Nyberg 2007). Sargassum muticum and Gracilaria vermiculophylla , can form vegetative cover in areas that were previously lacking such cover, attracting fishes and other species that e nhance local biodivers ity (Nyberg 2007).

The complexity of an invaded system can also be altered due to changes in flow regimes and other factors (Wallentinus 2002, 2006b, a). Particular ly significant for juvenile salmon is that some of their invertebrate prey are adapted to take advantage of seasona l spikes in available detritus, which invasives like purp le loosestrife ca n disturb (Grout et al. 1997). Some invasive or nuisance alga species suppress the growth of native species via release of chemicals (Guerriero et
al. 1992, Friedlander et al. 1996, Hellio et al. 2004, Råberg et al. 2005, Nyberg 2007, Wallentinus and Nyberg 2007). C. taxifolia produces toxic metabolites that negatively influence other plants (Wallentinus and Nyberg 2007). Overall, C. taxifolia is viewed as highly im pactful

to habitats it invades due to its ability to reduce the overall biodiv ersity of native species in the area (Molnar et al. 2008).

A red alga, Gracilaria salicornia , that established in Hawaii has led to significant changes in

coral reefs there. G. salicornia reduced irradiance by 99%, in creased sedimentation, and changed diurnal dissolved oxygen and pH cycles on the surfaces it established on. Some species may benefit from this, as it creat es cover for several filter-feedi ng and detritivore organisms, but smothered benthic sessile organisms, notably the most common coral species in these habitats, as they form the basis for the continued existence of coral reefs that all other species in the habitat ultimately rely upon (Martinez 2012). Two other red algae, Acrothaninion preissii and Womersleyella setacea , form dense mats that can impede sedimentation (Wallentinus and330 Nyberg 2007). Another red alga in Hawaii, Hypnea musciformis , competes with native coral for space on coral reefs (Lafferty and Kuris 1996).

Many algal species also alter ecosystems by reducing the amount of light available for other plants. The invasive brown alga, Sargassum muticum , can shade large areas and reduce available light for native algal taxa (Britton-Simmons 2004, Sanchez et al. 2005, Wallentinus and Nyberg 2007). As a result, some kelp specie s in the San Juan Islands can be outcompeted (reductions of 50-75%) and green se a urchin abundance can decline significantly with the loss of this primary food, potentially de stabilizing kelp ecosystems th at entire biological communities are dependent upon (Britton-Simmons 2004) (Ambrose & Nelson 1982) . Invasive freshwater plants in Sweden have reduced available li ght, outcompeted native plants, altered water chemistry, and reduced biodiversity where they have invaded (Josefsson and Andersson 2011).

Once intentionally planted in Hawaii, mangroves spread from the island of Molokai throughout the main Hawaiian Islands (Allen 1998). An unintended consequence of the mangrove introduction was that they completely fill and sh ade the shallow pools they establish in, leading to these pools filling in much more quickly than they otherwise would (up to several centimeters of sediment annually), likely heavily impacting the algal communities that may occur there and altering (either positively or nega tively) dissolved oxygen levels. These changes can affect other aspects of the ecosystem including decreased bi odiversity, impeding the recovery of endangered waterbirds, and disproportionally providing habitat for invasive species over native taxa (Allen 1998).

Other invasive plant species ca n be beneficial for native ta xa in other ways (Posey 1988) 

### (Bulleri

et al. 2006 . Invasive marine plants can stabilize sediments with their roots, increase the oxidizing capacity of sediments, reduce wave di sturbances, improve wate r quality, and transform beaches into marshes (Posey 1988, Allen 1998, Ronnback 1999, Bruno 2000, Bruno and Kennedy 2000, Gribsholt and Kristens en 2002)(Cuddington and Hastings 2004)

. The invasive green alga Codium fragile promoted recruitment of native mussels (Bulleri et al. 2005).

The genus Spartina has invaded wetlands throughout the world, causing diverse impacts to its new habitats (Wallentinus and Nyberg 2007). Lower biodiversity has been associated with Spartina spp. introductions (Car penter 1987). In Britain, S. anglica has invaded eelgrass habitat, reducing the abundance of eelgrass and thus food availability fo r waterfowl (Daehler and Strong 1996). In addition, the introduction of a variety of Spartina marshgrass to England via ballast water discharge where a similar native species occurred resulted in a hybrid species, which is now the dominant species within its habitat (Thompson 1991). In the Netherlands, along Oregon, and in San Francisco and Humbolt Bays, this species as well as S. alterniflora

, S. densiflora , and S. patens have outcompeted native plan ts (Daehler and Strong 1996)(Chen et al. 2004). Because Spartina spp. can occupy habitat that native grasses cannot, it establishes in

mudflats that can increase avai lable habitat for benthic inverteb rates (Paulo da Cunha Lana 1991, Daehler and Strong 1996).

Spartina spp

. and Zostera japonica stabilize sediments in higher energy environments (Wallentinus and Nyberg 2007). As a result, Spartina spp. frequently transform higher-energy habitats like beaches into lower-energy e nvironments (marshes) (Carpenter 1987, Bruno 2000,331 Bruno and Kennedy 2000). Z. japonica displaces native eelgrass, speeds decomposition rates,

and limits nutrient release into the water colu mn along the US west coast (Wallentinus and Nyberg 2007). Since sediment is trapped to a greater degree within Spartina -dominated areas than many other grasses or mud flats, tidal fl ows can be decreased in these areas (Ranwell 1964, 1967, Christie 1974, Carpenter 1987, Daehler and St rong 1996). This shift from high energy environments to low energy environments has obvious implications for ecosystem change, which are not always detrimental to natives. For exampl e, more rooted plants can establish in lowerenergy environments (Bruno 2000, Bruno and Kennedy 2000, Wallentinus and Nyberg 2007).

Fish. Trophic cascades can result from the intro duction of invasive fishes (Power 1990, 1992,

Ruiz et al. 1999). Round gobies introduced to the Gulf of Gdansk in the Baltic Sea on or before

1987 via ballast water became the dominant shallow wa ter fish in at least the western portion of the Gulf within a decade and continue to spread through the region, using all available solid substrates (Sapota and Skora 1999, 2005). This may have been aided by the anthropogeniclyinduced loss of predators (Sapota and Skora 1999) . Round gobies have also established in the Great Lakes, where they compete for prey re sources with native species, prey upon native competitor eggs and young directly, and aggressively exclude native fishes from their individual

territories (Thomas 1997, French and Jude 2001, Janssen and Jude 2001, Carman et al. 2006, Bergstrom et al. 2008). Although the round goby has a specialized diet in its native range, it has shown the ability to shift to be a general pr edator in some invasi ve populations, including foraging on invasive zebra mussels in the Great Lakes (Thomas 1997, Carman et al. 2006,

Lederer et al. 2006). By outcompeting native species, there is a shift of nutrien t availability from the pelagic zone to the littoral zone.

The sea lamprey ( Petromyzon marinus ) was introduced into the Great Lakes and has seriously altered the abundance of lake trout and lake whitefish(Janssen and Jude 2001). The Eurasian

ruffe was released into the Great Lakes via ba llast water discharge and has become a major predator of native benthic invertebrates and comp etes for available prey and habitat with native taxa, such as yellow perch and preys upon th e eggs of lake herring (Lovell and Stone 2005)(Savino and Kolar 1996, Selgeby 1998, Fullerton et al. 2000, Drake 2005)

. Invasive fishes can compose a large portion of fish taxa in at least some areas, including new Zealand where 53% of fish taxa are exotic, Puerto Rico where invasive fish are 91% of the total species, and Brazil where they are 13% of the total (Lövei 1997).

Habitat Effects. Invasive species have been referred to as one of the top four threats to the world’s oceans consistently ranked behind hab itat degradation and alteration (Raaymakers and Hilliard 2002, Raaymakers 2003, Terdalkar et al. 2005, Wambiji et al. 2007) (Pughiuc 2010). In most cases, habitat is directly a ffected by human alterations, as identified in the baseline section, such as hydromodification, mining, dredging, dril ling, and construction. However, invasive species, facilitated by human comme rce, have the ability to dir ectly alter ecosystems upon which listed species rely.

The spiny water flea, Bythotrephes longimanus , is another example of an invasive species that causes extensive ecosystem disruption (Grout et al. 1997). T his invader, first introduced by ballast water from northern Europe to Lake On tario in the early 1980s may eventually occupy thousands of lakes throughout North America (Johannsson et al. 1991, Grout et al. 1997, Kerfoot332 et al. 2011). Bythotrephes is an important contribut or to its native habitat, including as prey to salmon; however, in the Great Lakes, they redu ce the fitness of many fish that are prey to salmonids (Hessen et al. 2011). Bythotrephes preys heavily upon plankton species, severely reducing not only their abundance, but has also ca used their diversity to decline by roughly 20% (Foster and Sprules 2009, Kerfoot et al. 2011)(Rennie et al. 2011). As a result, rotifers decline

because of reduced diatom food resour ces and phytoplankton increase because Bythotrephes feeds on their competitors (Kerfoot et al. 2011) (Hovius et al. 2006; Strecker et al. 2011). Further tertiary effects include elevation of contaminant levels in higher-level predators due to

extensions in the food web that allow for a dditional contaminants to accumulate in the underlying prey base (Kerfoot et al. 2011, Rennie et al. 2011). Other macroinvertebrate predators and fishes are also like ly adversely impacted by this disr uption of their prey base, with less prey available to them (Parker Stetter et al. 2005, Foster and Sprules 2009). These alterations to ecosystem food webs app ear to be stable and persistent (Yan et al. 2008). Through these mechanisms, Bythotrephes alone represents a significant th reat to the biodiversity within temperate North American aquatic environments (Grout et al. 1997).

Dreissena mussels introduced to the Baltic Sea alte red sediment composition from soft bottom into shell gravel and even hard substrate due to the abundance of their shells in the benthos (Olenin and Leppäkoski 1999). Furthermore, Dreissena and other benthic invaders remove

nutrient resources from the water column and transf er them to the sediment in their excrement. In many cases, this decreases the nutrient outfl ow from some embayments and causes an accumulation of organic material, leading to incr eased eutrophication and fundamentally altering the ecological function of the habitat. Sometimes increasing the nutrients available in an area works to the advantage of predatory fishes for which additional prey become available (Olenin and Leppäkoski 1999).

An invasive polychaeate, Ficopomatus enigmaticus , established burrows in an Argentinean lagoon altering water flow and sedimentation (W allentinus and Nyberg 2007). However, as above, this habitat modification is not always to the detriment of native taxa. The ascidian, Pyura praeputialis , invaded coastal waters along Chile and formed dense matrices in the intertidal zone, providing habitat that was exte nsively used by native invertebrates and having double the species richness in these habitats than surrounding unalte red habitat (Castilla et al. 2004, Wallentinus and Nyberg 2007).

Other invertebrates can also have major impact s on the ecosystems they invade. The invasive bryozoan Membranipora membranacea established on kelp fronds in the Gulf of Maine, m aking them brittle and inflexible and causing them to drop off prematurely (Lambert et al. 1992). The invasive polychaete, Marenzelleria viridis , was introduced to coastal Baltic Sea lagoons in the mid 1980s as the first polychaete in the ecosystem . Within five years, the abundance of this species resulted in an increase in the bioma ss present within the benthos by an order of magnitude (Zmudzinski 1996, Zettler 1997, Olenin and Leppäkoski 1999, Leppaekoski and Olenin 2001, Panov et al. 2003). This species digs holes in the benthos that are much deeper than the native polycheate (deeper areas), thereby introducing much more oxygen into the benthos and increasing the denitr ification process causing change s to the native habitat (Panov et
al. 2003, Wallentinus and Nyberg 2007). These changes have a dversely affected the prey species for native predators.333 Asian crabs ( Hemigraspus sanguineus ) compete for space (burrows) with native fiddler crabs along Connecticut (Wallentinus a nd Nyberg 2007). Mitten crabs form extensive burrows that can enhance mud bank erosion (Rudnick et al. 2000, Chinese Mitten Crab Working Group 2003). The Asian copepod Pseudodiaptomus inopinus has become the dominant

macroinvertebrate in some areas of the Pacifi c Northwest after establ ishing there via ballast water discharge (Kipp et al. 2010). Oriental shrimp ( Palaemon macrodactylus ) were likely

introduced into New York State waters via ball ast water and may act as competitors of other species within the ecosystem, although no ecosystem damage has been documented by this invasive species here so far (Warkentine and Rachlin 2010).

Echinoids (sea urchins) cause si gnificant habitat alteration where they invade (Himmelman and Steele 1971, Dean et al. 1983, Kitching and Thain 1983, Haupt et al. 2010), becoming the dominant benthic grazers. They feed heavily upon kelp beds, diminishing their abundance and the habitat they offer for entire marine communities.

The introduced periwinkle, Littorina littorea , ranging along the Atlantic Coast from Canada to the mid-Atlantic, is highly-influential in th e sedimentation process; because individuals cumulatively engage in so much grazing, some bottom habitats have become dominated by hardbottom instead of soft bottom as they former ly were (Bertness 1984, Wa llentinus and Nyberg 2007)(Carlton 1999). Significant declines in soft-sed iment habitats and fringing salt marshes are attributed at least partially to the invasion of this species, possi bly due to consumption of marsh grasses, such as S. alterniflora (Bertness 1984). Species normally adapted to living in softbottom systems are gradually replaced by species better adapted for hard-bottom substrates.

Case Studies of Invasive Species Impacts to Habitat

San Francisco Bay As a result of the large numbers of introduc tions into San Francisco Bay, much of the biologically-available energy that was available to native taxa is now being exploited by invasive species (Cohen and Carleton 1995). With at l east 234 established non-native populations in San Francisco Bay, there are many well studied adverse effects to the area. However, the vast majority of the invasive species appear to have benign effects on fi sh species, such as salmonids, and habitat generally.

Direct Effects to Listed Species. The invasive Atlantic ribbed marsh mussel Arcuatula demissa has contributed significantly to the decline of loca l clapper rails by trapping the bird’s beak in its

shell if the rail inserts it into the open shell; rails ca nnot pull their beaks out and die of starvation (Cohen and Carleton 1995). Adult ra ils also lose toes by the musse l closing on them as the rails walk through the marsh, although the result fo r young rails is likely mortality (Moffitt 1941, Josselyn 1983, Takekawa 1993)(Cohen and Carleton 1995). However, the invasive mussel also appears to be prey of the rail to a significan t extent (Moffitt 1941)(Cohen and Carleton 1995).

