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 

1 
2

consultation assesses the probable exposure and possible risks associated with the ballast water 

management standard and then reach conclusions on the USCG’s ballast water management 

program generally defined in the final rule (FR) for the allowable concentration 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 attached to the hull of the vessel 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 implementing 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 “destruction or adverse modification of critical habitat” 

at 50 CFR 402.02. Instead, we relied on the statutory provisions of the ESA to complete our 

analysis of the effects of the action 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 conducted 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 requesting initiation of 

consultation, the final Programmatic Environmental 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 final U.S. recovery plans for smalltooth sawfish, 

Atlantic salmon, Upper Columbia River spring Chinook salmon, Puget Sound Chinook salmon, 

Hood Canal chum salmon, gulf sturgeon, shortnose sturgeon, Ozette Lake sockeye salmon, 

Upper Columbia River steelhead, Middle Columbia River steelhead, blue whale, fin whale, 

humpback whale, southern resident killer whale, North Atlantic right whale, sperm whale, 

Hawaiian monk seal, eastern and western Steller sea lions, Northwest Atlantic population of 

loggerhead turtles, Pacific populations of the loggerhead turtle, Atlantic populations of green 

turtles, Pacific populations of the green turtle, hawksbill 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 reports from prior ballast 

water reporting and recordkeeping requirements. 

Consultation History

Prior to discussing the final PEIS and potential 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 that 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 the USCG letter, providing a complete species 

list and requesting a consistent use of the environmental baseline for ESA purposes. 

On March 29, 2011, NMFS and USCG had a teleconference 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 conclusion of not likely 

to adversely affect listed species or their critical habitat. Previous determinations regarding 

ballast water exchange (BWE) had addressed uptake 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 Organization (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 discuss 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 draft request for initiation and asked NMFS 

to provide comments. 

On May 23, 2011, NMFS responded that the information 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 another draft reflecting agreements reached on 

May 24, 2011. 

On August 1, 2011, NMFS provided additional comments 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, signed 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 effect” listed species and their critical habitat, 

but left it up to NMFS to determine which species 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 determination of effects. 
4

A series of emails between November 14, 2011, and November 22, 2011, discussed NMFS’ 

intentions of initiating formal consultation to produce 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 baseline and how that led to a likely to adversely 

affect determination. There was also discussion about the ESA timeline and when consultation 

could be completed. Additionally, all three agencies 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 address 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 after the holidays on January 

5, 2012. 

On December 21, 2011, the USCG requested that NMFS continue consultation under the 

programmatic scope as was initially requested. 

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

case law relevant to the ESA baseline 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 summary 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 met 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 where the disagreement originated and whether the 

agencies could reach agreement. NMFS also requested information related to monitoring and the 

USCG provided information related to the oil water 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 information 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 section 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 they 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 incorporating USCG’s comments, NMFS provided the USCG with a 

second final draft. 

On May 15, 2012, the USCG provided a second round of comments focusing on the difference 

between the practicability review 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 practicability review and periodic review. 

Legal and Policy Framework for Ballast Water Regulation by the USCG 

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), the USCG is the 

lead federal agency in implementing regulations to reduce or prevent the introduction of 

nonindigenous species via shipping activities in United States (US) waters. The purposes of 

NANPCA are to (1) prevent unintentional introduction and dispersal of non-indigenous species 

(NIS) into waters of the US through ballast water management and other requirements, 

(2) coordinate federally conducted, funded, or authorized research; and to federally coordinate 

prevention control, information dissemination, and other activities regarding the zebra mussel 

and other aquatic NIS, (3) develop and carry out environmentally sound control methods to 

prevent, monitor, and control unintentional introductions of NIS from pathways other than 

ballast water exchange, (4) understand and minimize ecological and economic impacts of NIS 

that become established, including the zebra mussel, and (5) establish a program of research and 

technology development and assistance to states in the management and removal of zebra 

mussels. Under NANPCA, Congress directed the USCG to develop a national program of 

specific regulations and guidelines 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 equipped 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 mandated the use of ballast water 

management practices by ships entering the Great Lakes after operating outside the exclusive 

economic zone (EEZ). On December 30, 1994, the USCG published a final rule, “Ballast Water 

Management for Vessels Entering the Hudson River” [59 FR 67632], which expanded the 
6

mandatory requirements for the Great Lakes to portions of the Hudson River (north of the 

George Washington Bridge) that connect to the Great Lakes. The management practices 

mandated for use in the Great Lakes and portions of the Hudson River are found in 33 CFR Part 

151, Subpart C. 

NISA reauthorized and amended NANPCA and mandated the USCG to issue ballast water 

management guidelines and regulations applicable to all US waters. Under NISA, Congress 

authorized the USCG to establish a mandatory ballast water management program [69 FR 

44952], if shipping industry compliance under the initially established 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 prevent or reduce invasive species introductions or 

control infestations, that minimizes adverse impacts to the structure and function of an 

ecosystem, and minimizes adverse effects on nontarget organisms and ecosystems (USCG 2003). 

On May 17, 1999, the USCG published the interim rule 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 requirements remained in effect, and mandated reporting 

requirements. Both the voluntary management guidelines and the mandatory reporting 

requirements applied to vessels entering US waters after operating 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 District of Columbia, Puerto Rico, Guam, American 

Samoa, the US Virgin Islands, and the Trust Territories of the Pacific Islands). 

As a result of several years of public comments 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 management standard. In the 

meantime, the USCG also published the final rule 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 following management practices: (1) 

conduct mid-ocean ballast water exchange 200 nautical 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 ballast 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 specified in the final PEIS, as preventing 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 regulations to establish a type-approval requirement 

for ballast water management systems, which will undergo, if necessary, separate consultation 

tiered from this programmatic consultation. 
7

Actions Authorized by the Program 

The USCG’s ballast water management program proposes to establish a numerical standard and 

allows for a stricter standard to be established via a series of practicability 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 (m3

) of ballast 

water. 

(2) For organisms less than 50 micrometers and greater than or equal to 10 micrometers: 

discharge must include fewer than 10 living organisms per milliliter (mL) of ballast 

water. 

(3) Indicator microorganisms 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 enterococci: a concentration of fewer than 100 cfu per 100 mL. 

Practicability Review and Periodic Review 

NISA requires the USCG to assess and, if necessary, revise USCG’sballast water management 

regulations not less than every three years based onthe best scientific information available at the 

time of that review (16 U.S.C.4711(e)(1). The FR establishes apracticability review, completed 

no later than January 1, 2016, to determinewhether technology and testing protocols are available 

to practicably implementa more stringent ballast water discharge standard. The practicability 

review (33 CFR 151.2030(c)) will evaluate the capability of any identified technology to achieve 

a more stringent ballast water management standard; the effectiveness of any identified 

technology in the shipboard environment; the compatibility 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 technology; the availability, 

accuracy, precision and cost effectiveness of methods to measure concentrations of organisms or 

chemicals per volume of treated ballast water; and 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 management regulations in reducing the introduction 

and spread of aquatic nuisance species by vessels 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/NISA. Understanding the best scientific information available 

changes with the advent of time, the periodic reviews will likely include assessments of the 

practicability of technology or ballast 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 capability of testing methods to allow for enforcement of a more 

stringent standard, the effectiveness of approved technologies in the shipboard environment, and 
8

whether the use of any new technologies may have an adverse impact on the environment. To 

fulfill the NISA/NANPCA statutory mandate of ensuring 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 assessment practices that could be implemented to 

achieve greater reductions in the risks of vessel-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). These 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). Information on ballast water management must be 

retained by the owner and/or operator for 24 months, with the option of retaining information 

electronically for all but the most recent six months. 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. The 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 information, information on ballast water to be 

discharged, information on sediment to be discharged, 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 tonnage, 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 water information will be total ballast water capacity, 

total volume of ballast water onboard, total number of ballast tanks, and total number of ballast 

tanks in ballast. The required ballast water management 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 tank, and identify whether there is a ballast 

management plan onboard and whether it was used. The required information on ballast water to 

be discharged will be the origin of the ballast water, including dates, locations, volumes, and 

temperatures; treatment information will include dates, locations, volumes, method, percent 

exchanged (if using ballast water exchange), and sea 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 location of the facility 

where sediment disposal will take place. 

Compliance Monitoring 

The USCG identifies the components of their compliance monitoring in 33 CFR 151.2075. The 

compliance monitoring allows the USCG to verify that vessels are equipped with approved 
9

technology. This section of the FR also authorizes 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 meeting the ballast water numerical standard. This 

monitoring is established to function in essentially the same way the oil-water separator 

monitoring program works. 

The USCG has been monitoring vessels under the 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 ballast water treatment 

method, the USCG sampled ballast tanks to determine if the water in the 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, and continue to conduct on-board monitoring to verify that ships 

are complying with the ballast water management program. 

All vessels are required to install 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 technologies capable of meeting the proposed ballast water standard 

were installed. Two additional goals 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 water and sediment samples from vessel ballast tanks to ensure 

the technology is being maintained and functioning properly after installation. 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) describing appropriate mitigation measures that 

ships should take in addition to all required ballast water 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 regularly, 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 chains, and remove fouling organisms from vessel 

hulls, pipes, and tanks regularly. 

Limitations of Coverage 
10

Certain vessels, including offshore floating platforms, are exempted from this FR by NISA. The 

vessel exemptions are identified 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 water standard, reporting, or record keeping. 

Additionally, non sea going vessels; sea going vessels of less than 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 form to the USCG, but non-exempted vessels are 

required to report, regardless of whether and how they discharged ballast water. 

Implementation Dates 

Non-exempt vessels are required to meet the ballast water management standard as early as 

December 1, 2013 and all others by the first scheduled 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 capacities of 1,500-5,000 m3

) 

or January 1, 2016 (vessels with ballast water capacities of less than 1,500 m3

 and greater than 

5,000 m3). 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 penalties are codified in the NISA and the FR that may be enforced 

against violators of the statute 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 analyses 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, including changes in the spatial extent over time. The 

results of this step represent the action area for the consultation. The second step of our analyses 

identifies the listed and proposed species and designated or proposed critical habitat that are 

likely to co-occur with these effects 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 stage, and gender of these resources that are likely 

to be exposed to an action’s effects and the populations or subpopulations those individuals 

represent. Once we identify the listed and proposed resources that are likely to be exposed to an 

action’s effects and the nature of that exposure, we examine whether and how those resources are 

likely to respond given their exposure (these represent our response analyses). 

The final steps of our analyses—establishing the risks those responses pose to listed and 

proposed resources—are different for listed species and designated critical habitat (these 

represent our risk analyses). Our jeopardy determinations 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, subspecies, or distinct population segments of 

vertebrate species. Because the continued existence of listed species depends on the fate of the 

populations that comprise them, the viability (probability of extinction or probability of 

persistence) of listed and proposed species depends 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 comprise 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 modification determinations must be based on an action’s effects on the 

conservation values of the essential features of critical habitat. 

A programmatic review, however, typically analyzes the general environmental consequences of 

a broad scope of policy alternatives under consideration by an agency program, in this case, 

managing a ballast water program to aid in controlling the introduction and spread of non-native 

species from ship’s ballast water. Subsequent consultations 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 ballast 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 ballast water management systems could be used to 

achieve the ballast water standard, so tiered consultations will evaluate any further impacts that 

may occur due to chemical, biological, or mechanical treatment of ballast water. All subsequent 

section 7 consultations conducted by NMFS personnel will be designed to determine whether 

and to what degree the specific action under review fits within the general pattern identified in 

the “parent” or national programmatic consultation, 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 consultations by recognizing that an agency’s program normally 

represents the agency’s decision to authorize, fund, or carry out a suite or class of activities that 

may require subsequent decision-making. When we conduct programmatic examinations of 

proposals such as the USCG’s national 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, standards and criteria that NMFS considers during 

consultations on specific actions. We also ask whether 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 endangered or threatened individuals, increase the 

extinction risks of the population(s) those individuals represent, or increase the extinction risks 

of the species those populations comprise. Specifically, we ask: 

1. Has USCG structured the program so they will know or be able to reliably estimate the 

probable individual and cumulative effects of the action? For example, has USCG structured the 

program so it will know or be able to reliably estimate the probable number of non-native 

organisms that will be introduced under the authority of the program? Has USCG structured the 

program so it will know or be able to reliably 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 authorized 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 with the conditions, restrictions, 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 designated critical habitat are likely 

to be exposed to (a) chemical, physical, or biological stressors associated with ballast water 

management, or; (b) to the ecological consequences (habitat alterations or indirect effects to 

species) as a result of ballast water management in the United States? 

5. Has USCG structured the program so it will continuously identify, collect, and analyze 

information to detect impacts from ballast water management to endangered or threatened 

species or designated critical habitat 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 organisms or primary constituent elements of critical habitat? 

6. Has USCG structured the program to employ an analytical methodology that considers (a) 

the status and trends of endangered or threatened species or designated critical habitat; (b) the 

demographic and ecological status of populations and individuals of those species given their 

exposure to pre-existing stressors in different drainages and watersheds; (c) the direct and 

indirect pathways by which endangered or threatened species or designated critical habitat might 

be exposed to the consequences of organisms carried in ballast to waters of the United States; 

and (d) the physical, physiological, behavior, sociobiological, 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-existing 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 concentrations, durations, or frequencies that are potentially harmful to 

individual listed organisms, populations, or the species, or; (b) to ecological consequences that 

are potentially harmful to individual listed organisms, populations, the species or primary 

constituent elements of designated critical habitat? How quickly 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 habitat exposure to stressors authorized by the 
13

proposed action. 

Our assessment focused on whether and to what degree USCG structured the program to 

minimize endangered or threatened species or critical habitat that has been designated for those 

species from being exposed to stressors resulting from ballast discharges because such exposures 

commonly trigger a cascade of events with ultimate 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 associated with the exposure, their 

sensitivity to the stressor, other physical, chemical, or biotic stressors that they are exposed to 

concurrently, their pre-existing physiological state, and their individual fitness. 

When individual listed plants or animals experience reductions in fitness, those reductions may 

reduce the abundance, reproduction rates, or growth rates (or increase variance in one or more of 

those rates) of the populations those individuals represent (see Stearns 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 viability of the populations those individuals 

represent or the species those 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 included sufficient safeguards to ensure that the 

actions they authorize, fund, or otherwise carry out would not reduce 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 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 population 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 species or destroy or adversely modify their critical 

habitat, we assess the probable consequences of the specific actions that will be authorized, 

funded or carried out by the program 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 numerical ballast water standard; (b) the geographic 

distribution of species introductions via ballast water; (c) the frequency of non-native species 

introductions originating from similar geographic areas; (d) the stressors typically associated 
14

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 discharges, listed species 

and designated critical habitat. 

These analyses describe spatial overlap and any specific evidence (reports or studies) that 

particular endangered or threatened species or designated critical habitat have been or are likely 

to be exposed to the chemical, physical, or biological stressors resulting from those activities. 

However, this does not represent a detailed exposure analysis: we are merely establishing 

whether or to what degree endangered or threatened species or designated 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). Given spatial and temporal overlap, we then have 

reason to ask whether or to what degree USCG's program can insure that these species or critical 

habitat are not likely to be exposed. 

3. We conduct a detailed review of the literature available on the physical, physiological, 

behavioral, social, and ecological responses of endangered or threatened 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 than discuss the literature for each species, we 

organized the data using species groups (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 threatened species and designated critical habitat. 

Rather than discuss the literature for each species, it is only necessary to discuss the risks of 

exposing species groups (for example, Pacific salmon; sturgeon; sea turtles; etc.). 

In this Opinion, we will present the results of these analyses before we present the results of our 

review and evaluate the program using the sequence of seven questions we identified 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 designated for those species. 

Evidence Available for this Consultation 

To conduct our analyses we considered the information contained in the USCG’s letter 

requesting initiation of consultation, the final PEIS, the FR, 33 CFR 151, NANPCA, NISA, and 

monitoring data from previous ballast water rules. We supplemented this information using 

electronic searches of literature published in English 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 sites, doctoral dissertations and master’s theses for literature 

on the biological, ecological, and medical sciences. Particular databases we searched for this 

consultation included BasicBiosis, Dissertations, ArticleFirst, Proceedings, Aquatic Sciences 

and Fisheries Abstracts, BioOne Abstracts & Indexes, and Water Resources Abstracts. Some of 
15

the databases provide access to documents published from the 1960s through present, although 

references for many scientific journals contained 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 trends of the threatened and endangered species considered in 

this Opinion, and the impacts of invasive species on freshwater, estuarine, 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 titles and abstracts, appeared to comply with our keywords. If a 

reference’s title appeared relevant, we acquired and reviewed the reference. We supplemented 

our electronic searches by searching the literature cited sections and bibliographies of references 

we retrieved electronically to identify additional papers that had not been captured in our 

electronic searches. 

Collectively, this information provided the basis for our determination as to whether and to what 

degree listed resources under our jurisdiction are likely to be effected by the USCG’s ballast 

water management program, and whether and to what degree the USCG can insure that its ballast 

water management program is not likely to jeopardize the continued existence of proposed, 

endangered, or threatened species or result in the destruction or adverse modification of proposed 

or designated critical 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 technologies to understand the best available treatment 

technologies at the time of the proposal. Based on the feasibility of the technologies, the ballast 

water standard is established. The USCG also conducts annual 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 national 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 and practicability review that the USCG analyzes 

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

ballast water standards. We evaluated this information to determine 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, individually or cumulatively, jeopardize the continued existence of 

proposed or listed species or result in the destruction or adverse 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 Pro,rom l 

Periodic Tiered ESA Con.u t~ ion 

"",w I~ / Re ,n,tia te Co ul t~t n 

\ As.eo. T:Chno loiV / 

b .. o Chinee/A<:tion I l APP,olle Te ch no loe v j 

Propos e Pr acticable / 

BWD5 

Mo re rotec tl~e 

MeHure< Ale l ESA Proer~mm~tic Rev iew J 

Ac hievible, 

Practic able 

-------. ~me Tec hno loiY 

Not Meetine 

Numeric II t BWDS Rep eued l)' 

ter Stind.,d 

Eq uip Vessels With 

App roved Tech noloilV 

I \. 

Complia nce Effect ive ne ss 

Mo nitori nl Mon itorinl 

./ ~ Does WUer/Sedlme M 

l Are Vesse ls Eq uipped? . ... I, Tec hno loty ' - " . ...... - Samplin e Suppo rt Tech- " '. Seinli Us ed? . '-'- / .... noloe)' Ach ie vine BWDS? 

)( • " ' -.. , ... ", 

Enforc eme nt .-----8"" '" .... ;.",-,,; ..... , .... 
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 insured 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 threatened species, or result in the 

destruction or adverse modification of proposed or designated critical 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 program. 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 alternative, we assessed the ballast water program’s monitoring 

programs: three-year periodic reviews, compliance monitoring, effectiveness monitoring, and 

environmental monitoring. The results of those monitoring actions will feed back into the 

decision making process for future potential numerical standards. 

In addition, we asked: how does the USCG determine a numerical standard, how frequently 

does the USCG conduct program reviews, what is the criteria that triggers an adaptive 

management adjustment to ballast water standards, 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 determine the technology still meets the 

standard after it has been installed, is the USCG able to determine if the technology is able to 

meet the standard after having been used for years, is the USCG able to determine whether 

implementation of the numerical standard has effectively reduced or prevented new invasive 

species, and what is the ability of the USCG to monitor newly introduced species? 

Through the course of this consultation we learned that the decision to identify the proposed 

ballast water standard was dependent on the “maximum extent practicable” mandate of NISA. 

This decision is informed by vessel and human safety, water quality considerations, NANCPA, 

NISA, economic impacts to commerce and society, 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 standard. 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 the responses of listed species, we will consider 

both listed species and introduced species as functional groups, such as salmonids, pinnipeds, 

and sea turtles or diseases, predators, and competitors, respectively. Once we have established 

the likely rate of exposure to listed species, using the response analysis, we will determine the 

proportion of invasive species that would likely be harmful. Using the information from the 

exposure and response analyses, we can determine the risks ballast water discharge at the 

proposed standards may pose to listed species and their critical habitat in order to reach a 
18

conclusion about the USCG’s program and its ability to insure the USCG is not likely to 

jeopardize any listed species or destroy/adversely modify their critical habitat. 

During this consultation, NMFS considered the various ballast water discharge standards 

proposed by different states and authorized, funded, or carried out by other Federal agencies 

under different legal mandates. NMFS did not include those other actions in the analysis, but 

instead evaluated: whether the approved state standards are achievable, the outcomes of legal 

action against those enacted state standards, and the approval process for EPA vessel discharge 

permit program in order to understand how the 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 the USCG determines there 

may be an effect as a result of new technologies being developed to achieve the proposed 

numerical standard, NMFS will evaluate the effects of those technologies on proposed and listed 

resources. However, future consultations with EPA on its vessel 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 management program. Even if the 

ballast water 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), the reinitiation criteria 

at the end of this Opinion could require 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 ballast water at US ports with specific 

emphasis on San Francisco Bay, Puget Sound, and Chesapeake 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 listed individuals may be exposed to uptake of 

ballast water or non-native introduced species, but (a) the probability of exposure 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 individual listed species would respond to exposure, (c) there is no 

possibility or only a small probability of a negative response even if there is a response to uptake 

in ballast water tanks or to introduced non-native species, or (d) there is no possibility or only a 

small probability that the individual would experience a reduction in individual performance (or 

fitness), then we concluded that the USCG’s action is “not likely to adversely affect” those 

proposed or listed resources. 

If proposed or listed resources or their critical habitats may be lethally or sub-lethally affected by 

actions the USCG’s ballast water management program authorizes, funds, or otherwise carries 

out, NMFS examines if the program includes sufficient safeguards to insure that the incidental 
19

take of individuals does not occur 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 population’s extinction risks). Given their status and the environment 

in which the species exist, are those species likely to be adversely affected 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 the “Action Area” of a Federal action as all areas 

to be affected, directly or indirectly, and not merely the immediate area involved in the action 

(50 CFR 402.02). This Opinion assesses the consequences of the USCG's establishment of 

ballast water standards in all waters of the US. The USCG's proposed rule establishes the action 

area as all navigable waters within the US extending offshore to the outer edge of the EEZ, 200 

miles offshore of the US coastline except where this region overlaps with the EEZ of another 

nation (Mexico, Canada, Russia, Cuba, the Federation 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. Asterisks denote critical habitat in the action area. 

Double asterisks denote proposed critical habitat in the action area.

Common name (Distinct population 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 Turtles 
20

Common name (Distinct population 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 (Northern 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) Endangered 
21

Common name (Distinct population 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 critical habitats that occur 

exclusively (or nearly so) in offshore or deepwater marine environments. Based upon review of 

available literature detailed 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 spread to offshore habitats. Inshore, estuarine, 

and freshwater habitats are fundamentally different in ecological terms and we discount the 

ability of an invasive species to establish in offshore 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 whale critical habitat. 

Several listed or proposed resources also occur exclusively (except for extralimital occurrences) 

in areas where ballast water discharge is so limited that the probability of establishment of nonnative species and subsequent exposure of listed resource is discountable. 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 DPS 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 cosmopolitan 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 breed 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 Iceland to Norway in the Barents Sea. Whales 

migrate to the western coast of Africa and the Caribbean Sea during the winter. Humpback 

whales aggregate in four summer feeding areas: 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 Mediterranean 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 breeding aggregations occur off the Greater 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 North 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 “open” structure of humpback whales, a humpback 

whale migrated from the Indian Ocean to the South Atlantic Ocean, demonstrating that 

interoceanic movements can occur (Pomilla and Rosenbaum 2005). Genetic exchange at lowlatitude breeding groups between Northern and Southern Hemisphere individuals and widerrange movements by males has been suggested to explain 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-identification studies, the NMFS currently 

recognizes four stocks, likely corresponding to populations, 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, Tomilin 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 population winters in the waters around Hawaii while the eastern North 

Pacific population (also called the 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 areas of other populations, 

highlighting the potential fluidity of population structure. Humpback whales were recently found 

to migrate to the northwestern Hawaiian Islands, 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 Hawaiian 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 further mixing on feeding areas in Alaska) and 
23

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, Ogasawara 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 1998a, Eldredge 2003). During summer, whales 

from this population migrate to the Kuril Islands, Bering Sea, Aleutian Islands, Kodiak, 

Southeast Alaska, and British Columbia to feed. 

Humpback whales from both the eastern, western, 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 year-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 Island in recent surveys, with a density of 54 

individuals per 1,000 km2

 (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, Pakistan, 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 calving 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 from these stocks winter and breed in separate 

areas and are known to return to the same areas. However, the degree (if any) of gene flow (i.e., 

adult individuals wintering in different breeding locations) is uncertain (Carvalho et al. 2011). 

Individuals from breeding grounds in Ecuador are 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 coast 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 genetic 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 groups). 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 generally occurs during winter at lower 

latitudes. Gestation takes about 11 months, followed 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 (Glockner-Ferrari and Ferrari 1985 as cited in NMFS 

2005b, Clapham and Mayo 1987, 1990, Weinrich et al. 1993). Males appear 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 male dominance polygamy 

(Clapham 1996). Calving occurs in the shallow coastal waters of continental shelves and 

oceanic islands worldwide (Perry et al. 1999). Males “cort” females 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 appear to exist between 

males and females, mature females do pair with other females; those individuals 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 various 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 Atlantic 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, Clapham 1996). Humpback whales are generally believed to 

fast while migrating and on breeding grounds, but some individuals apparently feed while in 

low-latitude waters normally believed to be used exclusively for reproduction and calf-rearing 
25

(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 support 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 Island is two to four individuals, although larger groups 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 listed as endangered in 1970 (35 FR 

18319), and this status remains under the ESA. (Winn and Reichley 1985) argued that the global 

humpback whale population consisted of at least 150,000 whales in the early 1900s, mostly in 

the Southern Ocean. In 1987, the global population of humpback whales was estimated at about 

10,000 (NMFS 1987). Although this estimate is outdated, 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 North Atlantic abundance comes from 1992-

1993 mark-recapture data, which generated an estimate of 11,570 humpback whales (Stevick et 

al. 2003a). Historical estimates have ranged from 40,000-250,000 (Smith and G.Pike 2009). 

Smith and Reeves (2010) estimated that roughly 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 Nichols 1982). Several researchers report an 

increasing trend in abundance for the North Atlantic population, which is supported by increased 

sightings within the Gulf of Maine feeding 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 estimated 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 that the eastern and 

northeastern waters off Iceland are areas of significant humpback utilization for feeding, 

estimating nearly 5,000 whales in 2001 and proposing an annual growth rate of 12% for the area. 

The authors suggest that humpback whales in the area had probably recovered from whaling. 

However, recent data suggest that the upward growth 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 estimates are 6,000-8,000 whales (Rice 1978, Calambokidis et al.

1997). It is estimated that 15,000 humpback whales resided in the North Pacific in 1905 (Rice 

1978). However, from 1905 to 1965, nearly 28,000 humpback whales were harvested in whaling 

operations, reducing the number of all North Pacific humpback whale to roughly 1,000 (Perry et 

al. 1999). Estimates have risen over time from 1,407-2,100 in the 1980s to 6,010 in 1997 (Baker 

1985, Darling and Morowitz 1986, Baker and Herman 1987, Calambokidis et al. 1997). Because 

estimates vary by methodology, they are not directly comparable and it is not clear which of 
26

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 model assumptions. Tentative estimates 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). However, 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 whales, not counting calves (Barlow et al. (2009) provided a 

bias-corrected estimate of 20,800 individuals) and the population was growing at 4.9% annually. 

Almost half of these whales likely occur in wintering areas around the 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 present humpback whale abundance.

*Note: Confidence Intervals (C.I.) not provided by the authors were calculated from Coefficients of Variation (C.V.) 

where available, using the computation 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 likely 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, or

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 of

Scotian Shelf

-- -- 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 Arabian 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 60o

S -- -- 4,660 2,897-6,423 (IWC 1996)
27

Southern Hemisphere. Approximately 42,000 Southern Hemisphere humpbacks can be found 

south of 60° S during the austral summer feeding 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 humpback whales were harvested from Antarctic 

waters from 1946 through 1973, largely from management areas IV, V, and VI (Clapham et al.

2009). One-third of these catches occurred from 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 breeding 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 individuals divided amongst various 

populations/subpopulations (Constantine et al. 2010). A 2009 spike in calf mortality along 

western Australia brings into question whether 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 summering 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 included 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 Australia to be 10.9 and 10.1% annually, respectively (0.23 and 4.69 SE, 

respectively). 

Natural threats. Natural sources and rates of mortality of humpback whales are not well 

known. Based upon prevalence of tooth marks, attacks 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 (Steiger et al. 2008). 

Juveniles appear to be the primary age group targeted. Humpback whales engage in grouping 

behavior, flailing tails, and rolling 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 with attack (Ford and Reeves 2008). 

Parasites and biotoxins from red-tide blooms are 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 whales that stranded along Cape Cod between 

November 1987 and January 1988 indicate they apparently died from a toxin produced by 

dinoflagellates during this period. 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. Historically, whaling represented the greatest threat 

to every population of whales and was ultimately responsible for listing several species as 

endangered. 

Humpback whales are also killed or injured during 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 fisheries 
28

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 entangled 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 and 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 humpback whales that stranded along the Atlantic 

coast of the US between 1975 and 1996, 10 (8.1%) showed 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 rules 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 been identified from humpback whale blubber 

(Gauthier et al. 1997b). Higher PCB levels have been observed 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 Gulf of Maine and off Southern California tend to have the 

highest PCB concentrations, overall levels are on par with other 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 bioaccumulating 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 compared to blue whales. Humpback whales feed 

higher on the food chain, where prey carry higher 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 whales compose a single population. 

Although not all individuals undergo the same migratory pattern, no subpopulation structuring 

has been identified. 

Distribution. Right whales occur in sub-polar to temperate waters in all major ocean basins in 

the world, with a clear migratory pattern of high latitudes in summer and lower latitudes in 

winter (Cummings 1985, Rice 1998b, Perry et al. 1999). The historical range of North Atlantic 
29

right whales extended as far south 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 Atlantic 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 were designated as critical habitat for the North Atlantic right 

whale. 

North Atlantic right whales have been observed from the mid-Atlantic Bight northward through 

the Gulf of Maine year-round, but are primarily found along the northeast 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 Cod 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 Bay 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 general patterns of movement by reducing peak 

numbers, stay durations and specific locales (Brown et al. 2001, Kenney 2001). In particular, 

large changes in the typical pattern of food abundance will dramatically change the general 

pattern of North Atlantic right whale habitat use (Kenney 2001). 

Migration and movement. North Atlantic right whales exhibit 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/December. 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 within 30 nautical miles of the coastline and in waters less than 

160 feet deep. Although this pattern 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 North Atlantic right whales far offshore, including 

those from Dutch ships indicating some individuals occur between 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 offshore (greater than 30 miles) during surveys flown 

off the coast of northeastern Florida and southeastern 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 Maine into deeper waters off the continental shelf 

(Mate et al. 1997). The frequency with which North Atlantic right whales occur in offshore 

waters in the southeastern US remains unclear. Occasionally, individuals are observed in distant 

locations, including the Gulf of Mexico, Bermuda, the Gulf of St. Lawrence, Newfoundland, 
30

Greenland, Iceland, and northern Norway (an area known as a historical 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 habitat areas not presently well described. 

Reproduction and demography. Data through the 1990s suggests that mean calving interval 

increased since 1992 from 3.67 years to more than five years, a significant 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 studied populations of southern right whales (Best et 

al. 2001b). This has been attributed to several possible causes, including higher abortion or 

perinatal losses (Browning et al. 2009). An analysis of the age structure of North Atlantic right 

whales suggests that the population contains a smaller proportion of juvenile 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 reproductive senescence on the part 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 accidentally and soon after 

birth) and subsequent adoption through weaning has been found (Frasier et al. 2010). Although 

North Atlantic right whales historically separated from their calves within one year, a shift 

appears to have taken place around 2001 where mothers (particularly less experienced mothers) 

return to wintering grounds with their yearling 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 foraging activities where physical oceanographic features 

such as water depth, current, and mixing fronts combine to concentrate copepods (Wishner et al.

1988, Murison and Gaskin 1989, Mayo and Marx 1990, Baumgartner et al. 2003). 

Feeding. North Atlantic right whales fast during the winter and feed during the summer, 

although some may opportunistically feed during migration. North Atlantic right whales use 

their baleen to sieve copepods from dense patches, found in highly variable and spatially 

unpredictable locations in the Bay of Fundy, Roseway 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 zooplankton, especially shrimp-like copepods such as Calanus

(Kenney et al. 1985a, Beardsley et al. 1996). North Atlantic right whales feed largely by 

skimming these prey from the ocean surface (Pivorunas 1979, Mayo and Marx 1990), but may 

feed anywhere in the water column (Watkins and Schevill 1976, 1979, Goodyear 1993, Winn et 

al. 1995). Feeding behavior has only been observed in northern areas and not on calving 

grounds or during migration (Kraus et al. 1993). 

Status and trends. The Northern right whale was originally 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 evidence that North Atlantic and 

North Pacific right whales are separate species. Following a comprehensive status review, 
31

NMFS concluded that North Atlantic and North Pacific 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 abundant, with an estimated 5,500 individuals present 

in the 16th century throughout the North Atlantic (Reeves 2001, Reeves et al. 2007). A review of 

the photo-id recapture database in June 2006, indicated that only 313 individually recognized 

North Atlantic right whales were observed during 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 calculated for the population. 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 recovery. However, work by Caswell et al. (1999) suggested 

that crude survival probability declined from about 0.99 in the early 1980’s to about 0.94 in the 

late 1990s. Additional work conducted in 1999 showed that survival had indeed 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 suggested 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 Atlantic right whales overlapped with the foraging 

grounds of sei whales. Both species feed preferentially on copepods. Reeves et al. (1978) noted 

that several species of whales feed on copepods in the eastern North Pacific, so that the foraging 

pattern and success of right whales would be affected by other whales as well. Mitchell (1975) 

argued that the North Atlantic right whale population had been depleted 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 competition with the sei 

whale population impedes or prevents the recovery of the right whale population. 

Other natural factors influencing right whale recovery are possible, but unquantified. Right 

whales have been subjects of killer whale attacks and, because of their robust size and slow 

swimming speed, tend to fight killer 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 significant problems in the recovery of right whales because of their small 

population size. 

Anthropogenic threats. Several human activities are known to threaten North Atlantic right 

whales: whaling, commercial fishing, shipping, and environmental contaminants. Historically, 

whaling represented the greatest threat to every population of right whales and was ultimately 

responsible for listing right whales as an endangered species. As its legacy, whaling reduced 

North Atlantic right whales to about 300 individuals 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 whales, entanglement in commercial fishing gear 
32

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 North 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 five 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 slower 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 less 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 its effect on the ability of North Atlantic right 

whales to recover (Greene et al. 2003b). Specifically, the variations in oceanography resulting 

from current shifts and water temperatures can significantly affect the occurrence of the North 

Atlantic right whale’s primary food, copepod crustaceans. 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 distribution, 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 marine mammals, are exposed to numerous toxins in 

their environment, many of which are introduced by humans. Levels of chromium in North 

Atlantic right whale tissues are sufficient to be mutagenic and cause 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 samples 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 than 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 right whale health and recovery (Weisbrod et al. 2000). Flame 

retardants such as PBDEs (known to be carcinogenic) have also been measured in North Atlantic 

right whales (Montie et al. 2010). 

Critical habitat. Critical habitat is designated 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 critical habitats for feeding cover portions of the Great South 

Channel (east of Cape Cod), Massachusetts Bay and Cape Cod Bay, and Stellwagen Bank. 

Northern critical habitat was designated because of the concentration of right whales that feed in 

the area, apparently associated with complex oceanographic features that drive prey density and 

distribution. This area has come under considerable scrutiny within 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 habitat where North Atlantic right whale mortality 
33

occurs. Shipping traffic has generally increased in the recent past and could be considered to 

degrade the habitat due to the additional mortality and injury risk now present in the area. 