Introduced fishes have also greatly affected the San Francisco Bay ecosystem. Striped bass, largemouth bass, smallmouth bass, bluegill, and gr een sunfish are all introduced species to the area, although largely through m eans other than ballast water discharge (Cohen and Carleton334 1995). Cohen and Carleton (1995) documented that these fishes have led to the extinction of four native fish species not only in the Bay, but throughout their range ei ther directly through predation or indirectly through co mpetition for prey and/or breedi ng sites. These predatory fish also impact listed salmonids. And ironically, striped bass themselves are impacted by invasive species, with juvenile abundance declining in association with d eclines in their primary prey species, mysid shrimp, likely due to effects caused by other introduced species in the Bay (Nobriga and Feyrer 2008).

American bullfrogs, along with some of the fi sh species above, have predated heavily upon native red-legged and yellow-legged frogs in the Bay and Delta region (Cohen and Carleton 1995). Introduced silverside minnows have also caused the decline in de lta smelt via predation on the smelt’s eggs and larvae (Cohen and Carleton 1995).

Indirect Effects to Listed Species. The Asian and soft shell clams, Potamocorbula amurensis

and Mya arenaria , respectively, are now highly influentia l within the bay system, occurring in such high abundance so as to transfer the phyt oplankton biomass in the water column to the benthos that is then exploite d by benthic invertebrates and bo ttom-feeding fishes, instead of pelagic species (Alpine and Cloern 1 992, Cohen and Carleton 1995, Grosholz 2002). Mya spp. alone is estimated to be capable of filteri ng all phytoplankton in the bay daily (Cohen and Carleton 1995). Like the Asian and soft shell clam, in the southern portion of San Francisco Bay, introduced clams of the genera Gemma

, Venerupis , and Musculista control phytoplankton growth with their filtering capaci ties (Cohen and Carleton 1995).

Overbite clam ( P. amurensis ) abundance is high enough to reduce the availability of plankton in the Sacramento-San Joaquin River Delta (C ohen and Carleton 1995). The Sacramento–San Joaquin River Delta is estimated to have lost 43% of its annual primary productivity between 1975 and 1995, partly as a result of this invasive species (Jassby et al. 2002, Feyrer et al. 2003,

Falkner et al. 2009), leading to the decline of native fi sh species. Alpine and Cloern (1992) provided an even larger estimated decline from 106 g C/m 2 to 39 g C/m 2

.

The introduced snail, Batillaria attramentaria , is slowly out-competing its native counterpart, Cerithidea californica . The invasive Atlantic mudsnail Ilyanassa spp. has displaced Certhidea snails from mudflats to saltmarsh habitats, with an associated decline in the native species’ abundance (Cohen and Carleton 1995). Several sp ecies including the Asiatic clam, polyzoan, and hydroid ( Corbicula

, Urnatella , and Cordylophora spp., respectively) have led to increased sedimentation in canals, reducing flow rates a nd probably flushing within estuaries (Cohen and Carleton 1995). Native crayfish have been displaced by introduced varieties (Cohen and Carleton 1995). While these are all examples of shifts in species abundance, it is unclear if these shifts have any indirect effect to listed fish in San Francisco Bay.

Direct Effects to Habitat. The introduction of invasive species almost always has an unintended impact to the ecosystem. The direct a nd indirect effects to listed species described above impact the ecosystem to some extent. Ad ditionally, there are specie s that may not impact other species in the area, but th ey may alter habitat in such a way that the introduction make some species less fit. The Australian-New Zealand boring isopod Sphaeroma quoyanum creates
small holes in mud bank habitat. Since its intr oduction, this small animal has altered mud banks335 throughout San Francisco Bay with these bore holes. Sphaeroma ’s holes are now considered the primary mechanism of bank erosion, causing banks to recede by several meters in some locations

since its introduction ov er a century ago (Cohen and Carleton 1995).

Several invasive plants also impact Sa n Francisco Bay’s ecosystem. Peppergrass ( Lepidium latifolium ) may act to displace rare native marsh plants, such as the mudflat quillplant ( Lillaeopsis masoni ) (Cohen and Carleton 1995). The Atlantic cordgrass, Spartina alterniflora is invasive throughout many regions of the world and, in San Francisco Bay, may benefit some species like clapper rails, which occur in their highest densities in native cordgrass habitat. Expansion of cordgrass habitat may provide mo re protective habitat fr om predators, although competition between the native and invasive cordgr ass species may well lead to the decline of the native cordgrass (Cohen and Carleton 1995). Some properties of Atlantic cordgrass lend themselves to providing less habitat for native wetland animals and infauna (Cohen and Carleton 1995). The invasive Spartina may also alter sedimentation ra tes, reduce light levels by its increased canopy density, and loss of shorebird feeding habitat through col onization of mudflats (Cohen and Carleton 1995). Spartina alteriflora has hybridized with the native S. foliosa in the Bay (Daehler and Strong 1996, Anttila et al. 1998).

Chesapeake Bay Like San Francisco Bay, there are many intr oduced species in the Chesapeake Bay. A comprehensive review of the impacts of invasi ve species to the Chesapeake Bay was conducted by Ruiz et al. (1999). With at least 196 established non-native populations in the Chesapeake Bay, it is surprising that most of the impacts of invasive species on the Chesapeake Bay are generally undocumented. The authors found that 20% of the 196 documented invasive species had significant ecological impacts, while most of the other invasi ve species had not been studied for their impacts. Of the 39 species with si gnificant ecological impacts, 69% did so through competition with native species, 38% altered habita t, 44% served as prey, 15% were predators of native species, 21% engaged in ex tensive herbivory, 8% produced hybrids with native taxa, and 8% were parasitic (Ruiz et al. 1999). Plants and fish were the largest ta xonomic groups represented in the known invasive species of the Chesapeake Bay, representing 23% and 18% of the invasive species by taxa, respectively.

Direct Effects to Listed Species. There is limited information on direct effects to listed species. The only listed species under NMFS jurisdiction in the Chesapeake Bay are sturgeon and sea turtles. Juvenile sturgeon could be prey to introduced piscivorous fish. There are several introduced fish species (including smallmouth bass, and sunfishes) blamed for declines in native taxa (Ruiz et al. 1999) and while undocumented, may impact sturgeon.

Indirect Effects to Listed Species. In the case of the Chesapeake Bay, while the invasive species have not been well studie d, it appears the best documentati on of effects may be indirect to sturgeon or sea turtles via alteration of food web dynamics and food availability. Two protistan pathogens, Haplosporidium nelsoni and Perkinsus m arinus , are significant contributors to a 90% reduction in oyster abundance in the Chesapeake Bay over the past century, causing secondary effects such as reduced oyster reef habitat and alte red food webs (Ruiz et al. 1999). The rapa whelk is now an abundant predator of native clams and oysters in the Bay (Deacutis and Ribb 2002) with similar ecological impacts to the protest pathogens. Mud crabs have also336 declined as a result of the invasive parasitic barnacles, Loxothylacus panopaei , which causes reproductive failure in the host (Van Engel et al. 1966, Hines et al. 1997, Ruiz et al. 1999).

The Asiatic clam ( Corbicula fluminea ) is so abundant in the Potomac River that it is estimated this species alone can filter the total phytoplankton biomass in three to four days and can

constitute 90% or more of the bivalve biomass in some Bay areas. Such efficient conversion of energy from the pelagic to the benthic envi ronment likely benefits shortnose sturgeon by increasing worms and chironomids, two of their prey items. As a result of this invasion, between 1981 and 1993, water clarity tripled, subseque ntly increasing aquatic vegetation 50%, and ultimately increasing abundance of fish populatio ns, slowing currents, increasing sedimentation, as well as altering benthic community composition and sediment characteristics through its large production of pseudofeces (Cohen et al. 1984, Phelps 1994, Ruiz et al. 1999).

Direct Effects to Habitat. Direct and indirect effects to li sted species also generally have impacts to habitat. For instance, the descripti on of indirect effects cau sed by the Asiatic clam also have impacts to habitat in the Chesapeake Bay. The reed, Phragmites auatralis , also outcompetes local plants and ha s become widespread and dominant within the Chesapeake Bay, altering habitat parameters and animal abundances (Marks et al. 1994, Ruiz et al. 1999). Typha angustifolia has similar impacts, outcompeting local sp ecies, reducing flow rates, increasing sedimentation, and altering sediment chemistry (Ruiz et al. 1999).

Two invasive aquatic plants, Hydrilla verticillata and Myriophyllum spicatum , have received significant attention in the Chesapeake Bay. They form dense mats, alter aquatic chemical and habitat characteristics, fish and invertebrate communities, compete with native plants, and change the food base available for local waterfowl and fishes (Ruiz et al. 1999). Also noteworthy is that the cover provided by Hydrilla spp. provides additional refuge for smaller fishes, which can increase the populations of larger predatory species (Killgore et al. 1989, Ruiz

et al. 1999). Trapa natans, a floating plant, at one time also outcompeted native plant species to the detriment of fishes and waterfowl, but has no t recovered from an eradication program in the 1930s (Ruiz et al. 1999).

As has been documented in other regions, there appears to be a correl ation between habitat disturbance and the susceptib ility to invasions. Jewett et al. (2005) experimentally found that low dissolved oxygen levels (common in the Ch esapeake Bay) resulted in invasive species establishing in a new area, to the detriment of dominant native taxa prior to low dissolved oxygen levels. Many researchers have also documented the eutrophic conditions of the Chesapeake Bay influence the propensity of species to invade. This situat ion tends to lead to low light lev els, which make some invasive species, such as th e submerged and emergent aquatic plants outlined above more competitive with native varieties (Barko and Smart 1981, Grace and Harrison 1986, Marks et al. 1994, Ruiz et al. 1999).

Puget Sound Although we searched for information addressing the impacts of invasive species on Puget Sound, we did not find any significant informa tion documenting biologica l, physical, chemical, or ecological responses of species or hab itats in this area to invasive species.337 Integration and Synthesis

Consequences of Specific Proposed Actions under the Program

Direct effects to listed species. Little to no information is ava ilable supporting direct effects that ballast-mediated invasive species have on species listed under NMFS’ jurisdiction. The only direct connection we found supported a benefici al connection of inva sive algae providing a significant new food source for Hawaiian green s ea turtles (Russell and Balazs 2009). By assessing the adverse effects of di sease, predation, and competition to similar species or similar functional groups, we were able to conclude the likely affects of new ballast-mediated invasive species to NMFS trust resources.

It is reasonable to expect that ballast-mediated invasive diseases or para sites can have significant and catastrophic effects, such as that which ha s occurred to Atlantic salmon in Norway as a

result of an invasive fluke be ing introduced, white pox disease in Acroporid corals, withering syndrome in California abalone, and harmful alga l blooms that poison resident marine mammals, sea turtles, and fishes (Hallegraeff and Bolch 1992, McMinn et al. 1997, Anguiano-Beltrán et al. 1998, Hallegraeff 1998, Friedman et al. 2000, Hamer et al. 2000, Hamer et al. 2001, Ben-Haim and Rosenberg 2002, Lilly et al. 2002, Smith et al. 2003, Aguirremacedo et al. 2008, Bower 2009, Cohen 2010)(Hallegraeff et al. 1995)(Josefsson and Andersson 2011). When ballastmediated invasive species establish in an area occupied in the same time and place as listed species and the result is a disease or parasite, we expect a range of effects from beneficial to harm of an unpredictable number of listed individuals.

Predation on early life stages (particularly egg or larval stages) of listed salmonids, sturgeon, rockfish, eulachon, corals, and abalone is by numerous examples of zooplankton-feeding invasive species such as musse ls, silverside minnows, and numer ous piscivorous fish has been documented and is a likely consequence of future invasions. When ballast-mediated invasive species establish in an area occupied in the same time and place as listed species and the result is predation, we expect a range of effects from bene ficial to harm of an unpredictable number of

listed individuals.

Competition is another means by which invasive sp ecies are likely to directly affect listed species. Particularly susceptible to direct e ffects of competition are the listed invertebrates (abalone and corals) because f ilter feeding and photosynthetic organisms are more likely to become established than larger predatory fish species. Invasive filter feeders along the California coast would compete with abalone for both energy and space, while an invasive autotroph like an alga could grow over coral, out-competing corals for sunlight. While less likely, benthivores and piscivores have established in the past and that is likely to happen again in the future. The response to such ballast -mediated invasions would be competition with sturgeon, sawfish, or salmonid species. When balla st-mediated invasive species establish in an area occupied in the same time and place as list ed species and the result is competition, we expect a range of effects from beneficial to harm of an unpredictable number of listed

individuals.338 Effects to designated critical habitat. Our critical habitat analysis determines whether the consequences of the proposed action will destr oy or adversely modify critical habitat for ESAlisted species by examining any ch ange in the conservation value of the essential features of critical habitat. This analysis does not rely on the regulatory definition of “adverse modification

or destruction” of critical hab itat. Instead, this analysis focu ses on statutory provisions of the ESA, including those in Section 3 that define “critical habita t” and “conservation,” those in

Section 4 that describe the de signation process, and those in Section 7 setting forth the substantive protections and proced ural aspects of consultation.

The Response analysis documents indirect effects to list ed species via habitat and ecosystem alteration as the major mechanism by which ball ast-mediated invasive species can adversely affect listed species. While this is not the result of every i nvasion, it is a probable outcome over time. In many cases, this also constitutes adverse affects to cr itical habitat (specifically of eulachon, salmonids, corals, abalone, Cook Inlet be luga and southern resident killer whales, green and hawksbill sea turtles, and sturgeon) through impairment of PCEs associated with quantity and quality of available prey or suitability of habitat for growth.

The PCEs for each listed species, where they ha ve been designated, are described in the Status of

Listed Resources section of this Opinion. The PCEs identify those physical or biological

features that are essent ial to the conservation of the species that may require special management considerations or protections. The species addressed in this Op inion have similar life history characteristics and therefore, many of the same PCEs.

The PCEs for gulf and green sturgeon focus on life history stages, re aring, migration, and spawning and contain physical and biological features essential to the conservation of the species or DPS. The PCEs for gulf and green sturgeon are:

1. Abundant food resources for all larval, juvenile, sub-adult, an d adult life stages; 2. Water flow necessary for normal behavior , growth, and survival of all life stages; 3. Water quality, including temperature, sa linity, oxygen content, and other chemical characteristics, necessary for normal behavior , growth, and viability of all life stages; 4. Amigratory pathway necessary for the safe and timely passage of Southern DPS fish

within riverine habitats and between riverine and estuarine habitats; 5. Water depth in holding pools over five me ters deep for both upstream and downstream holding of adult or sub-adult fish, with ade quate water quality and flow to maintain the physiological needs of the holding adult or sub-adult fish; and 6. Sediment quality ( i.e., chemical characteristics) necessa ry for normal behavior, growth, and viability of all life stages.

These PCEs for salmonids and Pacific eulachon in clude sites essential to support one or more life stages (sites for spawning, rearing, migr ation and foraging) a nd contain physical or biological features essentia l to the conservation of the ESU/DPS, such as:

1. Freshwater spawning sites w ith water quantity and quality conditions and substrate supporting spawning, incubation and larval developm ent;339 2. Freshwater rearing sites with water quanti ty and floodplain connectivity to form and maintain physical habitat conditions and support juvenile growth and mobility; water quality and forage supporting juvenile deve lopment; and natural cover such as shade, submerged and overhanging large wood, log jams and beaver dams, aquatic vegetation, large rocks and boulders, side channels, and undercut banks;

3. Freshwater migration corridors free of obstr uction, along with water quantity and quality

conditions and natural cover such as s ubmerged and overhanging large wood, aquatic vegetation, large rocks and boulders, side channels, and undercut banks supporting juvenile and adult mobility and survival;

4. Estuarine areas free of obstruction, along with water quality, water quantity, and salinity conditions supporting juvenile and adult phys iological transitions between fresh and saltwater; natural cover such as submerged and overhanging large wood, aquatic vegetation, large rocks and boulders, and side channels; and juvenile and adult forage, including aquatic invertebrates and fish es, supporting growth and maturation;

5. Nearshore marine areas free of obstruction w ith water quality and quantity conditions and forage, including aquatic invertebrates and fishes, supporting growth and maturation; and

natural cover such as subm erged and overhanging large wo od, aquatic vegetation, large rocks and boulders, and side channels; and

6. Offshore marine areas with water quality conditions and forage, including aquatic invertebrates and fishes, suppor ting growth and maturation.

These PCEs for southern resident killer whales include sites essential to s upport one or more life stages and contain physical or biol ogical features essential to th e conservation of the DPS, such

as:

1. Water quality to support growth and development;

2. Prey species of sufficient qua ntity, quality, and availability to support individual growth, reproduction, and development, as well as overall populat ion growth; and

3. Passage conditions to allow for migration, resting, and foraging.

The PCE for leatherback sea turtles essential to support population fitne ss and the conservation of the DPS, is the occurre nce of prey species, primarily scyphomedusae of the order Semaeostomeae (Chrysaora, Aurelia, Phacelloph ora, and Cyanea), of sufficient condition, distribution, diversity, abundance and density necessary to su pport individual as well as population growth, reproduction, and development of leatherbacks.

These PCEs for Cook Inlet beluga whales are essential to their conservation:

1. Intertidal and subtidal waters of Cook Inlet with depths <30 feet MLLW and within 5 miles of high and medium flow anadromous fish streams;

2. Primary prey species consisting of four sp ecies of Pacific salm on (Chinook, sockeye,

chum , and coho), Pacific eulachon, Pacifi c cod, walleye pollock, saffron cod, and yellowfin sole;

3. Waters free of toxins or othe r agents of a type and amount harmful to Cook Inlet beluga whales;

4. Unrestricted passage within or betw een the critical habitat areas; and340 5. Waters with in-water noise below levels resu lting in the abandonment of critical habitat areas by Cook Inlet beluga whales.

These PCEs for Hawaiian monk seal are essent ial to ensuring the health, well-being, and continued viability of their populations:

1. Pupping and major hauling beaches includi ng the vegetation backing the beaches;

2. Shallow protected water adjacent to the above;

3. Deeper inner reef areas and lagoon waters;

4. Other waters surrounding the northwest Hawaii an Islands to at least 80 fathoms; and

5. Banks and shoals without emergent lands and pelagic waters.

We can draw several generalizations from the Response analysis that summarize the effects that ballast-mediated invasive species have on their ne w habitat. Benthic in vertebrates, such as mussels, polychaetes, and hydroids become dominant filter feeders, greatly reducing the amount of organic energy that is availabl e to native taxa in the water co lumn. This transfer of energy

from the water column into the benthos fundame ntally alters the ecol ogy of the host habitat, resulting in less food/prey available for other filter feeders. At the same time, this may increase resource availability for stur geon species. Adverse effects of this include reduced body condition, growth, survival, and/ or reproduction of native pelagi c organisms at the same or similar trophic level as the invader if the native competitor cannot adapt to another food source. These changes would be manifested to a greate r or lesser degree up the food chain to higher trophic level organisms in the habitat, including listed salmonids and resident, coastal marine

mammals and sea turtles.

In addition to extensive filtration, invasive bent hic invertebrates also ca use significant physical and chemical changes within their new habitats . As an invasive population grows, it also produces large amounts of feces or pseudofaeces. This provides abundant food sources for bacteria, which supports population e xplosions of bacteria that expl oit this resource. This often leads to anoxia or hypoxia, resulting in wide-scale mortality of benthic inhabitants. Invasive mussel and oyster populations have also slowed flow rates thr ough shallow water habitats and juvenile salmonid rearing areas, leading to increase d sedimentation that fills in estuarine habitat.

Introduced diseases and parasites can be expected to impact not only listed species directly, but also species in the food chain on which they rely. Numerous studies identified in the Response analysis show how both fishes and invertebrates ar e susceptible to intr oduced parasites and diseases, sometimes with devastating levels of mo rtality as a result. Even moderate levels of morbidity or mortality within a prey or prey base species can lead to less or lesser quality prey available for listed fishes, coastal marine mammals and sea turtles that can in turn lead to reduced body condition, growth, su rvival, and reproductive resour ces for listed individuals.

Invasive plants can alter their host environments in a number of ways to the detriment of listed species. Harmful algal blooms resulting for in troduced dinoflagellates can not only produce toxins that adversely affect list ed individuals directly, but also produce massive die-offs of fish and invertebrate species resu lting in less available prey. Estuarine plants, such as Spartina spp. and m angroves can rapidly alter the physical en vironment of an estuary, reducing wave energy,341 increasing sedimentation, and leading to estuaries f illing in that listed fishes rely upon. Plants, such as hyacinth and C. taxifolia , can clog waterways also leading to increased sedimentation or outright impairment of fish movement where they have established. Invasive algae on the ocean bottom frequently form a barrier to sedimentatio n or particle movement into the water column, reducing habitat complexity and, ultimately, th e species that can survive within a habitat (potentially prey species of listed organisms or the listed species themselves).

Future Invasive Species Impacts to Li sted Species and Critical Habitat We have established that, regardless of the m echanism by which invasive species have been introduced, they can have a wide range of effect s on the environments in which they establish. More importantly, ballast-mediated invasive spec ies have a similar affect on the environment and the species within the environment as do all i nvasive species generally. It is possible to anticipate the probability of a ballast-mediated invasive species establishing at various ports around the US, but it is impossible to anticipating the specific species that will be established at a specific port and therefore which listed species may be impacted by that specific species. For that reason, NMFS conducted a general assessment of invasion risk from ballast water to determine the probable exposure risk to each of our listed species. The next step is determining the probable response to those exposures.

Some introductions can be bene ficial, providing food for listed s ea turtles (Russell and Balazs 2009), increasing biodiversity (N yberg 2007), and establishing new habitat used by native species (Paulo da Cunha Lana 1991, Daeh ler and Strong 1996, Crooks 1998, Grosholz 2002,

Castilla et al. 2004, Wallentinus and Nyberg 2007). We have also found instances where the

establishment of an invasive has had catastr ophic effects on native species and ecosystems, causing high mortality or extinction (Vinogradov et al. 1989, Chew 1990, Tsikhon-Lukanina and Reznichenko 1991, Ciguarria and Elston 1997, Zeitsev and Ozturk 2001, Minchin and Rosenthal 2002, Moncheva and Kamburska 2002, Nell 2002, Lleonart et al. 2003, Nehring 2006b, Haupt et
al. 2010, Josefsson and Andersson 2011) or funda mentally altering an ecosystem through a single introduction (Colwell 1996, Castilla et al. 2004, Hecky et al. 2004, Wallentinus and Nyberg 2007). Far more often, an invasive speci es usually suppresses an ecologically similar native species low in the food web and the impacts of that invasion have a negligible effect to listed species or their critical habitat.

A thorough review of the literature reveals that just because an inva sive species establishes, it is not appropriate to assume there will be an adve rse effect to the receiving environment or the

listed resources within it. For th is reason, we determined what th e probability would be that an invasive species establishing in US waters would adversely affect the environment into which it

is introduced. We identified three different source s of information to evalua te this question: (1) evaluations of the impacts of invasive spec ies in terrestrial and freshwater environments written by the US Congress Office of Technology and Assessment (Anttila et al. 1998), (2) impacts of invasive species in the Chesapeake Bay published by Ruiz et al. (Ruiz et al. 1999), (3) as well as im pacts of invasive species in San Francisco Bay (McNickle et al. 2001).

While none of these studies identified only ballast-mediat ed invasive species, the Response Analysis in this Opinion shows that ballast-mediated invasive species elicit similar responses as other invasive species. Therefor e, NMFS expects that any invasive species that are introduced as342 a consequence of the USCG’s Ballast Water Management Program will have impacts similar to other introduced species with effects ra nging from beneficial to harmful.

US Technology Office of Technology and Assessment. This report (Anttila et al. 1998) summarized information on non-indigenous species (p rimarily terrestrial a nd freshwater) in the United States, as well as what proportion of them were ecologically or economically harmful, beneficial, neutral, or unknown. This non-exha ustive but reasonably comprehensive report used a panel of experts to identify non-indigenous species and the impa cts they caused (or failed to cause). In total, 15.3% of non-indigenous species caused “great harm” defined as “significant and difficult-to-control pests of agriculture, range lands, or forests; seri ously foul waterways, irrigation systems, and power plants; cause wide -scale disruption of indigenous ecosystems; or threaten indigenous species with extinction.” We believe that all of th ese categories at least represent species that are competitors, predators, diseases, or habitat modifiers of native species or habitats and would be impactful to the e nvironments into which they were introduced. Overall, 16% were considered harmful, 33% had unknown effects, 26% were considered beneficial, and 25% had both bene ficial and harmful effects. The survey included purposefully introduced species of agriculture, hunting, and co mmercial use, which should upwardly bias the

data in the beneficial and bene ficial and harmful categories. This is supported by one of the report’s findings that only 12% of intentional in troductions were considered harmful versus 44%

of unintentional introductions. Because this assessment includes terrestrial invasions and intentional introductions, NMFS believes the probabil ity of harm identified in this assessment is

the least accurate of the three.

Chesapeake Bay. A comprehensive review of invasive, cryptogenic, and impactful invasive and possibly invasive species was conduc ted in the Chesapeake Bay by Ruiz et al. (Ruiz et al. 1999). The authors compiled information on 196 species that were either confirmed as invasive for were cryptogenic (unclear whether they were invasive or native) over the past four centuries. Only cryptogenic species for which significant informa tion existed to support a strong possibility of being invasive were included in these 196 wh ile hundreds of others were excluded. The definition of whether a species was impactful wa s based exclusively upon available literature that qualitatively or quantitatively ascribed a pa rticular species as being impactful. Additional information on impact for some of these specie s was available outside of the Chesapeake Bay, but the authors excluded this information from these non-Chesapeake Bay locations because it was possible this information may not apply in th e Bay itself. Impacts were assigned to nine categories: competition, habitat change, food- prey, predation, herbivory, hybridization, parasitism, toxicity, or biotur bation which caused “significant a nd measurable changes in the abundance or distribution of resident species or habitats.” This de finition, unfortunately,

excluded some effects that we would consider significant, such as hybr idization or introduced predators for which information was lacking on population-level effects to native prey; this likely causes a slight negative bias in the numbe r of invasives we would consider having an adverse affect. We also expect that the total number of invasives is ac tually higher than the 196 included, as an unknown subset of those that we re excluded are likely actually non-native, causing a positive bias in the estimate. However, we have no mechanism(s) to account for these biases and expect that these caveats do not alter th e results so much as to make the estimates of invasive species impacting the Chesapeake Bay unreasonable.343 Overall, Ruiz et al. (Ruiz et al. 1999) found that 39 of 196 (19.9 %) invasive or cryptogenic species impacted the ecology of the Chesapeake Bay. Of the remaining species, 51% lacked information on impact and 29% may have an imp act but significant information was lacking. Cryptogenic species represented six (15%) of im pactful invasives. Ba sed upon this, cryptogenic species compose a relatively small proportion of im pactful invasives, but the actual number of impactful invasives is likely significantly higher than the 20% reported here, as a small majority

of invasives lacked information to assess thei r impacts. The ecological categories in which impacts occurred were: competition—69%; habi tat change—38%; food/prey—44%; predation— 15%; herbivory—21%; hybrid ization—8%; parasitism—8%; toxi city—0%; and bioturbation— 3%. These proportions of imp act types roughly correspond with the types of impacts we found globally for invasive species presently in the Response analysis

.

San Francisco Bay. We also examined available information on the number of invasive species

that have established in San Fr ancisco Bay/Delta and the number of these that we found in our information searches that we found to impact th e area. The number of invasives, was adopted from Cohen and Carlton (1998). These authors assembled data from the 16 th century to the time of publication from published, unpubl ished, and their personal resear ch on the introduced aquatic organisms of San Francisco Bay/Delta, includi ng some cryptogenic species. The authors identified 234 exotic species as well as 125 crypt ogenic species. We expect that some of the cryptogenic species identified by Co hen and Carlton (1998) are actu ally exotic, but we are not aware of any methodology to further characterize the origin of these species that Cohen and Carlton (1998) have not already done. This proportion may be low, as a 2010 survey in San Francisco Bay identified 92 cryptogenic, of which fi ve were considered exo tic and the rest were of unclear origin (Foss 2011). Therefore, in the present analysis, we exclude the 125 cryptogenic species as exotics and accept 234 ex otic species as the number of invasive species present. This is roughly corroborated by Foss et al. (2007). We expect that this may positively bias our results to some degree, but should be reasonably repr esentative of the number of invasive species present. Furthermore, this assumption would lead to an overestimation of the proportion of invasive species adversely imp acting San Francisco Bay/Delta.