Although voluntary regulations are in place, these are frequently ignored and mandatory 

regulations are under consideration. The southern critical habitats are along Georgia and 

northeastern Florida coasts (waters from the coast out 15 nautical miles between the latitudes of 

31°15’ N and 30°15’ N and from the coast out five 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 generally 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 whales 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 Washington, Oregon, and 

Vancouver Island and are known to travel as far south 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 mobile 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 Charlotte 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, Strait 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 Islands, 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 differ 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 distribution are largely unknown for the Southern Resident population. 

While the Southern Residents are in inland waters during the warmer months, all of the pods 

concentrate their activities in Haro Strait, Boundary 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 differences 
34

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 Southern Resident killer whales relate to those of 

their preferred prey, salmon. Pods commonly seek out and forage in areas where salmon occur, 

especially those associated with migrating salmon (Heimlich-Boran 1986, 1988, Nichol and 

Shackleton 1996). 

Members of different pods do interact, but members generally remain within their matrilinear 

group (Parsons et al. 2009). However, additional interaction between pods has occurred over the 

past two decades, possibly in association with the decline of the Southern Resident population as 

a whole (Parsons et al. 2009). 

Feeding. Southern Resident killer whales are fish 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 show 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 because 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 (Ford and Ellis 2006). Throughout inland waters from May to 

September, Southern resident killer whale 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 Southern 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 Oregon 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 killer whales appear to be 

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

An area to the southwest of San Juan Island 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 greater calving 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 between 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 believed to occur from May to 

October (Nishiwaki 1972, Olesiuk et al. 1990a, Matkin et al. 1997). However, conception 
35

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 offspring 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 have been listed as endangered since 2005 

(70 FR 69903). In general, there is little information available regarding the historical 

abundance of Southern Resident killer whales. Some evidence suggests that, until the mid- to 

late-1800s, the Southern Resident killer whale 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 Southern Resident population resembles that of the 

Northern Residents (Barrett-Lennard 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 smaller numbers along Washington’s outer coast. 

Olesiuk et al. (1990a) estimated the Southern Resident population 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 has continued to fluctuate in numbers. After 

growing to 98 whales in 1995, the population declined by 17% to 81 whales in 2001 (-2.9% per 

year) before another slight increase to 84 whales in 2003 (Ford et al. 2000, Carretta et al. 2005). 

The population grew to 90 whales in 2006, although 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, unstable population status, and population structure (e.g., 

few reproductive age males and non-calving adult females) continue to be causes for concern. 

Moreover, it is unclear whether the recent increasing trend will continue. The relatively low 

number of individuals in this population makes it difficult to resist/recover from natural spikes in 

mortality, including disease and fluctuations in prey availability (NMFS 2008f). Although 

disease outbreaks have not been identified in this population, increased contaminant load (see 

below) may increase the susceptibility 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 of 
36

salmon along the West Coast (NRC 1996c, Slaney et al. 1996, Gregory and Bisson 1997, 

Lichatowich 1999, Lackey 2003, Pess et al. 2003, Schoonmaker et al. 2003). Several salmon 

species are currently protected under the ESA, and are generally well below their former 

numbers. A 50% reduction in killer 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 drainage 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 bioaccumulate these toxins in their tissues, potentially 

leading to numerous physiological changes such as skeletal deformity, lowered disease 

resistance, and enzyme disruption (Krahn et al. 2009). Presently, the greatest contaminant 

threats are organochlorines, which include PCBs, pesticides, dioxins, 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 petroleum industry in Puget Sound, which has the low 

potential to create a catastrophic oil spill, or more likely, small but chronic releases of 

petrochemicals. As southern resident killer whales are normally exposed to high levels of whale 

watching, engine exhaust has been assessed asa possible threat and, in some cases, may 

contribute to health affects (Lachmuth et al. 2011). 

Vessel activity also has been identified 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 whale-watching vessels. In 2005, a US vessel 

participating in sonar exercises apparently caused significant behavior changes 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 potential remains for these disruptions to occur, or 

as in other areas, the potential for auditory trauma, stranding, and death. The increase in 

“background noise” resulting from vessel traffic and coastal development activities, although not 

directly traumatic, has the potential to influence or disrupt the acoustic system that Southern 

Resident killer whales use to navigate, communicate, 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 whale-watching in the region focuses primarily on 

Southern Resident killer whales and has increased dramatically 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, stress, 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 whales was designated 

on November 29, 2006 (71 FR 69054). Three specific areas were designated; (1) the Summer 
37

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 approximately 2,560 square miles of marine habitat. 

Three essential factors exist in these areas: water quality 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 passage conditions to 

allow for migration, resting, and foraging. Water quality has declined in recent years due to 

agricultural run-off, urban development resulting in additional treated water discharge, industrial 

development, oil spills. The primary prey of southern 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 growing 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 widely distributed in Arctic and subarctic waters, 

and in Alaska five putative populations exist (Beaufort 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. Mitochondrial 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 demonstrating the strong evidence of genetic isolation from the 

other Alaskan populations and other populations demonstrating weak 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-round 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 estuaries 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 focus 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 sightings 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 appeared to increase during winter 

months, and exhibited more widely dispersed patterns both within and among individuals (Hobbs

et al. 2005). Distribution during all months is likely influenced by prey distribution, where 

salmon and eulachon are concentrated in river mouths during summer months 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 wider dispersal of the whales. 

Based on past studies of the summer 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 recent surveys (1993-2007) observed more than 

90% of the beluga whales in upper Cook Inlet 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 of 
38

Anchorage (Point Woronzof), whereas more recently (1998-2007) 90% were detected within 20 

nmi. Although the precise reason for the range contraction is not known, the shrinking summer 

distribution likely reflects the reduction in the population size over the same intervals and the 

beluga whales’ preference for dense aggregations of preferred prey species. Through the 

distribution has shrunk since the 1970s, possibly due to human hunting in excess of 50 

individuals per year in the northern 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 northern reaches of Cook Inlet is 

superior habitat to southern areas, even with the anthropogenic stressors present there (Rugh et 

al. 2010). Overall, Cook Inlet beluga whale range has shrunk from greater than 7,000 km2

 to 

less than 3,000 km2 (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 experiences some of the 

most extreme tidal fluctuations in the world (see Hobbs et al. (2008) for a discussion), and 

beluga whales in the inlet have adapted to these tidal cycles and seemingly take advantage of 

them, although the precise causal reasons are not well known. Presumably, the feeding 

opportunities these tidal cycles proffer the beluga whale are a contributing factor. Beluga whales 

move further into inlets and arms during tidal floods and move back out during ebb tides in a 

relatively predictable manner, although highly-localized differences do exist (Ezer 2010). Tidal 

state is also believed to influence travel and milling behavior (Garner and Mckee. 2010). Aerial 

surveys and predictive models of habitat us indicate 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 habitat attributes to causative reasons for this 

preference. Besides feeding, studies have suggested this preference for tidal mudflats may also 

be attributed to calving and breeding, molting, or shelter from predators 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 produce a single calf every 2 

to 3 years following a 14-month gestation. Of ten 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 (Calkins 1983, Huntington 2000a, Mcguire et al. 2010a). Areas 

specific to calving and rearing have 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 northern side of Kachemak Bay in April and May, off the mouths 

of the Beluga and Susitna Rivers in May, and in Chickaloon Bay and Turnagain 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. Neonates 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 calving 
39

period. 

Calculation of beluga whale age is based on growth layers in teeth. Some debate exists as to 

whether a beluga whale tooth contains 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 layers rather than years (Table 3 from (Hobbs et al.

2008)). 

Table 3. Review of Female beluga life history parameters 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 1st

 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)c
40

Parameter Data Sources 

 0.11d (Burns and Seaman 1986b) 

 0.13d (Sergeant 1973) 

 0.09d

 (Brodie 1971) 

 0.09-0.12 d

 (Braham 1984) 

 0.09-0.14e

 (Braham 1984) 

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

 0.08-0.14e (Davis and Evans 1982) 

 0.06-0.10e (Davis and Finley. 1979) 

 0.08-0.10e

 (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 years of age (Burns and Seaman 1986a). 

Feeding. Analyses of beluga whale stomach contents indicate that beluga whales are 

opportunistic feeders (including including octopi, squids, crabs, shrimps, clams, mussels, snails, 

sandworms, and fishes such as capelin, cod, herring, smelt, flounder, sole, sculpin, lamprey, 

lingcod, salmon, trout, whitefish, northern pike, grayling, and 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, are 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 entering the estuary in May and again in June, and 

represents a significant biomass of prey, with estimates of several thousand fish entering the 

river in the first wave and several million entering 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 burned like a candle (Payne et al. 1999). This high fat content confers a 

significant source of energy for beluga whales, including 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 contained 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 composing the majority of the fish eaten (Hobbs

et al. 2008). Anadromous salmonids begin concentrating at the river mouths and intertidal flats 

in upper Cook Inlet in late spring and early summer 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 the greatest percent frequency of occurrence of the 

prey species found in Cook Inlet beluga whale stomachs(Hobbs et al. 2008). As salmonid 

numbers dwindle in the fall and winter, beluga whales return to feed on nearshore or deeper 

water species including cod, sculpin, flounder, sole, shrimp, crab and others (Hobbs et al. 2008). 

Feeding frequently occurs collaboratively. Significantly, 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 whales 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 Alaska Peninsula is likely an effective 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 particularly sensitive to decline. Because 

few robust data are available prior to the 1990s, historical population size is based upon surveys 

lacking strong statistical confidence. These estimates range from 150 to 450 individuals between 

1963 and 1988 (Klinkhart 1966, Sergeant and Brodie 1975, Murray and Fay 1979, Calkins 1984, 

Hazard 1988)(Table 4). However, the most robust 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 survey, 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 Estimate1

 CV 95% CI2

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 

20092

 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 abundance between 278 and 435 individuals, with a 

sharp 47% decline between 1994 and 1998 proceeding that (Hobbs et al. 2000, NMFS 
42

unpublished data). Data from June 2006 surveys supported an abundance of 302 individuals 

(Angliss 2007). Trends support a 71% probability that the population 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 annual 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 rate 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, and killer whale predation (Mahoney and Shelden 

2000, Angliss and Outlaw 2006, Angliss and Outlaw 2007b, Angliss 2007). However, Cook 

Inlet harvests have been severely restricted (1-2 whales annually) since 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 numbers. Although not currently 

factors, disease and habitat disturbance have the potential to significantly impact the Cook Inlet 

population. According to the best available population viability analysis, there is an 80% 

probability that the population is declining, a 26% probability that the population 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 significant 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 strandlings 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 have 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 been 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 takes in gillnet fisheries of roughly three to six 

individuals per year (Murray and Fay 1979, Burns and Seaman 1986a). During the early 1900s 

there was a short-lived commercial 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, areas of high beluga use coincide with regions of 

lowest ambient sound and have been known to evacuate high-use areas of humans (Huntington 

2000a). Significant behavioral changes have been observed in Cook Inlet belugas in response to 

small boat proximity (NMFS 2006c). Propeller scars 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 Huntington. 2010). In 1999 and 2000 there was a voluntary 

moratorium on subsistence harvest (Carter and Nielsen. 2010), and since substance harvest have 
43

been conducted under co-management 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 are 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 PCBs, 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 organic pollutants have increased 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 and 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 industrial 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 predictor 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 the Cook Inlet beluga whale on April 11, 

2011 (76 FR 20180). Two areas specific areas are 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. This area contains shallow tidal flats, river 

mouths or estuarine areas and is important as foraging 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 Turnagain Arm (Twenty-mile River 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 consists 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 Cook Inlet beluga whales in a dispersed fashion for fall and winter 

feeding and as transit waters. Area 2 includes near and offshore areas of the mid 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 constituent elements essential to the conservation of 
44

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 cod, 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 critical habitat areas; and (5) absence of in-water 

noise at levels result in the abandonment of habitat by Cook Inlet 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 (Channel Islands) off Southern California 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 (NMFS 1992). In the Bering Sea, the northernmost 

major rookery is on Walrus Island in the Pribilof 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 taxonomic basis at the sub-species 

level in both genetics and skull morphology (Phillips et al. 2009). 

Distribution. The eastern DPS of Steller sea lions includes 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 (Calkins and Pitcher 1982a, Raum-Suryan et al. 2002, Raum-Suryan

et al. 2004). Most adult Steller 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, Ban 2005, Call and Loughlin 2005). Adult males may 

disperse widely after the breeding season. Males that breed in California move north after the 

breeding season and are rarely seen in California 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 maturity and first breed between three and 

eight years of age and the average age of reproducing females (generation time) is about 10 years 

(Pitcher and Calkins 1981, Calkins and Pitcher 1982b, York 1994). They 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 maturity although there is a high rate of reproductive failures. 

The gestation period is believed to be about 50 to 51 weeks (Pitcher and Calkins 1981). The 

available literature indicates an overall reproductive (birth) rate on the order of 55% to 70% or 

greater (Pike and Maxwell 1958, Gentry 1970, Pitcher and Calkins 1981). However, natality 
45

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 three 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 and distance at first, then increase as the pup gets older (Merrick 

and Loughlin 1997, Milette 1999, Pitcher et al. 2001, Milette and Trites 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 dependent 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 nurse their offspring until they are one to two 

years old (Gentry 1970, Sandegren 1970, Pitcher 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 time 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 age (Pitcher and Calkins 1981). The sex ratio of 

pups at birth is assumed to be about 1:1 or biased toward slightly greater production of males, 

but non-pups are biased towards females (Pike and Maxwell 1958, Calkins 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 females may travel far out to sea into water 

greater than 1,000 m deep (Merrick and Loughlin 1997). Studies on immature Steller sea lions 

indicate three types of movements: 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 happen almost daily. Young 

individuals generally remain within 480 km of rookeries their 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 the 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 year 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 major 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 predators 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, and 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 never eat while on breeding grounds (Andrews et al. 2001, 

Loughlin 2002). 

Status and trends. Steller sea lions were originally listed as threatened under the ESA on 

November 26, 1990 (55 FR 49204), following a decline in the US of about 64% over the 

previous three decades. In 1997, the species was 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 population of Steller sea lions was between 

245,000 and 290,000 animals (including pups) in the late 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 population 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 worldwide 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 changed. 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 limitation (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 declined by over 50% in the 1980s, to approximately 

116,000 animals, with the entire decline occurring 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 the 1980s, showing a gradual increase in numbers since 1976 (NMFS 

2005d). Numbers in California, however, have declined to fewer than 2,000 non-pups, from 

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

from Central California in 2000 reached the second lowest recorded count of 349 non-pups (in 

1992 the count was as low as 276 non-pups). In Southeast 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 sites increased nearly 260% 

between 1982 and 2002 (NMFS 2005d). 

The NMFS considers this population stable or increasing, and multiplies pup counts by a factor 

of 4.5 (based on (Calkins and Pitcher 1982b) or 5.1 (Trites 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 population 
47

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 current 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 Canada 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 and Adkison 1994, York 1994, Holmes and York 

2003). 

Natural threats. Killer whale predation, particularly on the western DPS under reduced 

population size, may cause significant reductions in the stock (NMFS 2008g). Sleeper sharks are 

also significant predators of Steller sea lions. Frid et al. (2009) suggested that risk of predation 

in nearshore waters by killer whales 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 several pathogens, but disease levels are unknown (FOC 

2008). Similarly, parasites in this species are common, but mortality resulting 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 commercial exploitation and 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 entangled 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 harvest occur in Canada 

(FOC 2008). Additional mortality (362 from 1990 to 2003) has occurred from shooting of sea 

lions interfering in aquaculture operations along British Columbia (FOC 2008). 

Significant concern also exists regarding competition between commercial fisheries and Steller 

sea lions for the same resource: stocks of pollock, Pacific cod, and Atka mackerel. Significant 

evidence exists that supports the western DPS declining as a result of change in diet and resulting 

declines in growth, birth rates, and survival (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 a 
48

result of the Exxon Valdez oil spill and contained particularly high levels of PAH contaminants, 

presumably as a result of the spill. Blood testing confirmed hydrocarbon exposure. 

Subsequently, premature birth rates increased 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 lions in greater concentrations than 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 twice that in western Alaska pups (3.3 ng/g wet 

weight blood versus 1.6 ng/g wet weight). PCB levels in the kidneys 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 from pollack, which have been found to contain 

organochlorines throughout the Gulf of Alaska, but higher in regions occupied by the eastern 

DPS of Steller sea lions (NMFS 2008g). Heavy metals, including mercury, zinc, copper, 

metallothionien, and butyltin have been identified 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 higher significance, with liver levels 

being measured at levels above those necessary to impact fish (Holmes et al. 2008). However, 

contaminants leading to mortality in Steller sea lions have not been identified (NMFS 2008g). 

Contaminant burdens are lower in females than males, 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 higher levels of contaminants than their parents 

originally had. Steller sea lion contaminants are of additional concern because contaminants in 

the body tend to be mobilized as fat reserves are used, such as when prey availability is low; a 

situation that is likely occurring 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, Oregon, 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 aquatic areas. Specific 

terrestrial areas include rookeries and haul-outs where breading, pupping, refuge and resting 

occurs. More than 100 major haulouts are documented. The principal, essential aquatic areas 

are the nearshore waters around rookeries and haulouts, their forage resources and habitats, and 

traditional rafting sites. Air zones around terrestrial and aquatic habitats are also designated as 

critical habitat to reduce disturbance in these essential areas. Specific activities that occur within 

the habitat that may disrupt the essential life functions that occur there include: (1) wildlife 

viewing, (2) boat and airplane traffic, (3) research activities, (4) timber harvest, (5) hard mineral 

extraction, (6) oil and gas exploration, (7) coastal development and pollutant discharge, and 

others. 

In addition, British Columbia has established protective areas in which Steller sea lion rookeries 

occur at Triangle Island and Cape St. James (Canada 2008). Several other 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 Islands, where Steller sea lions haul-out and breed. 
49

Hawaiian monk seal 

Description of the species. Hawaiian monk seals are found primarily in the NWHI, which 

extend more than 2,000 km miles 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 and 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 population, with genetic connectivity high enough to maintain 

population-level genetic differentiation {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 substrates, 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 vegetation behind beaches as shelter from wind and 

rain. 

Pinniped movements are generally based on foraging. Oceanographic features, such as thermal 

changes that might concentrate prey densities, can affect individual seal foraging behavior (Field

et al. 2001). Hawaiian monk seal distribution, destinations, routes, food sources, and causes of 

movements when not traveling between islands are not well known. Approximately 10-15% of 

Hawaiian monk seals migrate among the breeding 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 Atoll (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 subpopulations 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 octopus, spiny lobster, eels, and bottom-dwelling 

and reef fish (Gilmartin 1983; Goodman-Lowe 1998; Rice 1960). Considered foraging 

generalists, monk seals exhibit significant differences in diet between islands, age, and sex 

groups (NMFS 2007b). No research or monitoring effort has been identified that will effectively 
50

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 reefs and banks (Pullin and Stewart 2006). Several 

recent studies of the foraging patterns of Hawaiian monk seals near rookeries in the NWHI 

provide insight into their diving behavior. Dive depths appear to differ slightly between 

rookeries as well as between age 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, most 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 much 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 towards night diving at FFS. 

Hawaiian monk seals tend to dive within the water 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 for resting and socializing, not feeding. Despite 

the reef fishes of the coral shallows, 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 atoll 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 rocks to find prey, rather than increasing dive 

duration or depth (Parrish et al. 2005). 

Reproduction. Hawaiian monk seals do not form breeding 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 patterns 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 are born annually (NMFS 2007b). Some islands, 

particularly in the southeastern Hawaiian Islands, have strong male-biased sex ratios, leading to 

female deaths from male aggression; a situation which has been significantly corrected by 

selective male removal (5:1, up to 25:1)(Johanos et al. 2010). 

Females typically give birth for the first time between ages of 5 and 10 (Antonelis et al. 2006). 

Pupping patterns vary greatly and not all females give birth in consecutive years (Johanos et al.

1994; Kenyon and Rice 1959). Females that do give birth in consecutive years pup later each 

season, while females 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 (Johanos and Ragen 1994, Ragen and Lavigne 1999). 

Females give birth from February to August, peaking in late March/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 surrounding 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 mother 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 between 43-73 kg (Kenyon and Rice 1959). Rice 

(1964) suggested that adult females weigh approximately 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 avoid 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 affected; however if a larger pup switches with a 

small one, the larger pup will have a longer nursing 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 because all of their populations are either extinct 

(Caribbean monk seal) or close to extinction (Mediterranean and Hawaiian monk seals). Two 

periods of anthropogenic decline have been reported; the first decline 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 increasing 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 Hermes 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 declined 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 monk seal is 1,176 seals (NMFS 2007b). Data 

collected in 2008 suggest that the species population 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). Trends in abundance vary considerably among 

the six main subpopulations. 

A recent five-year status review conducted by NMFS recommends that the Hawaiian monk 
52

seals’ endangered status should remain the same (72 FR 46966, August 2007). The population 

dynamics at the different subpopulations have varied considerably, and current demographic 

variability among the island populations probably reflects a combination of different histories of 

human disturbance and management (Gerrodette 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 development of a captive care program for juvenile 

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

The total of mean, non-pup, beach counts at the 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 declined 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 known minimum abundance. The fitted trend line reveals 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 high near 90% in the mid-1980s to a low of 8% in 

1997 (NMFS 2007b). Shark predation and prey availability are two potentially responsible 

factors. 
53

Populations at Laysan and Lisianski Islands declined sharply after the 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 beach counts increased from 1995 to 2000, but have 

declined in the following years, while the Lisianski subpopulation has remained relatively stable, 

yet low, since the 1970s. Marine debris and low fecundity are factors that 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 caused by male aggression, juvenile 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 increasing. At Midway, beach counts increased from 

1995 to 2000, and have since declined. The subpopulation at Pearl and Hermes Reef increased 

after the mid-1970s. Prior to 1999, beach count increases of up to 7% per year were observed. 

This is the highest estimate of the maximum net productivity rate observed for this species. 

Since 2000, low juvenile survival, thought to be due largely to food limitation, has been 

widespread with rare exceptions in the NWHI, resulting in the population 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 have occurred on at least three occasions at the remote Pacific 

location of Johnston Atoll (excluding nine adult males relocated there from Laysan Island in 

1984). 

The decrease in survival rates of immature animals, including a decline in survival from birth to 

weaning, and survival from weaning to age 2 years 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-specific analyses do support girth and year as 

predictors of first-year survival at each location. When conditions for survival are worse, the 

relationship between size and survival strengthens. The simplest explanation for 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 increasing in the MHI, although systematic surveys were not conducted 

before 2000, and counts do not represent total abundance, as they do not account for seals in the 

water, and not every seal on land is detected. In 2000, the count in the MHI 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 toxoplasmosis from wild and domestic animals, and 

increased interactions with humans, including fishermen, boaters, and divers raise conservation 

concerns which do not apply to the NWHI (NMFS 2007b). The 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 exhibit higher girths and lengths compared to pups 

from the NWHI, as a result of pre- and post-partum maternal investment – a partial reflection of 

prey availability – contradicting the studies that indicate better foraging conditions in the NWHI. 

Suggested explanations for this include a higher per capita availability of prey in the MHI, 

similar absolute preferred prey densities when apex predators are not included in the biomass, 

and increased prey availability due to reduced competition from apex predators (Baker and 

Johanos 2004). 
55

Critical habitat. Critical habitat was originally designated on April 30, 1986 (51 FR 16047), 

and was extended on May 26, 1988 (53 FR 18988; CFR 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 ocean 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 primarily as feeding areas for Hawaiian monk seals, 

while terrestrial habitat serves as pupping and nursing habitat for mothers and pups. Both 

components are currently under significant degradation pressure. Because the marine critical 

habitat is in the Papahānaumokuākea Marine National Monument, fishing is forbidden in the 

critical habitat. A result of this is the establishment 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 critical 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-latitude leatherback range includes 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 Wing 2000). 

Leatherbacks also occur in Mediterranean and Indian Ocean waters (Casale 2003, Casale et al.

2003, Hamann et al. 2006a). Associations exist 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 nesting 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 basin there are at least 

three genetically different nesting populations: the St. Croix nesting population (US Virgin 
56

Islands), the mainland nesting Caribbean population (Florida, Costa Rica, Suriname/French 

Guiana), and the Trinidad nesting population (Dutton et al. 1999). Further genetic analyses 

using microsatellite markers in nuclear DNA along 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, Western 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 waters north to Alaska within the North 

Pacific and is the most common sea turtle in the 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). Leatherback 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, and Australia, leatherback sea 

turtles have only been known to nest in low densities and scattered sites. Leatherback nesting 

aggregations occur widely in the Pacific, including China, Malaysia, Papua New Guinea, 

Indonesia, Thailand, Australia, Fiji, the Solomon 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 Japanese shores, two nests were identified in 

the central Ryukyu Islands in 2002 (Kamezaki et al. 2002). 

Nesting beaches also occur in Mexico and Costa Rica (nesting occurs October through March) 

are a separate population from the western Pacific beaches. In Costa Rica, leatherbacks nest at 

Playa Naranjo in Santa Rosa National Park, the 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 Langosta and Playa Grande and contains the largest 

colony of leatherbacks in the Pacific (Spotila 2004a). Females typically lay six clutches per 

season (average nine days between nests), which incubate for 58–65 days (Lux et al. 2003). 

Limited nesting also occurs along Nicaragua, Panama, El Salvador, Vanuatu, and Guatemala. 

Growth and reproduction. It has been thought that leatherbacks reach sexual maturity 

somewhat faster than other sea turtles (except Kemp’s ridley), with an estimated range of 3-6 

years (Rhodin 1985) to 13-14 years (Zug and Parham 1996). However, recent research suggests 

otherwise, with western North Atlantic leatherbacks possibly not maturing until as late as 29 

years of age (Avens and Goshe 2007). Female leatherbacks 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 per nesting season (Schultz 1975). However, up to ~30% of the 

eggs can be infertile. Thus, the actual proportion of eggs that can result in hatchlings is less than 

this seasonal estimate. The eggs incubate 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 dependent upon reproductive and feeding 

cycles and the oceanographic features that concentrate prey, such as frontal systems, eddy 

features, current boundaries, and coastal retention areas (Collard 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 exposure 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 waters 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 between 5o and 15o N in the Mauritania upwelling, 

south of the Cape Verde islands, over the Guinea Dome area, and off Venezuela, Guyana and 

Suriname. 

Migration and movement. Leatherback sea turtles migrate throughout open ocean convergence 

zones and upwelling areas, along continental margins, and in archipelagic waters (Morreale et al.

1994, Eckert 1998, 1999). In a single year, a leatherback 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 movements 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 waters 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 returning 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 during 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 females remaining in tropical waters for the first year subsequent to 

nesting and moving to northern latitudes during summer/fall and spending winter and spring in 

latitudes of 40-50o N. 

Satellite tracking data reveal that leatherback females leaving 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 southwest Pacific appears to favor sites with higher wind and wave 
58

exposure, possibly as a means to aid hatchling dispersal (Garcon et al. 2010). Individuals 

nesting in Malayasia undergo migrations to tropical 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 aloing the US west coast; individuals take 10-12 

months to migrate here (Benson et al. 2011). Individuals nesting 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 leatherback populations thus far studied. An 

examination of strandings and in-water sighting data 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-adult and adult size distribution. Leatherback sex 

determination is affected by nest temperature, with higher temperatures producing a greater 

proportion of females (Mrosovksy 1994, Witzell et al. 2005b). 

Feeding. Leatherbacks may forage in high-invertebrate prey density areas formed by favorable 

features (Ferraroli et al. 2004, Eckert 2006). Although leatherbacks forage in coastal waters, 

they appear to remain primarily pelagic through all life stages (Heppell et al. 2003b). The 

location and abundance of prey, including medusae, siphonophores, and salpae, in temperate and 

boreal latitudes likely has a strong influence on leatherback distribution in these areas (Plotkin 

1995). Leatherback prey are frequently found in the deep-scattering layer in the Gulf of Alaska 

(Hodge and Wing 2000). North Pacific foraging grounds contain individuals from both eastern 

and western Pacific rookeries, although leatherbacks from the eastern Pacific generally forage in 

the Southern Hemisphere along Peru and Chile (Dutton et al. 1998, Dutton et al. 2000, Dutton 

2005-2006). Mean primary productivity in all foraging areas of western Atlantic females is 

150% greater than in eastern Pacific 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 Oregon (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 Conservation 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 later refined to ~115,000 (Pritchard 1971, 1982). 

Spotila et al. (1996) estimated 34,500 females, but later 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 (perhaps up to 20,000 females annually) of leatherbacks are known to 
59

nest in West Africa (Fretey 2001b). The population of leatherbacks nesting on Gabon beaches 

has been suggested as being the world’s largest, 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 leatherbacks likely number 34,000-94,000 individuals, with females numbering 18,800 

and the eastern Atlantic segment numbering 4,700 (TEWG 2007a). Trends and numbers include 

only nesting females and are not a complete demographic or geographic cross-section. In 1996, 

the entire Western Atlantic population was characterized as stable 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 nesting beaches in the 

Americas, the Caribbean, and West Africa, totaled 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 metapopulation whose limits remain unknown (Rivalan et al.

2006). Heppell et al. (2003a) concluded that leatherbacks 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 indicating 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 nesting numbers have recently increased from more 

than 10,000 nests annually since 1999 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 since the early 1980s has increased 

~0.3% and 7.5% per year, respectively, but lags behind the French Guiana coast and elsewhere in 

magnitude (NMFS/SEFSC 2001). This positive growth 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 modeling and regression analyses, the TEWG 

(2007b) determined that the Southern Caribbean/Guianas stock had demonstrated a long-term, 

positive population growth rate (using nesting 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 and regression analyses indicated that the nesting population likely 

was not growing during 1995-2005 (TEWG 2007b). Other modeling of the nesting data for 

Tortuguero indicates a 67.8% decline between 1995 and 2006 (Troëng et al. 2007). 

In Puerto Rico, the primary nesting beaches are at Fajardo and on the island of Culebra. Nesting 

between 1978 and 2005 ranged between 469-882 nests, 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 Wildlife 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 along the east coast of Florida. 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) estimated 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 increased at all beaches with trends ranging 

from 3.1%-16.3% per year, with an overall increase 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-and-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 leatherback 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 1970s 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 adult female leatherback turtles 

has declined to slightly more than 200 individuals during 1998-1999 and 1999-2000 (Sarti et al.

2000). Spotila et al. (2000) reported the decline of the leatherback turtle 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 could 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 estimated for Central America (Spotila et al. 2000). The number 

of leatherback turtles nesting in Las Baulas National Park declined rapidly during the 1990s, 

from about 1,500 females during the 1988–89 nesting 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 extant leatherback nesting 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 metapopulation harbors the last remaining nesting 

aggregation of significant size in the Pacific with approximately 2700–4500 breeding females 

(Dutton et al. 2007, Hitipeuw et al. 2007). The total number of nests 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, Malaysia is 1% of that 

in 1950s (Chan and Liew 1996). The South China Sea and East Pacific nesting colonies have 

undergone catastrophic collapse. Overall, Pacific populations have declined from an estimated 
61

81,000 individuals to <3,000 total adults and subadults (Spotila et al. 2000). The number of 

nesting leatherbacks has declined by an estimated 95% over the past 20 years in the Pacific 

(Gilman 2009). Drastic overharvesting of eggs and 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 estimated 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 sharks 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 hatching 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 environments are facing increasing 

impacts through widespread development and 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 construction, 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 available 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 nesting adult behavior and is often fatal to emerging hatchlings as they are 

drawn to light sources and away from the sea (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 gastrointestinal tracts leading to death (Mrosovsky et al.

2009). Along the coast of Peru, 13% of 140 leatherback carcasses were found to contain plastic 

bags and film (Fritts 1982). Although global warming 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 collection is widespread and attributed to catastrophic 

declines, such as in Malaysia. Harvest of females along nesting beaches is of concern 

worldwide. 

Bycatch, particularly 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 and 2008, at least 85,000 sea turtles were 

captured as bycatch in fisheries worldwide. This 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 leatherbacks. Donoso and Dutton (2010) 

found that 284 leatherbacks were bycaught between 2001 and 2005 as part of the Chilean 

longline fishery, with two individuals 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/Oregon drift gillnet fishery; 500 leatherback turtles 

are estimated to die annually in Chilean and Peruvian fisheries; 200 leatherback turtles are 
62

estimated to die in direct harvests in Indonesia; and, before 1992, the North Pacific driftnet 

fisheries for squid, tuna, and billfish 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 longline 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 leatherback sea turtles and longline gear used in the HMS fishery (Garrison 

and Stokes 2011a, 2011b). All leatherbacks were released alive, with all gear removed for the 

majority of captures. While 2010 total estimates 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 observed takes (Garrison and Stokes 2010). 

Lewison et al. (2004) estimated that 30,000-60,000 leatherbacks were taken in all Atlantic 

longline fisheries in 2000 (including the US Atlantic tuna and swordfish longline fisheries, as 

well as others). Finkbeiner et al. (2011) estimated that annual bycatch interactions total 1,400 

individuals annually for US Atlantic fisheries (resulting in roughly fourty mortalities) and one 

hundred interactions in US Pacific 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 captured (16 lethally) in drift gillnets set in offshore 

waters from Maine to Florida during this period. Observer coverage 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 coastal waters of Trinidad in 2000. 

Half or more of the gravid turtles in Trinidad and Tobago waters 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 them 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 have high cadmium concentrations, is likely the 

cause (Caurant et al. 1999). Organochlorine pesticides have also been found (Mckenzie et al.

1999). PCB concentrations are reportedly equivalent to those in some marine mammals, with 

liver and adipose levels of at least one congener being exceptionally 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 from 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 since 1979, when tourism increased significantly, 

bringing nesting habitat and people into close and frequent proximity. However, studies do not 

currently support significant critical habitat 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 km2

) and 

California (Point Arena to point Vincente; 119,400 km2

). The primary constituent elements of 
63

these areas include (1.) the occurrence of prey species, primarily scyphomedusae of the order 

Semaeostomeae (Chrysaora, Aurelia, Phacellophora, and Cyanea) of sufficient condition, 

distribution, diversity, and abundance 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 the Atlantic, Indian, and Pacific oceans. 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 haplotypes 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 involve 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 where turtles occur and how long they remain oceanic. Nesting site 

selection in the southwest Pacific 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 straight carapace length) have been found in association with 

Sargassum spp. in both the Atlantic and Pacific oceans (Musick and Limpus 1997) and 

observations of newly hatched hawksbills attracted 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 other 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 USFWS 2007e). Eastern Pacific adult females 

have recently been tracked in saltwater mangrove forests along El Salvador and Honduras, a 

habitat that this species was not previously known to occupy (Gaos et al. 2011). Individuals of 

multiple breeding 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 with 
64

relatively high exposure to wind or wind-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 (Rabalais and Rabalais 1980, Hildebrand 1983, Witzell 1983, 

NMFS and USFWS 1993, Rester and Condrey 1996). Individuals stranded in Texas are 

generally young (hatchlings or yearlings) originating from Mexican 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 known, but are presumed to involve migrations to their nesting 

beach or to courtship stations along the migratory 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 fidelity to their nest sites. 

The life history of hawksbills consists of a pelagic stage that lasts 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 material (Bjorndal 1997b). 

Status and trends. Hawksbill sea turtles received protection on June 2, 1970 (35 FR 8495) 

under the Endangered Species Conservation Act and since 1973 have been listed as endangered 

under the ESA. Although no historical records 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). Worldwide, an estimated 21,212-28,138 hawksbills nest 

each year among 83 sites. Among the 58 sites for with historic trends, all show a decline during 

the past 20 to 100 years. Among 42 sites for which 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 Island), the US Virgin Islands, the Dominican 

Republic, Sao Tome, Guadaloupe, Trinidad and Tobago, Jamaica, Martinique, Cuba (Doce 

Leguas Cays), Mexico (Yucatan Peninsula), Costa Rica (Tortuguero National Park), Guatemala, 

Venezuela, Bijagos Archipelago, Guinea-Bissau, and Brazil. 

Population increase has been greater in the Insular Caribbean than along the Western Caribbean 

Mainland or the eastern Atlantic (including Sao Tomé and Equatorial Guinea). Nesting 

populations of Puerto Rico appeared to be in decline until the early 1990s, but have universally 

increased during the survey periods. 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 history 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. Mortimer 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 and USFWS 2007e). Additional populations are 

known from the eastern Pacific (potentially extending from Mexico through Panama), 

northeastern Australia, and Malaysia (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 Pacific 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 populations of hawksbill sea turtles (Spotila 

2004a, Hutchinson and Dutton 2007). These include western Australian, Andaman and Nicobar 

islands, Maldives, Seychelles, Burma, East Africa, Egypt, Oman, Saudi Arabia, Sudan, and 

Yemen. 

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 be lethal. The only other significant natural threat to 

hawksbill sea turtles is from hybridization of hawksbills 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, groupers, feral cats, and foxes (Bell et al. 1994, Ficetola 2008). In 

some areas, nesting beaches can be almost completely destroyed and all nests can sustain some 

level of depredation (Ficetola 2008). 

Anthropogenic threats. Threats to hawksbill sea turtles are largely anthropogenic, both 

historically and currently. Impacts to nesting beaches include the construction of buildings and 

pilings, beach armoring and renourishment, 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 particularly impacted by beachfront development and 

clearing of dune vegetation (Mortimer and Donnelly in review). The presence of lights on or 
66

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 stopped importing tortoiseshell in 1993 in order to 

comply with CITES (Limpus and Miller 2008). 

In addition to impacting the terrestrial zone, anthropogenic disturbances also threaten coastal 

marine habitats. These impacts include contamination from herbicides, pesticides, oil spills, and 

other chemicals, as well as structural degradation from excessive boat 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 the world’s most endangered marine ecosystems 

(Wilkinson 2000). Although primarily spongivorous, bycatch of hawksbill sea turtles in the 

swordfish fishery off South Africa occurs (Petersen et al. 2009). Finkbeiner et al. (2011) 

estimated that annual bycatch interactions total 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 hawksbill 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 temperatures 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 Rico (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 between 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 the US (TEWG 2000a). However, recent records support Kemp’s ridley sea 

turtles distribution extending into the Mediterranean 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 move along coastal migratory corridors either to the north or south 

from the nesting beach (Byles 1989b, Byles and Plotkin 1994, Renaud 1995b, Renaud et al.

1996, Shaver 1999, 2002). These migratory corridors appear to extend throughout the coastal 

areas of the Gulf of Mexico and most turtles appear to travel in waters less than roughly 164 feet 

in depth. Turtles that headed north and east traveled 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 settle into resident feeding areas for several 

months (Byles and Plotkin 1994, Morreale et al. 2007). Females may begin returning along 

relatively shallow migratory corridors toward the nesting 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). Females initially ovulate 

within a few days after successful mating and lay the first clutch approximately two to four 

weeks later; if a turtle nests more than once per season, subsequent ovulations occur within 

approximately 48 hours after each nesting (Rostal 2007). 

Approximately 60% of Kemp's ridley nesting occurs 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 (several 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 arribadas (Bernardo and Plotkin 2007). The period 

between Kemp's ridley arribadas averages approximately 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 eggs 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 size 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 must attain a size of approximately 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 would be required for this growth from benthic 

immature to mature size (Schmid and Witzell 1997b, Snover et al. 2007b). Currently, age to 

sexual maturity is believed to range from approximately 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 coastal Gulf of Mexico 

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

et al. 2007). Key foraging areas in the Gulf of Mexico include Sabine Pass, Texas; Caillou Bay 
68

and Calcasieu Pass, Louisiana; Big Gulley, Alabama; Cedar Keys, Florida; and Ten Thousand 

Islands, Florida. Foraging areas studied along the Atlantic coast include Pamlico Sound, 

Chesapeake Bay, Long Island Sound, Charleston Harbor, and Delaware Bay. Near-shore waters 

of 120 feet or less provide the primary marine habitat for adults, although it is not uncommon for 

adults to venture into deeper waters. 

Benthic coastal waters of Louisiana and Texas seem to be preferred foraging areas for Kemp’s 

ridley sea turtles (particularly passes and beachfronts), although individuals may travel along the 

entire coastal margin of the Gulf of Mexico. Sightings are less frequent during winter and 

spring, but this is likely due to lesser sighting effort during 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 source 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 was 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 (Hildebrand 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 roughly 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 number of nests for all the beaches in Mexico 

estimated at about 12,000 nests, which amounted 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 approximately 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 turtles over a three-day period at Rancho Nuevo (P. 

Burchfield, pers. comm. in NMFS and USFWS 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 was estimated at 13% per year between 1991 and 

1995 (TEWG 1998a). In 2008, there were 17,882 nests in Mexico (Gladys Porter Zoo 2008), 

and nesting in 2009 reached 21,144 (Gladys Porter 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 modelling used by the TEWG 

(2000b) projected that Kemp’s ridleys could reach the recovery plan’s intermediate recovery 

goal of 10,000 nesters by the year 2015. Recent calculations of nesting females determined from 

nest counts show that the population trend is increasing towards that recovery goal, with an 
69

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 remained 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 experienced a decline similar to that seen in 

Mexico for 2010, with 140 nests (National 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, http://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 lethal 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 curtailed due to the virtual elimination of 

sea turtle and egg harvesting, as well as assistance in hatching and raising 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 ridley 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 total at least 98,300 individuals annually for US Atlantic 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 potential to cause deficiencies 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). Organochlorines, including DDT, DDE, DDD, and PCBs have 

been identified as bioaccumulative agents and in greatest concentration 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 PCB congeners composed most of the contaminants: 

153/132, 138/158, 180, 118, and 187 in order of concentration. PCBs have also been identified 

in the liver, ranging in concentration from 272 ng/g to 655 ng/g of wet weight, values that are 

several fold higher than in other sea turtle species (Lake et al. 1994). However, concentrations 

are reportedly 5% of that which causes reproductive failure in snapping turtles. DDE was 

identified to range from 137 ng/g to 386 ng/g wet weight. Trans-nonachlor was found at levels 

between 129 ng/g and 275 ng/g wet weight. Blood 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 concentrations 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 habitat are linked to PFC 

bioaccumulation. 

Oil can also be hazardous to Kemp’s ridley turtles, 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 copper (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 identified (Orvik 1997, Innis et al. 2008). It is likely that blood samples can 

be used as an indicator of metal concentration. Mercury has been identified in all turtle species 

studied, but are generally an order of magnitude lower than toothed whales. The higher level of 

contaminants found in Kemp’s ridley sea turtles are likely due to this species tendency to feed 

higher on the food chain than other sea turtles. Females from sexual maturity through 

reproductive life should have lower levels of contaminants than males because contaminants are 

shared with progeny through egg formation. 

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

Green sea turtle 

Distribution. Green sea turtles have a circumglobal distribution, occurring throughout tropical, 

subtropical waters, and, to a lesser 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 distinct 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 foraging area originate from Galapagos Islands nesting beaches, 

while those in the Gulf of California originate primarily from Michoacán. Green turtles foraging 

in San Diego Bay and along the Pacific coast of Baja California originate primarily from 

rookeries of the Islas Revillagigedos (Dutton 2003a). 

Table 6. Locations and most recent abundance estimates of threatened 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 Sea 
71

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. 2006) 

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

have been attributed to their largely plant-eating diet (Bjorndal 1982). Growth rates of juveniles 
72

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 growth (declining growth rate with size) in the Atlantic and nonmonotonic growth (growth spurt in mid size classes) in the Pacific, although 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 maximum size just under 100 cm in carapace length 

(Tanaka 2009). A female-bias has been identified from studies of green sea turtles (Wibbels 

2003). 

Consistent with slow growth, age-to-maturity for green sea turtles 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, Limpus 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 females returning to nest ranges from 2 to 5 

years (Hirth 1997b), these reproductive longevity estimates suggest that a female may nest 3 to 

11 seasons over the course of her life. Each female 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 usually have 2-4 or more years between breeding 

seasons, whereas males may mate every year (Balazs 1983). Based on reasonable means of three 

nests per season and 100 eggs per nest (Hirth 1997b), a female may deposit 9 to 33 clutches, or 

about 900 to 3,300 eggs, during her lifetime. Nesting sites appear to be related to beaches with 

relatively high exposure to wind or wind-generated waves (Santana Garcon et al. 2010). 

Once hatched, sea turtles emerge and orient towards 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 apparent that during years of 

heavy nesting activity, density dependent factors (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 significant in sex determination, with greater nest 

moisture resulting in a higher proportion of males (Leblanc and Wibbels 2009). Green sea 

turtles often return to the same foraging areas following nesting migrations (Godley et al. 2002, 

Broderick et al. 2006). Once there, they move within specific areas, or home ranges, where they 

routinely visit specific localities 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 habitats 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 than 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 their 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 habitats 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 (Limpus 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, juveniles 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 protected bays and lagoons. 

While in these areas, green sea turtles rely on marine 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 seagrass 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 displacement 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 warm-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 preferentially in drift 

lines or surface current convergences, probably because 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 sites include coral recesses, the underside of 

ledges, and sand bottom areas that are relatively 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 juveniles 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 Florida 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 seagrass and macroalgae abundance, such as in the 

Laguna Madre of Texas. However, green sea turtles may also occur in offshore regions, 

particularly during migration and development. 

Feeding. While offshore and sometimes in coastal habitats, green sea turtles are not obligate 

plant-eaters as widely believed, and 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 individuals 
74

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 significant enough to alter habitat and ecological 

parameters, such as species composition (Lal et al. 2010). 

Status and trends. Federal listing of the green sea turtle 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 the green turtle as “endangered.” 

No trend data are available for almost half of the important nesting sites, where numbers are 

based on recent trends and do not span a full green sea turtle generation, and impacts occurring 

over four decades ago that caused a change in juvenile 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 measure of the possible trend of 

populations. 

Table 12 summarizes nesting abundance for 46 nesting sites worldwide. These include both 

large and small rookeries believed to be representative of the overall trends for their respective 

regions. Based on the mean annual reproductive 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 stable, and four are decreasing. Long-term 

continuous datasets of 20 years are available for 11 sites, all of which are 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 nesting in the Atlantic/Caribbean include: (1) 

Yucatán Peninsula, Mexico; (2) Tortuguero, Costa Rica; (3) Aves Island, Venezuela; (4) Galibi 

Reserve, Suriname; (5) Isla Trindade, Brazil; (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 considered to be stable or increasing with the exception 

of Bioko Island and the Bijagos Archipelago where the lack of sufficient data precludes a 

meaningful trend assessment for either site (NMFS and USFWS 2007a). Seminoff (2004a) 

reviewed green sea turtle nesting data for eight sites in the western, eastern, and central Atlantic. 

Seminoff (2004a) concluded that all sites in the central and western Atlantic showed increased 

nesting, with the exception of nesting at Aves Island, Venezuela, while both sites in the eastern 

Atlantic demonstrated decreased nesting. These sites are not inclusive of all green sea turtle 

nesting in the Atlantic. However, other sites are not believed to support nesting levels high 

enough that would change the overall status of the species in the Atlantic (NMFS and USFWS 

2007a). 
75

By far, the most important nesting concentration for green sea turtles in the western Atlantic is in 

Tortuguero, Costa Rica (NMFS and 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, Galibi Reserve, and Isla Trindade number in the 

hundreds to low thousands, depending on the 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 nesting in Florida has been 

increasing since 1989 (Florida Fish and Wildlife Conservation Commission, Florida Marine 

Research Institute Index Nesting Beach Survey Database). Since establishment of index beaches 

in 1989, the pattern of green turtle nesting shows biennial peaks in abundance with a generally 

positive trend during the ten years of regular monitoring. This is perhaps due to increased 

protective legislation throughout 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 substantiate the dramatic increase in nesting. 

In 2007, there were 9,455 green turtle nests found just on index nesting beaches, the highest 

since index beach monitoring began in 1989. The number fell back to 6,385 in 2008, further 

dropping under 3,000 in 2009, but that consecutive 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 Beach 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, North 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 nests in Georgia (nesting 

databases maintained on www.seaturtle.org). Increased nesting has also been observed along the 

Atlantic coast of Florida, on beaches where only loggerhead nesting was observed in the past 

(Pritchard 1997). Recent modeling by Chaloupka et al. (2008a)using data sets 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%, and 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, information on incidental captures 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 water intake structures 

has increased significantly. Green sea turtle annual captures averaged 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 captures in 2007, 299 in 2008, 38 in 2009 (power 

output was cut—and cooling water intake concomitantly reduced—for part of that year) and 413 

in 2010. Ehrhart et al. (2007) documented a significant increase in in-water abundance of green 

turtles in the Indian River Lagoon area. 

Pacific Ocean. Green turtles are thought to be declining throughout the Pacific Ocean, with the 

exception of Hawaii, from a combination 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 occur in Australia and Malaysia, with smaller colonies throughout 

the area. Indonesian nesting is widely distributed, but has experienced large declines over the 

past 50 years. Hawaii green turtles are genetically distinct and geographically isolated, and the 

population appears to be increasing in size despite the prevalence of fibropapillomatosis and 

spirochidiasis (Aguirre et al. 1998). 

The East Island nesting beach in Hawaii is showing a 5.7% annual growth rate over >25 years 

(Chaloupka et al. 2008a). In the Eastern Pacific, mitochondrial DNA analysis has indicated 

three key nesting populations: Michoacán, Mexico; Galapagos Islands, Ecuador; and Islas 

Revillagigedos, Mexico (Dutton 2003b). The number 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 nested in Michoacán alone (Clifton et al. 1982, NMFS and USFWS 

2007a). Thus, the current number of nesting females is still far below historical levels. Datasets 

over 25 years in Chichi-jima, Japan; Heron Island, 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 Indian Ocean showed evidence of increased nesting 

(Seminoff 2004a). 

Natural threats. Herons, gulls, dogfish, and sharks prey upon hatchlings. Adults 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 be 

lethal. For unknown reasons, the frequency of a disease called fibropapillomatosis is much 

higher in green sea turtles than in other species 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 birds, 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 probability 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 beach temperatures for nesting (Ackerman 1997). 

Structural impacts to nesting habitat include the 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 indirectly, through changing thermal 

profiles and increasing erosion, serve to decrease the amount of nesting area available to nesting 

females, and may evoke a change in the natural behaviors of adults and 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 night. The presence of lights on or 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 
77

(Witherington and Bjorndal 1991). In addition to impacting the terrestrial zone, anthropogenic 

disturbances also threaten coastal marine habitats, particularly areas rich in seagrass and marine 

algae. These impacts include contamination from herbicides, pesticides, oil spills, and other 

chemicals, as well as structural degradation 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 have 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 species of green sea turtles (De Weede 1996). Very 

few green sea turtles are bycaught in US fisheries (Finkbeiner et al. 2011). However, a legal 

fishery operates in Madagascar that harvested about 10,000 green turtles annually in the mid1990s. 

Sea level rise may have significant impacts upon green turtle nesting on Pacific atolls. These 

low-lying, isolated locations could 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 temperatures would be a much greater threat in the long term to the 

hatching success of sea turtle turtles in general and green sea turtles along northeastern Australia 

particularly. Green sea turtles emerging from nests 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, causing widespread failure of nests (Fuentes et al. 2010). Although the timing 

of loggerhead nesting depends upon sea-surface temperature, green sea turtles do not appear to 

be affected (Pike 2009). 

Green sea turtles have been found 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 what is fit for human consumption (van de Merwe et 

al. 2009). The heavy metals copper, lead, manganese, cadmium, and nickel have also been 

found in various tissues and life stages (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, developmental, 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-resistant strains of bacteria 

(Al-Bahry et al. 2009). DDE has not been found to influence sex determination at levels below 

cytotoxicity (Podreka et al. 1998, Keller and McClellan-Green 2004). To date, no tie has been 

found between pesticide concentration 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 theorized that dinoflagellates of the 

genus Prorocentrum that produce the tumorogenic compound okadoic acid may influence the 

development of fibropapillomatosis (Landsberg et al. 1999). 
78

Critical habitat. On September 2, 1998, critical habitat for 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 oceans. 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, North Indian Ocean, North Pacific Ocean, northeast 

Atlantic Ocean, northwest Atlantic Ocean, South Pacific Ocean, and southeast Indo-Pacific 

Ocean as endangered (75 FR 12598). 

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

locations from the Yucatán Peninsula to North Carolina (Addison and Morford 1996, Addison 

1997, Marcovaldi and Chaloupka 2007). This group comprises five nesting subpopulations: 

Northern, Southern, Dry Tortugas, Florida Panhandle, and Yucatán. Additional nesting occurs 

on Cay Sal Bank (Bahamas), Cuba, the Bahamian Archipelago, Quintana Roo (Yucatan 

Peninsula), Colombia, Brazil, Caribbean Central America, Venezuela, and the eastern Caribbean 

Islands. Genetic studies indicate that, although females routinely return to natal beaches, males 

may breed with females from multiple populations and facilitate gene flow Bowen et al. (2005). 

In the eastern Atlantic, we know of five rookeries from Cape Verde, Greece, Libya, Turkey, and 

the western Africa coast. 

Indian Ocean. Loggerhead sea turtles nest along the Indian 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, New Caledonia, New Zealand, Indonesia, Japan, and the Solomon 

Islands. 

Population structure in the Pacific is comprised of a northwestern Pacific nesting aggregation in 

Japan and a smaller southwestern nesting aggregation in Australia and New Caledonia (NMFS 

2006e). Genetics of Japanese nesters suggest that this subpopulation is comprised 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 these subpopulations and the 

loss of nesting at a beach means a significant loss of diversity and the beach is unlikely to be 
79

recolonized (NMFS 2006e). 

Reproduction and growth. Loggerhead nesting is confined to lower latitudes temperate and 

subtropic zones but absent from tropical 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 juveniles, subadults, novice breeders, first year 

emigrants, and mature breeders (Crouse et al. 1987). Hatchling loggerheads migrate to the ocean 

(to which they are drawn by near ultraviolet light Kawamura et al. 2009), where they are 

generally believed to lead a pelagic existence for as long as 7-12 years (NMFS 2005c). 

Loggerheads in the Mediterranean, similar to those 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 neritic habitats, although 

growth is generally monotypic (declines with age Casale et al. 2009a, Casale et al. 2009b). 

Individually-based variables likely have a high impact on individual-to-individual 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 years 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 prior 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 average of 4.1 nests per season (Murphy 

and Hopkins 1984) , although recent satellite telemetry 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 authors 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 nesting sites for loggerheads in the North Pacific, 

with nesting occurring on the Japanese mainland, 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 returning every two to three years (Iwamoto et al. 1985). 

Nesting females lay at least three nests of 60-115 eggs per nest each season, with roughly two 

weeks between nests (Iwamoto et al. 1985, Eckert 1993b, Nishimura 1994). Between nests, 

females appear to swim offshore into the Kuroshio 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 Florida, 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 associated 
80

with Sargassum spp. habitats, driftlines, and other convergence zones (Carr 1986). After 14-32 

years of age, they shift to a benthic habitat, where immature 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 migrations from nesting beaches to foraging 

grounds (TEWG 1998b). In the Gulf of Mexico, larger 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 immature individuals, this movement occurs in two patterns: a northsouth movement over the continental shelf with migration south of Cape Hatteras in winter and 

movement north along Virginia for summer foraging, and a not-so-seasonal 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 latitudinal basis that corresponds 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 Pacific also show wide-ranging movements. Loggerheads hatched on 

beaches in the southwest Pacific travel have been found to range widely in the southern portion 

of the basin, with individuals from populations nesting in Australia 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 California 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 reside 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 nutrient-rich oceanic fronts (Nichols et al. 2000, Polovina et al.

2000, Kobayashi et al. 2008). Individuals bycaught and satellite tracked in Hawaii longline 

fisheries show individual movement north and south within a thermal 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 areas for juvenile loggerheads (Polovina et al. 2004). The 

Kuroshio Current off Japan may be significant for juvenile and adult loggerheads as a wintering 

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

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

residents of the Gulf of Mexico, 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 developmental 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 winter in the eastern Gulf of Mexico, 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 Dodd (1988a) estimated the maximum 

female life span at 47-62 years. Heppell et al. (2003a) estimated annual survivorship to be 0.81 

(southeast US adult females), 0.78-0.91 (Australia adult females), 0.68-0.89 (southeast US 

benthic juveniles, and 0.92 (Australia benthic juveniles). Survival 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 vegetation 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 invertebrates such as gastropods, mollusks, 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 organisms on turtle shells has recently shown that 

although a loggerhead population may feed on a variety 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 information 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 may 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 stability at this life stage, even though there are doubts about the 

ability to estimate the overall 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. The global abundance of 

nesting female loggerhead turtles is estimated at 43,320–44,560 (Spotila 2004b). 

Atlantic Ocean. 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, researchers 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 Mediterranean 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 turtles as well (Hutchinson and Dutton 2007). A 

recently discovered nesting site along the southern 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 apparent 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 females lay 53,000-92,000 nests per year in the 

southeastern US and the Gulf of Mexico, and the total number of nesting females is 32,000-

56,000. All of these are currently in decline or data are insufficient to access trends (TEWG 

1998a, NMFS 2001b). Loggerheads from western North Atlantic nesting aggregations may or 

may not feed in the same regions from which they hatch. Loggerhead sea turtles from the 

northern nesting aggregation, which represents about 9% of the loggerhead 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 aggregation occur in higher frequencies 

in the Gulf of Mexico (where they represent ~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 Carolina, and North Carolina has a forty-year 

time-series trend showing an overall decline in nesting, but the shorter comprehensive survey 

data (20 years) indicate a stable population (GDNR, NCWRC, and SCDNR nesting data located 

at www.seaturtle.org). NMFS scientists have estimated that the northern subpopulation produces 

65% males (NMFS 2001b). 

The peninsular Florida recovery unit is the largest loggerhead nesting 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 database). An analysis of index nesting 

beach data shows a 26% nesting decline between 1989 and 2008, and a mean annual rate of 

decline of 1.6% despite a large 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 index nesting number is the largest since 2000. 

With the addition of data through 2010, the nesting trend for the nowrthwestern Atlantic DPS is 

slightly negative and not statistically different from zero (no trend)(NMFS and USFWS 2010). 
83

Preliminary, unofficial reports indicate that 2011 nesting may be a high nesting year on par with 

2010. 

Because of its size, the south Florida 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 from 1978-1990, and was initially increasing at 3.9-4.2% after 1990. 

An analysis of nesting data from 1989-2005, a period of more consistent and accurate surveys 

than in previous years, showed a detectable trend and, more recently (1998-2005), has shown 

evidence of a declining trend of approximately 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 Carr 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 size1

. 

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 extinction probabilities of loggerhead sea turtle 

populations in the Atlantic are only slightly lower than those of populations in the Pacific. 

Zurita et al. (2003) found a statistically significant increase in the number of nests on seven of 

the beaches on Quintana Roo, Mexico, from 1987-2001, where survey effort was consistent 

during the period. However, nesting has declined since 2001, and the previously reported 

increasing trend appears to have been temporary (NMFS and USFWS 2008). 

Mediterranean Sea. Nesting in the Mediterranean is confined almost exclusively to the eastern 

basin. The highest level of nesting in the Mediterranean occurs in Greece, with an average of 

3,050 nests per year. 

Pacific Ocean. Abundance has declined dramatically over 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 and Japan, although low level nesting may occur 

outside of Japan in areas surrounding the South China Sea (Kamezaki et al. 2003, Chan et al.

2007). Eastern Australia supported one of the major global loggerhead nesting assemblages until 

recently (Limpus 1985). Now, less than 500 females 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 Japan and the surrounding region is uncertain, but 

approximately 1,000 female loggerhead turtles may 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 nests 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 initial peak of approximately 6,638 nests in 

 

1 While this is a long period of decline relative to the past observed nesting pattern at this location, aberrant ocean 

surface temperatures complicate the analysis and interpretation of these data. Although caution is warranted in 

interpreting the decreasing nesting trend given inherent annual fluctuations in nesting and the short time period over 

which the decline has been noted, the recent nesting decline 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 again 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 Hawaiian 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 species is considered rare or vagrant 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 aggregation occurs on Masirah and Kuria Muria 

Islands in Oman (Ross and Barwani 1982). Extrapolations 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 recorded 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 loggerheads nest annually on Dirk Hartog Island 

beaches along Western Australia (Baldwin et al. 2003). 

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 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 loggerheads) found cold-stunned. Most survived, but 

several hundred were found dead or died after being discovered in a cold-stunned state. Eggs are 

commonly eaten by raccoons and ghost 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 loggerhead nesting habitat are 

numerous: coastal development and construction, placement of erosion control structures, 

beachfront lighting, vehicular and pedestrian traffic, sand extraction, beach erosion, beach 

nourishment, beach pollution, removal of native vegetation, and planting of non-native 

vegetation (Baldwin 1992, USFWS 1998, Margaritoulis et al. 2003, Mazaris et al. 2009b). 

Surprisingly, beach nourishment also hampers nesting success, but only in the first year postnourishment before hatching success increases (Brock et al. 2009). Loggerhead sea turtles face 

numerous threats in the marine environment as well, including oil and gas exploration, marine 

pollution, trawl, purse seine, hook and line, gill net, pound net, longline, and trap fisheries, 
85

underwater explosions, dredging, offshore artificial lighting, power plant entrapment, 

entanglement in debris, ingestion of marine debris, marina and 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 include 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 loggerhead 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. However, more recent estimates from suggest 

interactions and mortality has decreased from pre-regulatory 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 resulting 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). Prior to this closure, the longline 

fleet interacted with an estimated 1,160-2,174 loggerheads annually, with nearly all (89 percent) 

of the takes resulting in mortalities(Peckham et al. 2008). Offshore longline tuna and swordfish 

longline fisheries are also a serious concern for the survival and recovery of loggerhead sea 

turtles and appear to affect the largest individuals more than 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 loggerheads for their meat, shells, and 

eggs has declined from previous exploitation levels, but still exists and hampers recovery efforts 

(Lino et al. 2010). In the Pacific, loggerhead turtles are captured, injured, or killed in numerous 

Pacific fisheries including: 

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

 direct harvest and commercial fisheries off Baja California, Mexico 

 commercial and artisanal swordfish fisheries 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 and 2008, at least 85,000 sea turtles were 

captured as bycatch in fisheries worldwide. This 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 stranded or bycaught had injected marine debris in a Mediterranean 

study, with possible mortality resulting in some cases (Lazar and Gračan 2010). 

Climate change may also have significant implications on loggerhead populations 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 sex while incubating. Ambient temperature increase 
86

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 grounds 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 partial 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 feeding 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 contain very high levels of organochlorines 

chlorobiphenyl, chlordanes, lindane, endrin, endosulfan, dieldrin, PFOS, 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 system 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 endocrine, 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 arsenic, barium, cadmium, chromium, iron, lead, nickel, selenium, 

silver, copper, zinc, and manganese, have also been found in a variety 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 metals 

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 levels than any other sea turtle studied, but 

concentrations are an order of magnitude 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 salmon 
87

Description of the species. Gulf of Maine (GOM) DPS Atlantic salmon occur along the 

Atlantic coast from the Androscoggin River (Maine) in the south to the St. Croix River on the 

US-Canadian border. The lower Penobscot River has three primary tributaries that contain 

Atlantic salmon: Cove Brook, Kenduskeag Stream, Kennebec and Ducktrap rivers. The 

estimated population of Atlantic salmon in the lower Penobscot River and its tributaries is less 

than 20 adult Atlantic salmon. Atlantic 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 anadromous 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 Ungava 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 coarse gravel or rubble substrate (up to 3.5 

inches in diameter) with adequate water circulation to keep the buried 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 where upwellings of groundwater occur (Moir et al. 1998). The 

annual egg production is approximately 240 eggs per 1,075 feet2

 of fluvial habitat (Chaput et al.

1998). 

Movement, growth, and reproduction. Adult Atlantic salmon ascend the rivers of New 

England beginning in the spring and continuing into the fall, with peak numbers occurring in 

June. Although spawning does not occur until late fall, the majority of Atlantic salmon in Maine 

enter freshwater between May and mid-July (Meister 1958, Baum 1997). Salmon that return in 

early spring spend nearly 5 months in the river before spawning, often 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 approximately 10 pounds (USFWS 2005b). 

A minority (10 to 20%) of Maine salmon return as smaller fish, or grilse, after only one winter at 

sea and still fewer return after three years at sea. A spawning run of salmon with representation 

of several age groups ensures some level of genetic exchange among generations. Once in 

freshwater, adult salmon cease feeding during their up-river migration. Spawning occurs in late 

October through November. Spawning sites are positioned 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 positioned at the head of a riffle (Beland et al.

1982), the tail of a pool; or the upstream 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 depositing all of her eggs. Female anadromous 

Atlantic salmon produce a total of 1,500-1,800 eggs per kilogram of body weight, yielding an 
88

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 salmon 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 survives 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 rivers where adult returns are monitored--mainly the 

Penobscot River--were repeat spawners (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 alevins. Alevins remain in the redd for about six 

weeks and are nourished by their yolk sac. Alevins emerge from the gravel about mid-May, 

generally at night, and begin actively feeding (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, disease, 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 gradient, overwintering temperatures and water flows, 

and the level of predation and competition (Bley and Moring 1988a). 

Within days, the free-swimming fry enter the parr stage, moving downstream to areas with 

adequate cover (rocks, vegetation, overhanging banks, and woody debris), water depths ranging 

from approximately four to 24 inches, velocities 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 successfully spawn with sea-run adult females. 

Water temperature, appetite, parr density, photoperiod, 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 three and nine parr per 1,075 feet2

, with 

years up to 16 parr per 1,075 feet2

 not uncommon (Beland 1996). There is no evidence of 

density-dependent limitations at densities of 13 parr per 1,075 feet2 (Whalen et al. 2000). Parr 

feed on larvae of mayflies, stoneflies, caddisflies, chironomids, blackflies, annelids, and 

mollusks, as well as numerous terrestrial insects 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 (Schaffer and Elson 1975). This process, called 

smoltification, prepares parr for the dramatic change 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, including New England, most parr smolt after 

one year, but in cooler areas, they may take two to four years in freshwater before smolting 

(McCormick et al. 1998). Most smolts in New England rivers enter the sea during May and June 
89

to begin their ocean migration. Maine rivers produce approximately three smolts per 1,075 feet2

of habitat. 

Atlantic salmon of US origin are highly migratory, undertaking 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 marine 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 has been documented moving in 

small, loosely aggregated schools near the surface (Dutil and Coutu 1988). After entering the 

nearshore waters of Canada, the post-smolts become 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 adult salmon, they travel 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 Atlantic salmon was listed by the USFWS 

and NMFS as an endangered species on November 17, 2000 (65 FR 69495). The GOM DPS 

encompasses all naturally reproducing remnant populations of Atlantic salmon downstream of 

the former Edwards Dam site on the Kennebec River 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, Kenduskeag Stream, and Cove Brook. Naturally 

reproducing Atlantic salmon in the Penobscot River and its tributaries downstream of the former 

Bangor Dam are listed as endangered. The USFWS’s GOM DPS river-specific 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 estimated 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 were initiated in the mid1800s, but there was little success (Stolte 1981). There was a brief period of success in the late 

19th century when limited runs were reestablished in the Merrimack and Connecticut Rivers by 

artificial propagation, but these runs were extirpated by the end of the century. By the end of the 

19th century, three of the five largest salmon 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 depicted in Figure 5. Such counts typically 

underestimate the actual returns of Atlantic salmon, but can give an idea of trends over time for 

index reaches and watershed. Juvenile smolt production 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

year

adult 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 Analysis (PVA) techniques to determine 

the conservation status of Atlantic salmon in the 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 Narraguagus River, and adult spawner estimates for the 

GOM DPS. Using two time series, 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 individuals) within 100 years is 

61% and 75% (or 28% and 45% in 40 years), for each respective data set. 

Natural threats. Geographic features, such as waterfalls, pose natural barriers to salmon 

migration to spawning habitat. A variety of diseases 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, porpoises, 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 source pollution, both biological (bacteria) and 

chemical. Riverine environments are becoming 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 osmoregulation failure. This is because Atlantic 

salmon are highly sensitive to pH changes and 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 and its immigration into Maine waterways is also of concern. 
91

For example, atrazine can significantly impair water balance in salmon even at low 

concentrations, resulting in a reduced ability for salmon to move between 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 1998). Thus, male reproduction activity is not 

cued to that of females and has the potential to severely reduce recruitment. Nonylphenols are 

also severely detrimental to juvenile salmon. These chemicals also reduce the ability of smolts 

to transition between fresh and salt water, leading 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 aquatic waterways can degrade salmon habitat 

and the habitat of their invertebrate 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 extensive 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 decline originated largely from manmade barriers 

across rivers preventing movement to and from spawning and marine habitats. Although many 

of these barriers have since been modified or removed, modern construction (bridges, culverts, 

etc.) that do not consider Atlantic salmon needs can hinder recovery efforts (Holbrook et al.

2009). When water temperatures exceed 22o C during spawning runs, Atlantic salmon tend to 

have poorer success in passing 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, and 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 deleterious 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 edge of the ESU’s range. To survive, populations 

have adapted to distinct physical and environmental conditions here (Saunders 1981). Climate 

models predict significant, extended warming (IPCC 2001b). Although periods of North 

Atlantic warming and cooling have occurred, changes have not been uniform as global warming 

is, changing sea temperatures, wind currents, fresh water input, and mixing of the ocean’s 

surface layer. Small thermal changes can critically affect biological functions, such as protein 

metabolism, response to aquatic contaminants, reproductive performance, smolt development, 

and species distribution limits (Keleher and Rahel 1996, McCormick et al. 1997, Reid et al.

1997, Somero and Hofmann 1997, Van der 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 opportunities for growth in spring and summer could 

increase the percentage of fish that enter the upper size distribution of a population and smolt the 

following spring (Thorpe 1977, Thorpe et al. 1980, Thorpe 1994). In addition, warmer rearing 

temperatures during the late winter and spring have been shown 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 degradation and weight loss will increase with rising 

water temperature (McCarthy and Houlihan 1997). 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 established 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 support 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 temperatures 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 accommodate 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, chironomids, caddisflies, blackflies, aquatic annelids, and mollusks, 

as well as numerous terrestrial invertebrates, alewives, dace, or minnows) to support 

growth and survival of Atlantic salmon parr. 

 Freshwater and estuary migratory 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 migration sites with pool, lake, and instream habitat that provide 

cool, oxygenated water and cover items (e.g., boulders, woody debris, and vegetation) to 

serve as temporary holding and resting areas during upstream migration 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 migration sites with sufficiently cool water temperatures and 

water flows that coincide with diurnal cues to stimulate smolt migration. 

 Freshwater migration sites with water chemistry (particularly pH) needed to support sea 

water adaptation of smolts. 