The next step is to identify the number of invasi ve species impacting San Francisco Bay/Delta. To do this, we summed the number of species we identified as impacting this area under the San Francisco Bay case study presented earlier in the Response analysis . We also considered the findings of San Francisco Bay invasive sp ecies monitoring programs. Based upon our nonexhaustive review in the San Francisco Bay cas e study, we identified 21 i nvasive species that have adversely affected native speci es and/or habitats in San Franci sco Bay/Delta. We identified another 22 species negatively impacting this area from Cohen and Moyle (2004), producing a total of 43 adversely affecting i nvasive species out of 234 (18.4%).

Calculated Rate of Harm. The Congressional study found 1 5.3% of terrestrial and aquatic invasive species were harmful, while the Chesap eake Bay study found 19.9% of aquatic invasive species were harmful and the San Francisco Bay study identified 18.8% of invasive species as

harmful. The main consideration identified by all three methods is the overall lack of information, obligatorily biasing th e data in some direction depe nding on the definition of harm and what has been published on the effects of non- native species. Of these three estimates, NMFS decided to assess potential impacts to listed species and critical habitat using the mean344 probability of harm (19.9%) identif ied in the two aquatic studies on the Chesapeake Bay and San Francisco Bay to determine the probability of future ballast-mediated invasive species likely to cause adverse impacts to the environment. Th is method reflects the most probable rate of invasions likely to cause harm to listed species and thei r critical habitat. Although each analysis has specific caveats that make each more or less va luable on specific points, we believe that each of these probabilities are relevant and equally significant in determining the likelihood of an invasive species adversely affecting an environment at a global scale.

Estimating Effects of Exposure In this Opinion, NMFS identified the listed species within the action area th at could be affected by an invasive population of organisms. In the Effects Analysis , NMFS further identified the probable rate of exposure for thes e species and the probable effects of that exposure. When an invasive species has established, we mean that it has formed one or more reproductive population(s) in a host location a nd generally occur throughout the year. Although listed specieswould generally co-occur with the invasive species in time and space, the requirement for cooccurrence in space and time is not a prerequisite, as invasive species can ha ve indirect effects to habitat (such as alteratio n of physical conditions) and that imp act can have a significant, if not primary impact to listed species. While it is not possible to anticipate the exact species that will invade a particular location, it is possible to anticipate the prob ability of an establishment. Through the Response Analysis, we found that lik ely responses of liste d species to future invasions would be expected to be similar to past invasions, which ca n be diverse (including competition for prey/nutrition and space, preda tion, disease, genetic hybridization, and habitat alteration) and not necessarily negative (suc h as providing additional prey resources).

Having established that the effects range from beneficial to harmful, NMFS estimated the proportion of invasive species that have been harmful using seve ral hundred invasive species in the analysis to ensure there was very little chance of there being an impact we had not considered in our analysis. Furthermore, with the range of harmful impact rate s being between 18.8% and 19.9% in the two aquatic analyses we conducted, NM FS is confident this is the rate of harmful impacts that will be seen in the future, although we cannot predict which species will establish in the future and the specific effects that those sp ecies would have. By using the mean rate of harmful impacts, NMFS can assess the adverse imp acts of this action in the most probable way. However, because listed species and critical ha bitats often occur over multiple bioregions, to determine the effects to listed sp ecies and critical hab itat for the conclusions of this Opinion, we

must summarize the information from this section in the next sect ion. While the probability of an invasive species establishing in a specific bioregion may be negligible, the cumulative effect of those probabilities to listed species a nd critical habitats may be significant.

Furthermore, in order to determine the probabili ty of an invasive sp ecies establishing through time, NMFS estimated when these numerical stan dards will be required based on the history of ballast water regulations. NANPC A was initially passed in 1990. After six years, that law was updated and NISA was passed. After eight ye ars of working under th e initial framework established by NISA, the USCG issued a final ru le that mandated the use of BWE in 2004. And after eight years of mandatory BWE, the USCG is again reducing the probability of ballastmediated invasions by introducing a num erical standard. Given the history of regulation change,

it is not unreasonable to anticipate another regulation change in seven to eight years. Although345 past performance shows us that changes may well occur eight years from present, no change is required in eight years and we would run the ri sk of underestimating e xposure at the proposed

numeric standard if we limited our review to ei ght years. However, given the history of this program, NMFS anticipates within a duration of tw o historic cycles (16 years) there will be program differences not reflected in the assumpti ons of this Opinion (Table 42). We therefore do not consider effects beyond this timeframe. We are aware that effect s of invaders now and

within the next decade and a half will exte nd beyond this period, but have no information available to use to determine what, where, wh en, and how those effects will impact listed

resources and their critical habitats . As such, any analysis of th ese effects would be arbitrary.

Table 42. Probabilities of ballast-mediated invasions in each bioregion annually as well as over the next 16

years.

Bioregion Approximate

Discharge Volume

(tonnes) Probability of

Annual Invasion Probability of

Invasion over 16

Years Probability

of Harm Probability of

Harm over 16

Years

NA-S1 32000 0.0017 0.0269 0.0003 0.0053

NEP-I 8,540,000 0.4558 0.9999 0.0882 0.7718

NEP-II 603,000 0.032 0.4057 0.0062 0.0947

NEP-IV 1,810,000 0.0964 0.8025 0.0187 0.2607

NEP-V 6,540,000 0.3491 0.9991 0.0675 0.6731

NEP-VI 8,890,000 0.4745 0.9999 0.0918 0.7858

CAR-I 94,010,000 1.0 1.0 0.9707 1.0

CAR-IV 10,190,000 0.544 1.0 0.1053 0.8314

NA-ET2 579,000 0.0309 0.3948 0.006 0.0918

NA-ET3 25,560,000 1.0 1.0 0.2639 0.9926

CAR-VII 3,650,000 0.1946 0.9687 0.0377 0.4593

SP-IX 5,679 0.0003 0.0048 0.00005 0.0007

SP-XII 62,433 0.0033 0.0515 0.0006 0.0096

SP-XXI 507,000 0.0271 0.3557 0.0052 0.0800

Probability of Harmful Invasions by Bioregion

Ballast Water Bioregion NA-S1. NA-S1 includes all ports along the northern coast of Alaska. Within this bioregion, newly esta blishing invasive species have th e potential to impact southern DPS spotted seal.

Between 2005 and 2010 in NA-S1, the corrected av erage ballast water di scharge volume was approximately 32,000 tonnes. The reported volume was corrected to account for incomplete reporting. If this five year disc harge rate is similar for the ne xt 16 years, NMFS would expect the probability for a ballast-medi ated invasion in this bioregio n to be approximately 0.0017 per year under the numerical standard. Assuming a 16 year timeframe before another rulemaking introduces a new management plan, the probabil ity of an invasion in this bioregion is approximately 0.0269. Additionally, the probability of a ballast-mediated invasion causing harm to a listed species would be approximately 0.0003 per year but 0.0053 over 16 years.346 Ballast Water Bioregion NEP-I. NEP-I includes all ports of western Alaska and around the southern coast to Juneau. Within this bioreg ion, newly establishing invasive species have the potential to impact southern DPS spotted seal and western DPS Steller sea lions.

Between 2005 and 2010 in NEP-I, the corrected average ballast water discharge volume was approximately 8.54x10 6 tonnes. The reported volume was co rrected to account for incomplete reporting. If this five year disc harge rate is similar for the ne xt 16 years, NMFS would expect the probability for an invasion in this bioregion to be approxi mately 0.4558 per year under the numerical standard. Assuming a 16 year timeframe before another rulemaking introduces a new management plan, the probability of an inva sion in this bioregion is approximately 0.9999. Additionally, the probability of a ballast-mediated invasion causing harm to a listed species would be approximately 0.0882 per y ear but 0.7718 over 16 years.

Ballast Water Bioregion NEP-II. NEP-II includes ports of the Aleutian Islands including

Dutch Harbor and Cook Inlet. Within this bior egion, newly establishing invasive species have the potential to impact Cook Inlet DPS beluga wh ale, southern DPS spotted seal, and western

DPS Steller sea lions and their critical habitat.

Between 2005 and 2010 in NEP-II, the corrected av erage ballast water discharge volume was approximately 603,000 tonnes. The reported volume was corrected to account for incomplete reporting. If this five year disc harge rate is similar for the ne xt 16 years, NMFS would expect the probability for an invasion in this bioregion to be approxi mately 0.032 per year under the numerical standard. Assuming a 16 year timeframe before another rulemaking introduces a new management plan, the probability of an inva sion in this bioregion is approximately 0.4057. Additionally, the probability of a ballast-mediated invasion causing harm to a listed species would be approximately 0.0062 per y ear but 0.0947 over 16 years.

Ballast Water Bioregion NEP-IV. This bioregion covers the coasts of Oregon and Washington as well as ports in the Columbia River and P uget Sound. Within this bioregion, new non-native species have the potential to impact SONCC coho salmon and their cr itical habitat, lower Columbia River coho salmon, Oregon coast coho salmon and their critical habitat, Pacific

eulachon and their critical habita t, lower Columbia River Chi nook salmon and their critical habitat, upper Columbia River spring-run Chino ok salmon and their critical habitat, Snake River

fall run Chinook salmon and their critical habi tat, Snake River spring/summer run Chinook salmon and their critical habitat, upper Willa mette River Chinook salmon and their critical habitat, Hood Canal summer run chum salmon and their critical habitat, Columbia River chum salmon and their critical habitat, southern DPS gr een sturgeon and their cr itical habitat, Snake River sockeye salmon and their critical habitat, Ozette Lake sockeye sa lmon and their critical habitat, Puget Sound steelhead, Snake River basi n steelhead and their cr itical habitat, upper

Columbia River steelhead and their critical habi tat, middle Columbia River steelhead and their critical habitat, lower Columbia River steelhead and their critical habita t, upper Willamette River

steelhead and their critical ha bitat, Puget Sound/Georgia Basin DPS yelloweye rockfish, Puget Sound/Georgia Basin DPS canary rockfish, Puget Sound/Georgia Basin DPS bocaccio,

leatherback sea turtles and their critical habitat, eastern DPS Stel ler sea lion and their critical habitat, Southern resident killer whale and their critical habitat, and humpback whales.347 Between 2005 and 2010 in NEP-IV, the corrected average ballast water discharge volume was approximately 1.81x10 6 tonnes. The reported volume was co rrected to account for incomplete reporting. If this five year disc harge rate is similar for the ne xt 16 years, NMFS would expect the probability for an invasion in this bioregion to be approxi mately 0.0964 per year under the numerical standard. Assuming a 16 year timeframe before another rulemaking introduces a new management plan, the probability of an inva sion in this bioregion is approximately 0.8025. Additionally, the probability of a ballast-mediated invasion causing harm to a listed species would be approximately 0.0187 per y ear but 0.2607 over 16 years.

Ballast Water Bioregion NEP-V. This bioregion covers northe rn California including the major ports of SF Bay. Within this bioregion, newly establishing invasive species have the potential to impact white abalone , black abalone and their critical habitat, California coastal Chinook salmon and their critical habitat, centr al valley spring-run Chinook salmon and their critical habitat, Sacramento River winter-run Ch inook salmon and their crit ical habitat, central

California coast coho salmon and their critical hab itat, southern Oregon/northern California coast (SONCC) coho salmon and their critical habitat, gr een sturgeon and their cri tical habitat, Pacific eulachon and their critical habitat, central California coast steelhead and thei r critical habitat,

northern California steelhead and their critical ha bitat, south-central California coast steelhead and their critical habitat, Califor nia central valley steelhead and th eir critical habita t, leatherback

sea turtles and their critical hab itat, north Pacific Ocean DPS logge rhead sea turtles, eastern DPS

Steller sea lions and their critical habitat, and humpback whales.

Between 2005 and 2010 in NEP-V, the corrected av erage ballast water di scharge volume was approximately 6.54x10 6 tonnes. The reported volume was co rrected to account for incomplete reporting. If this five year disc harge rate is similar for the ne xt 16 years, NMFS would expect the probability for an invasion in this bioregion to be approxi mately 0.3491 per year under the numerical standard. Assuming a 16 year timeframe before another rulemaking introduces a new management plan, the probability of an inva sion in this bioregion is approximately 0.9991. Additionally, the probability of a ballast-mediated invasion causing harm to a listed species would be approximately 0.0675 per y ear but 0.6731 over 16 years.

Ballast Water Bioregion NEP-VI. Bioregion NEP-VI includes all ports from sout hern California south along the Mexican Coast, though this Opinion only considers the ports in southern California (Long Beach, Los Angeles, and San Diego). Within this bioregion, newly establishing invasive species have the potential to impact southern California steelhead and their crit ical habitat, eastern DPS Steller sea lions and their critical habitat, green sea turtles, leatherb ack sea turtles and their critical habitat, North Pacific DPS loggerhead sea turtles, olive ridley sea turtles, humpback whales.

Between 2005 and 2010 in NEP-VI, the corrected average ballast water discharge volume was approximately 8.89x10 6 tonnes. The reported volume was co rrected to account for incomplete reporting. If this five year disc harge rate is similar for the ne xt 16 years, NMFS would expect the probability for an invasion in this bioregion to be approxi mately 0.4745 per year under the numerical standard. Assuming a 16 year timeframe before another rulemaking introduces a new management plan, the probability of an inva sion in this bioregion is approximately 0.9999.348 Additionally, the probability of a ballast-mediated invasion causing harm to a listed species would be approximately 0.0918 per y ear but 0.7858 over 16 years.

Ballast Water Bioregion CAR-I. Bioregion CAR-I includes all US ports in the en tire Gulf of Mexico and around the southern tip of Florida. Within this bioreg ion, newly establishing invasive species have the potenti al to impact gulf sturgeon and th eir critical hab itat, green sea turtle, hawksbill sea turtle, Kemp’s ridley sea tur tle, leatherback sea turtles, Northwest Atlantic Ocean DPS loggerhead sea turtles, elkhorn coral and thei r critical habitat, staghorn coral and their critical habitat, a nd Johnson’s sea grass and th eir critical habitat.

Between 2005 and 2010 in CAR-I, the corrected av erage ballast water di scharge volume was approximately 94.01x10 6 tonnes. The reported volume was co rrected to account for incomplete reporting. If this five year disc harge rate is similar for the ne xt 16 years, NMFS would expect approximately 5.01 ballast-mediated invasive species to establish each year, which equates to a probability of 1.0. Assuming a 16 year timeframe before another rulemaking introduces a new management plan, the probability of an invasi on in this bioregion is 1.0. Additionally, the probability of a ballast-mediated invasion causing harm to a listed species would be approximately 0.9707 per year bu t 1.0 over 16 years.

Ballast Water Bioregion CAR-IV. CAR-IV includes all ports ar ound the US Virgin Islands

and Puerto Rico. Within this bi oregion, newly establishing invasive species have the potential to impact green turtles and critical habitat, Hawksb ill turtle and their criti cal habitat, leatherback

sea turtles and their critical hab itat, Northwest Atlantic Ocean D PS loggerhead sea turtles, olive ridley sea turtles, smalltooth sa wfish, elkhorn coral and th eir critical habitat, and staghorn coral

and their critical habitat.

Between 2005 and 2010 in CAR-IV, the corrected average ballast water discharge volume was approximately 10.19x10 6 tonnes. The reported volume was co rrected to account for incomplete reporting. If this five year disc harge rate is similar for the ne xt 16 years, NMFS would expect the probability for an invasion in this bioregion to be approxi mately 0.544 per year under the numerical standard. Assuming a 16 year timeframe before another rulemaking introduces a new management plan, the probability of an inva sion in this bioregion is approximately 0.9999. Additionally, the probability of a ballast-mediated invasion causing harm to a listed species would be approximately 0.1053 per y ear but 0.8314 over 16 years.

Ballast Water Bioregion NA-ET2. NA-ET2 includes all ports fr om Maine to Massechusetts. Within this bioregion, newly esta blishing invasive species have th e potential to im pact Gulf of Maine, Atlantic salmon, Gulf of Maine DPS of Atlantic sturgeon, Ne w York Bight DPS of Atlantic sturgeon, Chesapeake Bay DPS of Atlantic sturgeon, Caro lina DPS of Atlantic sturgeon, South Atlantic DPS of Atlantic sturgeon, shor tnose sturgeon, Kemp’s ridley sea turtle, leatherback sea turtles, and Northwest A tlantic Ocean DPS loggerhead sea turtles.

Between 2005 and 2010 in NA-ET2, the corrected av erage ballast water discharge volume was approximately 579,000 tonnes. The reported volume was corrected to account for incomplete reporting. If this five year disc harge rate is similar for the ne xt 16 years, NMFS would expect the probability for an invasion in this bioregion to be approxi mately 0.0309 per year under the349 numerical standard. Assuming a 16 year timeframe before another rulemaking introduces a new management plan, the probability of an inva sion in this bioregion is approximately 0.3948. Additionally, the probability of a ballast-mediated invasion causing harm to a listed species would be approximately 0.006 per ye ar but 0.0918 over 16 years.

Ballast Water Bioregion NA-ET3. NA-ET3 includes all ports so uth of Massechusetts to the southern edge of the Chesapeake Bay. Within this bioregion, newly establishing invasive species have the potential to im pact Gulf of Maine DPS of A tlantic sturgeon, New York Bight DPS of Atlantic sturgeon, Ches apeake Bay DPS of Atlantic st urgeon, Carolina DPS of Atlantic sturgeon, South Atlantic DPS of Atlantic sturgeon, shortnos e sturgeon, green sea turtles, hawksbill sea turtles, Kemp’s ridley sea turtle, leatherback sea turtles, and Northwest Atlantic Ocean DPS loggerhead sea turtles.

Between 2005 and 2010 in NA-ET3, the corrected av erage ballast water discharge volume was approximately 25.56x10 6 tonnes. The reported volume was co rrected to account for incomplete reporting. If this five year disc harge rate is similar for the ne xt 16 years, NMFS would expect approximately 1.364 new ballast-mediated invasi ve species per year, which equates to a probability of 1.0. Assuming a 16 year timeframe before another rulemaking introduces a new management plan, the probability of an inva sion in this bioregion is approximately 1.0. Additionally, the probability of a ballast-mediated invasion causing harm to a listed species would be approximately 0.2639 per y ear but 0.9926 over 16 years.

Ballast Water Bioregion CAR-VII. Bioregion CAR-VII include s all ports between northern Florida and North Carolina. Within this bioreg ion, newly establishing invasive species have the potential to impact Gulf of Maine DPS of Atla ntic sturgeon, New York Bight DPS of Atlantic sturgeon, Chesapeake Bay DPS of Atlantic sturgeon, Carolina DPS of Atlantic sturgeon, South Atlantic DPS of Atlantic sturgeon, shortnose st urgeon, green sea turtles, hawksbill sea turtles, Kemp’s ridley sea turtle, leatherback sea turtle s, Northwest Atlantic Ocean DPS loggerhead sea turtles,

Between 2005 and 2010 in CAR-VII, the corrected average ballast water discharge volume was approximately 3.65x10 6 tonnes. The reported volume was co rrected to account for incomplete reporting. If this five year disc harge rate is similar for the ne xt 16 years, NMFS would expect the probability for an invasion in this bioregion to be approxi mately 0.1946 per year under the numerical standard. Assuming a 16 year timeframe before another rulemaking introduces a new management plan, the probability of an inva sion in this bioregion is approximately 0.9687. Additionally, the probability of a ballast-mediated invasion causing harm to a listed species would be approximately 0.0377 per y ear but 0.4593 over 16 years.

Ballast Water Bioregion SP-IX. Bioregion SP-IX includes all ports of American Somoa. Within this bioregion, newly esta blishing invasive species have th e potential to impact green sea turtles, hawksbill sea turtles, leatherback sea tu rtles, south Pacific DPS loggerhead sea turtles, olive ridley sea turtles, and humpback whales.

Between 2005 and 2010 in SP-IX, the corrected av erage ballast water discharge volume was approximately 5,679 tonnes. The reported volume was corrected to account for incomplete350 reporting. If this five year disc harge rate is similar for the ne xt 16 years, NMFS would expect the probability for an invasion in this bioregion to be approxi mately 0.0003 per year under the numerical standard. Assuming a 16 year timeframe before another rulemaking introduces a new management plan, the probability of an inva sion in this bioregion is approximately 0.0048. Additionally, the probability of a ballast-mediated invasion causing harm to a listed species would be approximately 0.00005 per y ear but 0.0007 over 16 years.

Ballast Water Bioregion SP-XII. Bioregion SP-XII includes all th e ports of Guam. Within this bioregion, newly establishi ng invasive species have the poten tial to impact green sea turtles, hawksbill sea turtles, leatherback sea turtles, south Pacific DPS loggerhead sea turtles, olive ridley sea turtles, and humpback whales.

Between 2005 and 2010 in SP-XII, the corrected av erage ballast water di scharge volume was approximately 62,433 tonnes. The reported volume was corrected to account for incomplete reporting. If this five year disc harge rate is similar for the ne xt 16 years, NMFS would expect the probability for an invasion in this bioregion to be approxi mately 0.0033 per year under the numerical standard. Assuming a 16 year timeframe before another rulemaking introduces a new management plan, the probability of an inva sion in this bioregion is approximately 0.0515. Additionally, the probability of a ballast-mediated invasion causing harm to a listed species would be approximately 0.0006 per y ear but 0.0096 over 16 years.

Ballast Water Bioregion SP-XXI. Bioregion SP-XXI is an isolat ed area in the center of the Pacific Ocean surrounding the islands of Hawaii. Within this bioregion, newly establishing invasive species have the potential to impact Hawaiian monk seal and their critical habitat, green sea turtles, hawksbill sea turtles, leatherback sea turtles, North Pacific DPS loggerhead sea turtles, olive ridley sea turt les, and humpback whales.

Between 2005 and 2010 in SP-XXI, the corrected av erage ballast water discharge volume was approximately 507,000 tonnes. The reported volume was corrected to account for incomplete reporting. If this five year disc harge rate is similar for the ne xt 16 years, NMFS would expect the probability for an invasion in this bioregion to be approxi mately 0.0271 per year under the numerical standard. Assuming a 16 year timeframe before another rulemaking introduces a new management plan, the probability of an inva sion in this bioregion is approximately 0.3557. Additionally, the probability of a ballast-mediated invasion causing harm to a listed species would be approximately 0.0052 per y ear but 0.08 over 16 years.

Ballast Water Discharges to the coastal US. While there are not NMFS protected resources in all waters of the US, there are NMFS species in all marine waters. Between 2005 and 2010, the corrected average ballast water discharge volume was approximately 182.6x10 6 tonnes. The reported volume was corrected to account for incomp lete reporting. If this five year discharge rate is similar for the next 16 years, NM FS would expect approximately 9.7 new ballastmediated invasive species per year under the numerical standard. Therefore, new invasive species are still expected to es tablish in waters of the US annually, but at a reduced rate compared to BWE. Additionally, given the rate of invasions that is expected to harm listed species, approximately 1.877 harmful invasions will occur each year in the US.351 EFFECTS OF THE PROPOSED ACTION:

EFFECTS OF THE BALLAST WATER MANAGEMENT PROGRAM

Overview of the Program

Adaptive Management from Monitoring and Enforcement The USCG ballast water program ha s been functioning for over 20 year s. While this is the first

time the USCG has chosen to have NMFS and US FWS consult on the pote ntial impacts of its program to listed species and thei r critical habitat, there are examples of how successful the ballast water management program monitoring has been in the past. Furthermore, the USCG has other programs with similar monitoring plans that have also proven successful. These monitoring efforts have led to changes within the programs being monitored, showing that the monitoring in place is not only effective but also led to changes within the programs to make the programs more successful.

Voluntary BWE to mandatory BWE. When NISA was passed in 1996, the USCG implemented voluntary BWE as a practicable ballas t water treatment techni que. However, there

were several conditions that would allow vessels to discharge ballast water at a port without having conducted BWE. When the USCG evalua ted the voluntary program in a 2002 report to Congress, the report concluded that low partic ipation in the voluntary program resulted in

insufficient data for an accurate assessment of its effectiveness. This finding triggered the requirement in NISA that the voluntary ballast water management program become mandatory. Two years later in July 2004, the USCG publis hed the final rule making BWE mandatory, requiring all vessels equipped with ballast water tanks and bound for ports or places of the United States to conduct a mid-ocean BWE, retain their ballast water onboard, or use an alternative environmentally sound BWM method approved by the Coast Guard.

Also in 2004, the USCG published penalties for failure to comply with the reporting requirements located in 33 CFR 151 and broade ned the applicability of the reporting and recordkeeping requirements to a majority of vessels bound for ports or places of the United States. The penalties related to record keepi ng are another deterrent to vessels considering avoiding ballast water management actions. Since being passed eight years ago, several criminal cases have been initiated related to record keeping.

Successful Ballast Water Monitoring and Enforcement. The USCG monitors every ship that enters US waters annually. This amounts to approximately 10,000 foreign vessels and 12,000 domestic vessels. After BWE was made mandatory, the USCG was able to test the salinity of the water to determine whether open ocean exchange had occurred.

The primary means of monitoring and enforcement is to analyze the records vessels are required to maintain. When the USCG boards a vessel, they analyze many different aspects of ship safety and statutory requirements, includi ng ballast water treatment compliance. To determine whether ballast water has been treated, the USCG will look at the record book to make sure all treatments were recorded, assess the crew’s ability to operat e the treatment device, and inspect the treatment equipment to determine if it is being maintaine d. This cursory inspection leads to more thorough352 inspections if it is apparent the equipment is not being used. Vessels th at have been cited for violations previously will receiv e more thorough investigations of their equipment to ensure they are no longer in violation of the USCG’s ballast program. In the event violations are discovered, the USCG will initiate criminal prosecution of th e vessels. The USCG is currently in the midst of a prosecution for presenting false information in the balla st water record book to USCG inspectors. The false statement charge was for the crew's attempt to conceal the fact that fuel oil was leaking into the forepeak ballast tank.

Oil-water separator monitoring and enforcement. The oil water separator program is similar to the proposed ballast water mon itoring program in that every ve ssel is boarded at least once a year and as part of the Act for Preventing Pollution from Ships monitoring. While the monitoring goals and methods of the programs may be different, both rely on inspecting record books and physically sampling the vessel.

In calendar year 2011, the USCG has uncovered 18 in stances of criminal ac tions relating to oil water separator inspections. This is down from 22 criminal prosecutions in 2010. In addition, the USCG has initiated two civil penalty cases for oil water separator-related incidents resulting in one unproven case and one still under review by the USCG Hearing Officer. The proven case resulted in a civil penalty of $13,000 and the one under review is re questing a penalty of $19,000. Finally, USCG inspectors issued 268 deficien cies for oil water separator equipment in calendar year 2011 with the large majority of those being rectif ied during the course of the inspection. The oil water separator program acc ounted for a small fraction (0.70%) of the total

vessel related deficiencies identified on vesse ls related to other programs in 2011 (38,568).

Evaluation of Those Components of the Ballast Water Management Program

Assessment of the Program The USCG requested programmatic consulta tion with NMFS, which will result in NMFS determining the effect of the program on listed sp ecies and their critical habitat as opposed to NMFS determining the effect of the action of that program on listed species or critical habitat. This distinction is important because many programs authorize actions with adverse effects that cannot be further mitigated. When that occurs, the programs include feedback loops and adaptive management that allow the effects to li sted species and their critical habitat to be reduced promptly in the future after further mitigation is developed.

Ballast water standards are an example of an au thorized activity under th e USCG’s ballast water management program that are likely to have adve rse effects to listed species and their critical habitat without any further mitigation being availa ble at this time. As previously noted, the probability of an invasion occurring depends on: (1 ) the volume of ballast water discharged and (2) the number of non-native organisms in that discha rge. If no ballast wate r is discharged there is no exposure, or if no organisms are released in a ballast water discharge there is no exposure. Any other scenario has varying levels of affect depending on those two factors.