These PCEs have undergone significant degradation over in the recent past. Dams, along with 
93

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 habitat. Agriculture and urban development largely affect the 

lower third of the Penobscot River below the Piscataquis River sub-basin by reducing substrate 

and cover, reducing water quality, and elevating 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 predator/prey relationships. Similar 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 migration of Atlantic salmon and other diadromous fish and either 

reduce or eliminate access to historically accessible spawning and rearing habitat. In addition to 

hydropower dams, agriculture and urban development 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, significantly degrade habitat quality throughout the 

Merrymeeting Bay recovery unit by altering natural 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 quality 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 small ice dam on the lower 

Narraguagus River, limit access to spawning and rearing 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 populations associated with impoundments. In the 

Pleasant River and Tunk Stream, which collectively 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 rivers contain the highest quality habitat and 

collectively account for approximately 40 percent of the spawning 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 Ventura River in California to Point Hope, Alaska in North 

America, and in northeastern Asia from Hokkaido, Japan to the Anadyr River in Russia (Healey 

1991a). In addition, Chinook salmon have been reported in the Canadian Beaufort Sea (McPhail 

and Lindsey 1970). We discuss the distribution, status, and critical habitats of the species2 of 

 

2 We use the word “species” as it has been defined in section 3 of the ESA, which include “species, subspecies, and 

any distinct population segment of any species of vertebrate fish or wildlife which interbreeds when mature (16 

U.S.C 1533).” Pacific salmon that have been listed as endangered or threatened were listed as “evolutionarily 
94

endangered and threatened Chinook salmon separately, and summarize their common 

dependence on waters of the United States. However, because Chinook salmon in the wild are 

virtually indistinguishable between listed species, and are the same biological species we begin 

this section describing those characteristics common across the listed species. 

Of the Pacific salmon species considered herein, Chinook salmon exhibit arguably one of the 

most diverse and complex life history strategies with multiple races within 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 (Healey 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 spans fresh and marine waters, with one 

reproductive event per adult (Chinook salmon are semelparous and die after spawning). 

Spawning migrations generally occur 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 influence 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 general latitudinal cline is 

apparent across the species range with spawning typically occurring earlier in the spring/summer 

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

On the spawning grounds, mate competition is intense with males competing to fertilize eggs and 

females competing for optimal nest site selection. 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 within a redd, excavating stream rock as she moves 

upstream, increasing the size of her redd 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 spawning. Body size can be 

important in determining reproductive success in terms of nest selection and mating competition 

(Foote 1990). Chinook salmon age at maturity ranges 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 incubation. Incubation generally takes a couple of 

months or more. Alevin (also called “yolk-sac” 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 swimming fry. Egg to fry survival can also vary 

widely across basins, years, and habitat conditions within a basin. In general, the survival of 

eggs and alevin, and the fitness of emerging fry are affected by sediment 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 populations 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 coastal waters where they rear in estuaries for a few weeks to 

months, while others migrate directly to ocean waters. Stream-type Chinook salmon do not 

migrate to sea until the spring following emergence, 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 are often found in the lower 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 velocity. 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 production, 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 begins long before fish enter saltwaters, including 

exposure to heavy metals and elevated water temperatures (Wedemeyer et al. 1980). 

Life at sea varies according to population, race, and age-class. Chinook salmon tend to remain at 

sea between 1 and 6 years, with most fish returning to freshwater after 2-4 years at sea. Fishery 

catches indicate that ocean- and stream-type fish exhibit divergent 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 are 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 within 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 features (PFMC 2000). Individuals 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 coastal migratory corridors 

(PFMC 2000). 

Generally, Chinook salmon outmigrants (termed smolts) are about 5-13 cm long when they enter 

saline (often brackish) waters. The process of smoltification is physiologically demanding, 

involving osmoregulation—the maintenance of osmotic pressure as the fish enters waters of 
96

increased salinity, which is necessary to maintain 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 this 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 found in waters below 5° C (MBC 1987, PFMC 2000). 

Chinook salmon originating from the same freshwater region have similar age-dependent marine 

distributions (even if associated with different runs), which are distinct 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 counterparts (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 smolts 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 juveniles feed on adult and larval insects 

including terrestrial and aquatic insects such as dipterans, beetles, stoneflies, chironomids, and 

plecopterans (Healey 1991a). During the first 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 herring 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 marine mammals, including harbor seals, sea lions, 

and killer whales. There have been recent concerns 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 evolutionary 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 river hydrology; gravel mining that impedes their 

migration and alters the dynamics (hydrogeomorphology) of the rivers and streams that support 

juveniles; water diversions that deplete water levels in rivers and streams; destruction or 

degradation of riparian habitat that increase water temperatures in rivers and streams sufficient to 

reduce the survival of juvenile Chinook salmon; and land use practices (logging, agriculture, 

urbanization) that destroy wetland and riparian ecosystems while introducing sediment, nutrients, 

biocides, metals, and other pollutants into surface and ground water and degrade water quality in 

the freshwater, estuarine, and coastal ecosystems throughout the Pacific northwest (Buhle et al.

2009). Of several habitat factors for spring-run Columbia River Chinook salmon, reductions in 
97

fine sediments may be particularly important 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 establishment of waterway obstructions such as dams, power plants, 

and sluiceways for hydropower, agriculture, flood control, and water storage. These structures 

have blocked salmon migration to spawning habitat or resulted in direct mortality and have 

eliminated entire salmon runs as a result. While some 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 barriers alter the natural hydrograph of basins, both 

upstream and downstream of the structure, and significantly reduce the availability and quality of 

spawning and rearing habitat (Hatten and Tiffan 2009). Many streams and rivers, particularly in 

urban or suburban areas, suffer from streamside development, which contributes sediment, 

chemical pollutants from pesticide applications and automobile or industrial activities, altered 

stream flows, loss of streamside vegetation and allochthonous materials to name a few. These 

factors can directly cause mortality, 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 there are potential benefits to the artificial production of salmon 

as well. Adverse effects of artificial propagation include: a decline in the natural population 

from the taking of wild brood stock 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 spawned salmon populations. Potential benefits to 

artificial propagation include the bolstering of the numbers of naturally spawning fish in the 

short-term, the conservation of genetic resources, and guarding against the catastrophic loss of 

naturally spawned populations at critically 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 uneven exploitation of certain 

populations and size classes (Mundy 1997, Reisenbichler 1997). Targeted fishing of larger 

individuals results in excluding the most fecund individuals from spawning (Reisenbichler 

1997). Genetic changes that promote smaller body sizes have occurred in heavily exploited 

populations in response to size-selective harvest pressures (Mundy 1997, Reisenbichler 1997, 

Swain et al. 2007). Age at maturity can also be accelerated 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 stages, but is likely particularly 

significant for freshwater life stages. Organic pollutants, especially PCBs, DDT and its 

congeners, pesticides, and endocrine disruptor,s are particularly concerning. These chemicals 

can inhibit smell, disrupt reproductive behavior and physiology, impair immune function, and 

lead to mortality through impairment of water balance when traveling between fresh- and 

saltwater systems (Varanasi et al. 1993a, Varanasi et al. 1993b). Diffuse and extensive 

population centers contribute increase contaminant volumes and variety from such sources as 

wastewater treatment plants and sprawling development. Urban runoff from impervious surfaces 
98

and roadways often contains oil, copper, pesticides, 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 Chinook 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 existed (NMFS 2011g). Two artificial 

propagation programs are included in this ESU: winter-run Chinook salmon 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 natural populations than would be expected between closely related 

populations within this ESU. 

This ESU consists of a single spawning population that enters the Sacramento River and its 

tributaries in California from November to June and spawns from late April to mid-August, with 

a peak from May to June (Table 7). Sacramento River winter-run Chinook salmon historically 

occupied cold, headwater streams, such as the upper reaches of the Little Sacramento, McCloud, 

and lower Pit Rivers. Young winter-run Chinook 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 

Abundancea 

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 commercial 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 intervals 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 historic 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 periods of spawning, incubation and rearing. 

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

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

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

have created conditions suitable 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 to 
99

the mainstem Sacramento River below Keswick 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 available for the period before 1996, the completion of 

Red Bluff Diversion Dam. However, 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 the county facility at the Red Bluff Diversion 

Dam (1967 to 1969), the spawning run of winter-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 the period of 1989 to 1991 was 388 fish. Based on 

the Red Bluff Diversion Dam counts, the population has been growing 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 for the ESU is an average of 10,000 

female spawners per year and a population growth rate >1.0, calculated over 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 for 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 Sacramento 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 waters. These areas are important for the species’ 

overall conservation by protecting quality growth, reproduction, and feeding. Factors 

contributing to the downward trends in this ESU include reduced access to spawning/rearing 

habitat, possible loss of genetic integrity through population bottlenecks, inadequately screened 

diversions, predation at artificial structures and by nonnative 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 propagation programs are part of this ESU: The 

Humboldt Fish Action Council (Freshwater Creek), Yager Creek, Redwood Creek, Hollow Tree, 

Van Arsdale Fish Station, Mattole Salmon Group, and Mad River Hatchery fall-run Chinook 

hatchery programs. These artificially propagated populations are no more divergent relative to 

the local natural populations than would be expected between closely related populations within 

this ESU. 

California Coastal 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 artificially propagated fish to the 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 

Abundancea 

Mean Number 

of Spawners 

(Range)b 

Percent 

Hatchery 

Contributionc

Long-term Trendd

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 opinion and evaluation of habitat 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 spawners 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 degraded 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 retained their threatened status on June 28, 2005 

(70 FR 37160). California Coastal Chinook salmon were listed due to the combined effect of 

dams that prevent them from reaching spawning habitat, logging, agricultural activities, 

urbanization, and water withdrawals in the river drainages that support 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 sizes. Sproul Creek abundance (part of the lower Eeel River 
101

population) has been on a positive trend since 1995, but remains negative over longer time series 

(NMFS 2011a). The Russian River probably contains some natural production, but the origin of 

those fish is not clear because of a number of introductions of hatchery fish over the last century. 

Average number of returning spawners in the Russian River has slowly declined since 2001 

(NMFS 2011a). The Eel River contains a substantial fraction of the remaining Chinook salmon 

spawning habitat for this species. Since its original listing and status review, little new data are 

available or suitable for analyzing trends or estimating changes in this population’s growth rate 

(Good et al. 2005b). Adult spawning returns in Prairie Creek (part of the Redwood Creek 

population) have been on the decline since they 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, the 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 unquantified 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 for California Coastal Chinook salmon on 

September 2, 2005 (70 FR 52488). Specific geographic areas designated include the following 

CALWATER hydrological units: Redwood Creek, Trinidad, Mad River, Eureka Plain, Eel 

River, Cape Mendocino, Mendocino Coast, and the Russian River. These areas are important for 

the species’ overall conservation by protecting quality growth, 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 rearing sites, freshwater migration corridors, nearshore marine habitat 

and estuarine areas. The physical or biological features that characterize these sites include 

water quality and quantity, natural cover, forage, adequate passage conditions, and floodplain 

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 that were excluded from 

designation. 

In total, California Coastal Chinook salmon occupy 45 watersheds (freshwater and estuarine). 

The total area of habitat designated as critical includes about 1,500 miles of stream habitat and 

about 25 square miles of estuarine habitat, mostly within Humboldt Bay. 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 defined as the bankfull elevation. In estuarine areas the lateral extent 

is defined by the extreme high water because extreme high tide areas encompass those areas 
102

typically inundated by water and regularly occupied by juvenile salmon during the spring and 

summer, when they are migrating in the nearshore zone and relying on cover and refuge qualities 

provided by these habitats, and while they are foraging. Of the 45 watershed reviewed in NMFS' 

assessment of critical habitat for 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 summer and winter rearing 

habitat, as well as marginal spawning habitat. Compared to historical conditions, there are fewer 

pools, limited cover, and reduced habitat complexity. The limited instream cover that does exist 

is provided mainly by large cobble and overhanging vegetation. Instream large woody debris, 

needed for foraging sites, cover, and velocity refuges is especially lacking in most of the streams 

throughout the basin. NMFS has determined that these degraded habitat conditions are, in part, 

the result of many human-induced factors affecting critical habitat including dam construction, 

agricultural and mining activities, urbanization, stream channelization, water diversion, and 

logging, among others. 

Lower Columbia River Chinook salmon 

Distribution. The Lower Columbia River Chinook salmon ESU includes all naturally spawned 

populations of Chinook salmon from the Columbia River 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 than would be expected between 

closely related populations within this ESU. 
103

Figure 6. Population boundaries, dams, accessible areas, and extirpated 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 fall runs (tules), late 

fall runs (brights), and spring-runs. Spring and fall runs have been designated 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 Chinook typically enter the 

Columbia River in August through October to spawn 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 reaffirmed 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 risk of extinction (NMFS 2011d). Spring-fun 

individuals are at particularly high risk, as they have been cut off from extensive spawning areas 
104

by dams (NMFS 2011d). Historical records of Chinook salmon abundance are sparse, but 

cannery records suggest a peak run of 4.6 million fish (43 million pounds) in 1883 (Lichatowich 

1999). Although fall-run Chinook salmon are still present throughout much of their historical 

range, they are still subject to large-scale hatchery production, relatively high harvest, and 

extensive habitat degradation. The Lewis River late-fall-run population is the healthiest and has 

a reasonable probability of being self-sustaining. Abundances largely declined during 1998-

2000 and trend indicators for most populations are negative, especially if hatchery fish are 

assumed to have a reproductive success equivalent to that of natural-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 

abundancea

Mean number 

of spawners 

(range)b

Percent 

hatchery 

contributionc

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 Rivere

 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 

47,591f

 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 Rivere

 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 Rivere

 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 performance based on reach-specific habitat attributes. Estimates are provided 
105

as a means of comparing the historical abundance of populations relative to current abundance. See Good et al.

(2005) for a discussion about the uncertainty associated with these estimates. b

Recent geometric mean number of spawners as reported in Good et al. 2005 c

Recent average hatchery-origin spawners (%) 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 reproductive success in the first estimate. In the second estimate 

hatchery fish are assumed to have the same relative reproductive success as natural-origin fish. Growth rates were 

not calculated for all populations, as adequate 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. 

feCombined estimate of Lewis River fall run (East Fork only) and Lewis River brights (Good et al. 2005) 

Most populations for which data are available have a long-term declining population trend 

(Table 15). Currently, the spatial extent of populations in the Coastal and Cascade fall runs are 

similar to their respective historical conditions. New data include spawner abundance estimates 

through 2001, new estimates of the fraction of hatchery spawners, and harvest estimates. In 

addition, estimates of historical abundance have been provided by the Washington Department 

of Fish and Wildlife. The Willamette/Lower Columbia River Technical Review Team estimated 

that 8-10 historic populations have been extirpated, most of them spring-run populations. Near 

loss of that important life history 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 hatchery 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 increased 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 critical habitat includes all Columbia River 

estuarine areas and river reaches proceeding upstream to the confluence with the Hood River as 

well as specific stream reaches in a number of tributary subbasins. These areas are important for 

the species’ overall conservation by protecting quality growth, 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. Specific sites include freshwater spawning sites, 

freshwater rearing sites, freshwater migration corridors, nearshore marine habitat and estuarine 

areas. The physical or biological features that characterize these sites include water quality and 

quantity, natural cover, forage, adequate passage conditions, and floodplain connectivity. Of 52 

subbasins reviewed in NMFS' assessment of critical 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 industrial harbor and port 

development, and urban development. Limiting factors identified for this species include 
106

reduced access to spawning/rearing habitat in tributaries, hatchery 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, hydropower projects, including 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 development are also significant influences 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 channel within the mainstem Columbia 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 channel of the Lower Columbia 

River is now maintained at a depth of 43 feet and a width of 600 feet. The Lower Columbia 

River supports five ports on the Washington State 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 sediment chemicals, such as arsenic and polycyclic 

aromatic hydrocarbons (PAHs), have been identified in Lower Columbia River watersheds in the 

vicinity of the ports and associated industrial facilities. 

The most extensive urban development in the Lower 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 water quality issues with urban development and 

residential septic systems include higher water temperatures, lowered dissolved oxygen, 

increased fecal coliform bacteria, and increased chemicals associated with pesticides and urban 

runoff. 

The Columbia River estuary has lost a significant amount of the 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 the Columbia River inundated the 

margins and floodplains along the estuary, allowing juvenile salmon and steelhead access to a 

wide expanse of low-velocity marshland and tidal channel habitats. In general, the riverbanks 

were gently sloping, with riparian and wetland vegetation at the higher elevations of the river 

floodplain becoming habitat for salmon and steelhead during flooding river discharges or flood 

tides. Sherwood et al. (1990) estimated that the Columbia River 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 within the estuary, and other factors affecting salmon population 

structure and life histories, have altered the estuary’s capacity to support juvenile salmon 
107

(Bottom et al. 2005, Fresh et al. 2005, NMFS 2006d, LCFRB 2010). Diking and filling activities 

have reduced the tidal prism and eliminate emergent and forested wetlands and floodplain 

habitats. These changes likely have reduced the estuary’s salmon-rearing capacity. Moreover, 

water and sediment in the Lower Columbia River and its tributaries have toxic contaminants that 

are harmful to fish and wildlife (LCREP 2007). Contaminants of concern include dioxins and 

furans, heavy metals, polychlorinated biphenyls (PCBs) and organochlorine pesticides such as 

DDT. Simplification of the population structure and life-history diversity of salmon possibly is 

yet another important factor affecting juvenile salmon viability. Restoration of estuarine habitats, 

particularly diked emergent and forested wetlands, reduction of avian predation by terns, and 

flow manipulations to restore historical flow patterns may have begun to enhance the estuary’s 

productive capacity for salmon, although historical changes in population structure and salmon 

life histories may prevent salmon from making full use of the productive capacity of estuarine 

habitats. 

Species-specific threats and limitations to recovery. Lower Columbia River Chinook salmon 

face numerous hindrances to their survival and recovery, including degraded estuarine and nearshore marine habitat resulting from cumulative impacts of land use and flow management by the 

Columbia River hydropower system and floodplain connectivity and function, channel structure 

and complexity, riparian areas, stream substrate, stream flow, and water quality have been 

degraded as a result of cumulative impacts of agriculture, forestry, and development (LCFRB 

2010, NMFS 2011d). Further habitat-based impacts to the species result 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 productivity resulting from sediment and nutrient-related changes 

in the estuary. Also hampering the species is reduced access to spawning and rearing habitat 

mainly as a result of tributary hydropower projects, hatchery-related effects, 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 Chinook 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 propagation program, the Feather River Hatchery 

spring-run Chinook salmon program. This artificially 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 spawning 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 while they mature over the summer. 
108

Table 10. Central Valley spring-run Chinook salmon populations and selected measures of population viability. 

Population

Historical 

Abundancea 

Mean Number of 

Spawners (Range)b 

Percent 

Hatchery 

Contributionc

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 current 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 population 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 classification this species retained on June 

28, 2005 (70 FR 37160). This species was listed because dams isolate them from most of their 

historic spawning habitat and the habitat remaining to them is degraded. Historically, spring-run 

Chinook salmon were predominant throughout the Central Valley occupying the upper and 

middle reaches (1,000 to 6,000 feet) of the San Joaquin, American, Yuba, Feather, Sacramento, 

McCloud and Pit Rivers, with smaller populations in most tributaries with sufficient habitat for 

over-summering adults (Stone 1874{Rutter, 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 already 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 the operation of Folsom Dam. 

Naturally spawning populations of Central Valley 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, Clear 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 respectively, well above the 1991 to 2001 average. 

Evaluating the ESU as a whole, however, masks significant changes that are occurring among 

populations that comprise the ESU (metapopulation). For example, the mainstem Sacramento 

River population has undergone a significant decline 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 for 
109

the period 1991 to 2001, while the average abundance of Sacramento River tributary populations 

increased from a low of 1,227 to a high of 5,925 over the same periods. 

Abundance time series data for Mill, Deer, Butte, and Big Chico creeks 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 ocean 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 positive population growth. 

However, since 2001, escapement has declined for all three populations, possibly resulting from 

2007-2009 drough conditions in California’s central valley 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). Although the populations are small, Central Valley 

spring-run Chinook salmon have some of the highest 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). Specific geographic areas designated include the 

following CALWATER hydrological units: Tehama, Whitmore, Redding, Eastern Tehama, 

Sacramento Delta, Valley-Putah-Cache, Marysville, 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 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 rearing sites, freshwater migration corridors, nearshore marine habitat and 

estuarine areas. The physical or biological features that characterize these sites include water 

quality and quantity, natural cover, forage, adequate passage conditions, and floodplain 

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 that 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 includes about 1,100 miles of stream 

habitat and about 250 square miles of estuarine habitat in the San Francisco-San Pablo-Suisun 

Bay complex. 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 defined as the bankfull 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 are migrating in the nearshore zone and relying 

on cover and refuge qualities provided by these habitats, 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 rating of conservation value, three received a 

medium rating, and 27 received a high rating of conservation value for the species. 

Upper Columbia River Chinook salmon 
110

Distribution. The Upper Columbia River spring-run Chinook salmon ESU includes all naturally 

spawned populations in all river 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 artificial propagation programs are part of this 

ESU. These artificially propagated populations are no more divergent 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 Rock Island Dam: the Wenatchee, 

Entiat, and Methow Rivers. Table 11 identifies 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 extirpated 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. Juvenile 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 

contributionb

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 (-23.2d

) 

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 reaffirmed 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, spawning escapements for the Wenatchee, Entiat, 

and Methow rivers have declined an average of 5.6, 4.8, and 6.3% per year, respectively, since 

1958. In the most recent 5-year geometric mean (1997-2001), spawning escapement for 

naturally produced fish was 273 for the Wenatchee 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 for this ESU is estimated at 0.93 (meaning the 

population is not replacing itself) (Fisher and 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 extinction: a 75 to 100% probability of extinction 

within 100 years (given return rates for 1980 to present). Since 2005, abundance estimates have 

increased throughout the ESU, although a large portion of returns are hatchery origin fish 

(NMFS 2011i). 

Hatchery influence and genetic diversity are significant 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 Coulee Fish Maintenance Project 

from 1939-1943. Stray hatchery fish and a high proportion of hatchery fish during spawning 

have contributed to the high genetic diversity risk. 

Critical habitat. NMFS designated critical habitat for Upper Columbia River spring-run 

Chinook salmon on September 2, 2005 (70 FR 52630). The designation includes all Columbia 
112

River estuaries and river reaches upstream to Chief Joseph Dam and several tributary subbasins. 

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 highwater line is not defined the lateral extent is defined as the bank full elevation. These areas are 

important for the species’ overall conservation by protecting quality growth, 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. Specific sites include freshwater 

spawning sites, freshwater rearing sites, freshwater migration corridors, nearshore marine 

habitat, and estuarine areas. The physical or biological features that characterize these sites 

include water quality and quantity, natural cover, 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. The 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 loss of in-river wood, altered tributary floodplain 

and channel morphology, reduced tributary stream flow and impaired passage, and harvest 

impacts. 

Migratory habitat quality in this area has been severely affected 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 Idaho, 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 structure leading to increased rates of piscivorous and 

avian predation on juvenile salmon and steelhead, and delayed migration for both adult and 

juveniles. Physical features of dams such as turbines also kill migrating fish. In-river survival is 

inversely related to the number of hydropower projects encountered by emigrating juveniles. 

Similarly, development and operation of extensive irrigation systems and dams for water 

withdrawal and storage in tributaries have drastically altered hydrological cycles. A series of 

large regulating dams on the middle and upper Deschutes River affect flow and block access to 

upstream habitat, and have extirpated one or more populations from the Cascades Eastern Slope 

major population (ICTRT 2003b). Similarly, operation 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 habitat are over-allocated under state water law, with 

more allocated water rights than existing streamflow conditions can support. Withdrawal of 

water, particularly during low-flow periods that commonly overlap with agricultural 

withdrawals, often increases summer stream temperatures, blocks fish migration, strands fish, 

and alters sediment transport (Spence et al. 1996). Reduced tributary stream flow has been 
113

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 habitat are listed on the state of Oregon’s Clean 

Water Act section 303(d) list for water temperature. Many areas that were historically suitable 

rearing and spawning habitat are now unsuitable due to high summer stream temperatures. 

Removal of riparian vegetation, alteration of natural stream morphology, and withdrawal of 

water for agricultural or municipal use all contribute to elevated stream temperatures. 

Contaminants such as insecticides and herbicides from agricultural runoff 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-run Chinook salmon (Upper Columbia 

Salmon Recovery Board 2007, NMFS 2011i). Mainstem 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, forestry, 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 factors 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 and streams flowing into 

Hood Canal, South Sound, North Sound and the Strait of Georgia in Washington (Figure 8). 

Twenty-six artificial propagation programs are part of the ESU. These artificially propagated 

populations are no more divergent relative to the local natural 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 historical 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 their 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, significant variation occurs in 

residence time in freshwater and estuarine environments. For example, Hayman et al. (1996) 

described three juvenile Chinook salmon 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 

abundancea

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 

abundancea

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 about 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 reproductive 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 type) that were either independent historical 

populations or major components of the remaining 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 trends 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 populations is declining. The overall 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 split between populations that 

are growing and those that are 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 spawners then the largest estimated decline occurs in the Green 

River. Populations with the largest positive short 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 recent 
116

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 populations 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 portion of the ESU. Natural-origin pre-harvest 

recruit escapements remained fairly constant from 1985-2009 (Ford et al. 2010). Returns (preharvest run size) from the natural spawners were highest in 1985, 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 1997. 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; presently 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 returns 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 numbers. During a recent 5-year 

period, the geometric mean of natural spawners in populations of Puget Sound Chinook salmon 

ranged from 222 to just over 9,489 fish. Most populations 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 fraction of hatchery fish. The populations with 

the greatest estimated component of hatchery fish tend to be in mid- to southern Puget Sound, 

Hood Canal, and the Strait of Juan de Fuca regions. Estimates of the historical equilibrium 

abundance, based on pre-European settlement habitat conditions, range from 1,700 to 51,000 

potential spawners per population. 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 habitat 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, Stillaguamish River, Skykomish River, Snoqualmie 

River, Lake Washington, Green River, Puyallup River, White River, Nisqually River, Hamma 

Hamma River and other Hood Canal watersheds, the Dungeness/Elwha Watersheds, and 

nearshore marine areas of the Strait 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 defined as the bank full elevation. 

The designation for this ESU includes sites necessary to support one or more life stages. These 

areas are important for the species’ overall conservation by protecting quality growth, 

reproduction, and feeding. Specific primary constituent elements include freshwater spawning 

sites, freshwater rearing sites, freshwater migration corridors, nearshore marine habitat, and 

estuarine areas. The physical or biological features 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 rated as having a medium conservation value, 12 were rated as 
117

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 Puget Sound Chinook salmon. Factors contributing 

to the downward trends in this ESU are hydromorphological changes (such as diking, 

revetments, loss of secondary channels in floodplains, widespread blockages of streams, and 

changes in peak flows), degraded freshwater and marine habitat affected by agricultural activities 

and urbanization, and upper river tributaries widely affected by poor forest practices. Changes in 

habitat quantity, availability, diversity, flow, temperature, sediment 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 stream sedimentation. Unpaved 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 riparian trees near stream channels. Subsequent 

agricultural and urban conversion permanently altered riparian vegetation in the river valleys, 

leaving either no trees, or a thin band of trees. The riparian zones along many agricultural areas 

are now dominated by alder, invasive canary grass 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 remove smaller substrate particles and large wood. 

The loss of side-channels, oxbow lakes, and backwater 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 acres 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 surface and groundwater 

in complex stream and wetland systems helps to moderate stream temperatures. Forest wetlands 

are estimated to have diminished by one-third in Washington State ((FEMAT) 1993, Spence et 

al. 1996, SSPS 2007). 

Loss of riparian habitat, elevated water temperatures, elevated levels of nutrients, increased 

nitrogen and phosphorus, and higher levels of turbidity, presumably from urban and highway 

runoff, wastewater treatment, failing septic systems, and agriculture 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 soils on residential and agricultural lands, simplified and extended 

drainage networks, loss of wetlands, 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 mortality 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 juvenile salmonids normally find refuge. When 
118

diversion headgates are shut, access back to the main channel is cut off and the channel goes dry. 

Mortality can also occur with inadequately screened 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 projects are major habitat problems in many Puget 

Sound tributary basins (SSPS 2007). 

The nearshore marine habitat has been extensively altered and armored 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, eliminating natural cover along the 

shore and natural recruitment of beach sand (SSPS 2007). 

Dams constructed for hydropower generation, irrigation, or flood control have substantially 

affected Puget Sound Chinook salmon populations in a number of river systems. The 

construction and operation of dams have blocked access to spawning and rearing habitat (e.g., 

Elwha River dams block anadromous fish access to 70 miles of potential habitat) changed flow 

patterns, resulted in elevated temperatures and 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 often change flow rates, 

stranding and killing fish, and reducing aquatic invertebrate (food source) productivity (Chappell 

1980). 

Species-specific threats and limitations to recovery. Residential and commercial development 

has reduced the amount of functioning nearshore 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 rearing opportunities in nearshore and estuarine 

areas. Floodplain connectivity and function, channel structure and complexity, riparian areas 

and large wood supply, stream substrate, and water quality have also been degraded for adult 

spawning, embryo incubation, and rearing as a result of cumulative impacts of agriculture, 

forestry, and development. Total fishery exploitation rates have decreased 14 to 63% from rates 

in the 1980s, but weak natural-origin Chinook salmon populations in Puget 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 distinct life histories 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 entering 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 spawned 
119

populations of fall-run Chinook salmon in the mainstem 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 are part of this ESU. These 

artificially propagated populations are no more divergent relative to the local natural populations 

than would be expected between closely related populations within this ESU. 

Figure 9. Population boundaries, accessible areas, extirpated reaches, and dams of Snake River fall run Chinook 

distribution, incorporating critical habitat boundaries (dark black line). Taken from NMFS (2011l). 

Historically, the primary fall-run Chinook salmon 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 significantly reduced spawning and rearing habitat. Consequently, 

salmon now reside in waters that are generally cooler than pre-dam habitats. Currently, natural 

spawning occurs at the upper end of Lower Granite Reservoir to Hells Canyon Dam, the lower 

reaches of the Imnaha, Grande Ronde, Clearwater, and Tucannon rivers, and small mainstem 

sections in the tailraces of the lower Snake River hydroelectric dams. 

Adult Snake River fall-run Chinook salmon enter the 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 ocean-type and reservoir-type occur (Connor et al. 2005a). The 

reservoir-type juveniles overwinter in pools created by dams before migrating to sea; this 
120

response is likely due to early development in cooler temperatures, which prevents rapid growth. 

Phenotypic characteristics have shifted in apparent response to environmental changes from 

hydroelectric dams (Connor et al. 2005a). Migration downstream appears to be influenced by 

flow velocity within both river 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 average 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 their 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 fish 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 exhibited an upward trend in returns over Lower 

Granite Dam since the mid 1990s. Returns classified as natural-origin spawners exceeded 2,600 

fish in 2001, compared to a 1997-2001 geometric mean natural-origin count of 871 (35% of the 

proposed delisting abundance criteria 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)(NMFS 2011l). Harvest impacts declined 

after listing and have remained relatively constant in recent years. Mainstem conditions for 

subyearling Chinook migrants from the Snake River have generally improved since the early 

1990s. The hatchery component, derived from outside the basin, has decreased 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 increase in the Lyons Ferry component, 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 population due to three major components: natural-origin fish 

(which may be progeny of hatchery fish), returns 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 extinction within 100 years (NMFS 2011l). 

Critical habitat. The NMFS designated critical habitat for Snake River fall-run Chinook 

salmon on December 28, 1993 (58 FR 68543). This critical habitat encompasses the waters, 

waterway bottoms, and adjacent riparian zones of specified lakes and river reaches in the 

Columbia River that are or were accessible to listed individuals (except reaches above 

impassable natural falls, and Dworshak and Hells Canyon Dams). These areas are important for 

the species’ overall conservation by protecting quality growth, reproduction, and feeding. 

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 channel or from the shoreline of a standing body of 
121

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 accessible. Limiting factors include: 

mainstem lower Snake and Columbia hydrosystem 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 channel morphology (NMFS 2005b). 

Species-specific threats and limitations to recovery. Floodplain connectivity and function, and 

channel structure and complexity have been degraded as a result of cumulative impacts of 

agriculture, forestry, and development have hindered Snake River fall-run Chinook salmon 

recovery, as has degraded freshwater, estuarine, and marine habitat, harvest and hjatchery 

effects, as well as mainstem Columbia River and Snake River hydropower facilities (NMFS 

2011l). 

Snake River spring/summer-run Chinook salmon 

Distribution. The Snake River spring/summer run Chinook 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 propagation programs are part of the ESU. These artificially propagated 

populations are no more divergent relative to the local natural populations than would be 

expected between closely related populations within this ESU. The Interior Columbia Basin 

Technical Recovery Team identified 32 populations 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 species. Historic populations above Hells Canyon 

Dam are considered extinct (ICBTRT 2003) Table 13 identifies extant populations within the 

Snake River spring/summer Chinook salmon ESU, their abundances, and the relative 

contribution of hatchery fish. 
122

Figure 10. Population boundaries, accessible areas, extirpated reaches, and dams of Snake 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 incubate over winter and hatch in late winter and 

early spring of the following year. Juveniles mature in the river for one year before migrating to 

the ocean in the spring of their second year. Larger outmigrants have a higher survival rate 

during outmigration (Zabel and Williams 2002, Zabel and Achord 2004). Depending on 

tributary and the specific habitat 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 years in the ocean. A small fraction 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 

contributionb

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 -9.4 
123

Current populations, Mean number of 

spawners (range)a 

Percent hatchery 

contributionb

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. 

bcReported 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 Snake River spring/summer-run Chinook 

salmon as threatened on April 22, 1992 (57 FR 14653), and reaffirmed their status as threatened 

on June 28, 2005 (70 FR 37160). Although direct estimates of historical annual Snake River 

spring/summer Chinook salmon returns are not available, returns may have declined by as much 

as 97% between the late 1800s and 2000. According to Matthews and Waples (1991), total 

annual Snake River spring/summer Chinook salmon 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 run component at Lower Granite Dam was slightly 

more than 6,000 fish, compared to the geometric 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 originated in hatcheries (Good et al. 2005b). Good et al. (2005b) reported 

that risks to individual populations within the ESU may be greater than the extinction risk for the 

entire ESU due to low levels of annual abundance and the extensive production areas within the 

Snake River basin. Recent spawner returns have been higher, but low natural spawner 

abundance and the ability of the populations to cope with low ocean survival remain concerns 

(NMFS 2011l). 
124

Critical habitat. The NMFS designated critical habitat for Snake River spring/summer-run 

Chinook salmon on October 25, 1999 (64 FR 57399). This critical habitat encompasses the 

waters, waterway bottoms, and adjacent riparian zones of specified lakes and river reaches in the 

Columbia River that 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 channel 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 upstream to Hells Canyon Dam; the Palouse River from its confluence with the Snake 

River upstream to Palouse Falls, the Clearwater River from its confluence with the Snake River 

upstream to its confluence with Lolo Creek; the North Fork Clearwater River from its confluence 

with the Clearwater river upstream to Dworshak Dam. Critical habitat also includes several river 

reaches presently or historically accessible. These areas are important for the species’ overall 

conservation by protecting quality growth, reproduction, and feeding. Limiting factors identified 

for this species include hydrosystem mortality, reduced stream flow, altered channel morphology 

and floodplain, excessive fine sediment, and degraded water quality (NMFS 2006i). 

Species-specific threats and limitations to recovery. As with other listed salmonids, 

floodplain connectivity and function, channel structure and complexity, riparian areas and large 

wood supply, stream substrate, elevated water temperature, 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 (NMFS 2006i). Additional threats to recovery exist 

from mainstem Columbia River and Snake River 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, Oregon (7 populations in all)(NMFS 2011j)(Figure 

11). Seven artificial propagation programs are part of the ESU. These artificially propagated 

populations are no more divergent relative to the local natural populations than would be 

expected between closely related populations within this ESU. 
125

Figure 11. Population boundaries, dams, accessible areas, and extirpated reaches, 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 those entering the Clackamas River, must pass 

Willamette Falls. In the past, this ESU included 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 Creek. 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 other 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 reaffirmed 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 (approximately 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 River 
126

returns had been increasing in abundance until peaking in 2004, but have since dropped to 

historical levels (NMFS 2011j). Until recent years, interpretation of abundance levels has been 

confounded by a high but uncertain fraction of hatchery-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 analysis of the population trend; 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 remaining naturally reproducing population identified in 

the McKenzie River. Unfortunately, recent short-term declines in abundance suggest that this 

population may not be self-sustaining (Meyers et al. 1998, Good et al. 2005b). Abundance in 

this population has been relatively low (low thousands) with a substantial 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 are unpredictable, and the 

likelihood of long-term sustainability for this population 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 fisheries, particularly off southeast Alaska and 

northern Canada, and in the mainstem Columbia and Willamette rivers during spring. 

Critical habitat. The NMFS designated critical habitat for Upper Willamette River Chinook 

salmon on September 2, 2005 (70 FR 52630). Critical habitat includes defined areas within 

subbasins of the middle fork Willamette River, upper Willamette River, McKenzie River, 

Santiam River, Crabtree Creek, Molalla River, and 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 line is not defined the 

lateral extent is defined as the bank full elevation. 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 and rearing sites, 

freshwater migration corridors. The physical or biological features that characterize these sites 

include water quality and quantity, natural cover, forage, adequate passage conditions, and 

floodplain connectivity. Of 65 subbasins reviewed in NMFS’ assessment of ESU critical habitat, 

19 subbasins were rated as having a medium conservation 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 access to spawning/rearing 

habitat in tributaries, hatchery impacts, altered water quality and temperature in tributaries, 

altered stream flow in tributaries, and lost or degraded floodplain connectivity and lowland 

stream habitat. 

Land management activities have severely degraded stream habitat conditions in the Willamette 

River mainstem above Willamette Falls and associated subbasins. In the Willamette River 

mainstem and lower sub-basin mainstem reaches, high density urban development and 
127

widespread agricultural effects have reduced aquatic 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 other 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 channelized and stripped of large wood. Development 

began to encroach on the riparian forest beginning in the 1870s (Sedell and Froggatt 1984). 

Gregory et al. (2002d)(calculated that the total mainstem Willamette River channel area 

decreased from 41,000 to 23,000 acres between 1895 and 1995. They noted that the lower reach, 

from the mouth of the river to Newberg (RM 50), is confined within 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 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 and productivity of aquatic 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 river, reducing structural features, organic inputs 

from litter fall, entrained allochthonous materials, and flood flow filtering capacity. Extensive 

changes began before the major dams were built, 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 organic matter during flood pulses, food sources for 

macroinvertebrates, and slow-water refugia for fish during flood events. These forests also 

cooled river temperatures as the river 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 brush, marsh, and ash tree openings maintained by 

annual flood inundation. Below the City of Newberg, the most noticeable change was that 
128

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 reduced river shading and the potential for 

recruitment of wood to the river, reducing channel complexity and the quality of rearing, 

migration and spawning habitats. 

Hyporheic flow in the Willamette River has been examined through discharge measurements and 

found to be significant in some areas, particularly those with gravel deposits (Wentz et al. 1998, 

Fernald et al. 2001). The loss of channel complexity and meandering that fosters creations of 

gravel deposits decreases the potential for hyporheic flows, as does gravel mining. Hyporheic 

flow processes water and affects its quality on reemerging into the main channel, stabilizing 

variations in physical and chemical water characteristics. Hyporheic flow is important for 

ecological functions, some aspects of water quality (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 with 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 several factors, including reduced access to spawning 

and rearing habitat because of tributary dams, degraded freshwater habitat, 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 result of both tributary dams and the cumulative 

impacts of agriculture, forestry, and urban development, hatchery-related effects, and ccean 

harvest rates of approximately 30 percent (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 hindering recovery of the species. 

Chum salmon 

Description of the species. Chum salmon are more widely distributed than other salmon and 

may have at one time made up nearly 50% of the Pacific salmon biomass in the Pacific Ocean 

(Salo 1991). Historically, chum salmon were distributed throughout the coastal regions of 

western Canada and the United States, as far south as Monterey Bay, California, to the Arctic 

coast and east to the Mackenzie River, 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 found only as far south as Tillamook Bay on the 

northern Oregon coast. In this section of the Opinion, we discuss 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 indistinguishable between listed ESUs, and are the 

same biological species sharing the same generalized life history, we begin this section 

describing those characteristics common across ESUs. 

Chum salmon exhibit obligatory anadromy (there are no recorded landlocked or naturalized 

freshwater populations), and like Chinook salmon, chum salmon are semelparous (die after one 
129

spawning event). Their general life cycle spans fresh and marine waters, although chum salmon 

are more marine oriented than other Pacific salmon in that they spend very little time rearing in 

freshwater. Chum salmon spend 2-5 years in feeding areas in the northeast Pacific Ocean, which 

is a greater proportion of their life history than other Pacific salmonids. Chum salmon distribute 

throughout the North Pacific Ocean and Bering Sea, although North 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 coast in a narrow coastal band that broadens in 

southeastern Alaska, although some data suggest that Puget Sound chum, including Hood Canal 

summer run chum, may not make extended migrations 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 euryhaline water at depths ranging from the 

surface to 250 m, although juveniles are primarily 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 migrate within the Gulf of Alaska coastal belt during 

their first summer at sea (Salo 1991). Maturing individuals are also distributed widely in the 

Gulf of Alaska during spring and summer (Salo 1991). 

Chum salmon usually spawn in the lower reaches of rivers, with redds usually dug in the 

mainstem or in side channels of rivers from just above tidal influence to roughly 80 km 

upstream. Juveniles outmigrate to seawater almost 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 salmon depend less on freshwater conditions (unlike 

stream-type salmonids which depend heavily on freshwater habitats) than on favorable estuarine 

conditions. Another behavioral difference between chum salmon and species that rear 

extensively in freshwater is that chum salmon form schools, presumably to reduce predation 

(Pitcher 1986), especially if their movements are 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 water temperatures can 

influence stream entry. Sexual differences in the timing of returns to spawning grounds are 

apparent, with males generally arriving early and females later in the run. Once on the spawning 

grounds mate competition is intense with males 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 environment 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 (latitudinal trends are 

evident with northern population having lower absolute and relative fecundities)(Salo 1991). 

Size and age at maturity is partially under genetic control, but can be influenced by environment 
130

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 4 or 5 egg pockets 

(Salo 1991). Chum salmon fecundity is highly variable, and is correlated with body size and 

region (latitudinal trends are evident with northern population having lower absolute and relative 

fecundities; (Salo 1991). 

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 prolong incubation. Egg and alevin survival, and the fitness of emerging fry 

are affected by sediment loading, intergravel water flow and dissolved oxygen levels, gravel 

composition, spawning time and density, and water temperatures. 

Once they emerge from their gravel nests, chum salmon fry outmigrate to seawater almost 

immediately (Salo 1991). 

Feeding. Generally, chum fry emigrate to estuaries between March and May where they forage 

on epibenthic and neritic food resources. As food resources decline and the fish grow, they 

move further out to forage on pelagic and nektonic organisms (Simenstad and Salo 1982, Salo 

1991). The timing of juvenile entry 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 variable 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 Pacific Ocean, which is a greater 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 spaces can reduce egg survival. Other factors 

that reduce egg survival and larvae development include low dissolved oxygen, poor percolation, 

and extreme cold or warm temperatures. 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). Invasive fishes also threaten the survival and recovery of Pacific salmonids by 

competing directly for resources, altering food webs and trophic structures, and altering 

evolutionary trajectories (NMFS 2011l). 

Chum salmon have declined under the combined effects of overharvests in fisheries; competition 

from fish raised in hatcheries and native and non-native exotic species; dams that block their 

migrations and alter river hydrology; gravel mining that impedes their migration and alters the 

dynamics (hydrogeomorphology) of the rivers and streams that support juveniles; water 

diversions that deplete water levels in rivers and streams; destruction 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 (logging, agriculture, urbanization) that destroy 

wetland and riparian ecosystems while introducing sediment, nutrients, biocides, metals, and 
131

other pollutants into surface and ground water and 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-spawned populations of 

chum salmon in the Columbia River and its tributaries in Washington and Oregon (Figure 12). 

Three artificial propagation programs are part of the ESU. These artificially propagated 

populations are no more divergent relative to the local natural populations than would be 

expected between closely related populations within this ESU. 

Figure 12. Population boundaries, dams, accessible areas, and extirpated reaches of Columbia River chum 

distribution, incorporating critical habitat boundaries (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 immediately 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 Columbia (in Oregon) are considered 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 number of spawners in 

these areas are generally quite low (McElhany et al. 2007). Chum salmon return to the 

Columbia River in late fall (mid-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 River 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 years has averaged a few thousand per year, with returns limited to 

a very restricted portion of the historical range. Significant spawning occurs in only two of the 

17 historical populations (Table 14), meaning that 88% of the historical populations are 

extirpated, or nearly so. The two remaining populations are the Grays 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 appear 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 

abundancea

Recent spawner 

abundance 

Short-term median 

growth rate (λ)

c

Youngs Bay 

Gray’s River 7,511 331/704b 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 425b

 0.984 (0.883-1.096) 

Upper gorge tributaries >8,912 

a

Estimated total historical abundance for this ESU was about 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 accurate. Other estimates, long- and short-term trends, suggest 

the population is declining (see Good et al. 2005). 

Few Columbia River chum salmon have been observed 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 tributaries during the 2003 and 2004 spawning ground 

surveys. Most Columbia River chum populations have been functionally extirpated or are 
133

presently at very low abundance levels. Outmigrating 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 fishery in the first half 

of this century, which landed more than 500,000 fish 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 abundance of natural spawners for the lower 

Columbia Gorge, Washougal, and Grays River populations was below 4,000 adults. In 2002, 

however, the abundance of natural spawners exhibited a substantial increase at several locations 

(estimate of natural spawners is approximately 20,000 adults). The cause of this dramatic 

increase in abundance is unknown. However, long- and short-term productivity trends for 

populations are at or below replacement. The loss of off-channel habitat and the extirpation of 

approximately 17 historical populations increase this species’ vulnerability to environmental 

variability and catastrophic 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 for Columbia River chum salmon on 

September 2, 2005 (70 FR 52630). The designated includes defined areas in the following 

subbasins: Middle Columbia/Hood, Lower Columbia/Sandy, Lewis, Lower 

Columbia/Clatskanie, Lower Cowlitz, Lower Columbia subbasin and river corridor. 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 defined as the bank full elevation. 

The critical habitat designation for this ESU identifies primary constituent elements that include 

sites necessary to support one or more chum salmon life stages. These areas are important for 

the species’ overall conservation 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 spawning, freshwater rearing, freshwater migration, 

estuarine areas free of obstruction, nearshore marine 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, 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 altered channel form and stability in tributaries, excessive sediment 

in tributary spawning gravels, altered stream flow in tributaries and the mainstem Columbia 

River, loss of some tributary habitat types, and 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 recovery, including degraded estuarine and nearshore marine habitat resulting from cumulative impacts of land use and flow management by the 

Columbia River hydropower system and floodplain connectivity and function, channel structure 

and complexity, riparian areas, stream substrate, stream flow, and water quality have been 

degraded as a result of cumulative impacts of agriculture, hydropower, forestry, and 
134

development (LCFRB 2010, NMFS 2011d). Further habitat-based impacts to the species result 

from an altered flow regime and Columbia River 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. Also hampering the species is reduced access to 

spawning and rearing habitat mainly as a result of tributary hydropower projects, current or 

potential predation from hatchery-origin salmonids, including coho salmon, juvenile fish 

strandings that result from ship wakes, and contaminants affecting fish health and reproduction. 

Hood Canal summer-run chum salmon 

Distribution. The Hood Canal summer-run chum salmon ESU includes all naturally spawned 

populations of summer-run chum salmon in Hood Canal 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 populations 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 estuaries for 23 days; daily tidal migrations have 

not been observed, but prey availability does influence movement patterns (Bax 1983). Upon 

leaving their natal estuaries, individuals generally migrate through 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 this 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, which 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 Quilcene rivers. Hood Canal summer-run chum are the 

focus of an extensive rebuilding program developed and implemented since 1992 by state and 

tribal comanagers. Two populations (the combined Quilcene and Union River populations) are 

above the conservation thresholds established by the rebuilding plan. However, most 

populations remain depressed. Estimates of the fraction of naturally spawning hatchery fish 

exceed 60% for some populations, indicating that reintroduction programs are supplementing the 

numbers of total fish spawning naturally in streams. 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 productivity trends above replacement. Although the 

1994-2004 productivity trend for most populations was increasing, that trend generally reversed 

from 2005-2009 (NMFS 2011f). 

Of the eight programs releasing individuals 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 NMFS’ assessment of the effects of artificial 
136

propagation on ESU extinction risk concluded that these hatchery programs collectively do not 

substantially reduce the extinction risk of the ESU. 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 habitat 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 subbasin, and nearshore marine areas of Hood 

Canal and the Strait of Juan de Fuca from the line 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 (Finch, 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, pers. comm., Jan. 2010). The designation for Hood 

Canal summer-run chum, like others made at this time, 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 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 areas free of obstruction, 

nearshore marine 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, natural cover, forage, adequate passage 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 three were rated as having a medium value to the 

conservation. These areas are important for the species’ overall conservation by protecting 

quality growth, reproduction, and feeding. Limiting factors identified for this species include 

degraded floodplain and mainstem river channel structure, degraded estuarine conditions and 

loss of estuarine habitat, riparian area degradation and loss of in-river wood in the mainstems, 

excessive sediment in spawning gravels, and reduced stream flow in migration areas. 

Degradation of the near-shore environment has occurred in the southeastern areas of Hood Canal 

in recent years, resulting in late summer marine oxygen depletion and significant fish kills. 

Circulation of marine waters is naturally limited, and partially driven by freshwater runoff, 

which is often low in the late summer. However, human development has increased nutrient 

loads from failing septic systems along the shoreline, and from use of nitrate and phosphate 

fertilizers on lawns and farms. Shoreline residential 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 near-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 been altered by human activities (Brewer et al. 2007, 

NMFS 2011f). Nutrient loading has lowered dissolved oxygen concentrations, which can kill or 
137

stress marine organisms, including salmon. Residential and commercial development has 

reduced the amount of functioning habitat available for salmon rearing and migration. The loss 

of mudflats, eelgrass meadows, and macroalgae further limits salmon foraging and rearing 

opportunities in nearshore and estuarine areas. In addition, floodplain connectivity and function, 

channel structure and complexity, riparian areas and large wood supply, stream substrate, and 

stream flow have been degraded as a result of cumulative impacts of agriculture, 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 California 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 winter, 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 occurs through the species’ use of different areas 

of a river for reproduction and rearing, and the length of time they spend in these ecosystems. 

For instance, Chinook salmon spawn in fast flowing 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 slow-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 farther upstream (Sandercock 1991). 

Most Coho salmon enter rivers between September 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 provide 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 temperatures 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 there 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 occurs in a few third-order streams, but most 

spawning activity occurs in fourth- and fifth-order streams. As with other Pacific 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 exceptions; 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 within 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 years 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° to 
138

15.2° C, but optimal range is between 8° and 12° C (Emmett et al. 1991c). 

Rates of incubation are largely temperature dependent: colder water temperatures will slow 

development. Generally, in optimal temperatures 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 may move to streams with gradients of 4-5%. 

Juvenile Coho salmon are often found in small streams 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 channels that are groundwater fed). 

At not quite 2 years of age, Coho salmon will migrate downstream where they undergo the 

physiological transition to salt water. 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 prior to northward migration (PFMC 2000). This is 

particularly true for Coho originating 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 generally 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 waters. 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 aquatic and terrestrial insects. Smaller fish, such as fry, eat 

chironomids, plecopterans, and other 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 natural predation at each 

life stage. Most mortality, however, occurs in freshwater. Winter mortality may be significant 

for Coho salmon because they overwinter in freshwater, where they can be swept downstream 

from freshets or eaten by raccoon, cutthroat trout, or other small animals. Once Coho reach the 

ocean, survival is high (Sandercock 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 and double-crested cormorants annually in the Columbia River estuary (NMFS 

2011l). Invasive fishes also threaten the survival and recovery of Pacific salmonids by 
139

competing directly for resources, altering food webs and trophic structures, and altering 

evolutionary trajectories (NMFS 2011l). 

Coho salmon have declined under the combined effects of overharvests in fisheries; competition 

from fish raised in hatcheries and native and non-native exotic species; dams that block their 

migrations and alter river hydrology; gravel mining that impedes their migration and alters the 

dynamics (hydrogeomorphology) of the rivers and streams that support juveniles; water 

diversions that deplete water levels in rivers and streams; destruction 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 (logging, agriculture, urbanization) that destroy 

wetland and riparian ecosystems while introducing sediment, nutrients, biocides, metals, and 

other pollutants into surface and ground water and degrade water quality in the fresh water, 

estuarine, and coastal ecosystems 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 Francisco Bay, excluding the SacramentoSan Joaquin River system. Two natural populations have been genetically identified (NMFS 

2011a). Four artificial propagation programs are 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 populations once existsed (NMFS 2011a). Fisher 

Flats Conservation Program, Scott Creek Captive Broodstock Program, and the Noyo River Fish 

Station egg-take Program (recently terminated) coho hatchery programs. These artificially 

propagated populations are no more divergent relative to the local natural populations than 

would be expected between closely related populations within this ESU. 

Coho salmon in this ESU enter rivers to spawn very late (peaking in January), with little time 

spent in fresh water between river entry and spawning. This compressed adult freshwater 

residency appears to coincide with the single, brief peak of river 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 later 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 abundance for individual river systems. The best 

available data are from Lagunitas Creek, where red 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 for 1963, and more recent estimates suggest the 

ESU dropped to about one-fourth that size by the late 1980s and early 1990s (Good et al. 2005c). 

Where data are available, analyses of juvenile coho presence-absence information, juvenile 

density surveys, and irregular adult counts for the South Fork Noyo River indicate low 
140

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 harvest restrictions have contributed to increased returns in 2001 in 

streams in the northern portion of the ESU, as indicated by an increase in the observed presence 

of fish in historically occupied streams. Data are particularly lacking 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 coho salmon populations in several 

major river basins, and across most of the southern historical range of the ESU, represents a 

significant risk to ESU spatial 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 continues at the Noyo River and Scott Creek 

facilities, and two captive broodstock populations have recently been established. Genetic 

diversity risk associated with out-of-basin transfers appears to be minimal, but diversity risk 

from domestication selection and 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 the Russian River through the mid 1990s, but was 

terminated in 1996. Termination of this program 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 comparatively 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 this generates, was of specific concern to the 

biological review team. Informed by the VSP risk 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 “in danger of extinction.” The minority opinion 

was that this ESU is “likely to become endangered within the foreseeable future” (70 FR 37160). 

Accordingly, NMFS upgraded the status of central California coast coho ESU to endangered on 

June 28, 2005 (70 FR 37160). 

Central California Coast coho salmon populations continue to be depressed relative to historical 

numbers. Strong indications show that breeding 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 streams 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 recolonization 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 encompasses accessible reaches of all rivers 

(including estuarine areas and riverine reaches) between Punta Gorda and the San Lorenzo River 

(inclusive) in California, including two streams entering San Francisco Bay: Arroyo Corte 

Madera Del Presidio and Corte Madera Creek. This critical habitat designation includes all 

waterways, substrate, and adjacent riparian zones 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). These areas are important for the species’ overall 

conservation by protecting growth, 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 tributaries in Washington and Oregon, 

from the mouth of the Columbia up to and including the Big White Salmon and Hood Rivers, 

and includes the Willamette River to Willamette Falls, Oregon (Figure 14). Twenty-five 

artificial propagation programs are part of this ESU. 

Figure 14. Population boundaries, dams, accessible areas, and extirpated reaches of lower Columbia River coho 

distribution, incorporating critical habitat boundaries (dark black line). Taken from NMFS (2011). 

Two distinct runs distinguished by the timing of adult 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 September. Peak spawning of early returning adults occurs from 

mid-October to early November. Late returning adults exhibit a northern oceanic distribution, 
142

returning to the Columbia River from late September through December, and enter tributaries 

from October through January. Most late return adults spawn between 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 historical populations occur in the ESU; 21 are at 

high risk of extinction while the other three are fairing somewhat 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 so. Recent counts of natural-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 

counta

Geometric 

mean 

abundance 

 2000-2002b

Percent 

hatchery 

contributionsc

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 Clackamas 
143

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 this ESU produce a sizeable number of naturally 

spawned fish, the upper Sandy River population above Marmot Dam and the Clackamas River 

population above the North Fork Dam. Excluding 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 hatchery 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 for both the Sandy and Clackamas rivers, but the 

confidence intervals for each are very wide indicating there is a large amount of uncertainty. 

Both populations within the Sandy and Clackamas rivers have suffered 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 population, loss of diversity, and fragmentation and 

isolation of the remaining naturally-produced fish. 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 large 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 habitat-related factors 

inhibit the recovery of Lower Columbia River Coho salmon (LCFRB 2010, NMFS 2011d). 

Habitat impacts include degraded estuarine and near-shore marine habitat resulting from 

cumulative impacts of land use and flow management by the Columbia River hydropower 

system, fish passage barriers that limit access to spawning and rearing habitats, floodplain 

connectivity and function, channel structure and complexity, riparian 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. Additional 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 changes in the estuary. Other impacts 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 includes 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 hatchery 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. 

Basina

 Mean Spawner 

Abundanceb 

13-Year Spawner 

Trend (SE) c

Percent Hatchery 

Contributiond

Necanicum 1,889 1.169 (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 population. 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 contributions 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 was listed as a threatened species 

under the ESA on February 11, 2008 (73 FR 7816), the 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 assessed 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 (recruitment 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 abundance 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 occurred in many populations in the northern 

portion of the ESU, which were the most depressed at the time of the last review (NMFS 1997a). 

Although encouraged by the increase in spawner 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, the total abundance of natural 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-level productivity (evaluated in terms of 

the number of spawning recruits resulting from spawners 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 since 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 years. As of December 13, 2007, the average peak 

number of spawners per mile was below 2006 levels in 38 of 41 surveyed streams (ODFW 2007 

in 73 FR 7816). It is possible that 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 harvest rates, is of concern (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 specific primary constituent elements include: 

spawning sites with water and substrate quantity to support spawning, incubation, and larval 

development; freshwater rearing sites with water quantity and floodplain connectivity to form 

and maintain physical habitat conditions and support juvenile growth, foraging, behavioral 

development (e.g., predator avoidance, competition), and mobility; freshwater migratory 

corridors free of obstruction with adequate water quantity and quality conditions; and estuarine, 

nearshore and offshore areas free of obstruction with adequate water quantity, quality and 

salinity conditions that support physiological transitions between fresh- and saltwater, predator 

avoidance, foraging and other 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, almost all of it on Federal lands. The dominant 

disturbance now is logging on a cycle of approximately 30 to 100 years, with fires suppressed. 

The State of Oregon (2005) completed an assessment of habitat conditions in the range of 

Oregon coast coho salmon in 2005. Oregon’s assessment mapped how streams with high 

intrinsic potential for coho salmon rearing are distributed by land ownership categories. 

Agricultural lands and private industrial forests have by far the highest percentage of land 

ownership in high intrinsic potential areas and along all coho salmon stream miles. Federal lands 

have only about 20% of coho salmon stream miles and 10% of high intrinsic 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 that at the scale of the entire domain, pools 

are generally abundant, although slow-water and off-channel habitat (which are important 
146

refugia for coho salmon during high winter flows) are limited in the majority of streams when 

compared to reference streams in minimally-disturbed areas. Amounts of large wood in streams 

are low in all four Oregon Department of Fish and Wildlife monitoring areas 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 procedures) 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 scale, 42% of monitored sites had 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 conditions (six sites in excellent or good condition out of nine sites), 

and the Mid-South coast had the poorest conditions (no excellent condition sites, and only two 

out of eight sites in good condition). For the 10-year period monitored between 1992 and 2002, 

no sites showed a declining trend in water quality. The area with the most improving trends was 

the North Coast, where 66% of the sites (six out of nine) 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 anthropogenic factors, including floodplain connectivity and 

function, channel structure and complexity, riparian areas and large wood supply, stream 

substrate, stream flow, and water quality that have been degraded as a result of cumulative 

impacts of agriculture, forestry, instream mining, dams, road crossings, dikes, levees, etc. (Stout

et al. 2011). In addition, fish passage barriers limit access to spawning and rearing habitats and 

adverse climate, altered past ocean/marine productivity, 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 coast 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 Hatchery, Trinity River Hatchery, and Iron Gate 

Hatchery coho hatchery programs. The three major river systems supporting Southern Oregon – 

Northern Coastal California coast coho are the Rogue, Klamath (including the Trinity), and Eel 

rivers. 

Southern Oregon and Northern California coast coho 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 farther south. Because 

individuals enter rivers late, they spend much less time in the river. Coho salmon adults spawn 
147

at age 3, spending just over 1 year in fresh water and a year and a half in the ocean. 

Status and trends. Southern Oregon/Northern California coast coho salmon were listed as 

threatened on May 7, 1997 (62 FR 24588); they retained that classification when their status was 

reviewed on June 28, 2005 (70 FR 37160). Southern Oregon/Northern California Coast coho 

salmon extend from Cape Blanco in southern Oregon 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 ESU, 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 and 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, indicating that a smaller 

percentage of streams contained coho salmon compared to the percentage presence in an earlier 

study. However, it was unclear whether these new data represented actual trends in local 

extinctions or were biased by sampling effort. 

Data on population abundance and trends are limited for the California portion 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 guesses” made using a combination of limited catch 

statistics, hatchery records, and the personal observations of biologists and managers. Although 

quantitiative estimates of abundance in this ESU are rare, the best indications 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 three categories of populations: 

functionally independent populations, potentially independent populations and dependent 

populations. Functionally independent populations are populations capable of existing in 

isolation with a minimal risk of extinction. Potentially independent populations are similar but 

rely on some interchange with adjacent populations to maintain a low probability of extinction. 

Dependent populations have a high risk of extinction 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 for Southern Oregon/Northern California 

Coast coho salmon on May 5, 1999 (64 FR 24049). Critical 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 off-channel habitats) in 
148

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. These 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 floodplain and estuarine habitats, loss of riparian 

habitats and large in-river wood, reduced stream flow, poor water quality, temperature and 

excessive sedimentation, and unscreened diversions and fish passage structures. 

The Elk River flows through Curry County, and drains approximately 92 square miles (or 58,678 

acres)(Maguire 2001). Historical logging, mining, and road building have degraded stream and 

riparian habitats in the Elk River basin. Limiting factors identified for salmon and steelhead 

production in this basin include sparse riparian cover, especially in the lower reaches, excessive 

fine sediment, high water temperatures, and 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 Corps 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 estuary 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 riprap and the other north 

shore developments (Hicks 2005). Jetties constructed in 1960 to stabilize the mouth of the river 

and prevent shoaling have altered the Rogue River, 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 forest conditions, fish passage and over-wintering 

habitat. Limiting factors identified for the Upper Rogue River basin include fish passage barriers, 

high water temperatures, insufficient 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 jetties 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 high water temperature caused by lack of shade, 
149

especially in tributaries, high rates of sedimentation due to roads, poor over-wintering habitat 

due to a lack of large wood in tributaries and the mainstem, and poor quality estuary habitat 

(Maguire 2001). 

Species-specific threats and limitations to recovery. Numerous anthropogenic factors hinder 

recovery of Southern Oregon/Northern California Coast coho salmon (NMFS 2007c, 2011h). 

These include lack of floodplain and channel structure, impaired water quality, estuarine 

function, fish passage, adverse hatchery-related effects, predation, competition, and disease, 

altered sediment transport, degraded stream substrate, altered hydrologic function due to altered 

amount and timing of river flows, and degraded riparian forest conditions 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 the Sacramento River in California and northern 

Hokkaido in Japan, to as far north as far as Bathurst 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 lacustrine 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 velocity 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 juveniles of both life history types prey primarily upon 

insects. The presence of both life history types may be related to the energetic costs of 

outmigrating to sea, and the productivity of the lacustrine system they inhabit. In coastal lakes, 

where the migration to sea is relatively short and energetic costs are minimal, kokanee 

populations are rare. 

Habitat. Once smolts enter the Pacific Ocean, they distribute 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 freshwater move north along the Alaskan coastal belt beginning in mid-
150

May until early winter when they disperse into the Gulf of Alaska (Emmett et al. 1991b, NMFSAKR 2005). 

Thermoclines may also influence vertical distribution, with fish only mingling between surface 

and deeper waters when the boundary temperature difference is weak. Sockeye appear to prefer 

cooler waters relative to other salmon species, but younger salmon may prefer warmer sea 

surface temperatures (optimally 15° C) than larger, 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 stress fish by increasing their susceptibility to disease and 

elevating their metabolism (Brett 1979). Maturation 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 summer and autumn, but the precise time can 

vary greatly among populations. Age at maturity varies by region from 2 to 5 years, but is 

generally 2 to 4 years in Washington State (Burgner 1991). Males often arrive earlier than 

females on the spawning grounds, and will persist 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. Fecundity 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 rapidly downstream or upstream along the banks to 

the lake rearing area. Fry emerging from lakeshore 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 trophic structures, and altering evolutionary 

trajectories (NMFS 2011l). 

Ozette Lake sockeye salmon 

Distribution. This ESU includes all naturally spawned sockeye salmon in Ozette Lake, Ozette 

River, Coal Creek, and other tributaries flowing into Ozette Lake, Washington (Figure 15). 

Composed of only one population, the Ozette Lake sockeye salmon ESU consists of five 

spawning aggregations or subpopulations which are grouped according to their spawning 

locations. The five spawning locations are Umbrella and Crooked creeks, Big River, and 
151

Olsen’s and Allen’s beaches (NMFS 2009c). 

Figure 15. Population boundaries of Lake Ozette sockeye distribution, incorporating critical habitat boundaries 

(dark black line). Taken from NMFS (2011).

Adult Ozette Lake sockeye salmon enter Ozette Lake through the Ozette River from mid-April to 

mid-August, holding three to nine months in Ozette Lake prior to spawning in late October 

through January. Sockeye salmon spawn primarily in lakeshore upwelling areas 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 all 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 tributaries in November and December. Eggs 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+ smolts, 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 retained on June 28, 2005 (70 

FR 37160). This ESU includes all naturally spawned populations of sockeye salmon in Ozette 

Lake, Ozette River, Coal Creek, and other tributaries flowing into Ozette Lake, Washington. 

Two artificial propagation programs are considered part 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 related natural populations (70 FR 37160). 
152

The historical abundance of Ozette Lake sockeye 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 believed to have declined substantially from historical levels. In 

the first study of lake escapement 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 that they were likely conducted 

upstream from fisheries in or near to the Ozette River. Makah Fisheries Management (MFM 

2000) concluded that there appears to be a substantial decline in the tribal catch of Ozette Lake 

sockeye salmon beginning in the 1950s and a similar decline in the run size since the 1920s weir 

counts reported by Kemmerich (1945). 

An analysis of total annual Ozette Lake sockeye salmon abundance (based on adult run size data 

presented in Jacobs et al. 1996) indicates a trend in abundance 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 declining population trend identified for the Ozette Lake 

sockeye salmon population. The most recent (1996 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 estimates before 1998 are likely to be even more 

unreliable than recent counts, and new counting 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 aggregations. Although overall abundance 

appears to have declined from historical levels, 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 capacity, a viable sockeye salmon population in Lake 

Ozette watershed would range between 35,500 to 121,000 spawners (Rawson et al. 2009). 

There has been no harvest of Ozette Lake sockeye salmon for the past four brood-cycle years 

(since 1982). Prior to that time, ceremonial and 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, due to the early river entry timing of returning 

Ozette Lake sockeye salmon (beginning in late April, with the peak returns prior to late-May to 

mid-June), the fish are not intercepted in Canadian and United States marine area fisheries 

directed at Fraser River sockeye salmon. There are currently no known marine area harvest 

impacts on Ozette Lake sockeye salmon. 

Overall abundance is substantially below historical levels (Good et al. 2005b). Declines in 

abundance have been attributed to a combination of introduced species, predation, loss of 

tributary populations, a loss of quality of beach spawning habitat, temporarily unfavorable ocean 

conditions, habitat degradation, and excessive historical harvests (Jacobs et al. 1996). In the last 

few years the number of returning adults has increased, although some of these individuals are of 

hatchery origin. This produces uncertainty regarding natural growth rate and productivity of the 

ESU’s natural component. In addition, genetic integrity has perhaps been compromised due to 

the artificial supplementation that has occurred in this population, since approximately one 

million sockeye have been released into the Ozette watershed from the late 1930s to present 
153

(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 specific geographic areas designated as critical 

are the Hoh/Quillayute Subbasin, Ozette Lake and 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 specific primary constituent 

elements identified for Lake Ozette sockeye salmon are areas for spawning, freshwater rearing 

and migration, estuarine areas free of obstruction, nearshore marine 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, natural 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 tributary basin is 77 mi2

 and includes several large tributaries and numerous 

smaller tributaries. Currently, land ownership 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 hydrogeomorphic events, while contemporary 

disturbance additionally includes logging, road construction and maintenance, residential and 

agricultural development, stream channelization and direct and indirect stream wood clearance. 

These activities alter stream flow patterns and elevate of sediment loads and increased 

sedimentation within drainage basins. Wood removal has resulted in less hydraulic roughness, 

reduced instream water depths, and reduced backwater effects on Lake Ozette, which has thus 

altered the entire hydraulic control on Lake Ozette levels and changed the in-river stagedischarge relationship. More recently, deposition of sediment originating from Coal Creek at the 

lake outlet has further altered lake and river levels (Haggerty et al. 2009b). 

Private timber companies own approximately 93% of the four largest tributary 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 depended on stream size, gradient, and time after 

harvest. In high-energy coast streams, landslides and debris torrents often modify steep slope 

tributaries and the mainstem of creeks. Bank erosion 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 juvenile 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 forest canopy, have lower biomass of trout and fewer predator 

taxa than old-growth sites (Murphy and Hall 1981). Based on the quantity and quality of the 

physical and biological features, the conservation 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 complexity, riparian areas and large wood supply, lake beach spawning 

habitat, and stream substrate have been degraded as a result of cumulative impacts of forest 

practices, agriculture, and development (Haggerty et al. 2009a, USDC 2009a, Muir and Williams 
154

2011). Harbor seals and river otters, and predaceous non-native and native fish species, are 

reducing the abundance of adult fish that successfully 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 sockeye 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 separated 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 substantially lower elevations (Wenatchee 

at 1,870 feet and Okanagon at 912 feet) and occupy 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 individuals from the Redfish Lake Captive Broodstock 

Program. 

Figure 16. Population boundaries, accessible areas, and extirpated 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 originally 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 population in the Snake River basin at the time of 

listing was that in Redfish Lake. Other lakes in the Snake River basin historically supported 
155

sockeye salmon populations, including Wallowa Lake (Grande Ronde River drainage, Oregon), 

Payette Lake (Payette River drainage, Idaho) 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 the time of listing (1991). Subsequent to the 1991 

listing, a residual form of sockeye residing in Redfish 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 salmon ESU. In 1993, the NMFS determined that 

residual sockeye salmon in Redfish Lake were part of the Snake River sockeye salmon. Also, 

artificially propagated sockeye salmon from the Redfish Lake Captive Propagation program are 

considered part of this species (70 FR 37160; June 28, 2005). 

The NMFS has determined that this artificially propagated population is genetically no more 

than moderately divergent from the natural population (NMFS 2005f). Five lakes in the Stanley 

Basin historically contained sockeye salmon: Alturas, Pettit, Redfish, Stanley and Yellowbelly 

(Bjornn et al. 1968). It is generally believed that adults were prevented from returning to the 

Sawtooth Valley from 1910 to 1934 by Sunbeam Dam. Sunbeam Dam was constructed on the 

Salmon River approximately 20 miles downstream of Redfish Lake. Whether Sunbeam Dam 

was a complete barrier to adult migration remains unknown. It has been hypothesized that some 

passage occurred while the dam was in place, 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 were 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 Department of Fish and Game eradicated sockeye 

salmon from Pettit, Stanley, and Yellowbelly lakes, and built permanent structures on each of the 

lake outlets that prevented re-entry of anadromous sockeye salmon (Chapman and Witty 1993). 