Under the NISA of 1996, the USCG is tasked with preventing inva sive species from entering US waters; however they are tasked with doing so pr acticably. Therefore, th e USCG is required to consider economic and environmental impacts of any method to prevent an invasion. It is not353 economically practicable to eliminate ballast water discharges. And it is not technologically possible to eliminate all living organisms from ballast water discharge without considerable environmental impacts. Therefore, while it ap pears the action being pr oposed by the USCG may adversely affect listed species and their critical habitat, the actual ques tions NMFS must address are whether their proposal is th e best practicable solution to invasive species, whether the program allows for the lowest probability of a new invasion occurring given the current state of technology, and whether there is anything the USCG could do to further protect listed species and their critical habitat that they are not currently proposing to do.

Many action agencies request consultation on a nu mber of small actions with relatively minor effects that all fall under a single national leve l program. In many cases, programs authorize activities that could be described as “death by one thousand cuts;” allowing many very small projects with small adverse impacts to occur w ithout determining whethe r those small projects could cumulately have large impacts. For inst ance, if the USCG requested consultation with

NMFS on each individual vessel’s ba llast water management, while cumbersome under the ESA, the cumulative effects analysis should identify th e impacts of all of those smaller actions, which may appear to have an individually negligible impact to listed speci es or their critical habitat.

Therefore, the analytical process of assessing the function of a program is different than analyzing the effects of an indi vidual or collection of projects and in many ways, requires more of a long-term approach. Analyzing the US CG’s program allows NMFS to take into consideration the best available technology, acknowledge its imperfections, analyze the legal framework of the USCG’s program, and reach a conclusion about the program instead of the cumulative effect of the individual proposed discharges. NMFS uses the same consultation criteria to assess all programs, including the USCG ’s ballast water management program. In the Approach to the Assesssment section, NMFS identifies seven criteria that the USCG’s program would need to propose within thei r program, demonstrate an ability to achieve and monitor those aspects of their program, and report to NMFS their success in order to insure their program is not likely to jeopardize listed sp ecies or adversely modify th eir critical habitat.

USCG’s Ballast Water Program’s Ability to Estimate the Individual and Cumulative Effects of Ballast Water Discharges

The USCG has monitored ballast water for over two decades. Because the USCG authority for managing ballast water comes fr om the NISA, they are routin ely assessing the impacts of invasive species. To assess i nvasive species impacts, the USCG gathers information on ballast water through reporting, record keeping, and mo nitoring. In addition to program monitoring activities, the USCG also reviews invasive sp ecies literature and continues, as mandated by NISA, to review ballast water treatment perf ormance to reduce individual and cumulative impacts from ballast water discha rge. While the proposed action is the best numerical standard that is technologically achievable at this ti me, the USCG has implemented practicability and periodic reviews with the goal of identifying technology that ca n practicably achieve a more stringent ballast water standard in order to furt her reduce the probability of an invasive species establishing.

The probability of an invasive species establishing depends on two variables. The first, the number of organisms discharged per volume of water, is regulated by the USCG and the USCG354 has proposed the most stringent discharge standa rd that is technologica lly achievable. The second, the total volume of ballast water discha rged, cannot be regulated by the USCG because NISA does not give them that au thority. Although we predict a probability of invasion(s), we cannot anticipate which species would compose these invasions.

The USCG, as mandated by NISA, is required to limit the threat of invasive species as much as is practicable. To achieve this mandate, the USCG has opted to identify the most stringent numerical standard that is achievable, which is the subject of its recent FR. However, eight states have established more stringent ballast water numeric standards than the USCG is proposing. Of those eight, Califor nia has proposed the most string ent standard of any state and New York has proposed the most stringent standard on the East Coast. Additionally, EPA is proposing essentially the same standards to be achieved under their Vessel General Permit. One major difference between the Vessel General Perm it and the USCG standard is that the US Congress exempted certain vessel types under NISA but did not exempt them from discharges under the Clean Water Act, so more vessels will be required to meet the same national standards under the Vessel General Permit than under the Ballast Water Management Program. Additionally, the USCG st andards evaluate whether a specific technology can be installed on all types of vessels or whether it would only fit so me vessel types and designs, whereas, the Vessel

General Permit does not take into account the ac hievability of the standard on various vessel
types.

Those states with more stringent standards have allowed for onshore discharge, no discharge, or discharging potable water just like the USCG; howe ver, if ballast water is to be discharged,

vessels in those states must meet a more stri ngent standard. The manner in which the numeric

standards for those states was derived is different than the way the USCG determined the

national standard. In a ll cases, evaluations of current tec hnology were conducted but for instance in California, a technology is c onsidered potentially achievable , and therefore acceptable as a treatment system, if it meets the standard just once out of all samples taken. Of 46 technologies in California, only eight were pot entially able to achieve a more stringent standard and none of the technologies were capable of meeting a more stringent standard in every test. Only three technologies tested in California have proven to be able to meet the more stringent standard more

than 50% of the time. While the USCG rule is not numerically as stringent, the standard set will be met during every test to be cleared for use in US waters. Additionally, a group of Great Lakes states has voiced concern about the ability for technology to consistently achieve California’s more stringent standard. Even California is concerned about the long-term goals of its ballast water standard of zero organisms by 2020 and have established a review in 2016 that will assess whether that standard is possible.

Like California, New York has established st andards 1000 times more stringent than those proposed by the USCG. However, because no technology exists to meet the standards introduced by the state of New York in 2008, in February 2012, the stat e extended the deadline to meet the New York standard, meaning until th e date that technology is capable of meeting a standard 1000 times more stringent than the US CG’s, New York effectively has no ballast water

standard.355 Based on sampling conducted by foreign nations, it a ppears there is technology available that can consistently meet the standard proposed by th e USCG. Furthermore, because it appears technology is not yet capable of achieving a more stringent sta ndard, the USCG monitors the total ballast water discharge through their reporting and reco rd keeping requirements. Additionally, the USCG has establis hed periodic reviews every thr ee years with the intention of making the ballast water standard more stringe nt if practical tec hnology is discovered.

USCG’s Ability to Estimate Stressors Resu lting from the Ballast Water Discharges Many publications have attempted to predict whic h non-native species are likely to invade and what their impacts will be (Guisan and Thuiller 2005, Vaclavik and Meentemeyer 2009)(Williamson 1996, Williamson 1999, Enserink 1999, Peterson and Vieglais 2001, Kolar and Lodge 2001), but those same studies have dete rmined that it is not currently possible to identify individual non-native species with the highest probability of invading an area next, nor is it possible to anticipate sp ecific locations those species ma y establish viable populations. Because it is not possible to predict which specie s are likely to become established in a given area or what the anticipated impacts of an invasion could be, the USCG and NMFS have determined that the range of ecological effects documented over the past 50 years of invasive species science are likely to be similar to the ecological eff ects of future invasions.

Given the unpredictability of invasive speci es establishment, the USCG has proposed a numerical standard that will limit, to the ma ximum extent practicable, the probability of an invasion occurring. Furthermore, the USCG acknowl edges that the total volume discharged is critical, but the agency has no authority to re gulate the volume of ballast water discharged. Because of this, the USCG requires reporting and record keeping that monitor all aspects of ballast water from its uptake to discharge. The USCG has also implemented three year practicability and periodic revi ews to evaluate the practicabil ity of new technologies being developed with the goal of moving the numerical ballast water standard from the current

standard to one that is more stringent, ultimately eliminating the threat of invasive species.

USCG’s Ability to Estimate Compliance wi th the Ballast Water Management Program The USCG has established monitoring, reporting, a nd record keeping requirements of all vessels to ensure that all vessels and vessel operators ha ve complied with the conditions, restrictions, and mitigation measures required by the proposed acti on. The records are checked when monitoring takes place; the USCG boards ever y vessel that enters US waters at least once each year. As previously described, a vessel is re quired to report ballast water discharges prior to making them

and maintain a record for at leas t 24 months of information about the ballast water in the tanks, treatment, and discharge. Furthermore, mitigation measures are established by the FR and while reporting and record keeping are not specifically required to verify the vessels have conducted the mitigation, the reports and records will identify the location of ballast discharge to ensure vulnerable areas are avoided and protected. The FR also contains specific compliance monitoring to ensure that all ve ssels have been equipped with ballast water treatment technology that was approved by the USCG and that the technology is being maintained and continues to

meet the numeric standard. In the event that non-compliance is identif ied, the NISA provides tough enforcement regulations that allow the USCG to bring charges against the vessel operators. Under the similar oil water separator monitoring program, approximately 20 criminal charges are discovered every year and hundreds of other violations are docu mented and violating vessels356 fined. While it is more difficult to prove a vess el is not complying with BWE, the USCG has brought two cases related to record keeping a nd BWE. Under the proposed numerical ballast water standard, the USCG expects monitoring wate r and sediment samples to lead to increased compliance and easier enforcement.

USCG’s Ability to Determine Exposures of List ed Resources to the Effects of the Proposed

Action The USCG has compliance monitoring, reporting, and record keeping requirements that allow them to reliably estimate whether and to what degree specific endangered or threatened species and their critical habitat are likely to be exposed to the direct or indirect effects of the discharges to be authorized by the proposed action. The reporting and record keeping required by each

vessel is maintained at the National Ballast In formation Clearinghouse. Because every vessel is boarded at least once a year a nd those vessels are required to report their ballast information electronically, the NBIC data is kept up to date annually. Based on years of records, the USCG is well-informed of the reporting rates, the reporting re liability, and the volume of ballast water discharged annually. Furthermore, annual compliance monitoring ensures that vessels are equipped with ballast water trea tment technology and that those t echnologies continue to meet the ballast water standard after being installed.

Based on the mandatory monitoring requirements in the FR, the USCG can monitor the volume of treated ballast water being discharged to vari ous ports, estuaries, and marine bioregions to determine the listed species in the area that may be affected by the proposed ballast water discharges. The goal of the USCG, as authorized by NISA, is to reduce the risk of invasive species from ballast water discharge as much as is possible. Reducing the risk of invasive species is accomplished in one of two ways: (1) reducing the number of organisms per volume of discharge or (2) reducing the volume of ballast water discharged. The USCG does not have the authority to manage the volume of ballast wate r discharged and theref ore can only reduce the number of organisms per discharge to the maxi mum extent practicable. Furthermore, they currently monitor the volume of ballast water discharged by each ship, while also conducting practicability reviews an d periodic reviews to identify new technology capable of achieving a more stringent standard. In these ways, the US CG is able to ensure the exposure of listed resources to ballast water discharg e is not greater than has been analyzed in this Opinion. All these monitoring efforts will allow the USCG to understand the ballast water discharges and the potential exposure of listed sp ecies to ballast water discharges and undertake adaptive management measures if problems are identified in their program.

USCG’s Ability to Monitor Adverse Effects from Authorized Activities As described above, it is not possible to predict which species may invade or what those impacts may be. As a result, NMFS and USCG have dete rmined that the likely effects of new listed species will be similar to the e ffects seen by the 500 non-native species currently inhabiting the waters of the US. Furthermore, there is genera lly a delay between when a species is introduced and when its population is first identified as be ing invasive, the point at which adverse effects could potentially occur. This is because, as the final PEIS show s, there is a high likelihood that most species that are introduced will go extinct rath er than establish in their new environment.357 Ballast water is not the only source of non-native species in waters of the US. In fact, Molner et
al. (2008) identify all sources of invasive species and identify shipping as responsible for no more than 69% of total invasions, and of that number, ballast water accounts for between 31 and 61% of shipping related invasive species. Therefore, simple environmental monitoring would

not be able to detect whether invasions were re sulting from this proposed action or from one of the many more numerous sources of invasive species.

To address the uncertainties about non-native species and their effects, the USCG has implemented compliance monitoring and effectiven ess monitoring that will ensure ballast water treatment technology has been inst alled in all vessels and is e ffectively meeting the proposed numerical standard. That coupled with monitori ng the total volume of ballast water discharge

allows the USCG to monitor the risk of new invasive species resulting from ballast water discharge and to anticipate any adverse eff ects caused by those invasions. All of these monitoring efforts will allow the USCG to undert ake adaptive management measures if problems are identified in their program.

USCG’s Consideration of Species’ Status and Population Effects The USCG began working with NMFS and USFWS in 2006 to evaluate the effects of a ballast water standard on listed species and their critical habitat. The US CG’s primary intent during this process was reducing the risk of invasive species establishing and therefore the potential adverse effects of invasive species. While this new numerical standard is a clear improvement over BWE, both in terms of reducing the risk of inva sion and increasing the enforceability of ballast water regulations, there is still a po tential risk to listed species.

The USCG proposed the most stringent practicabl e ballast water management solution to protect the environment. Furthermore, the USCG progr am includes a practicability review and later periodic reviews that allow the USCG to increas e the stringency of the standard as soon as technology is developed that can practicably ach ieve a greater stringen cy. Additionally, the USCG included three alternative discharge t echniques that are acceptable: (1) no ballast discharge, (2) discharge to an on-site facility or barge, and (3) using potable water in ballast tanks. The USCG expects that some vessels may d ecide not to discharge ballast water due to the

increased risk of enforcement and cost of in stalling treatment technol ogies; however, most vessels are unable to operate safely without taking on and discha rging ballast water. There are currently no on-site discharge facilities in the US, but some ports may choose to develop discharge sites. And using potable ballast wate r, while an option, may increase the strain on domestic drinking water resources and have an adverse economic impact.

In addition to the periodic and practicability reviews, the USCG will also conduct compliance monitoring to ensure all vesse ls are equipped with treatment technologies and that those technologies are being used. And as part of those reviews, the USCG will also ensure the technologies are still performing in compliance with the numerical standard. In addition to these on-ship monitoring efforts, the USCG will also re quire vessels to report ballast water discharges and maintain records of those discharges for 24 months. All of these monitoring efforts will

allow the USCG to understand the ballast water di scharges and the risks associated with them and undertake adaptive management measures if problems are identified in their program.358 Summary of USCG’s Ability to Prevent Harmful Exposure The USCG is unable to prevent harmful exposure, which would have resulted in a no effect determination to listed species and their critic al habitat. The USCG has minimized to the maximum extent practicable the lik ely adverse effects associated w ith this action. As noted in the approach to the assessment section, minimizi ng adverse effects to the maximum practicable extent may still not avoid jeopardizing listed species or adversely modifying their critical habitat because adverse effects are still expe cted as a result of this action.