In 1985, 1986, and 1987, 11, 29, and 16 sockeye, respectively, were counted at the Redfish Lake 

weir (Good et al. 2005b). Only 18 natural origin sockeye 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 broodstock program returned to the 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 sockeye salmon in the Stanley Basin have been 

extremely low. No natural origin anadromous adults have returned since 1998 and the 

abundance of residual sockeye salmon in Redfish Lake is unknown. This species is entirely 

supported by adults produced through the captive propagation program at the present time. 

Current smolt-to-adult survival of sockeye originating 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 biological review team members that the species 

remains in danger of extinction (Good et al. 2005b). However, the most recent run abundances 

from 2008 and 2009 have been more robust, with 650 and 809 individuals, respectively, 
156

returning to the Sawtooth Hatchery (NMFS 2011l). 

Critical habitat. Critical habitat for these salmon was designated on December 28, 1993 (58 FR 

68543), and encompasses the waters, waterway bottoms, and adjacent riparian zones of specified 

lakes and river reaches in the Columbia River that 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 channel 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 upstream to the confluence of the Salmon River; all 

Salmon River reaches to Alturas Lake Creek; Stanley, 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 River. 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 the species because it provides spawning and juvenile 

rearing habitat, areas for juvenile growth and development, and migration corridors for smolts to 

the ocean and adults to spawning habitat from the Pacific Ocean. Limiting factors identified for 

Snake River sockeye include: reduced tributary stream flow, impaired tributary passage and 

blocks to migration, and mainstem Columbia 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. Portions 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 adult returns to Lower Granite 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 salmon smolts reaching the estuary, and piscivorous 

fish consume an estimated 8% of migrating 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 northwestern 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 type, summer steelhead, enters freshwater in a sexually immature 

condition and requires several months in freshwater to mature and spawn; and the oceanmaturing type, winter steelhead, enters freshwater with well-developed gonads and spawns 

shortly after river entry. Variations in migration timing exist between populations, and some 

river basins have both summer and winter steelhead, while others only have one race. 

There is a high degree of overlap in spawning timing between populations regardless of run type 
157

(Busby et al. 1996d). Difficult field conditions 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 populations 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, although 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 migration in early spring to natal 

streams, and then spawn in January and February (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 (Barnhart 1986). Some adults, however, do not enter 

some coastal streams until spring, just before spawning (Meehan and Bjornn 1991). 

As with other salmonids, the larger the fish, the more eggs produced. Egg and hatching success 

are related to the conditions within the redd and 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 actively feeding. After emerging from the gravel, fry 

usually inhabit shallow water along banks of perennial streams. Fry occupy stream margins 

(Nickelson et al. 1992). Summer rearing takes place primarily in the faster parts of pools, 

although young-of-the-year are abundant 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 their first summer and occupy a range of habitats 

featuring moderate to high water velocity and variable depths (Bisson et al. 1988). Steelhead 

hold territories close to the substratum where flows 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 freshwater from 1 to 4 years, then smolt 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 migrate directly offshore during their 

first summer from whatever point they enter the ocean rather than migrating along the coastal 

belt as salmon do. Steelhead typically reside in marine waters for 2 or 3 years prior to returning 

to their natal stream to spawn as 4- or 5-year olds; 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 (chironomids, baetid mayflies, and hydropsychid 
158

caddisflies (Merz 1994). Adults feed on aquatic and terrestrial insects, mollusks, crustaceans, 

fish eggs, minnows, and other small fishes (including greenling and other trout; (Chapman and 

Bjornn 1969, Stolz and Schnell 1991)). Survival at smoltification 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 and decreases with age. For example, Puget 

Sound steelhead leaving freshwater and estuarine 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 natural predation each life stage. The highest 

mortality occurs between the egg stage and smolt 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 and 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. Invasive fishes also threaten the 

survival and recovery of Pacific salmonids by competing directly for resources, altering food 

webs and trophic structures, and altering evolutionary trajectories (NMFS 2011l). 

Steelhead have declined under the combined effects of overharvests in fisheries; competition 

from fish raised in hatcheries and native and non-native exotic species; dams that block their 

migrations and alter river hydrology; gravel mining that impedes their migration and alters the 

dynamics (hydrogeomorphology) of the rivers and streams that support juveniles; water 

diversions that deplete water levels in rivers and streams; destruction 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 introducing sediment, nutrients, biocides, metals, and 

other pollutants into surface and ground water and degrade water quality in the fresh water, 

estuarine, and coastal ecosystems 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 and substrate quantity to support spawning, 

incubation, and larval development. 

 Freshwater rearing sites with water quantity and floodplain connectivity to form and 

maintain physical habitat conditions and support juvenile growth, foraging, behavioral 

development (e.g., predator avoidance, competition), and mobility. Specific features 

include forage supporting juvenile development 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 migration corridors free of obstruction with water quantity and natural cover 

such as submerged and overhanging large wood, aquatic vegetation, large rocks and 
159

boulders, side channels, and undercut banks to support mobility and survival. Without 

these features, juveniles cannot avoid high flows and predators, successfully compete, 

begin the behavioral and physiological changes needed for marine life, or out-migrate. 

 Estuarine areas free of obstruction with water quality and salinity conditions supporting 

juvenile and adult physiological transitions between fresh- and saltwater. Estuaries must 

also include natural cover (submerged and overhanging large wood, aquatic vegetation, 

large rocks and boulders), side channels as well as prey for both juveniles and adults. 

These features are essential to conservation because without them juveniles cannot reach 

the ocean in a timely manner and use the variety of habitats that allow them to avoid 

predators, compete successfully, and complete the behavioral and physiological changes 

needed for life in the ocean. 

 Nearshore marine areas free of obstruction with water quality and quantity conditions and 

forage, including aquatic invertebrates and fishes, supporting growth and maturation; and 

natural cover such as submerged and overhanging large wood, aquatic vegetation, large 

rocks and boulders, and side channels. No areas 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 and maturation. These features are essential 

for conservation because without them juveniles cannot forage and grow to adulthood. 

Central California coast steelhead 

Distribution. The Central California Coast steelhead DPS includes all naturally spawned 

anadromous steelhead populations below natural and manmade impassable barriers in California 

streams from the Russian River (inclusive) to Aptos Creek (inclusive), and 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 Sacramento-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 the 
160

specific demographics and life history characteristics 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 juveniles 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 California Coast steelhead DPS and their estimated 

run sizes. 

Table 19. Central California coast steelhead populations and their estimated abundances. 

Basin Estimated abundancea 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 here 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” (see 
161

(Good et al. 2005b) for a discussion). Other 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 the DPS are more recent (Table 26). Two rivers 

thought to have contained the largest populations within the DPS were the Russian River, and the 

San Lorenzo River. Based on run size estimates from the 1990s, the Russian River is still likely 

the largest run within the DPS, albeit estimates suggest the population has declined between 90-

96% from the 1970 level. 

No current estimates of total population size are available for this DPS, and consequently there is 

no time series data available to evaluate the central California coast steelhead population trends. 

Rather, a general dearth of data on adult steelhead within the DPS 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 portion 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 data from a number of watersheds (presumed populations). As a 

result, the team derived trend estimates for five populations: the San Lorenzo River, Scott 

Creek, Waddell Creek, Gazos Creek, and Redwood Creek in Marin County (see (Good et al.

2005b) for a detailed discussion of the approach). All populations exhibited downward trends in 

abundance. Accordingly, provided the 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 the biological review team 

(69%) considered this DPS likely to become endangered thus supporting the renewal of the 

threatened status for central California coast steelhead. Notably, 25% of the team voted that the 

DPS be upgraded to endangered status (voted the DPS as in danger of extinction)(Good et al.

2005b). Abundance and productivity were of relatively high concern (as a contributing factor to 

risk of extinction), and spatial structure was also of concern. 

Since the original status review, fishing regulations have changed that probably reduces 

extinction risk for Central California Coast steelhead. Ocean sport harvest is prohibited, and 

ocean harvest is considered rare. Although freshwater 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-and-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 Mateo, Bay Bridge, Santa Clara, San Pablo, and Big 

Basin. These areas are important for the species’ overall conservation by protecting quality 

growth, reproduction, and feeding habitat. The critical habitat designation for this ESU identifies 
162

primary constituent elements that include sites necessary to support one or more steelhead life 

stages. Specific sites include freshwater spawning sites, freshwater rearing sites, freshwater 

migration corridors, nearshore marine habitat and estuarine areas. The physical or biological 

features that characterize these sites include water quality and quantity, natural cover, forage, 

adequate passage conditions, and floodplain 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 that were excluded from designation. 

In total, Central California Coast steelhead occupy 46 watersheds (freshwater and estuarine). 

The total area of habitat designated as critical includes about 1,500 miles of stream habitat and 

about 400 square miles of estuarine habitat (principally Humboldt Bay). 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 defined 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 are migrating in the nearshore zone and relying 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 habitat, 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 inaccessible 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 greatly 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 ascended 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 drainages of the Feather, American, Yuba, and 

Stanislaus Rivers which are now inaccessible (McEwan and Jackson 1996a, Yoshiyama et al.

1996). 

Existing wild steelhead populations in the Central 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 and 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) indicate that steelhead are present in Clear Creek (Good et al. 2005a). 

Because of the large resident O. mykiss population in Clear Creek, steelhead spawner abundance 

has not been estimated. Until recently, steelhead were thought to be extirpated from the San 

Joaquin River system. Recent monitoring has detected small self-sustaining populations of 

steelhead in the Stanislaus, Mokelumne, Calaveras, and other streams previously thought to be 
163

devoid of steelhead (McEwan 2001). On the Stanislaus River, steelhead smolts have been 

captured in rotary screw traps at Caswell State Park and Oakdale each year since 1995 (Demko 

and Cramer 2000). It is possible that naturally spawning 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 the only migration route to the drainages of the 

Sierra Nevada and southern Cascade mountain ranges for anadromous fish. The CDFG 

considers all steelhead in the Central Valley as winter steelhead, although “three distinct runs,” 

including summer steelhead, may have occurred there as recently as 1947 (McEwan and Jackson 

1996a)(CDFG 1995). Steelhead in these basins travel extensive distances in freshwater (some 

exceed 300 km to their natal streams), making these the longest freshwater migrations of any 

population of winter steelhead. The upper Sacramento 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 extend 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 data, 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). Over the past 30 years, the naturally spawned 

steelhead populations in the upper Sacramento River have declined substantially. Hallock et al.

(1961) estimated an average of 20,540 adult steelhead 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 Dam counts; (McEwan and 

Jackson 1996a, McEwan 2001). The five-year geometric mean, however, is just under 2,000 

steelhead (Table 20), and the long-term trend suggests that the population is declining. 

Table 20. California Central Valley steelhead and their long-term trend. 

Population 5-year mean (min – max)a λ Long-term trendb

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 available on steelhead numbers in the 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 early 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 Valley steelhead populations have generally 

declined, with hatchery fish forming a larger 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 hatchery 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 River over nearly the same time frame. A very 

small number (40-70) of redds have been counted 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 habitat in California’s Central Valley has 

been lost, largely due to mining and water development activities. They also noted that declines 

are “due mostly to water development, inadequate instream flows, rapid flow fluctuations, high 

summer water temperatures in streams immediately below reservoirs, diversion dams which 

block access, and entrainment of juveniles into unscreened or poorly screened 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 (agriculture and forestry) and 

urbanization have also contributed to population declines. It is unclear how harvest has affected 

California’s Central Valley steelhead, although it is likely a continuing threat. A CDFG creel 

census in 2000 indicated that most fish are caught and released, but due to the size of the catch 

and release fishery (more than 14,000 steelhead were caught and released according to the 

survey) even a small amount of mortality in this fishery could cause declines in the populations. 

Critical habitat. The NMFS designated critical habitat for California Central Valley steelhead 

on September 2, 2005 (70 FR 52488). Specific geographic areas designated include the 

following CALWATER hydrological units: Tehama, 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-Mendota Canal, North Diablo Range, and the San 

Joaquin Delta. These areas are important for the species’ overall conservation by protecting 

quality growth, reproduction, and feeding habitat. The critical habitat designation for this ESU 

identifies primary constituent elements that include sites necessary to support one or more 

steelhead life stages. Specific sites include freshwater spawning sites, freshwater rearing sites, 

freshwater migration corridors, nearshore marine habitat and estuarine areas. The physical or 

biological features that characterize these sites include water quality and quantity, natural cover, 

forage, adequate passage conditions, and floodplain 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 that were excluded from designation. 

In total, California Central Valley steelhead occupy 67 watersheds (freshwater and estuarine). 

The total area of habitat designated as critical includes about 2,300 miles of stream habitat and 

about 250 square miles of estuarine habitat in the San Francisco-San Pablo-Suisan Bay estuarine 

complex. 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 defined 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 are migrating in the nearshore zone and relying 

on cover and refuge qualities provided by these habitats, and while they are foraging. Of the 67 
165

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 Cowlitz 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 identifies the populations that comprise Lower Columbia River steelhead 

(23 historically) and summarizes several measures available to characterize population viability. 

Figure 17. Population boundaries, dams, accessible areas, and 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 

abundancea

Mean number 

of spawners 

Percent 

hatchery 

contribution 

Median shortterm growth 

rate (λ)

b

Winter Cispus River 

 Tilton River 2,787c 73 

 Upper Cowlitz River 

 Lower Cowlitz River 1,672 
166

Life 

history 

Population Historical 

abundancea

Mean number 

of spawners 

Percent 

hatchery 

contribution 

Median shortterm growth 

rate (λ)

b

 Coweeman River 2,243 466d 50 0.920, 0.787 

 South Fork Toutle River 2,627 504d 2 0.933, 0.929 

 North Fork Toutle River 3,770 196d

 0 1.038, 1.038 

 Kalama River 554 726d

 0 0.984, 0.922 

 North Fork Lewis River 713 

 East Fork Lewis River 3,131 

 Salmon Creek 

 Washougal River 2,497 323d

 0 

 Clackamas River 560e

 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 756f

 52 

Summer Wind River 2,288 472g

 5 0.995, 0.903 

 Hood River 931f

 83 Unknown 

 Washougal River 1,419 264g

 8 1.029, 0.960 

 East Fork Lewis River 422 434g

 25 

 North Fork Lewis River 

 Kalama River 3,165 474g

 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 reproduce 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 sexually immature to the Columbia River from May to November, and 

spend several months in freshwater prior to spawning. Winter steelhead enter freshwater from 

November to April, are close to sexual maturation during freshwater entry, and spawn shortly 

after arrival in their natal streams. Where both races spawn in the same stream, summer 

steelhead tend to spawn at higher elevations 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 status 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 naturally spawning steelhead populations. During this 

review the NMFS was unable to identify any natural populations that would be considered at low 

risk. 

All populations declined between 1980 and 2000, with sharp declines beginning in 1995. Those 

with adequate data for modeling are estimated to have a high extinction risk (Good et al. 2005b). 

Abundance trends are generally negative, showing that most populations are in decline, although 
167

some populations, particularly summer run, have shown 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 impassable 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 hypothesized to be sustained largely by hatchery 

production. Exceptions are the Kalama, the Toutle, and East Fork Lewis winter-run populations. 

These populations have relatively low recent mean abundance estimates with the largest being 

the Kalama (geometric mean of 728 spawners). 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 greatly 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 Columbia/Sandy subbasin, Lewis subbasin, Lower 

Columbia/Clatskanie subbasin, Upper Cowlitz subbasin, Cowlitz subbasin, Clackamas subbasin, 

Lower Willamette subbasin, and the Lower Columbia River corridor. These 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 constituent elements that 

include sites necessary to support one or more steelhead life stages. Specific sites include 

freshwater spawning sites, freshwater rearing sites, freshwater migration corridors, nearshore 

marine habitat and estuarine areas. The physical or biological features that characterize these 

sites include water quality and quantity, natural cover, forage, adequate passage conditions, and 

floodplain connectivity. The critical habitat designation (70 FR 52630) contains additional 

description of the watersheds that are included as part of this designation, and any areas 

specifically excluded from the designation. 

In total, Lower Columbia River steelhead occupy 32 watersheds. The total area of habitat 

designated as critical includes about 2,340 miles of stream habitat. This designation includes the 

stream channels within the designated stream reaches, and 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. Of the 32 watersheds reviewed in NMFS' assessment 

of critical habitat, two watersheds received a low rating of conservation value, 11 received a 

medium rating, and 26 received a high rating of conservation value for the species. Limiting 

factors identified for Lower Columbia River steelhead include: degraded floodplain and steam 

channel structure and function, reduced access to spawning/rearing habitat, altered stream flow 

in tributaries, excessive sediment and elevated water temperatures in tributaries, and hatchery 

impacts. 

Species-specific threats and limitations to recovery. Several direct and habitat-related factors 

inhibit the recovery of Lower Columbia River Coho salmon (LCFRB 2010, NMFS 2011d). 

Habitat impacts include degraded estuarine and near-shore marine habitat resulting from 
168

cumulative impacts of land use and flow management by the Columbia River hydropower 

system, fish passage barriers that limit access to spawning and rearing habitats, floodplain 

connectivity and function, channel structure and complexity, riparian 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. Additional 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 changes in the estuary. Other impacts also include hatchery- and 

harvest-related effects, juvenile fish strandings that result from ship wakes, avian and marine 

mammal predation in the lower mainstem Columbia 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 natural 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 propagation program are part of this DPS. These 

artificially propagated populations are considered no more divergent relative to the local natural 

populations than would be expected between closely related natural populations within the DPS. 

Figure 18. Population boundaries, dams, accessible areas, and 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 intermontane region of the Pacific northwest, 
169

which includes some of the driest areas in the region generally receiving 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 characterized 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 (ICTRT 2003a) this DPS is comprised of 16 putative 

populations in four major population groups (Cascades Eastern Slopes Tributaries, John Day 

River, Walla Walla and Umatilla Rivers, and Yakima River) and one unaffiliated independent 

population (Rock Creek). Table 22 lists extant (putative) populations that compose this DPS. 

There are two extinct populations in the Cascades Eastern Slope major population group, the 

White Salmon River and Deschutes Crooked River above the Pelton/Round Butte Dam complex. 

Present population structure is delineated largely on the basis of geographical proximity, 

topography, distance, ecological similarities or differences. Additional genetic studies are 

needed to describe the DPS substructure, as well as the fine-scale genetic 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. 

Populationa

 Major population 

groups 

Mean number of 

spawners (range)b

Percent 

hatchery 

contributionc

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, redds, 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 consists 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 values 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 Klickitat 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. Factors contributing to decline 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 reaffirmed 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 collected between 1997 and 2001 indicates 

that several populations within the DPS have increased since the 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 Yakima River, the Deschutes River and 

sections of the John Day River system were substantially higher compared to numbers surveyed 

between 1992 and 1997 (Good et al. 2005b). Similarly, spawner numbers substantially increased 

in the Umatilla River and Fifteenmile Creek relative to annual levels in the early 1990s. 

Nonetheless, most populations remain below interim target levels. For instance, the Yakima 

River returns are still substantially 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 within the DPS, given that the average population 

growth rate across all streams is negative (0.98 assuming 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 Tributaries 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 River 
171

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. These 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 constituent elements that include sites necessary to 

support one or more steelhead life stages. Specific sites include freshwater spawning sites, 

freshwater rearing sites, freshwater migration corridors, nearshore marine habitat and estuarine 

areas. The physical or biological features that characterize these sites include water quality and 

quantity, natural cover, forage, adequate passage conditions, and floodplain connectivity. The 

final rule (70 FR 52630) lists the watersheds that comprise the designated subbasins and any 

areas that are specifically excluded 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 miles of stream habitat. 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 defined as the bank full elevation. Of the 114 watersheds 

reviewed in NMFS' assessment of critical habitat for Middle Columbia River steelhead, nine 

watersheds received a low rating of conservation value, 24 received a medium rating, and 81 

received a high rating of conservation value for the species. Although pristine habitat conditions 

are still present in some wilderness, roadless, and undeveloped areas, habitat complexity has 

been greatly reduced in many areas of designated critical habitat for Middle Columbia River 

steelhead. Limiting factors identified for Middle Columbia River steelhead include: hydropower 

system mortality, reduced stream flow, impaired passage, excessive sediment, degraded water 

quality, and altered channel morphology and floodplain. 

Species-specific threats and limitations to recovery. Floodplain connectivity and function, 

channel structure and complexity, riparian areas, fish passage, stream substrate, 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 mainstem Columbia 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 Okanogan Rivers (Table 23). All Upper Columbia 

River steelhead are summer steelhead. Steelhead primarily use streams of this region that drain 

the northern Cascade Mountains of Washington State. This species includes hatchery 
172

populations of summer steelhead from the Wells Hatchery because it probably retains the genetic 

resources of steelhead populations 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, and 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 quickly to their natal tributaries. A portion of the returning run 

overwinters in mainstem reservoirs, passing over upper- to mid-Columbia dams in April and 

May of the following year. Spawning occurs in the late spring of the year following river entry. 

Juvenile steelhead spend one to seven years rearing in freshwater before migrating to sea. Smolt 

outmigrations are predominantly year class two and three (juveniles). Most adult steelhead 

return after from sea after 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 originally listed as endangered in 

1997 (62 FR 43937), after their status was reviewed, 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 Upper Columbia River steelhead have been 

based on extrapolations from mainstem dam counts and associated sampling information (e.g., 

hatchery/wild fraction, age composition). Returns of both hatchery and naturally produced 

steelhead to the upper Columbia River have increased 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 fish. 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 annual steelhead run over Priest Rapids Dam increased from an 

average of 1,040 (1992-1996), representing about 10% of the total adult count, to 2,200 (1997-

2001), representing about 17% of the adult count during this period of time (ICBTRT 2003). 

Since 2005, abundance estimates have increased in all four populations within the ESU, although 

a large portion of returns are hatchery origin fish (NMFS 2011i). 

In terms of natural production, recent population abundances for the Wenatchee, Entiat, and the 

Methow population remain well below the minimum abundance thresholds developed for these 

populations (ICBTRT 2005). A five-year geometric mean (1997 to 2001) of approximately 900 

naturally produced steelhead returned to the Wenatchee and Entiat rivers (combined). Although 

this is well below the minimum abundance thresholds, it represents an improvement (an 

increasing trend of 3.4% per year). However, the average percentage of natural fish for the 

recent five-year period dropped from 35% to 29%, compared to the previous status review. For 

the Methow population, the five-year geometric mean of natural returns over Wells Dam was 

358. Although this is well below the minimum abundance threshold, it is increasing trend of 

5.9% per year. In addition, the 2001 return (1,380 naturally produced spawners) was the highest 

single annual return in the 25-year data series. However, the average percentage of wild origin 

spawners dropped from 19% for the period prior to the 1998 status review to 9% for the 1997 to 

2001 returns. This DPS is failing to meet viability criteria in all four categories; productivity, 

abundance, spatial structure, and genetic diversity. Overall, long-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 spatial structure risk as “low” for the Wenatchee 

and Methow, “moderate” for the Entiat, and “high” for the Okanogan. 

Critical habitat. Critical habitat was designated for 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. The critical habitat designation for this DPS 

identifies PCEs that include sites necessary to support one or more life stages 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 value to the DPS. In addition, the Columbia 

River rearing/migration corridor 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 streamflow, (3) tributary riparian degradation and loss of in-river 

wood, (4) altered tributary floodplain and channel morphology, 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 anthropogenic factors, including impaired tributary fish 

passage, hatchery and harvest effects, completion, predation, disease, as well as fisheries 

management, including past introductions and persistence of non-native (exotic) fish species that 

continue to affect habitat conditions 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 cumulative 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 river basins as far west as the 

Elwha River and as far north as the Nooksack River. Puget Sound's fjord-like structure may 

affect steelhead migration patterns; for example, some populations of Coho and Chinook salmon, 

at least historically, remained within Puget Sound and did not migrate to the Pacific Ocean itself. 

Even when Puget Sound steelhead migrate to the high seas, they may spend considerable time as 

juveniles or adults in the protected marine environment of Puget Sound, a feature not readily 

accessible to steelhead from other areas of the Pacific northwest. This species is primarily 

composed of winter steelhead but includes several stocks of summer steelhead, usually in 

subbasins of large river systems and above seasonal hydrologic barriers. Life history attributes 

of Puget Sound steelhead (migration and spawn timing, smolt age, ocean age, and total age at 

first spawning) appear similar to those of other west coast steelhead. 

Figure 20. Population boundaries, dams, accessible areas, and extirpated reaches of Puget Sound steelhead 
175

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 stocks were of native origin and 

predominantly natural production. Their assessment of the status of these 31 stocks was 11 

healthy, three depressed, one critical, and 16 of unknown status. Their assessment of the status 

of the remaining (not native/natural) stocks was three healthy, 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 trends, with a range from 18% annual decline (Lake Washington 

winter-run steelhead) to 7% annual 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 available for summer-run steelhead. Most of these trends were based on 

relatively short data series. The Skagit and Snohomish River winter-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 modest overall upward trends at the time of the Busby et al. (1996d) report. 

Busby et al. (1996d) estimated five-year average natural 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 all but a few small populations were currently 

declining at a 3-10% rate annually, with a high risk 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—nearly 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—about 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 calculate 
176

(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 within 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 evaluation to designate future 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. Further impairment results from declining 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 reduced habitat quality through changes in river 

hydrology, temperature profile, downstream gravel recruitment, and reduced movement of large 

woody debris. Further habitat-based threats include 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 urban 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 and sinuosity, have increased the 

likelihood of gravel scour and dislocation of rearing juveniles. 

Northern California steelhead 

Distribution. Northern California steelhead includes steelhead in California coastal river basins 

from Redwood Creek south to the Gualala River, inclusive. Table 25 identifies populations 

within the Northern California Steelhead salmon ESU, their abundances, and hatchery input. 

Both winter and summer steelhead occur in this ESU, with immature half-pounders returning 

from the sea after only two to four months and overwintering 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 Unknown 
177

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 reviewed 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 1930s), the South Fork Eel River (Benbow Dam– 

annual average number of adults was 19,000 in the 1940s), and the Mad River (Sweasey Dam– 

annual average number of adults was 3,800 in the 1940s). Estimates of steelhead spawning 

populations for many rivers in this DPS totaled 198,000 by the mid-1960s. At the time of the 

first status review, adult escapement trends could be computed on seven populations. Five of the 

seven populations exhibited declines, while two exhibited 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 snorkel 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 model 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 trends were apparent (Good et al. 2005b). 

Critical habitat. Critical habitat was designated for this species on September 2, 2005 (70 FR 

52488). The critical habitat designation 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 inland 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, Pahsimeroi, Lemhi, Selway, Clearwater, Wallowa, 

Grande Ronde, Imnaha, and Tucannon Rivers. Table 26 identifies populations within ESU, their 

abundances, and hatchery input. 
178

Figure 21. Population boundaries, dams, accessible areas, and extirpated reaches, of Snake River steelhead 

distribution, incorporating critical habitat boundaries (dark black line). Taken from NMFS (2011l). 

Snake River steelhead occupy a basin that is annually warmer and drier than other steelhead 

DPSs. Snake River Basin steelhead are generally classified as summer run, based on their adult 

run timing pattern. Sexually immature 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 migrate upriver until they reach tributaries from 3,300-6,600 feet above 

sea level where they spawn between March and May of the following year. Adult spawner 

survival is much higher for individuals 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 typically 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 Unknown 
179

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 predominately one-year-old ocean 

fish. B-run steelhead are predominately two-year-old ocean fish. A-run populations are found in 

lower Clearwater River, upper Salmon River, lower 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 four major subbasins not occupied by A-run 

steelhead: two on the Clearwater 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 tributaries. 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 DPS (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 classification (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 steelhead 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 and warmer in these areas than in areas occupied by other 

steelhead species in the Pacific northwest. Snake 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 particularly depressed. 

The paucity of information on adult spawning escapement for specific tributary production areas 

for Snake River Basin steelhead made a quantitative 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 Ronde, and Imnaha Rivers. The 2001 return over 

Lower Granite Dam was substantially higher relative to the low levels seen in the 1990s; the 

recent 5-year mean abundance (14,768 natural returns) was approximately 28% of the interim 

recovery target level (52,000 natural spawners). The 10-year average for natural-origin steelhead 

passing Lower Granite Dam between 1996 and 2005 was 28,303 adults. Parr densities in natural 

production areas, which are another indicator of population status, have 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). Homogenization of hatchery stocks occur within basins, 

and some stocks exhibit high stray rates. Little change from the 2005 review was found in the 
180

2011 status review (NMFS 2011l). 

Critical habitat. Critical habitat was designated for this species on September 2, 2005 (70 FR 

52488). The critical habitat designation for this ESU identifies PCEs that include sites necessary 

to support one or more life stages of steelhead. The critical habitat designation for this ESU 

identifies PCEs that include sites necessary to support one or more steelhead 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 identified for Snake River Basin 

steelhead include: (1) hydrosystem mortality, (2) reduced stream 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 increased 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, 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, 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 Steelhead salmon ESU, their abundances, and hatchery input. 
181

Figure 22. Population boundaries, dams, accessible areas, and extirpated reaches, 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 migration past Willamette 

Falls also begins in early April and proceeds into early June, peaking in early- to mid-May 

(Howell et al. 1985a). Smolts generally migrate through 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 FR 
182

14517) and when their status was reviewed on January 5, 2006 they retained that classification 

(71 FR 834). A major threat to Upper Willamette River steelhead results from artificial 

production practices. Fishways built at Willamette Falls in 1885 have allowed Skamania-stock 

summer steelhead and early-migrating winter steelhead of Big Creek stock to enter the range of 

Upper Willamette River steelhead. The population of summer steelhead is almost entirely 

maintained by hatchery salmon, although natural-origin, Big Creek-stock winter 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 abundance studies support a recent upturn in population 

numbers until 2002, when the population peaked and has since returned to historical levels (2008 

abundance was 4,915 individuals passing Willamette Falls and 2,110 in 2009)(McElhaney et al.

2007, NMFS 2011j). Spatial structure for the North and South Santiam populations has been 

substantially reduced by the loss of access to the upper North Santiam basin and the Quartzville 

Creek watershed in the South Santiam subbasin 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 Calapooia 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 historical habitat, legacy effects of past winter-run hatchery 

releases, and the ongoing release of summer steelhead (McElhaney et al. 2007). 

Critical habitat. Critical habitat was designated for this species on September 2, 2005 (70 FR 

52488). The designation includes all Columbia River 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 Santiam, South Santiam, Middle Willamette, 

Molalla/Pudding, Yamhill, Tualatin, and Lower Willamette (NMFS 2005b). The critical habitat 

designation for this DPS identifies PCEs that include sites necessary to support one or more life 

stages of steelhead. Of 43 subbasins reviewed in NMFS' assessment of critical habitat for Upper 

Willamette River steelhead, 20 subbasins were rated 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 conditions in the Willamette 

River mainstem above Willamette Falls and associated subbasins. In the Willamette River 

mainstem and lower sub-basin mainstem reaches, high density urban development and 

widespread agricultural effects have reduced aquatic 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 other 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 channelized and stripped of large wood. Development 

began to encroach on the riparian forest beginning in the 1870s (Sedell and Froggatt 1984). 

Gregory et al. (2002d) calculated that the total mainstem Willamette River channel area 

decreased from 41,000 to 23,000 acres between 1895 and 1995. They noted that the lower reach, 

from the mouth of the river to Newberg (RM 50), is confined within 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 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 and productivity of aquatic 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 river, reducing structural features, organic inputs 

from litter fall, entrained allochthonous materials, and flood flow filtering capacity. Extensive 

changes began before the major dams were built, 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 organic matter during flood pulses, food sources for 

macroinvertebrates, and slow-water refugia for fish during flood events. These forests also 

cooled river temperatures as the river 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 brush, marsh, and ash tree openings maintained by 

annual flood inundation. Below the City of Newberg, 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 reduced river shading and the potential for 

recruitment of wood to the river, reducing channel complexity and the quality of rearing, 

migration and spawning habitats 

Hyporheic flow in the Willamette River has been examined through discharge measurements and 

found to be significant in some areas, particularly those with gravel deposits (Wentz et al. 1998, 

Fernald et al. 2001). The loss of channel complexity and meandering that fosters creations of 
184

gravel deposits decreases the potential for hyporheic flows, as does gravel mining. Hyporheic 

flow processes water and affects its quality on reemerging into the main channel, stabilizing 

variations in physical and chemical water characteristics. Hyporheic flow is important for 

ecological functions, some aspects of water quality (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 with 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 recovery of Upper Willamette River steelhead 

(Guimarães et al. 2011, ODFW and NMFS 2011). Floodplain connectivity and function, 

channel structure and complexity, riparian areas and large wood recruitment, and stream flow 

have been degraded as a result of cumulative impacts of agriculture, forestry, and development. 

Degraded water quality and altered temperature as a result of both tributary dams and the 

cumulative impacts of agriculture, forestry, and urban 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 spawning tributaries. Genetic effects of non-native 

summer steelhead hatchery programs and anthropogenic 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 Pajaro River (inclusive) to, but 

not including the Santa Maria River, California. Only winter steelhead are found in this ESU. 

Migration and spawn timing are similar to adjacent steelhead populations. 

Status and trends. South-Central California Coast steelhead were listed as threatened in 1997, 

when their status was reviewed on January 5, 2006 they retained that classification (71 FR 834). 

Historical data on the South-Central California Coast steelhead DPS are limited. In the mid1960s the California Department of Fish and Game estimated that the adult population at about 

18,000. We know of no recent estimates of the total DPS. However, five river systems, the 

Pajaro, Salinas, Carmel, Little Sur, and Big Sur, indicate that runs are currently less than 500 

adults. Past estimates for these basins were almost 5,000 fish. Carmel River time series data 

indicate that the population declined by about 22% per year 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 local steelhead population. Other possibilities were 

considered including that the substantial immigration or transplantation occurred, or that resident 

trout production increased as a result of improved environmental conditions within the basin. 

Critical habitat. Critical habitat was designated for this species on September 2, 2005 (70 FR 

52488). The critical habitat designation for this DPS identifies PCEs that include sites necessary 

to support one or more life stages of steelhead. The critical habitat designation 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 rivers 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 input. Migration and life history patterns of 

southern California steelhead are dependent on rainfall and streamflow (Moore 1980). Steelhead 

within this ESU can withstand higher temperatures than more northerly populations and the 

relatively warm and productive waters of the Ventura River permit more rapid 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 that classification (71 

FR 834). In many watersheds throughout Southern California, dams isolate steelhead from 

historical spawning and rearing habitats and alter the hydrology of the basin (e.g., Twitchell 

Reservoir within the Santa Maria River watershed, Bradbury Dam within the Santa Ynez River 

watershed, Matilija and Casitas dams within the 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 estimated 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 steelhead in the Ventura River, which has its 

headwaters in Los Padres National 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 genetically distinct but physically 

indistinguishable clades: a Northern DPS whose populations are relatively healthy, and a 

Southern DPS that has undergone 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. Both 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 estuary, Willapa Bay, Grays Harbor and southeast 
186

Alaska. In general, green sturgeon are more common north of Point Conception, California. 

This DPS occupies a limited number of river systems, 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 River is the only known spawning habitat 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 the 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, Rogue, and Columbia rivers of northern California 

and Oregon. Individuals have also rarely been captured in Southern California waters, but their 

abundance increases north of Point Conception. 

Upon outmigration from fresh water, subadult green sturgeon disperse 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 sturgeon distribution and habitat in estuaries and the North Pacific 

Ocean. Adults are known from coastal estuaries and bays, including Willapa and Nehalem bays, 

Grays Harbor, and the Columbia 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 and 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 estuaries and coastal marine waters (Wilson and McKinley 2004). 