There are two primary concerns with ballast wa ter discharge: (1) the number of organisms per volume of water and (2) th e total volume of ballast water di scharged. The USCG does not have the authority to limit total volume of ballast water discharge, but they are able to limit the number of organisms per volume that can be disc harged. Therefore, the USCG has selected a

numerical standard that is curre ntly the most practicably achieva ble. The USCG requires vessels subject to reporting requirements (identified above) to report their ballast water plans as they come in to port, prior to discharging ballast water. Mitigation measures are included in the FR to avoid environmentally sensitive areas and complia nce with these mitigation measures will be reviewed by the reports as we ll as required record keeping on board vessels for 24 months.

Vessel operators would be required to report and ma intain records of their discharges, but there is the potential for dishonesty. To make sure ve ssel operators are in fact treating ballast water and discharging it when they say they will, th e USCG boards all vessels at least once a year determined by a decision matrix. During these exer cises, the USCG determines that the vessel is in compliance with the proposed FR and has tr eatment technology installed. Furthermore, the USCG takes samples of the ballast water and balla st sediment to ensure the treatment technology is meeting pre-installation treatment success and being maintained to continue to meet the

numerical standard. If the vesse l is out of compliance, the USCG is able to take enforcement action against the vessel, as has been demonstrat ed previously with criminal prosecutions for falsified BWE recordkeeping and during complia nce monitoring of the oil-water separator

program.

In addition to reporting prior to discharge and keeping records of those discharges and treatments for 24 months, the USCG has implemented a practi cability review to be carried out by 2016 that will assess whether there are new technologies av ailable to increase the stringency of the standard. After 2016, the USCG will conduct pe riodic reviews every three years that will effectively be the same as the practicability review. The technology reviews will assess new technologies to determine if the numerical standard can be made mo re stringent. In the event of a technology that can allow vessels to practicab ly achieve a more stringe nt standard, the USCG

will determine whether that tec hnology will require an economic impact analysis. If the new technology is not economically si gnificant, the new practicable numeric standard could be implemented in less than a year and if it is economically significan t, then it would be implemented within three years.

Because the USCG was not able to avoid all possible harmful e xposure associated with ballast water discharges, NMFS must assess whether an d how quickly the USCG could implement new protective measures in the event th at the proposed numerical standard is unable to be achieved. The USCG is going to assess th e potential of all new technol ogies that may achieve the359 numerical standard under a tier ed consultation process due to the quantity and variety of treatment technologies, as well as the anticipation that new t echnologies will continually be developed until the technologies are able to achieve more stringent standards. Essentially, once a technology is developed and can prac ticably increase the stringency of the ballast water standard, all vessels operators will be re quired to equip their vessels wi th that technology until a new technology is developed that ca n achieve that same level of stringency. Similarly, if a technology is approved but through mon itoring is revealed to consiste ntly not be able to achieve the numerical standard, the USCG would reinit iate consultation on that technology or quickly determine that particular tec hnology is no longer approved for use and suggest one of the other technologies that is available that can meet th e numerical standard. Because of all of the monitoring and reporting required by the FR, as well as the mitigation measures, the USCG is able to quickly undertake adaptive management measures if problems are identified in their program to protect listed species, their critical habitat, and the general ecosystem health within

the coastal areas of the US.

Cumulative Effects

Cumulative effects include the effects of future St ate, tribal, local, or private actions that are reasonably certain to occur in the action area consid ered in this Opinion. Future Federal actions that are unrelated to the proposed action are not consider ed in this section because they require separate consultation pursuant to section 7 of the ESA.

Government and private actions may include ch anges in land and water use patterns, including ownership and intensity, any of which could affect li sted species or their habitat. It is difficult, and perhaps speculative, to analy ze the effects of such actions, considering the broad geographic landscape covered by this BO, the geographic and political variati on in the action area, extensive private land holdings, the uncertain ties associated with State a nd local government and private actions, and ongoing changes in the region’s economy. Adverse effects to riverine habitat from continued urbanization in the basin are reasonably certain to occur. However, state and local governments have regulations in place to mini mize these effects to listed species, including regulations regarding construc tion best management practices, storm water control, and

treatment of wastewater.

Although quantifying an incremental change in su rvival for the species considered in this consultation due to the cumulative effects is not po ssible, it is reasonably likely that those effects within the action area will have a small, long- term, negative effect on the likelihood of their survival and recovery.

Conclusion

After reviewing the current status of Gulf of Maine Atlantic salmon, south Pacific loggerhead sea turtles, North Pacific loggerhead sea turtle s, Hawaiian monk seals, and Cook Inlet beluga whales, the proposed action, the environmental base line, the action area, and cumulative effects, it is NMFS’ biological opinion that the program, ma y affect but is not likel y to adversely affect these threatened and endangered sp ecies or their critical habitat.360 NMFS reviewed the current status of Californi a coastal Chinook salmon, central valley Chinooksalmon, lower Columbia River Chinook salmon, Puget Sound Chinook salmon, Snake River fallrun Chinook salmon, Snake River spring/summer-run Chinook salmon, upper Columbia River spring-run Chinook salmon, Upper Willamette Rive r Chinook salmon, Sacramento River winterrun Chinook salmon, Hood Canal summer run ch um salmon, Columbia River chum salmon, lower Columbia River coho salmon, Oregon Co ast coho salmon, SONCC coho salmon, central California coast coho salmon, Snake River so ckeye salmon, Ozette Lake sockeye salmon, California central valley steelhead, lower Colu mbia River steelhead, middle Columbia River steelhead, Puget Sound steelhead, Snake River Ba sin steelhead, south-ce ntral California coast steelhead, upper Columbia River steelhea d, upper Willamette River steelhead, northern California steelhead, southern Ca lifornia steelhead, central Ca lifornia coast steelhead, green sturgeon, gulf sturgeon, Atlantic sturgeon, shor tnose sturgeon, smalltooth sawfish, yelloweye rockfish, bocaccio, canary rockfish, Pacific eu lachon, elkhorn coral, staghorn coral, white abalone, black abalone, Johnson’s se a grass, Steller sea lions, spo tted seals, green sea turtles, hawksbill sea turtles, loggerhead sea turtles, Kemp’s ridley sea turtles, olive ridley sea turtles, leatherback sea turtles, killer whales, and humpb ack whales, the environmental baseline for the action area, the effects of the proposed action, and the cumulative effects. Because the program has adaptive management capabili ties, feedback loops, monitori ng plans, record keeping and maintaining requirements, and enforcement capab ilities, it is NMFS’ Opinion that the USCG ballast water management program , is not likely to je opardize the continued existence of these endangered or threatened species.

NMFS also reviewed the current status of California coastal Chi nook salmon, central valley Chinook salmon, lower Columbia River Chi nook salmon, Puget Sound Chinook salmon, Snake River fall-run Chinook salmon, Snake Rive r spring/summer-run Chinook salmon, upper Columbia River spring-run Chinook salmon, Upper Willamette River Chinook salmon, Sacramento River winter-run Chinook salmon, H ood Canal summer run chum salmon, Columbia River chum salmon, Central California coho salmon, Oregon Coast coho salmon, SONCC coho salmon, Snake River sockeye salmon, Ozette Lake sockeye salmon, California central valley steelhead, lower Columbia River steelhead, middle Columbia River steelhead, Puget Sound steelhead, Snake River Basin steelhead, south-cent ral California coast steelhead, upper Columbia River steelhead, upper Willamette River steelh ead, northern California steelhead, southern California steelhead, central Ca lifornia coast steelhead, Pacifi c eulachon, green sturgeon, gulf sturgeon, smalltooth sawfish, Steller sea lions, southern resident killer whales, green sea turtles, hawksbill sea turtles, leatherback sea turtles, logg erhead sea turtle, elkhorn coral, staghorn coral, and black abalone, the environmen tal baseline for the action area, the effects of the proposed action, and the cumulative effects of this pr ogram. Because the program has adaptive management capabilities, feedback loops, mon itoring plans, record ke eping and maintaining requirements, and enforcement capabilities, it is NMFS’ Opinion that the USCG ballast water management program, is not likely to adversely modify critical habitat for these species.

INCIDENTAL TAKE STATEMENT

Section 9 of the Act and Federal regulation pursuant to section 4(d) of the ESA prohibit the take of endangered and threatened species, respectivel y, without special exemp tion. Take is defined as to harass, harm, pursue, hunt, shoot, wound, kill , trap, capture or collect, or to attempt to361 engage in any such conduct. Harm is further defined by NMFS to include significant habitat modification or degradation that results in death or injury to listed species by significantly impairing essential behavioral patterns, includ ing breeding, feeding, or sheltering. Harass is defined by the USFWS as intentiona l or negligent actions that create the li kelihood of injury to listed species to such an extent as to signifi cantly disrupt normal behavior patterns which include, but are not limited to, breedi ng, feeding, or sheltering. Incide ntal take is de fined as take that is incidental to, and not the purpose of, the carry ing out of an otherwise lawful activity. Under the terms of section 7(b)(4 ) and section 7(o)(2), taking th at is incidental to and not intended as part of the agency action is not considered to be prohi bited taking under the Act provided that such taking is in compliance with the terms and conditions of this Incidental Take

Statement.

Amount or Extent of Take For this proposed action, take is authorized for each species identified in this section. However, given the complexity and limitati ons of scientific knowledge of the invasion process, predicting invasions, invasive species, or th e expected effects is impossible at this time. B ecause there is no means of predicting specific invaders, the numbe r or species of listed species that will be affected, the proportion of the popula tion that will be taken, or the am ount or extent of impacts to critical habitat, NMFS decided to assess the risks and effects of invasive species at a geographic scale using a surrogate for populati on and habitat scale effects.

As was explained in the exposure analysis, usin g the invasion rate identified using BWE, applying the anticipated reduction in propagule pressure identified in the final PEIS, and then multiplying that rate by the volume of ballast wate r discharged to a bioregion, it was possible to conservatively anticipate the rate of invasions to either bioregions or the US as a whole in terms of the number of new viable populations establ ishing each year. The volume of water in the coastal, estuarine, and fr eshwater regions of the US is so va st that it is not economically possible

to conduct annual environmental assessments a nd reliably determine the number of newly established invasive species each year. It is more practicable to estimate the number of new invasions by using the best av ailable science and measuring i nvasions by the volume of water discharged geographically based on the propos ed numerical standard for ballast water management. Therefore NMFS will provide take in terms of the number of harmful invasive species establishing each year usin g volumes of ballast water disc harged as a surrogate for the take that is likely to affect li sted species. While the USCG do es not control the volume of water discharged, they do regulate the number of organi sms per volume of ballast water that is then discharged and therefore NMFS belie ves this is an appropriate pr oxy and easier to identify than the number of organisms discharged. The USCG currently requires all vessels to report all information about ballast water from uptake, trea tment, and discharge, and maintain records of those reports for 24 months.

This Opinion does not reach conclusions about the risk posed by the numerical ballast water standard but rather the ballast water management program and identifies the possible, not probable, effects resulting from any non-zero numerical standard. The probable rate of species

invasions is identified as is th e probable rate of harmful invasions . The incidental take expected annually under this program is id entified using the volume of balla st water discharged as a proxy for listed species and critical habitat affected. As the peri odic reviews allow the USCG to362 establish more stringent ballast wa ter standards, the amount and extent of take of lis ted species is expected to decrease. Assuming technology will advance in the future to allow for a new, practicable standard to be implemented, the impact s to the listed species and their critical habitat will be reduced, much as the reductions of thos e impacts are identified in the Opinion as the

USCG moves from BWE to a numeric standard. The new levels of take resulting from the program’s standard and authorized by NM FS would depend on the numerical standard implemented and the volume of ballast water being discharged by vessels, which may range between no take (no effect) and the amount of take authorized in this Opin ion, but not to exceed the amount in this Opinion.

Not all newly established invasive species are harmful to listed specie s or their critical habitat. In fact, some may be beneficial, while most are innocuous. Past research has concluded that the proportion of invasive species th at are harmful ranges from approximately 15% to 19.9%. Once an invasive species has established, co-occurring in time and space with a listed species and its critical habitat, the take that occurs is not lim ited to the introduction, bu t will persist as long as that non-native species continues to co-occur in time and space with the listed entity with harmful effects. Using the conservative figure of 19.9% of invasive species being harmful,

NMFS estimates:

California coastal Chinook salmon and their critical habitat will be harmed by no more than one

ballast-mediated harmful invasive species per year and no more than three in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion

NEP-V.

Central Valley Chinook salmon and their critical ha bitat are likely to be harmed by no more than one ballast-mediated harmful invasive species pe r year and no more than three in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-V.

Lower Columbia River Chinook salmon and their cr itical habitat are likel y to be harmed by no

more than one ballast-mediated harmful invasive species per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 per year tonnes to Bioregion NEP-IV.

Snake River fall-run Chinook salmon and their cr itical habitat are likely to be harmed by no

more than one ballast-mediated harmful invasive species per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Snake River spring/summer-run Chinook salmon and th eir critical habitat are likely to be harmed by no more than one ballast-mediated harmful invasi ve species per year and no more than one in the next 16 years, which would correlate to a volumetric ballas t discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Upper Columbia River Chinook salmon and their cr itical habitat are likel y to be harmed by no

more than one ballast-mediated harmful invasive species per year and no more than one in the363 next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Upper Willamette River Chinook salmon and their cr itical habitat are likely to be harmed by no

more than one ballast-mediated harmful invasive species per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Puget Sound Chinook salmon and their critical hab itat are likely to be harmed by no more than one ballast-mediated harmful invasive species per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes to Bioregion

NEP-IV.

Sacramento winter-run Chinook salmon and their cr itical habitat are likel y to be harmed by no

more than one ballast-mediated harmful invasive species per year and no more than three in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-V.

Columbia River chum salmon and their critical ha bitat are likely to be harmed by no more than one ballast-mediated harmful invasive species per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Hood Canal summer-run chum salmon and their crit ical habitat are likely to be harmed by no

more than one ballast-mediated harmful invasive species per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Lower Columbia River coho salmon are likely to be harmed by no more than one ballastmediated harmful invasive species per year and no more than one in the next 16 years, which

would correlate to a volumetri c ballast discharge of 94x10 6 tonnes per year to Bioregion NEPIV.

SONCC coho salmon and their critical habitat ar e likely to be harmed by no more than one

ballast-mediated harmful invasive species per year and no more than three in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregions

NEP-IV and NEP-V.

Central California coast coho salm on and their critical ha bitat are likely to be harmed by no more than one ballast-mediated harmful invasive specie s per year and no more than three in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-V.