Sturgeon, like salmon, possess strong site fidelity and 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 salt water in the fall early, often with the first 

increases in fall flows, although Northern DPS individuals 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 months 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 downstream migration (Erickson and 

Webb 2007a, Heublein et al. 2009). In the Sacramento River adult green sturgeon spawn in late 

spring and early summer above Hamilton City, above Red Bluff Diversion Dam, and possibly as 

far upstream as Keswick Dam (CDFG 2002, Heublein et al. 2009). It appears that specific 

habitat for spawning includes large cobblestones (where eggs can settle between), although 

spawning is known to occur over clean sand or bedrock. 

Adults likely return on a 2 to 5 year basis for spawning 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) while females on the spawning grounds are often older 
187

(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 for egg survival with temperatures 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), although they are believed to be 

physiologically adapted to sea water survival at 6 months of age (Allen 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 being moved downstream more than 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 sturgeon of the Northern DPS (Rogue 

River) have been captured in the Rogue River estuary 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-San 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 green sturgeon show that they consume benthic 

invertebrates including shrimp, clams, chironomids, 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 use 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 predominant sources, fisheries and tagging. Only three 

data sets were considered useful for the population time series analyses by NMFS’ biological 

review team: the Klamath Yurok Tribal fishery catch, a San Pablo sport fishery tag returns, and 

Columbia River commercial landings (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 Pablo data suggest that green sturgeon abundance may 

be increasing, but the data showed no significant trend. The data set is not particularly 

convincing, however, as it suffers from inconsistent effort and since it is unclear whether 

summer concentrations of green sturgeon provide a strong indicator of population performance 

(NMFS and USFWS 2005). Although there is not sufficient information available to estimate 
188

the current population size of southern green sturgeon, catch of juveniles during state and federal 

salvage operations in the Sacramento delta are low in comparison to catch levels before the mid1980s. 

Natural threats. Green sturgeon eggs and larvae are likely preyed upon by a variety of larger 

fish and animals, while sub-adult and adult sturgeon may occasionally be preyed upon by shark 

sea lions, or other large body predators. Physical barriers, changes 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 impassible barriers (e.g., Oroville, Shasta and 

Keswick dams). Other threatens include potentially lethal temperature limits, harvest, 

entrainment by water projects and toxins and invasive species (Adams et al. 2007, Erickson and 

Webb 2007b, Lackey 2009). Since this DPS is composed of a single spawning population within 

the Sacramento River, stochastic variation in environmental conditions and significant 

fluctuations in demographic rates increases the risk of extinction for this DPS. 

Climate change has the potential to affect sturgeon in similar, if not more significant ways it 

affects salmonids. Elevated air temperatures could lead to precipitation falling as rain instead of 

snow. Additionally, snow would likely melt sooner and more rapidly, potentially leading to 

greater flooding during melting and lower water levels at other times, as well as warmer river 

temperatures. Although sturgeon can spawn over varied 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 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 et al. 2000). Sturgeon are also 

sensitive 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 (ISAB 2007b). If river and lake temperatures 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 temperatures may disrupt 

habitat, including the availability 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 areas (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 likely 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 (Kruse and Scarnecchia 2001). Mercury has also 

been identified from white sturgeon of the lower Columbia River (Webb et al. 2006). Numerous 
189

organochlorines, including DDT, DDD, DDE, chlordane, 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 critical habitat for southern green 

sturgeon (74 FR 52300). The geographical area identified as critical habitat is based upon the 

overlapping distribution of the southern and northern DPS, and encompasses all areas where the 

presence of southern green sturgeon have been confirmed or where 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 River, 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 the entire biological 

species range within a depth of 60 fathoms. The principle biological or physical constituent 

elements essential for the conservation of southern green sturgeon in fresh water include: food 

resources; substrate of sufficient 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 corridors; water depth; and sediment quality. Primary constituent 

elements of estuarine habitat include food resources, water flow, water quality, migratory 

corridors, water depth, and sediment quality. The specific primary constituent elements of 

marine habitat include food resources, water quality, and migratory corridors. 

Critical habitat of the Southern DPS of green sturgeon is threatened by several anthropogenic 

factors. Four dams and several other structures currently are impassible for green sturgeon to 

pass on the Sacramento, Feather, and San Joaquin 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 reproductive health of Southern DPS green sturgeon 

through bioaccumulation. Other activities of concern include those that may disturb bottom 

substrates, adversely affect prey resources, or degrade water quality through re-suspension of 

contaminated sediments. Of particular concern are activities that affect prey resources. Prey 

resources can be affected by: commercial shipping and activities generating point source 

pollution and non-point source pollution that can discharge contaminants and result in 

bioaccumulation of contaminants in green sturgeon; disposal of dredged materials that can bury 

prey resources; and bottom trawl fisheries that can disturb the bottom (but may result in 

beneficial or adverse effects on prey resources for green sturgeon). In addition, petroleum spills 

from commercial shipping activities and proposed alternative energy hydrokinetic projects may 

affect water quality or hinder the migration of green sturgeon along 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, south to the St. John’s River in Florida. NMFS’ 

recovery plan (1998b) recognized 19 wild populations based on their strong fidelity to their natal 

streams (Table 29), and several captive populations (from a Savannah River broodstock) that are 
190

maintained for educational and research purposes (NMFS 1998b). Although these populations 

are geographically isolated, genetic analyses suggest that individual 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) 1998-2005 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 Riverd

 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 Rivere

 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 2006 
191

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 researchers based on their sampling effort and success. c

Range represents total population estimates using four different 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 migration) 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 rates 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 while 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 Delaware 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 transients 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 shortnose 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 preferred deep-water curves with sand 

and cobble to higher velocity runs, particularly during winter months, and shifted to channel 

habitat as water temperatures rose in summer months. Many fish also exhibited 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, shortnose sturgeon congregate in 

cool, deep, areas of rivers where adult and juvenile sturgeon can take refuge from high 

temperatures (Flournoy et al. 1992, Weber 1996a). In the Connecticut River and the Merrimack, 

Kynard et al. (2000) found shortnose generally used water about 3 meters deep, ranging from 
192

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 migrations), 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 shortnose sturgeon will migrate 

upstream to spawning areas in the spring or in the fall. Fish that migrate upstream in the fall 

generally overwinter in areas just downstream of spawning sites, while others including 

non-spawners will overwinter in estuarine waters. After spawning in the spring, spent 

(post-spawned) adults tend to migrate rapidly downstream to feeding areas in the estuary or to 

tidally influence fresh water during summer, when 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 occur year-round in estuarine waters, spending midOctober to mid-April in the upper estuarey and moving downstream 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 sturgeon move downstream after hatching, remaining in fresh water 

for about 1 year (Kynard 1997b). Initially, young shortnose sturgeon will reside 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 their year at the saltand fresh water interface of coastal 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, between 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 cued to decreasing river discharge following the 

peak spring freshet, when water temperatures range 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 for shortnose sturgeon and age at first spawning 

appears to vary along a latitudinal cline. According 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, varying on a latitudinal 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 maximum age reported for a 

shortnose sturgeon in the St. John River in New Brunswick is 67 years (for a female), 40 years 
193

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 mouths, structured for 

feeding on benthos. Foraging generally occurs in areas with abundant macrophytes, where 

juvenile and adult shortnose sturgeon feed 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 likely important to foraging, while 

vision is thought to play a minor role (Miller 2004). As adults, a significant portion of a 

shortnose sturgeon’s diet may consist 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 portion of the species’ range live longer than 

individuals in the southern portion of the species’ 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 the Pee Dee River, and 10 years in the Altamaha River (Dadswell

et al. 1984b). Male shortnose sturgeon appear to have shorter life spans than females (Gilbert 

1989a). 

Status and trends. Shortnose sturgeon were listed as endangered on March 11, 1967, under the 

Endangered Species Preservation Act (32 FR 4001) and remained on the endangered species list 

with enactment of the ESA of 1973, as amended. Pollution and overfishing, including bycatch in 

the shad fishery, were listed as principal reasons for the species' decline. Shortnose sturgeon are 

listed as an endangered species throughout all of their range. 

Northern shortnose sturgeon population abundances are generally larger than southern 

populations (Kynard 1997b). Updated population estimates also suggest that three of the largest 

populations (Kennebec, Hudson, and Delaware River) 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 based on available data exhibits appears to have increased (NMFS 

1998b, Bain et al. 2000). The most recent population estimate indicates this population consists 

of about 61,000-shortnose sturgeon (95% confidence interval [CI] was between 52,898 and 

72,191 fish (Bain et al. 2000)). A comparison of the Bain estimate to the 1979/1980 population 

estimate of spawning adults by Dovel et al. (1992); about 13,000 fish) led Bain et al. (2000) to 

conclude that the population had made a dramatic increase (about 400 % increase) between 1979 

and 1997. While still evidence of an increasing population, a comparison of total population 

estimates (30,000:60,000) would suggest the population has only doubled in size during the 

study years. Similarly, the Kennebec River population appears to be increasing. Early 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 population 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 the current population for the Delaware River is 12,047 adult fish 

(1999-2003; 95% CI: 10,757-13,589), which is similar 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 longevity, 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 interannual variation in juvenile recruitment 

would be expected as a result of stochastic factors 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 survival 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 stable equilibrium. 

South of Chesapeake Bay, populations are relatively 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 shown 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 smaller than that in the Altamaha River. The highest point estimate in 1993 

using a modified Schnabel technique resulted in a total population estimate 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 population may be declining. 

Annual variation in population estimates in many basins is due to changes in yearly capture rates, 

which are strongly correlated with weather conditions (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 shortnose sturgeon population in the basin. 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 assumptions used by the various studies. Long-term 

data sets and an open population model would likely provide for more accurate population 

estimates across the species range, and could provide the opportunity to more closely link strongyear classes to habitat 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 documented in the St. 

John’s River (FL), the St. Mary’s River, Chesapeake Bay, Potomac River, Piscataqua River, the 

Housatonic River, and others. Some basins probably previously contained shortnose 

populations, but recent sampling has been largely unsuccessful. Despite the occasional 

observations of shortnose sturgeon, populations may be extinct in several basins (e.g., St. John’s 

(FL), St. Mary’s, Potomac, Housatonic, and Neuse rivers). Those few fish that have been 

observed in these basins are generally presumed to be immigrants from neighboring basins. In 
195

some cases, (e.g. Chesapeake Bay) migratory information collected from tagged fish and genetic 

evidence confirms that fish captured in Chesapeake Bay were part of the Delaware River 

population (Grunwald et al. 2002, Wirgin et al. 2005b). 

Natural threats. Yellow perch, sharks, and seals are predators of shortnose sturgeon juveniles 

(NMFS 1998a). The effects of disease and parasites are generally unknown. 

Anthropogenic threats. Shortnose sturgeon have declined from 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 dams divide shortnose sturgeon 

spawning habitat (e.g., Connecticut River, Penobscot 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 Penobscot River in 1977 five river kilometers were 

opened to sturgeon and other anadromous fishes. With the recent signing of the Penobscot River 

Restoration Trust, access may be restored 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 twentieth 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 incidentally 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 others (COSEWIC 2005, Bahn and Peterson 2009). 

Habitat alterations from discharges, dredging or disposal of material into waterways, and other 

developmental activities along riverine and estuarine 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 structure, and water quality (increase in suspended sediments). 

Shoreline development of liquefied natural gas facilities and alternative power sources also alters 

coastal habitats through changes in benthic communities by dredging, changes in water quality 

and water temperatures, and may increase the potential 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 America, increasing interests in finding alternative energy sources is 

expected to result in an increase in the number of marine turbines along the coast. 

Fish kills have also been observed where estuaries are affected by urban and agricultural 

discharges that cause vegetative blooms and eutrophic conditions. Extreme declines in dissolved 

oxygen levels have occurred periodically throughout the species range. In the late 1960s and 
196

early 1970s, dissolved oxygen levels reached zero ppm in the Penobscot, Kennebec, and 

Androscoggin rivers and estuaries during the summer. Extreme low dissolved oxygen levels are 

have also plagued Chesapeake Bay. In most cases, 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 and other benthic organisms are regularly in direct contact with 

legacy pollutants, as well as a suite of common contaminants added from more current industrial 

and agricultural practices. Studies demonstrate that shortnose sturgeon carry a wide number of 

potentially hazardous contaminants. Individuals from the Delaware River contain numerous 

metals (mercury, aluminum, antimony, barium, cadmium, calcium, chromium, copper, iron, 

magnesium, manganese, nickel, potassium, sodium, 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 could lead to precipitation falling as rain instead of 

snow. Additionally, snow would likely melt sooner and more rapidly, potentially leading to 

greater flooding during melting and lower water levels at other times, as well as warmer river 

temperatures (ISAB 2007c). Although sturgeon can spawn over varied benthic habitat, they 

prefer localized depressions 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 spawning habitat that sturgeon 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 sensitive 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 (ISAB 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 habitat, including the availability of prey (ISAB 2007c). 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 areas (ISAB 2007c). 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 likely that global warming 

will increase pressures on sturgeon survival and recovery. 

Critical habitat. Critical habitat has not been established for shortnose sturgeon. 

Smalltooth sawfish 
197

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 circumtropical distribution. 

Although they are rays, sawfish physically more resemble sharks, with only the trunk and 

especially the head ventrally flattened. Smalltooth sawfish are characterized by their “saw,” a 

long, narrow, flattened rostral blade with a series of transverse teeth along either edge. 

Distribution. In the western Atlantic, the smalltooth 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 the Atlantic are truly smalltooth sawfish or 

closely related species is unknown (Adams and 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 habitat 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 and summer periods (May to August) when inshore waters reach 

appropriately high temperatures. The data also suggest that smalltooth sawfish may utilize warm 

water outflows of power stations as thermal refuges during colder months to enhance their 

survival or become trapped by surrounding cold water 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 typically well above ambient 

temperatures. 

Movement. Historic records of smalltooth sawfish indicate 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 two smalltooth sawfish have been recorded north 

of Florida since 1963 (the first was captured off of North Carolina in 1999 and the other off 

Georgia 2002) but it is unknown whether these individuals resided in Georgia and North 

Carolina waters annually or if they had migrated 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) hypothesize the population previously undertaking the 

summer migration has declined to a point where the migration is undetectable or does not occur. 

Most specimens captured along the Atlantic coast north of Florida have also been large (>2.7 m) 

adults and likely represent seasonal migrators, wanderers, or colonizers from a core population(s) 

to the south rather than being members of a continuous, even-density population (Bigelow and 

Schroeder 1953). It is likely that these individuals migrated southward toward Florida as water 
198

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, Florida, to Florida Bay (NMFS 2000, Simpfendorfer 

and Wiley 2004).3 

Habitat. Smalltooth sawfish are euryhaline, occurring in waters with a broad range of salinities 

from freshwater to full seawater (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 preferred habitat; juveniles also expand 

their ranges, whereas small individuals have very restricted 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. These banks are often very small and daily home 

range sizes can be of the magnitude of 100–1,000 m2 (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 (Wiley and Simpfendorfer 2007). Their occurrence in 

freshwater is suspected to be only in estuarine areas temporarily freshwater from receiving high 

levels of freshwater input. Many encounters are reported at the mouths of rivers or other sources 

of freshwater inflows, suggesting estuarine areas may be an important factor in the species 

distribution (Simpfendorfer and Wiley 2004). 

Information on juvenile smalltooth sawfish indicates 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 been relocated in the same general areas over periods of several 

months, suggesting a high level of site fidelity (Simpfendorfer 2003). The daily home ranges of 

these animals are considerably larger (1–5 km2

) 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, the 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 animals are more likely to be found in deeper waters. 

Poulakis and Seitz (2004b) reported that almost all of the sawfish <3 m in length were found in 

water less than 10 m deep and 46% of encounters 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 research {Poulakis, 2004 #1266;Seitz, 2002 #1426;Simpfendorfer, 

2004 #1465;Simpfendorfer, 2004 #1465}. 
199

occur regularly in waters in excess of 164 feet (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 and Wiley (2004), however, 

caution this may be an overestimate, with recent anecdotal information suggesting smaller litter 

sizes (about ten). Smalltooth sawfish mating and pupping seasons, gestation, and reproductive 

periodicity are all unknown. Gestation and reproductive periodicity, however, may be inferred 

based on that of the largetooth sawfish, sharing the same genus and having similarities in size 

and habitat. Thorson (1976) reported the gestation period for largetooth sawfish was 

approximately five months and concluded that females probably produce litters every second 

year. 

Bigelow and Schroeder (1953) describe smalltooth sawfish as generally about 61 cm long at 

birth and growing to a length of 5.5 m or greater. Recent data from smalltooth sawfish caught 

off Florida, however, demonstrate young are born at 76 to 87 cm (Simpfendorfer and Wiley 

2004), with males reaching maturity at approximately 2.7 m and females at approximately 3.6 m 

(Simpfendorfer 2002). A recent study by Simpfendorfer suggests rapid juvenile growth occurs 

during the first two years after birth (Simpfendorfer 2008). First year growth is 65-85 cm and 

second year growth is 48-68 cm. Growth rates beyond two years are uncertain; however, the 

average growth rate of captive smalltooth sawfish has been reported between 13.9 and 19.6 cm 

per year. The maximum reported size of a smalltooth 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 sawfish have been conducted to date, but growth 

studies of largetooth sawfish suggest 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 increase (Simpfendorfer 2000b). 

Simpfendorfer estimated intrinsic rates of natural 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, with 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 20th century, the smalltooth sawfish population declined 

dramatically during the middle and 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. Smalltooth sawfish were historically caught as bycatch 

in various fishing gears throughout their historic range, including gillnet, otter trawl, trammel 

net, seine, and to a lesser degree, handline. Frequent accounts in earlier literature document 
200

smalltooth sawfish being entangled in fishing nets 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 many 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 through 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) documented recent (1990 to 2002) 

occurrences of sawfish along the southwest coast of Florida, and in Florida Bay and the Florida 

Keys, respectively and includes a total of 2,969 smalltooth sawfish encounters. Mote Marine 

Laboratory also maintains a smalltooth sawfish public encounter database, established in 2000 to 

compile information on the distribution and abundance of sawfish. A total of 434 sawfish 

encounters have been validated since 1998, most from recreational 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 collected 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 decrease the greater the distance from the core area (Simpfendorfer 

and Wiley 2004). The capture of a smalltooth sawfish off Georgia in 2002 is the first record 

north of Florida since 1963. New reports during 2004 extend the current range of the species to 

Panama City, offshore Louisiana (south of Timbalier Island in 100 feet of water), southern Texas 

(unconfirmed), and the northern coast of Cuba. 

The abundance of juveniles encountered, including very small individuals, suggests that the 

population remains reproductively active and viable (Seitz and Poulakis 2002, Simpfendorfer 

2003a, Simpfendorfer and Wiley 2004). The declining numbers of individuals with increasing 

size is consistent with the historic size composition data (G. Burgess, pers. comm. in 

Simpfendorfer and Wiley 2004). This information and recent encounters in new areas beyond 

the core abundance area suggest that the population may be increasing. From 1989-2004, 

smalltooth sawfish relative abundance has increased by about 5 percent per year (Carlson et al.

2007). However, recovery of the species expected to be slow on the basis of the species’ life 

history and other threats to the species remaining (see below), the population’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 hundreds 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 population sizes, this biological parameter means that 

recovery, at best, will be slow, and that catastrophic 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, resulting in drowning or injury. Even when 

individuals that have been entangled are retrieved alive, individuals 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 sawfish. Although a concern, pollution impacts on 

particularly reproductive biology are unknown. However, habitat 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 southwest coast of Florida (74 FR 45353). Although PCEs were not 

identified, the mangrove and adjacent shallow euryhaline habitat are important 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 Mexican and California and is separated from a 

northern population by a region of apparent 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 responsible 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, California, and south of Point 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 largely 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 depths of 12-478 m, but tend to remain in 

shallow waters on the continental shelf (20-250 m), still associating 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). Artificial 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 occupy 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 grow larger (Hart 1973b, Echeverria 1987, Love et al.

2002). However, either sex has been known to attain sexual maturity as small as 35 cm or 3 

years of age and, in recent years 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 born 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 waters, 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). Individuals 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 larger 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 juveniles consume fish larvae, 

copepods, and krill, while older, nearshore juveniles and adults prey upon rockfishes, hake, 

sablefish, anchovies, lanternfish, and squid (Reilly et al. 1992, Love et al. 2002). 

Acoustics and hearing. Data regarding bocaccio hearing are not available. However, field 

measurements have recorded bocaccio calls, which 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 abundance 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 by 
203

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 failure of recruitment classes, 

possibly because of unfavorable climactic/oceanographic 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 capable 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 failures, 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, with 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 diverted resources from reproduction, potentially 

resulting in additional impairment to recovery. Overall, bocaccio have the highest variability of 

recruitment of any rockfish studied to date, with recruitment exhibiting a random walk and high 

temporal variability (MacCall and He 2002b, Tolimieri and Levin 2005). 

Although population estimates are not available for the northern population, the southern 

population has been estimated to number 1.6 million fish of 1 year of age or older in 2002 

(MacCall 2002a). Of these, 1.0 million were estimated to occur south of Pt. Conception, where 

recruitment has been stronger. However, individuals 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 Sound has been estimated at ~3% annually for the 

period 1965-2007. Various rebuilding estimates for bocaccio populations have predicted 

recovery, but require long periods (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 competition, predators, and climactic regimes are the primary 

natural factors that depress bocaccio numbers. Copper and quillback rockfish may compete with 
204

bocaccio in Puget Sound for available resources (NMFS 2008d). King salmon, lingcod, terns 

and other seabirds, harbor seals, and Steller sea 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 negatively influenced by El Niño conditions, possibly 

reducing available prey supply (Moser et al. 2000a, Harvey 2005). 

Anthropogenic threats. Although overfishing is the primary reason for bocaccio being 

proposed as a listed species, bycatch and habitat loss are also human-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 fishing, 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 intentionally. 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 habitats (reportedly, there are only 

217 km2

 in Puget Sound) being threatened by construction 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, exacerbate 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 composed of at least two populations 

and possibly more. Yamanaka et al. (2006) found that those individuals found within the 

Georgia Basin and Queen Charlotte Strait were genetically distinct from other samples from 

Oregon to Alaska. The Georgia Basin/Queen Charlotte Sound population is the one which has 

been proposed for listing in US waters. 

Distribution. Yelloweye rockfish occur from Baja California to the Aleutian Islands, but are 

most common from central California to Alaska (Love et al. 2002). 

Habitat. As with other rockfishes, yelloweye habitat varies based upon life stage. Larvae 

maintain a pelagic existence but as juveniles, move into shallow 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 debris or pilings (Lamb and Edgell 1986). As adults, yelloweye 

rockfish move in to deeper habitats. Individuals 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, particularly with caves and 

crevices, pinnacles, and boulder fields (Carlson 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, residential aggregations infrequently found (Coombs 1979, DeMott 

1983, Love et al. 2002). In the Puget Sound region, sport 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 with internal fertilization. Copulation 

occurs between September and April, with fertilization 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 believed 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 nearshore habitats and settling into 

rocky reef habitat at about 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 latitudinal gradient, with 

individuals in more northerly regions growing faster and larger. Year class strength appears to 

be most strongly linked to survival of the larval stage (Laidig et al. 2007). In general, sexual 

maturity appears to be reached by 50% of individuals 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 years)(Munk 2001). Maximum size has been reported as 910 cm, but 

assymptotic size in Alaskan waters for both males 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. Juveniles tend to begin life in 

shallow rocky reefs and graduate to deeper rocky 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 yelloweyes eat other rockfish (including members 

of their own species), sand lance, gadids, flatfishes, shrimp, 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 greater 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 importantly, age classes appear to have been truncated 

to younger, smaller fish, severely hampering the ability 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 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. Over the 

period of 1965-2007, it is estimated that rockfish species has declined by 3% per year. 

The most recent estimate of yelloweye rockfish abundance in the Puget Sound region was 3,000 

individuals, with low abundance through spawning areas (Palsson et al. 2008). 

Natural threats. Interspecies competition, predators, and climactic regimes are the primary 

natural factors that depress yelloweye rockfish numbers. Copper and quillback rockfish may 

compete with yelloweye rockfish in Puget 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, Lance 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 temperature 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 primary cause of yelloweye rockfish 

decline throughout their range, including 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 generation in the past few decades, an astonishing rate of decline. 

Habitat loss is also a factor in yelloweye decline, with rocky habitats (reportedly, there are only 

217 km2

 in Puget Sound) being threatened by construction of bridges, sewer 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 populations compose canary rockfish as a 

species. Genetic analysis have found that individuals south of Cape 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 populations is still lacking. 

Distribution. Canary rockfish are found from the northern Baja peninsula north to the western 

Gulf of Alaska, and with the greatest abundance 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 habitats based upon their life stage. Larvae and 

younger juveniles tend to occupy shallow waters at the beginning of their lives, but generally 

remain in the upper 100 m of the water column (Love et al. 2002). Juveniles initially settle into 

tide pools and rocky reefs (Miller 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, Boehlert 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 (Moser 1996b, Methot and Stewart 2005, Tissot et al. 2007). 

Mid shelf locations seem to have the highest concentrations of canary rockfish off Washington 

and Oregon (Weinberg 1994). Adults tend to occur in shallow areas in higher latitudes than their 

southern counterparts, although adults do appear to move into progressively deeper waters as 

they age (Vetter and Lynn 1997, Methot and Stewart 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, Bellingham, 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 rockfish 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 depths in late winter to 100-170 m in late 

summer (COSEWIC in press). 

Reproduction. Canary rockfish develop their young internally before giving birth to live young 

as larvae. During each annual spawning event, 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 gradient associated with number 

of eggs produced (Gunderson et al. 1980, Love et al. 2002). Birth takes place in Oregonian and 

Washingtonian waters between September through March, with a peak in December and 

January. The peak in British Columbian waters 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 (Waldron 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 reaching 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 Island 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, IGFA 
208

1991, Williams et al. 1999, Love et al. 2002, Methot and Stewart 2005). 

Foraging. Canary rockfish prey upon different species 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 lifestyle as well as eating euphasiids and other 

crustaceans. Adults consume other 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 for 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 frequently caught in the Puget Sound and Georgia Basin 

region since 1965 (NMFS 2008d). The rate of decline 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 individuals 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 caught greater than 55 cm since 2000, 

whereas a variety of large size classes had formerly been caught. There are 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 competition, predators, and climactic regimes are the primary 

natural factors that depress canary rockfish numbers. Copper and quillback rockfish may 

compete with canary rockfish in Puget Sound for available resources (NMFS 2008d). Predators 

of canary rockfish include other rockfishes, lingcod (for which rockfish is a particularly 

important dietary component), cabezon, seabirds, salmon, sharks, dolphins, seals, Steller sea 

lions, and perhaps river otters (Merkel 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 negatively influenced by El Niño conditions, possibly 

reducing available prey supply (Moser et al. 2000a, Harvey 2005). 

Anthropogenic threats. Overharvesting the primary cause of canary rockfish declines, but 

habitat loss is also important. Canary rockfish are considered overfished by the Pacific Fisheries 

Management Council and are not presently harvested intentionally. However, bycatch is still 
209

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 canary rockfish decline, with rocky habitats 

(reportedly, there are only 217 km2 in Puget Sound) being threatened by construction of bridges, 

sewer lines, cable and pipeline deployment, and dredge spoil (Palsson et al. 2008). Low oxygen 

levels as well as pollutant, chemical, and nutrient loading are also considered significant threats 

to canary rockfish recovery (NMFS 2008d). 

Critical habitat. Critical habitat has not been designated 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 timing 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 freshwater 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 drain (Hay and McCarter 2000, Beacham et al. 2005). Specific 

spawning rivers within the natal 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, California, 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 distribution has receded northward over the past several decades. The 

lower Columbia River mainstem provides spawning and incubation sites, and a large migratory 

corridor to spawning areas in the 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, Skamokawa, Elochoman, Kalama, Lewis and Sandy 

rivers. 

Growth and reproduction. Although primarily marine, eulachon return to freshwater to spawn. 

Adult eulachon have been observed in several rivers 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 Creek, 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, Elochoman, Kalama, Lewis, and Sandy rivers), with 

secondary levels in the Mad and Klamath rivers, as well as sporadic production in other 

Oregonian and Washingtonian rivers (Emmett et al. 1991b, Musick et al. 2000a, WDFW and 
210

ODFW 2001). Southern DPS fish likely take less time to mature than do fish from more 

northerly rivers and generally spawn 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 occurs 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 entry 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 northern California rivers. 

Additionally, smaller-scale resolution of British Columbia spawning supports earlier runs 

(February to early March) in northern territorial rivers (i.e., Nass River) and later spawning 

(April and May) in more southerly 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 hour can limit upstream migration (Lewis et al. 2002). 

Spawning normally occurs when water temperature is between 39o and 50o F, although Stikine 

and Nass River spawning is known to occur at cooler temperatures (Smith and Saalfeld 1955, 

Langer et al. 1977, Franzel and Nelson 1981, WDFW and 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). Adults 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 earlier than females and tend to stay longer, 

making them more susceptible to commercial and recreational fisheries (Hart and McHugh 

1944). However, males outnumber females by a roughly 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. Females lay eggs over 

sand, course gravel, or detritial substrate. However, like abalone, this method requires high 

eulachon density to ensure fertilization. Density may also impact the ability of eulachon to 
211

undergo synchronization of spawning, mate choice, and gonadal sterol levels. Eggs attach to 

gravel or sand and incubate for 30 to 40 days after which larvae drift to estuaries and coastal 

marine waters (Wydoski and Whitney 1979a). Larvae develop into juveniles and, after being in 

estuaries and marine waters for 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 offshore 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 freshwater, larvae and juveniles 

become thoroughly mixed in coastal waters generally less than 50 feet deep, but can be found as 

far down as 600 feet deep (Barraclough 1964, Hay and McCarter 2000). 

Foraging. Larval and post larval eulachon prey upon phytoplankton, copepods, copepod eggs, 

mysids, barnacle larvae, worm larvae, and other eulachon larvae until they 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 proposed 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 variety of factors, including predation, commercial and 

recreational fishing pressure (directed and bycatch), and loss of habitat. Population decline is 

anticipated to continue as a result 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 likely 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, abundant 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 rivers as late as the 1960s and 1970s (i.e., Klamath River, Mad 

River, and Redwood Creek) no longer occur (Larson and Belchik 2000). This decline likely 

began in the 1970s and continued until, in 1988 and 1989, the last reported 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 have again declined to the very low levels observed in the mid-
212

1990s (JCRMS 2009) , and since 2005, the fishery has operated at the most conservative level 

allowed in the management plan (JCRMS 2009). 

As mentioned, the Columbia River and its tributaries form the backbone of southern DPS 

spawning grounds. Historically, this 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 recreational catches have been implemented throughout southern 

DPS freshwater range, commercial catch records 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 metric tons of commercial catch, or roughly 300,000 

individuals)(Schweigert et al. 2007, DFO 2008). This is a decline 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, arrowtooth 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, Jeffries 1984, Drake and Wilson 1991, Yang and 

Nelson 1999, Willson et al. 2006). The high fat content of eulachon make them a particularly 

valuable prey for white sturgeon in the Columbia and Fraser rivers during winter (Willson et al.

2006). 

Anthropogenic threats. Eulachon face numerous anthropogenic effects 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 and ODFW 2001). Efforts to quantify fish that will be returning 

to the Columbia River to spawn in subsequent years have suggested significant 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 size 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, composing 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 survival and recovery is changing ocean conditions 

associated with global warming. Increases in ocean temperatures have already occurred and will 

likely to continue to impact eulachon and their habitats. For example, changes in glacial melt 

throughout the Alaskan, British Columbian, Washingtonian, and Oregonian coasts will likely 

change flow patterns in major rivers, such as the Fraser, that are important for eulachon 
213

spawning (Morrison et al. 2002). Alterations in the hydrology 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 temperatures will likely alter these communities, 

making it more difficult for eulachon and their larvae to locate or capture prey (Roemmich and 

McGowan 1995, Zamon and Welch 2005a). Warmer waters 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 eulachon 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 eulachon are particularly 

sensitive to (NMFS 2008c). Sediment retention structures 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 the benthic substrates they rely 

upon. Although poorly known, eulachon ecotoxicological 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 essential to the conservation of the 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 temperature 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 quality 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. Individuals are currently present in 36 rivers, and spawning occurs in at 

least 20 of these. Modern genetic analyses suggest that Atlantic sturgeon exhibit high fidelity to 

their natal rivers (Harwood 2010). Because of high natal river fidelity, 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 Hamilton 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 pattern of Atlantic sturgeon 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 three to five years for females (Bain 1997, 

Stevenson and Secor 1999, Gales et al. 2010, Schueller and Peterson 2010). 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 years, approximately 3-10 times longer than for 

other bony fish species examined (Boreman 1997). 

Sturgeon eggs are highly adhesive and are deposited on the bottom substrate, usually on hard 

surfaces (e.g., cobble)(Gilbert 1989b, Smith and Clugston 1997). Hatching occurs 

approximately 94-140 hrs after egg deposition, and larvae assume a bottom-dwelling existence 

(Anonmyous 2010a). The yolksac larval stage is completed in about 8-12 days, during which 

time larvae move downstream to rearing grounds over a 6 – 12 day period (Kynard and Horgan 

2002). During the daytime, larvae use benthic structure (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 years or more, but aging studies are limited by 

inaccuracy once individuals are older than 15 years 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 support 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 specific spawning locations have been identified in 

the United States, through genetic analysis, many rivers are known to support reproducing 

populations. Early life stage Atlantic 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-May 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 sturgeon undertake long marine migrations and utilize East Coast 
215

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 gravel 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 polychaetes, 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 individuals not foraging in spring or in freshwater, 

evidence also exists to support half to almost all individuals foraging in these circumstances 

(Brosse et al. 2002, Collins et al. 2008). 

Several authors have found that polychaetes constitute a major portion of Atlantic sturgeon diets. 

Brosse et al. (2002) reported that over 90% of Atlantic sturgeon diet was polychaetes during 

spring, summer, and winter. Savoy (2007) found 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 polychaetes. At the mouth of the 

Hudson River, Haley (1999) found that sturgeon fed on 47% polychaetes, 27% amphipods, and 

22% isopods. In North Carolina, Moser and Ross (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 the proportion of the sampled Atlantic sturgeon 

with each species in their guts and most guts contained 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 threatened (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 capacity. In the mid-1800s, 

incidental catches of Atlantic sturgeon in the shad and river herring haul seine fisheries indicated 

that the species was very abundant (Armstrong and 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 the Atlantic Coast (Smith 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 supported the largest population along the Atlantic 

Coast (Matthiopoulos and Aarts 2010). Ten years after peak landings, the fishery collapsed in 
216

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 Atlantic States Marine Fisheries 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 Atlantic sturgeon were present (ASMFC 1998). 

Currently, the only populations that have been studied well enough to provide an estimate of size 

are from the Hudson and Altamaha Rivers. These 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 natal rivers (Dovel and Berggren 1983, Bain et al. 1999). 

Peterson et al. (2000) reported that there were approximately 4,300 age-1 and -2 Atlantic 

sturgeon in the Hudson River between 1985 and 1995. Schueller and Peterson (2010) reported 

that age-1 and -2 Atlantic sturgeon population 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 the healthiest populations within the US. Thus, 

other spawning populations within the US are predicted to have fewer than 300 adults spawning 

per year. However, evaluating the status of the 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 catastrophic event from causing their extinction. 

Natural threats. Naturally, these are small populations and in some rivers because of variable 

spawning returns, Allee affects could be an issue. During all stages of development, Atlantic 

sturgeon are sensitive to temperatures above 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 poaching all contribute to the currently depressed 

populations of Atlantic sturgeon despite having very few natural predators. 