Oregon Coast coho salmon and their critical habita t are likely to be harmed by no more than one

ballast-mediated harmful invasive species per year and no more than one in the next 16 years,364 which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion

NEP-IV.

Snake River sockeye salmon and their critical ha bitat are likely to be harmed by no more than one ballast-mediated harmful invasive species per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Ozette Lake sockeye salmon and th eir critical habitat are likely to be harmed by no more than one ballast-mediated harmful invasive species per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

California central valley steelhead and their critical habitat are li kely to be harmed by no more than one ballast-mediated harmful invasive specie s per year and no more than three in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-V.

Lower Columbia River steelhead and their critical habitat are likely to be harmed by no more than one ballast-mediated harmful invasive specie s per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Middle Columbia River steelhead and their critical habitat are likely to be harmed by no more than one ballast-mediated harmful invasive specie s per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Northern California steelhead and their critical habitat are likely to be harmed by no more than one ballast-mediated harmful invasive species pe r year and no more than three in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-V.

Snake River Basin steelhead and their critical ha bitat are likely to be harmed by no more than one ballast-mediated harmful invasive species per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

South-central California coast stee lhead and their critical habita t are likely to be harmed by no

more than one ballast-mediated harmful invasive species per year and no more than three in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-V.

Southern California steelhead and their critical habitat are likely to be harmed by no more than one ballast-mediated harmful invasive species per year and no more than one in the next 16365 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Upper Columbia River steelhead and their critical habitat are likely to be harmed by no more than one ballast-mediated harmful invasive specie s per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Upper Willamette River steelhead and their critica l habitat are likely to be harmed by no more than one ballast-mediated harmful invasive specie s per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Puget Sound steelhead are likely to be harmed by no more than one ballast-mediated harmful invasive species per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Central California coast steelhead and their critical habitat are likely to be harmed by no more than one ballast-mediated harmful invasive specie s per year and no more than three in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-V.

Pacific eulachon and their critical habitat are likely to be harmed by no more than one ballastmediated harmful invasive species per year and no more than three in the next 16 years, which

would correlate to a volumetri c ballast discharge of 94x10 6 tonnes per year to Bioregions NEPIV and NEP-V.

Shortnose sturgeon are likely to be harmed by no more than one ballast-mediated harmful invasive species per year and no more than eight in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregions NA-ET2, NA-ET3, and

CAR-VII.

Gulf of Maine DPS Atlantic sturgeon are like ly to be harmed by no more than one ballastmediated harmful invasive species per year and no more than eight in the next 16 years, which

would correlate to a volumetri c ballast discharge of 94x10 6 tonnes per year to Bioregions NAET2, NA-ET3, and CAR-VII.

New York Bight DPS Atlantic sturgeon are likely to be harmed by no more than one ballastmediated harmful invasive species per year and no more than eight in the next 16 years, which

would correlate to a volumetri c ballast discharge of 94x10 6 tonnes per year to Bioregions NAET2, NA-ET3, and CAR-VII.

Chesapeake Bay DPS Atlantic sturgeon are likely to be harmed by no more than one ballastmediated harmful invasive species per year and no more than eight in the next 16 years, which

would correlate to a volumetri c ballast discharge of 94x10 6 tonnes per year to Bioregions NAET2, NA-ET3, and CAR-VII.366 Carolina DPS Atlantic sturgeon are likely to be harmed by no more than one ballast-mediated harmful invasive species per year and no more than eight in the next 16 years, which would

correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregions NA-ET2, NA-ET3, and CAR-VII.

South Atlantic DPS Atlantic sturgeon are like ly to be harmed by no more than one ballastmediated harmful invasive species per year and no more than eight in the next 16 years, which

would correlate to a volumetri c ballast discharge of 94x10 6 tonnes per year to Bioregions NAET2, NA-ET3, and CAR-VII.

Green sturgeon and their critical habitat are lik ely to be harmed by no more than one ballastmediated harmful invasive species per year and no more than three in the next 16 years, which

would correlate to a volumetri c ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-V.

Gulf sturgeon and their critical habitat are lik ely to be harmed by no more than two ballastmediated harmful invasive species per year a nd no more than 16 in the next 16 years, which

would correlate to a volumetri c ballast discharge of 188x10 6 tonnes per year to Bioregion CARI.

Yelloweye rockfish are likely to be harmed by no more than one ballast-mediated harmful invasive species per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Bocaccio are likely to be harmed by no more than one ballast-mediated harmful invasive species per year and no more than one in the next 16 year s, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Canary rockfish are likely to be harmed by no more than one ballast-med iated harmful invasive species per year and no more than one in th e next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Smalltooth sawfish and their critical habitat are lik ely to be harmed by no more than one ballastmediated harmful invasive species per year and no more than five in the next 16 years, which

would correlate to a volumetri c ballast discharge of 94x10 6 tonnes per year to Bioregion CARIV and the Florida ports of CAR-I.

Elkhorn coral and their critical habitat are likely to be harmed by no more than one ballastmediated harmful invasive species per year and no more than five in the next 16 years, which

would correlate to a volumetri c ballast discharge of 94x10 6 tonnes per year to Bioregion CARIV and the Florida ports of CAR-I.

Staghorn coral and their critical habitat are likely to be harmed by no more than one ballastmediated harmful invasive species per year and no more than five in the next 16 years, which

would correlate to a volumetri c ballast discharge of 94x10 6 tonnes per year to Bioregion CARIV and the Florida ports of CAR-I.367 White abalone are likely to be harmed by no more than one ballast-mediated harmful invasive species per year and no more than three in th e next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-V.

Black abalone and their critical habitat are lik ely to be harmed by no more than one ballastmediated harmful invasive species per year and no more than three in the next 16 years, which

would correlate to a volumetri c ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-V.

Eastern DPS Steller sea lions and their critical habitat are likely to be harmed by no more than one ballast-mediated harmful invasive species pe r year and no more than five in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregions NEP-IV and NEP-V.

Western DPS Steller sea lions and their critical ha bitat are likely to be harmed by no more than one ballast-mediated harmful invasive species pe r year and no more than four in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregions NEP-I and NEP-II.

Spotted seals are likely to be harmed by no more than one ballast-mediated harmful invasive species per year and no more than four in th e next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregions NEP-I, NEP-II, and NAS1.

Southern resident killer whales and their critical habitat are likel y to be harmed by no more than one ballast-mediated harmful invasive species per year and no more than one in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregion NEP-IV.

Humpback whales are likely to be harmed by no more than one ballast-mediated harmful invasive species per year and no more than 13 in the next 16 years, which would correlate to a volumetric ballast discharge of 94x10 6 tonnes per year to Bioregions SP-XXI, IX, XII, NEP-I, NEP-II, NEP-V, NEP-VI, NEP-IV, NA-ET3, NA-ET2, and CAR-VII.

Green sea turtles and their critical habitat are likely to be harmed by no more than two ballastmediated harmful invasive species per year a nd no more than 23 in the next 16 years, which

would correlate to a volumetri c ballast discharge of 188x10 6 tonnes per year to Bioregions SPXXI, SP-IX, SP-XII, NA-ET3, CAR-VII, CAR-IV, CAR-I, and NEP-VI.

Leatherback sea turtles and their critical habitat are likely to be harmed by no more than two

ballast-mediated harmful invasive species per year and no more than 24 in the next 16 years, which would correlate to a volumetric ballast discharge of 188x10 6 tonnes per year to Bioregions SP-XXI, SP-IX , SP-XII, NA-ET3, CAR-VII, CAR-IV , CAR-I, NEP-VI, and NEP-V.

Kemp’s Ridley sea turtles are likely to be harm ed by no more than two ballast-mediated harmful invasive species per year and no more than 23 in the next 16 years, which would correlate to a368 volumetric ballast discharge of 188x10 6 tonnes per year to Bioregions NA-ET3, CAR-VII, CAR-IV, and CAR-I.

Olive ridley sea turtles are likely to be harm ed by no more than two ballast-mediated harmful invasive species per year and no more than 18 in the next 16 years, which would correlate to a volumetric ballast discharge of 188x10 6 tonnes per year to Bioregions SP-XXI, SP-IX, SP-XII, CAR-I, and NEP-VI.

North Pacific loggerhead sea turtles are likely to be harmed by no more than one ballastmediated harmful invasive species per year and no more than five in the next 16 years, which

would correlate to a volumetri c ballast discharge of 94x10 6 tonnes per year to Bioregions SPXXI, NEP-VI, and NEP-V.

North Atlantic loggerhead sea turtles are likely to be harmed by no more than two ballastmediated harmful invasive species per year a nd no more than 23 in the next 16 years, which

would correlate to a volumetri c ballast discharge of 188x10 6 tonnes per year to Bioregions NAET3, CAR-VII, CAR-IV, and CAR-I.

Hawksbill sea turtles and their critical habitat are likely to be harmed by no more than two

ballast-mediated harmful invasive species per year and no more than 23 in the next 16 years, which would correlate to a volumetric ballast discharge of 188x10 6 tonnes per year to Bioregions SP-XXI, SP-IX, SP-XII, NA-ET3, CAR-VII, CAR-IV, and CAR-I.

This Opinion calculates the likely number of i nvasive species introduced using the volume of ballast water discharged as a surrogate, then pr ovides take coverage for harmful invasions by calculating the ratio of harmful invasive spec ies identified from seve ral hundred previous invasive species that have been studied. Under a new, more stri ngent ballast wate r standard, the number of non-native species per volume of ba llast water would be reduced. This will reduce the probability of a harmful inva sive species establishing, and th erefore require a new proxy for incidental take. However, as long as establishing a new more stringent numerical standard would operate within the requirements of the ballast water management program as were assessed in this Opinion, then identifying a more stringent numerical stan dard would not require reinitiation of consultation, but ra ther it would require NMFS to am end this Opinion to identify

the new volume of ballast water that would act as a surrogate for take under a new practicable

ballast water standard. NMFS would also be re quired to provide the amended Opinion with a

new incidental take statement to the USCG.369 Reasonable and Prudent Measures

Reasonable and Prudent Measures (R PMs) are non-discretionary meas ures to minimize take that may or may not already be part of the description of the pr oposed action. They must be implemented as binding conditions for the exempti on in section 7(o)(2) to apply. The USCG has the continuing duty to regulate the activities covered in this incide ntal take statement. If the

USCG fails to adhere to the terms and conditi ons of the incidental take statement through enforceable terms that are added to this document, or fails to retain the oversight to ensure compliance with these terms and conditions, the protective coverage of section 7(o)(2) may lapse. Activities which do not comply with all relevant RPMs will require further consultation.

NMFS believes that the following RPM is necessary and appropriate to minimize take of listed species and their critical habitat resul ting from implementation of this action.

The USCG shall monitor and report to NMFS Offi ce of Protected Resources various components of the ballast water management program.

Terms and Conditions

To be exempt from the prohibitions of section 9 of the ESA, the action must be implemented in compliance with the following terms and conditions, which implement the RPM described above. These terms and conditions are non-discretionary.

To implement the RPM (monitoring and reporting):

1. Provide the NBIC report summa rizing national patterns of ballast water delivery and
management.

2. Provide results of the practicability review.

3. Provide results of any peri odic programmatic reviews which may be undertaken as

mandated under NISA/NANPCA.

4. Provide an annual report detai ling the number of ships entering US waters, the number of inspections, ballast water vi olations found, and corrective actions taken to fix the

violation.

CONSERVATION RECOMMENDATIONS

Section 7(a)(1) of the Act direct s Federal agencies to utilize their author ities to further the purposes of the ESA by carrying out conservation programs for the benefit of endangered and threatened species. Conservation recommendations are discretionary agency activities to minimize or avoid adverse effects of a proposed ac tion on listed species or critical habitat, to help implement recovery plans, or to develop information.370 1. Address Other Sources of Invasive Species . The USCG should establish mandatory requirements of vessels that will, along with ballast water treatment, more effectively reduce the potential of invasive species. These requirements should focus on hull fouling, sea chests, anchor ch ains, and any other source for a vessel to bring non-native species into waters of the US.

2. Incentives . The USCG, to best implement the goals of NISA and the ESA, should establish economic incentives or grant program s with the goal of creating more stringent ballast water standards by developing new technologies and tes ting procedures.

3. Habitat Improvement Measures . The USCG should impl ement habitat restoration measures in areas where non-native species have established and are harmful to listed species or their critical habitat. The goal of these measures should be to directly counter

the adverse effects of the non-na tive species to mitigate the im pacts to listed species and

their critical habitat.

4. Alternative Discharge Options . The USCG should work with US port owners to create on-site (land based or vessel-ba sed) discharge facilities capab le of receiving ballast water discharges from vessels at the port. The USCG should consider funding these facilities under 7(a)(1) of the ESA in an effort to even tually eliminate ballast water discharges and invasive species introducti ons to waters of the US.

5. Eliminate Discharges . Several vessels were identified in the final PEIS that could operate without discharging ballast water. The USCG should consider making it mandatory that the vessels that can operat e without discharging ballast water always operate without discharging ballast water except when holding ballast water onboard

would create an unsafe situation.

6. Collaboration with EPA . The USCG should continue working with EPA to identify jurisdictional boundaries that can be unclear between the NISA and CWA, both within the federal and state context, so as to maximize compliance u nder NISA. In particular, a major pathway of new invasive species is from hull fouling, yet neither agency is currently regulating the discharges of those non-native propagules.

In order for NMFS to be kept informed of ac tions minimizing or avoi ding adverse effects or benefiting listed species or th eir habitats, the USCG should notify NMFS of any conservation recommendations they implement in their final action.

REINITIATION NOTICE

This concludes formal consultation on the USCG’s National Ballast Water Management Program. As provided in 50 CFR '402.16, reinitiati on of formal consultati on is required where discretionary Federal agency involvement or cont rol over the action has been retained (or is authorized by law) and if: (1) the amount or ex tent of take is exceeded; (2) new information reveals effects of the agency action that may aff ect listed species or cr itical habitat in a manner371 or to an extent not considered in this opinion; (3) the agency acti on is subsequently modified in a

manner that causes an effect to the listed species or crit ical habitat not consid ered in this opinion; or (4) a new species is listed or critical habitat designa ted that may be affected by the action. In the event USCG would authorize, fund, or carry out ballast water di scharge related activities that may affect resources in a manner or to an extent not considered in th is Opinion, USCG must reinitiate consultation to compensate for inform ation that was not available for consideration during this consultation.

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