The Atlantic Sturgeon Status Review Team (2007) determined Atlantic sturgeon in the Delaware 

River are at a moderately high risk of extinction because of ship strikes and sturgeon in the 

James River are at a moderate risk from ship strikes. 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 from 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 the bottom of ships and the benthos 

of the river. Because of the size of the propellers used on large ships, everything along the 
217

bottom is sucked through the propellers. Large sturgeon are most often killed by ship strikes 

because smaller fish often pass through the propellers without making contact but larger sturgeon 

get hit. As shipping increases in the future, as has been predicted 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, the 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 (ASSRT 2007). Dredging primarily affects 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 several fisheries both within river systems and along 

the coast. In the James River, bycatch in the striped 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 fisheries interactions occur in estuaries and along the 

coast, where sturgeon from all rivers 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 system 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 overfished 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 harvest period, landings remained low through the 20th century until 1996 when 

the fishery was closed due to concerns about the recovery of their populations. 

Atlantic sturgeon have also been impacted by industrialization, poor water 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 quality as a moderate risk to 

every DPS in the United States (ASSRT 2007). Atlantic 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 radiating from a central trunk that 

is firmly attached to the sea floor. The largest species of its genus, colonies can reach at least 6.6 

feet high and 13 feet in diameter (Veron 2000). Corallites (branches 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 Caribbean, including in the Florida Keys, 

Abaco Island (The Bahamas), Alacran Reef, Mexico, Belize, Colombia, Costa Rica, Guatemala, 

Honduras, Nicaragua, Panama, Venezuela, Bonaire, Cayman Islands, Jamaica, 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 distribution is reduced, largely due to water temperature and quality issues. 

Growth and reproduction. Elkhorn corals employ both sexual and asexual reproduction. 

Sexual reproduction is accomplished by releasing sperm and egg during spawning events. 

Colonies are referred to as simultaneous 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 during July, August, and/or September. In some 

populations, spawning is synchronous after a full moon. Annual egg 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, preferably coralline algae (Bak 1977, Sammarco 1980, Rylaarsdam 

1983). Initial calcification ensues and develop into daughter corallites. 

Studies indicate that larger colonies (as measured by surface area of the live colony) have higher 

fertility and fecundity rates; over 80% of the colonies larger than 620 inches2

 were fertile. 

Estimated colony size at sexual maturity was 248 inches2

 and the smallest reproductive colony 

observed was 6.3 inches2 by 3.15 inches2 (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 disturbance, physical oceanography, the structure of 

established coral assemblages, and chemical cues (Lewis 1974, Birkeland 1977, Goreau et al.

1981, Rogers et al. 1984, Baggett and Bright 1985, Harriott 1985, Hughes and Jackson 1985, 

Sammarco 1985, Morse et al. 1988, Fisk and Harriott 1990, Richmond 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 elkhorn coral to construct significant 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 that coral polyp 

growth maximizes exposure to available light (Kaniewska et al. 2009). Growth can also occur 

from fragmentation and dispersal (Tunnicliffe 1981, Bak and Criens 1982). A broken branch 

may be carried by waves and currents to another location and, if favorable, branches grow into a 

new colony. Rapid growth and fragment dispersal facilitate a competitive advantage for elkhorn 

coral relative to other coral and benthic species (Shinn 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 increases in water turbidity than are some 
219

other corals. Different strains of symbiotic zoozanthelle (Symbiodinium spp.) can confer 

different thermal and light tolerances 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 often 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). Colonies generally do not form thickets below 

16.4 feet, with maximum water depths of framework construction ranging from 10 to 39.4 feet 

(Lighty et al. 1982). Elkhorn coral thrive in shallow 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 intermediate wave energy conditions (Geister 1977). The 

preferred habitat of elkhorn coral is the seaward face of a reef (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 declines in the early 1980s throughout its range 

and this decline has continued. Although quantitative data on historical distribution and 

abundance are scarce, best available data indicate 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 previous event in 1997. In all, 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, episodic, and unpredictable in its occurrence 

and results in high mortality. This is primarily due 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 the individual coral (Ainsworth et al. 2008). Optimal water temperatures 

range from 77o to 84oF, with mortality observed at 61o and 96oF (Jaap 1979, Roberts et al. 1982). 

High light levels can also induce 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 mortality risk under high light and temperature 

regimes, possibly by reducing exposure 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 possibly hinder immune response 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. In 2005, 

wide-scale bleaching occurred throughout the Caribbean with wide-scale mortality, with some 

areas reaching 95% of coral colonies affected (Wilkinson and Souter 2008). The US Virgin 
220

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 experienced disease rates of 50% of coral colonies or greater. Bleaching was 

associated with unusually warm waters in the region. 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 hurricanes also caused extensive 

damage to coral reefs; the prevalence of hurricanes 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 recruitment in years following storm passage as well as physical 

damage to coral colonies themselves (Mallela and Crabbe 2009). 

Anthropogenic threats. Threats to elkhorn coral also include eutrophication, sedimentation, 

anchoring, which degrade coral condition and increase synergistic stress effects (e.g. bleaching). 

Critical habitat. NMFS published a final rule to designate 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 contour is the shoreward limit), to support successful 

larval settlement, recruitment, and reattachment of fragments. Four specific areas are proposed 

for designation: the Florida unit, which comprises approximately 1,329 square miles of marine 

habitat; the Puerto Rico unit, which comprises 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, 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 impacts. The lone PCE identified thus far is natural 

consolidated hard substrate or dead coral skeleton that are free from fleshy or turf macroalgae 

cover and sediment cover. This feature is essential to the conservation of these two species 

because of the extremely limited recruitment currently 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 colonies, with cylindrical, straight 

or slightly curved diverging branches. Branches are 0.1 inch to 0.6 inch in diameter and rarely 

may grow back together. Colonies in turbulent water are smaller than in calm water, with greater 

branch density. Branching is irregular and secondary branches form at 60 to 90 degree angles 

relative to a primary branch. Prominent axial corallites (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 monospecific thickets. Tissue color 

ranges from golden yellow to medium 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 Caribbean, including in the Florida Keys, 

Abaco Island (The Bahamas), Alacran Reef, Mexico, Belize, Colombia, Costa Rica, Guatemala, 

Honduras, Nicaragua, Panama, Venezuela, Bonaire, Cayman Islands, Jamaica, 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 distribution is reduced, largely due to water temperature and quality issues. 

Growth and reproduction. Staghorn corals employ both sexual and asexual reproduction. 

Sexual reproduction is accomplished by releasing sperm and egg during spawning events 

(Szmant 1986). Colonies are referred to as simultaneous hermaphrodites, meaning that a given 

colony contains both female and male reproductive sex organs. Spawning events are relatively 

short, with gametes released only a few nights during July, August, and/or September. In some 

populations, spawning is synchronous after a full moon. Annual egg 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, preferably coralline algae (Bak 1977, Sammarco 1980, Rylaarsdam 

1983). Initial calcification ensues and develop into daughter corallites. 

Studies indicate that larger colonies (as measured by surface area of the live colony) have higher 

fertility and fecundity rates; colonies with a branch 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 inches 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 disturbance, physical oceanography, the structure of 

established coral assemblages, and chemical cues (Lewis 1974, Birkeland 1977, Goreau et al.

1981, Rogers et al. 1984, Baggett and Bright 1985, Harriott 1985, Hughes and Jackson 1985, 

Sammarco 1985, Morse et al. 1988, Fisk and Harriott 1990, Richmond and Hunter 1990). 

Growth rates are relatively rapid, expressed as the linear extension of branches, ranging from 1.2 

to 4.3 inches annually, and have enabled staghorn coral to construct significant 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 that 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, Bak and Criens 1982). A broken branch may be 

carried by waves and currents to another location and, if favorable, branches grow into a new 

colony. Rapid growth and fragment dispersal facilitate a competitive advantage for staghorn 

coral relative to other coral and benthic species (Shinn 1976, Neigel and Avise 1983, Jaap et al.

1989). Larval recruitment is influenced by the type and availability of benthic substrate, with 

certain types of coral or rock substrates resulting in greater or lesser recruitment success (RitsonWilliams et al. 2009). 
222

Habitat. Historically, staghorn coral so dominated reef 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 Yucatan), 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 octocoral-dominated hard-bottom (Goldberg 1973, 

Davis 1982, Jaap 1984, Wheaton and Jaap 1988). Colonies 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 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 increases in water turbidity than are some 

other corals. Different strains of symbiotic zoozanthelle (Symbiodinium spp.) can confer 

different thermal and light tolerances 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 underwent precipitous declines in the early 1980s throughout its 

range and this decline has continued. Although quantitative data on historical distribution and 

abundance are scarce, best available data indicate declines in abundance (coverage and colony 

numbers) by greater than 97% (Figure 23). 

Staghorn corals still occupy their historic range, but localized range reductions and extirpations 

have occurred with most populations experiencing 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 contain staghorn corals, representing 2-3% of the coral cover in these 

areas (Rogers et al. 2002). Surveys of fragments of staghorn from nearshore areas of St. Thomas 

and outlaying cays indicate that colonies of these 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 m2

 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 depth). Based on analysis of point count data, 

staghorn coral had a mean total cover of 31.1% (range 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 outbreak of white pox disease at Los Corchos Reef 
223

in Culebra, Puerto Rico. Prior to the outbreak, coral cover on the 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 partial 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 that caused mass mortality and resulted in a 

net 20-60% decline in living coral cover at surveyed reefs of the 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 mortality associated with the 2005 bleaching event 

and current monitoring data does not indicate significant 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 study 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, episodic, and unpredictable in its occurrence 

and results in high mortality. This is primarily due to a disassociation of zoozanthelle from coral 

tissue. Just prior to this, coral epithelium and gastrodermis tissue begins to decay and die, likely 
224

as a result of stress to the individual coral (Ainsworth et al. 2008). Optimal water temperatures 

range from 77o to 84oF, with mortality observed at 61o and 96oF (Jaap 1979, Roberts et al. 1982). 

High light levels can also induce 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 mortality risk under high light and temperature 

regimes, possibly by reducing exposure 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 possibly hinder immune response 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 Oscillation cycle in 1997-1998 resulted in a large bleaching event in 

the Caribbean and the Atlantic, as well as massive losses of corals in the Indian Ocean and 

Western Pacific (Wilkinson and Souter 2008). However, 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 affected 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 experienced disease rates of 50% 

of coral colonies or greater. Bleaching was associated with unusually warm waters in the region. 

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 hurricanes also caused extensive damage to coral reefs; the prevalence of hurricanes 

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 

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 disease to date that has 

been documented to cause major changes in the composition and structure 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 disease in elkhorn corals. 

Reductions in long-term water clarity can also reduce the coral photosynthesis to respiration ratio 

(P/R ratio). Telescnicki and Goldberg (1995) and Yentsch et al. (2002) found that elevated turbidity 
225

levels did not affect gross photosynthetic oxygen production, but did lead to increased respiration 

that consumed the products of photosynthesis with little remaining for coral growth. 

Unfortunately, since staghorn corals are broadcast spawners once colonies become rare, the 

distance between colonies may limit fertilization success and there is substantial evidence to 

suggest that sexual recruitment of staghorn corals is currently compromised. Reduced colony 

density in some areas is compounded by low genotypic diversity, indicating that fertilization 

success and consequently, larval availability, is likely reduced. This can have long-term 

implications for genetic variability of remaining colonies due to the reduced potential for 

exchange of genetic material between populations that are spatially further apart (Bruckner 

2002). 

Data on levels of genetic diversity and population structure suggest that there is a population 

structure among islands, and even over spatial scales of no more than 20 km, as well as varying 

degrees of genetic diversity within local populations (Lirman 2002, Vollmer 2002). For 

instance, one clone of staghorn coral may dominate areas up to 10 m2

 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 exacerbated further by eutrophication, 

sedimentation, and anchoring, which degrade coral condition and increase synergistic stress 

effects (e.g. bleaching). Excessive sedimentation can smother corals and increased nutrient 

availability promotes algal growth on corals, leading to light blockage to zoozanthelle and death 

of corals (Acropora Biological Review Team 2005). Although 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 reducing 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 designate 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 contour is the shoreward limit), to support successful 

larval settlement, recruitment, and reattachment of fragments. Four specific areas are proposed 

for designation: the Florida unit, which comprises approximately 1,329 square miles of marine 

habitat; the Puerto Rico unit, which comprises 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, 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 impacts. The lone PCE identified thus far is natural 

consolidated hard substrate or dead coral skeleton that are free from fleshy or turf macroalgae 

cover and sediment cover. This feature is essential to the conservation of these two species 

because of the extremely limited recruitment currently 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 along the mainland coast in their northern and 

southern range, but are more frequently at the 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 (Guzmán 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 relatively 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 before, as adults, moving to more open habitats on 

the tops and sides of rocks where food is more plentiful (Haaker et al. 1986, Hobday and Tegner 

2000). Adult white abalone are found most abundantly at depths of 82 to 99 feet, but may occur 

in waters from 66 to 197 feet (Hobday and Tegner 2000). 

Feeding. White abalone eat algae, feeding as postlarvae 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 primary source of nutrition (Tutschulte 1976). 

Growth and reproduction. Abalone aggregate for spawning, but low numbers and physical 

barriers can prevent large spawning aggregations from forming (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 average 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 fertilization, a free-swimming larva emerges 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 juveniles 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 individual, 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 characterized in green lip abalone by Cennie et al. (2009a), 

where individuals belonged to either a sedentary or a wandering clade (which moved over larger 

areas and covered greater distances, occupied inner portions of hides more frequently, responded 

more rapidly to food odor, foraged for longer periods, and displaced conspecifics from food 

patches). 

Life span. Researchers believe that abalone are long-lived, but the average life span is unclear 
227

(Hobday and Tegner 2000). Models of growth predict 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 was 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-exploitation abundance at 2,221,800 abalone, but the 

population in 1996 to 1997 was estimated at 1,613 individuals, representing a 99.9% decline, and 

was estimated to disappear by 2010. White abalone recovery is hindered by low spawning 

densities, resulting in recruitment failure. White abalone along Mexico are believed to be 

depleted, but their status is generally 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 only based upon adults, as 

juveniles are infrequently observed. White abalone observed during surveys were of large size 

which corresponds to predicted ages near the end 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 the 1960s (Hobday and Tegner 2000). 

Natural threats. Natural pressures exist from sea otter predation (Johnson et al. 2009) and the 

low density of individuals during spawning events. Interspecific competition has not been 

studied in white abalone, but blacklip abalone face 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 spawning densities in many areas that are blamed for 

the inability of the species to recover. Although small-scale aquaculture takes have occurred to 

attempt captive breeding and recovery, these takes 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 bioaccumulate 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 environmental concentrations (Wang et al. 2009). 

Changes in sea surface temperatures 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 and from Cedros to Punta Asuncion along Baja, 

Mexico (Guzmán Del Proó 1992). Present occurrence 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 sperm density (Davis 1996). Fecundity increases 

exponentially with size, with small mature females producing a few hundred thousand eggs each 

year, but older individuals producing 10 to 15 million eggs (Hahn 1989). Fertilized eggs sink 

and hatch into free-swimming larvae within 72 hours. After one or two weeks, larvae settle 

(CDFG 1993). Larval mortality is assumed to be high (Leighton 1972). At 1.5 inches (roughly 3 

years of age), black abalone are considered 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 eating 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 rocky intertidal and shallow subtidal zones (Haaker et al. 1986). 

Depth distribution is usually from the shore to about 10 feet of water depth. 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 habitats 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 wandering clade (which moved 

over larger areas and covered greater distances, occupied inner 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 abalone are believed to survive for 30 years 

or more (Blecha et al. 1992). 

Status and trends. On January 11, 2008, NMFS published a proposed rule to list black abalone 

as endangered under the ESA (73 FR 1986). Species decline is reflected by the decrease in 

commercial catches until 1993, when commercial harvests were halted. Historic levels 

approached 2,200 tons in California in 1879 and declined to around 1000 tons in the 1970's. 

Commercial landings then decreased to 19.1 tons in the last year of harvests, when mortality 

from withering syndrome devastated remaining black abalone stocks throughout southern 
229

California (Haaker 1994). Over 20 years, densities of more than 100 individuals per cubic yard 

disappeared from most of their former range south of Point Conception (Davis 1993). A similar 

mass mortality was reported at Palos Verdes Peninsula 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 declines due to a temperature-related disease called 

withering syndrome. This bacteria-based disease prevents assimilation of nutrients in the 

digestive system and results in abalone that “wither” as individuals consume body tissues. The 

disease was first identified west of Santa Cruz and Anacapa islands in 1985 and 1986 before 

spreading to Santa Rosa Island and Santa Barbara 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-off 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 throughout 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 syndrome, 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 traditionally temperate regions. In red abalone, 

higher rates of infection and more prominent signs of infection 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 for 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 black 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 and 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 these 
230

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 during recovery. Other factors that can threaten black abalone 

include storms (crushing abalone between rocks and sedimentation in rocky habitat), fresh water 

input, sedimentation on gills leading to asphyxiation, and temperature impacts on reproduction 

and growth (Cox 1960, 1962). 

Anthropogenic threats. Although commercial harvests historically lead 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 bioaccumulate 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 environmental concentrations (Wang et al. 2009). 

Changes in sea surface temperatures 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 water to six meters water depth in the Farallon, 

Channel, and Año Nuevo islands, as well as the California coastline from Del Mar Ecological 

Reserve south to Government Point (excluding some stretches, such as in Monterey Bay and 

between Cayucos and Montaña de Oros State Park) 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 (bacterial and diatom films, crustose coralline algae, and a source of detrital 

macroalgae), juvenile settlement habitat (rocky intertidal habitat containing crustose coralline 

algae and crevices or cryptic biogenic structures (e.g., urchins, mussels, chiton holes, 

conspecifics, anemones)), suitable water quality (temperature, salinity, pH, and other chemical 

characteristics necessary for normal settlement, growth, behavior, and viability of black 

abalone), and suitable nearshore circulation 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 therefore no historical distribution 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 geographic distribution of any seagrass in the world 

(Kenworthy 1997). However, northern range extentions (likely temporary) have recently been 

observed (Virnstein and Hall 2009). The largest 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 discharges have been observed (Gallegos and 

Kenworthy 1996), although Torquemada et al (2005) noted that highly hypo- or hypersaline 

conditions can negatively impact growth. Intertidal patches may be completely exposed at low 

tides, suggesting tolerance to desiccation and wide temperature ranges (Kahn and Durako 2009). 

Growth and reproduction. Only female flowers have been observed; 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 present, 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 branching and fragmentation (Jewitt-Smith et al.

1997, Hammerstom and Kenworthy 2003). Consequently, genetic diversity is low (Freshwater 

and York 1999), putting Johnson’s seagrass at a potential 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 leaf-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 expand 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 feet2

 in size 

(Kenworthy 1997, Virnstein et al. 1997, Kenworthy 2000, 2003, Virnstein and Morris 2007). 

Patches can expand rapidly (nine feet2

 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 and 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 “pulsating patch,” with two to three consecutive 

summers of growth followed by a rapid decline. Johnson’s seagrass frequently undergo whole 

patch mortality followed by recolonization (Virnstein et al. 1997, Heidelbaugh et al. 2000, 

Greening and Holland 2003, Kenworthy 2003, Virnstein and Morris 2007). 

Johnson’s seagrass appears to be physiologically adapted to exploit unstable environments and 

unvegetated patches, with minimal resources allocated 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 otherwise compete (Dean and Durako 2007). These 
232

growth characteristics also help explain its patchy distribution (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 disturbance (Durako 2003). Johnson’s 

seagrass thrive in unstable or newly-created unvegetated 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 shallow root system, growth over large unsuitable 

patches may be unlikely, and its ability to recover 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 seagrass 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 artifact 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 individuals and scatter them widely. Although this can serve to 

disperse individuals into new habitats, it can also catastrophically eliminate established 

meadows. Subsequent siltation following 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. Johnson’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 runoff and turbidity), and (6) siltation, as well as 

climate change (Waycott et al. 2009). 

Critical habitat. Critical habitat for Johnson’s seagrass was designated on April 5, 2000 

(65 FR 17786) and includes (1) locations with populations that have persisted for 10 years; (2) 

locations with persistent flowering populations; (3) locations at the 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 compared 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 populations, allow for protective buffers along the 

distribution limits (i.e., edges of survival), and protect regions of high density that without 

further knowledge of species biology, appear to serve the needs 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, north of St. Lucie Inlet, a portion 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 Biscayne Bay. These regions occupy approximately 

22,574 acres or 9,139 hectares. Simply the nature of Johnson’s seagrass critical habitat makes it 
233

variable and prone to change. The growth of boating in Florida and development of coastal areas 

has resulted in trampling, propeller scarring, dredging, filling, shading, and altered water quality 

that has degraded these areas compared to historical conditions. Although many of the factors 

that can negatively affect Johnson’s seagrass and their habitat are generally well regulated and 

enforced, the species is still under 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 actions and other human activities in the action area, the anticipated 

impacts of all proposed federal projects in the action area that have already undergone formal or 

early section 7 consultation, and the impact of state or private actions that are contemporaneous 

with the consultation in process (50 CFR §402.02). The Environmental baseline for this Opinion 

includes the effects of several activities affecting the survival and recovery of proposed or listed 

species as well as their proposed or designated 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 consequences 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, wherever possible we have focused on 

common indicators of the biological, chemical, and physical health of the nation’s aquatic 

environments. The Environmental baseline for this consultation provides the backdrop for 

evaluating the effects of the action on listed and 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 section we described the biological and ecological characteristics of 

the region such as the climate, geology, and predominant vegetation to provide landscape context 

and highlight some of the dominant processes that influence the biological and ecological 

diversity of the region where proposed, threatened, and endangered species reside. We then 

described the predominant land and water uses within 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 environment tend to be much more ubiquitous than in freshwater 

ecosystems and thus we have not generally divided stressors in the marine environment into 

more specific components, although some areas are relatively unique in regards to some 

stressors, such as oil and gas industrial activities 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 narratives. 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 substantial direct and indirect effects on individuals, 

populations, species, and the structure and function of marine, coastal, and 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 accelerating 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-listed species occur) is warming 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 weakening of the thermohaline circulation 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 magnitude of these changes 

remain unknown. Species that are shorter-lived, of larger body size, or generalist in nature are 

liable to be better able to adapt to climate change over the long term versus 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 are already in tenuous positions (Isaac 2008). 

As such, we expect the risk of extinction to listed species to rise with the degree of climate shift 

associated with global warming. 

Some indirect effects of climate change would result from changes in the distribution of 

temperatures suitable for whale calving and rearing, 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 availability or habitat suitability. If either is 

disrupted, the timing of migration 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 seals during El Niño periods, 

when cooler, more productive waters are associated with higher first year pup survival 

(McMahon and Burton. 2005). Reduced prey availability resulting from increased sea surface 

temperatures has also been suggested to explain 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 wetter years than in drier ones (Twiss et al. 2007). For marine 

mammals considered in this Opinion, available data suggest sperm whale females have lower 

rates of conception following periods of unusually warm sea surface temperature (Whitehead 

1997). Marine mammals with restricted distributions linked to water temperature may be 

particularly exposed to range restriction (Learmonth et al. 2006, Issac 2009). MacLeod (2009) 

estimated that, based upon expected shifts in water temperature, 88% of cetaceans would be 

affected by climate change, 47% would be negatively affected, and 21% would be put at risk of 

extinction. Of greatest concern are cetaceans with ranges limited to non-tropical waters and 

preferences for shelf habitats (Macleod 2009). Variations in the recruitment of krill and the 
235

reproductive success of krill predators correlate to variations in sea-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 Antarctic 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 biomass 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 performance of predators 

(Antarctic fur seals, gentoo penguins, macaroni penguins, and black-browed albatrosses) that 

depend on krill for prey and concluded that these 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 population, particularly reduced recruitment into older krill age classes, 

which lowered the number of predators krill could sustain. The authors concluded that the 

biomass of krill within the largest size class was sufficient to support predator demand in the 

1980s but not in the 1990s. By 2055, severe reductions in fisheries catch due to climate change 

have been suggested to occur in the Indo-Pacific, Red Sea, Mediterranean Sea, Antarctic, and 

tropical areas worldwide while increased catches are 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 reason for a poleward shift in the Eastern Australian Current, shifting 

warm waters into the Tasman Sea and altering biotic features of the area (Poloczanska et al.

2009). Similarly, the Kuroshio Current in the western North Pacific (an important foraging area 

for juvenile sea turtles) has shifted southward as a result of altered long-term wind patterns over 

the Pacific Ocean (Poloczanska et al. 2009). 

Climate-mediated changes in the distribution and abundance of keystone prey species like krill 

and climate-mediated changes in the distribution of cephalopod populations worldwide is likely 

to affect marine mammal populations as they re-distribute throughout the world’s oceans in 

search of prey. If sea ice extent decreases, then larval krill may not be able to survive without 

access to underice algae to feed on. This may be a cause of decreased krill abundance in the 

northern western Antarctic Peninsula during the last decade (Fraser and Hofmann 2003). 

Meltwaters have also reduced surface water salinities, shifting primary production along the 

Antarctic Peninsula (Moline et al. 2004). Blue whales, as predators that specialize in eating krill, 

are likely to change their distribution 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 change their distribution or 

could not find the biomass of krill necessary to sustain their population numbers, their 

populations (and other large whales with similar life histories, such as humpback whales) would 

likely experience declines similar to those observed in other krill predators, including dramatic 

declines in population size and increased year-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 North 
236

Sea and an overall reduction in plankton biomass as warm-water species invade formerly coldwater areas. Productivity may increase in other areas, though, providing more 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 distribution 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 habitat suitability. If either is disrupted by changing 

ocean temperature regimes, the timing of migration can change or negatively impact population 

sustainability (Simmonds and Eliott. 2009). For sea turtles, warming ocean temperatures may 

extend poleward the habitat which they can utilize (Poloczanska et al. 2009). Seagrass habitats 

have declined by 29% in the last 130 years and 19% of coral reefs have been lost due to human 

degradation, reducing lower latitude habitat for some sea turtle species (Poloczanska et al. 2009). 

Primary production is estimated to have declined 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 potential aspect that climate change could influence. AcevedoWhitehouse and Duffus (2009) proposed that the rapidity 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 turtle 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 reproductive success, although average nesting and 

emergence dates have changed over the past several decades by days to weeks in some locations 

(Poloczanska et al. 2009). However, such a fundamental shift in population demographics 

causes a fundamental instability in population viability. 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 patterns will likely have profound effects on the coastlines of every 

continent by increasing sea levels and the intensity, 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 risen an average of 1.7 mm/year over the 20th 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 supported 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 destroyed 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 nests (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 attributes 
237

(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 (producing smaller hatchlings), and reduces nesting 

success due to exceeded thermal tolerances (Fuentes et al. 2009b, Fuentes et al. 2009c, Fuentes

et al. 2010). Smaller individuals likely experience 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 for the current and likely future status of 

salmon, but particularly so in the Pacific northwest, where snow melt into the Columbia River 

Basin has significant influence on regional hydrology. Recent studies, particularly by the 

Independent Scientific Advisory Board (ISAB), describe 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 temperatures. 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, possibly 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 could further imbalance salmon habitat. 

2. Variations in rainfall intensity may alter seasonal hydrography. With reduced 

snowpack and greater rainfall, the timing of stream flow will likely change, reducing 

spring and summer stream flow and increasing 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 and hydrologic effects of climate change, 

abundance Snake River spring/summer Chinook populations would be substantially 

decreased (20-50% decline from simulated 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 temperature would 

accelerate the rate of egg development of fall Chinook that spawn in the mainstem of 
238

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, increased fallback, and loss of energy reserves 

due to elevated metabolic demand. Thermal stress 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 coastal upwelling could cause these fish to enter the ocean 

before foraging conditions are optimal. The first few weeks in the ocean are thought 

to be critical to the survival of salmon off Oregon and Washington, so a growing 

mismatch between smolt migrations and coastal upwelling would likely have 

significant negative impacts on early ocean survival rates. 

7. Within the Columbia estuary, increased sea levels in conjunction with higher winter 

river flows could degrade estuary habitats. Numerous warm-adapted fish species, 

including several non-indigenous species, normally found in freshwater have been 

reported from the estuary and might expand their populations with the warmer water. 

Climate change also may affect the trophic 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 associated with warming include increases in 

temperature, increased water column stratification, and changes in the intensity and 

timing of coastal upwelling. These changes will alter primary and secondary 

productivity, the structure of marine communities, and, in turn, the growth, 

productivity, survival, and migrations of salmonids. 

9. Changing ocean temperatures may alter 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 intensified 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 carbonate 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 maturity, delaying the usual times of adult 

migrations into coastal water and rivers. 

Climactic shifts also occur due to natural phenomenon. In the North Atlantic, this primarily 

concerns fluctuations in the North Atlantic Oscillation (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 (Hurrell 1995, Curry and McCartney 2001, Stenseth et 

al. 2002). This interaction affects sea surface temperatures, wind patterns, and oceanic 

circulation in the North Atlantic (Stenseth et al. 2002). The NAO shifts between positive and 
239

negative phases, with a positive phase having persisted since 1970 (Hurrell 1995). North 

Atlantic conditions experienced 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 North Atlantic right whales due to its 

influence on the species primary prey, zooplankton of the genus Calanus, which are more 

abundant in the Gulf of Maine 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 Pacific 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 influence the northeastern Pacific (JOI/USSSP 2003, 

Stabeno et al. 2004). Typical changes include increased winter air temperature, precipitation, 

sea level, and downwelling favorable conditions (Royer and Weingartner 1999, Whitney et al.

1999). La Niña events tend to swing these conditions 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 varying degrees (Freeland 1990, Bailey et al. 1995, Brodeur et al.

1996a, Royer 2005). Eddies occur with greater frequency 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 (Freeland 2000, Coyle and Pinchuk 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 during 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 humpback whales have also been suggested 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 plankton 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 capable 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 years, are more prominent outside the tropics, and 

mechanisms controlling them are relatively unknown (Minobe 1997, 1999, Hare and Mantua 

2000, Mantua and Hare 2002). During positive PDOs, the northeastern Pacific experiences 
240

above-average SSTs while the central and western 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, euphausiid biomass remained the same and copepod 

abundance actually increased in the Pacific northwest; 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). Positive 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 under the influence of climactic change is also a 

significant concern. If water temperatures warm in marine ecosystems, native species may shift 

poleward to cooler habitats, opening ecological niches that can be occupied by invasive species 

introduced via ships ballast water 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 temporary, 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 intensify with a warming climate, resulting in 

roughly 13 more days of potential bloom conditions per year by the end of the 21st century. 

Invasive species that are better adapted to warmer water temperatures can also outcompete native 

species that are physiologically geared towards lower water temperatures; such a situation 

currently occurs along central and northern California, where the Mediterranean blue mussel has 

established and is displacing 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 Hampshire, and Massachusetts. The region is 

ecologically diverse, encompassing several broad ecoregions—according to Bailey’s (1995) 

Description of the Ecoregions of the United States this region encompasses the warm 

continental, the hot continental and the hot continental mountains divisions —these ecoregions 

can be further subdivided into provinces based on vegetation (Bailey 1995). This region 

encompasses the New England/Acadian mixed forests and the Northeastern Coastal Forests. 

In this section, we describe several basins and estuarine complexes to characterize the general 

ecology and natural history of the area, and past and current human activities and their impacts 

on the area. In certain instances we described some river basins in further detail to provide 

additional context for evaluating the influence of the environmental baseline on listed species 

under NMFS’ jurisdiction and the health of the environment. 

Natural History. This region encompasses drainages entering the Gulf of Maine, and 

encompasses all of Maine, parts of New Hampshire, Massachusetts the Canadian provinces of 
241

New Brunswick and Nova Scotia. Characterized by a temperate climate and a rocky coastline, 

the greater Gulf of Maine encompasses the 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 soils, 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 nutrient rich than waters to the 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 marshes, kelp and seagrass beds, 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 Northern right whales—where they are regularly 

observed in the spring and summer at regular 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 (mi2

) (Jackson et al. 2005). The river flows from the 

mountains of western Maine, including Maine’s highest peak, Mt. Katahdin 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 river or estuary. Despite being home to so many 

fish, there are only three nonnative species (Baum 1983(Jackson et al. 2005). The Penobscot 

estuary extends from Bangor downstream to Penobscot Bay in the Gulf of Maine, approximately 

31 miles, making it the largest estuary in Maine and one of the largest on the East Coast (PEARL 

2007). Downstream of Bangor, the river is a tidally influenced, salt-wedge estuary. The 

majority of the estuary is bedrock-based, and sediment deposits are limited to isolated coves and 

near marshes.

Merrymeeting Bay Basin. Merrymeeting Bay is the largest, 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 typically has the largest freshwater outflow to 

the Gulf of Maine, usually exceeding 15,000 cubic feet 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 surface area, some toxic substances were identified. Sediments 

associated with the Androscoggin River had higher levels of PAHs and mercury, while 

sediments from the Kennebec River had higher levels of chromium, arsenic, and selenium 

(Hayden 1998). The bay has more moderate levels of these toxins than the rivers themselves. 

Chilcote and Waterfield (1995) found that levels of arsenic are higher 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 and Kennebec than the Sheepscot River. Based 

on benthic samples taken in 1980 and again in 1991, it appears that all metals are declining in 

Merrymeeting bay except for copper, which showed an increase (Hayden 1998). Commercially 

important fish also have elevated metal concentrations 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 mi2

 (Jackson et al. 2005, Maine Rivers 2007). The Kennebec 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, including 10 nonnative species. 

The Androscoggin River travels 164 miles, with a watershed of 3,263 mi2 (Jackson et al.

2005)MaineRivers 2007c). The river flows from northwest Maine, into New Hampshire, and 

then back into Maine, where it meets the Kennebec River in Merrymeeting Bay. The 

Androscoggin has been Maine’s principle industrial river (MaineRivers 2007c). The average 

precipitation in the watershed is 43.7 inches per year, resulting in an average discharge at the 

mouth of the Androscoggin, entering Merrymeeting 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 mi2 (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 average 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 fish, including 5 nonnative 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 the 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 of 
243

33,000 people (Jackson et al. 2005). While there is not much urban development in the 

watershed, Doggett and Sowles (1989) report tanneries, metal finishing, pulp and paper mills, 

textile plants, chemical products, and municipal sewage contribute chromium, mercury, zinc, 

copper, lead, arsenic, hydrocarbons, dioxins, PAHs, pesticides, and other contaminants to the 

river. 

The Kennebec River watershed usage is 82% forest, 10% water, 6% agriculture, 2% developed 

(Jackson et al. 2005). The only major town in the watershed is Augusta, Maine, but there are 

approximately 39 people per square mile 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) (Harding Lawson Associates 1999, 2000). Since 

1990, the levels of dioxins in other rivers in Maine have been decreasing, but the levels in the 

Kennebec have remained constant (Kahl 2001). 

The Androscoggin River watershed usage is 5% agriculture, 86% forested, 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 20th century, textile mills, paper and pulp 

mills, and municipalities contributed large quantities 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, sedimentation, total suspended 

solids, PCBs, and dioxins are still considered too high (Chamberland et al. 2002). 

The Merrimack River watershed is composed of 75% forest, 13% urban, 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 and 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, nutrients, 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 other licensed dams located along the 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. Unfortunately, 6,000 to 10,000 salmon are required for a 

sustainable population, so the Penobscot run depends on fish from a local hatchery (Moore and 

Platt 1997). 
244

The Kennebec River has eight large hydroelectric dams on its mainstem, which restricts fish 

passage both up and downstream. In 1999, the Edwards Dam was removed, opening 17 

additional miles of habitat for fish and macroinvertebrates in the river. Removal of Edwards 

dam restored full access to historical spawning habitat for species like Atlantic sturgeon, 

shortnose sturgeon, and rainbow smelt, but not for species like alewife, American shad, and 

Atlantic salmon that migrated much further up the river. Since the removal 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 breached 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 installed 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, causing 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 the mainstem into a series of ponds (Dunn Jr 2002, 

Jackson et al. 2005). Flow alteration is considered a problem 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 materials. 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 Northeast 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 pollutants leaked by mining operations in this 

area are lead, mercury, arsenic, and selenium (Ayuso et al. 2006, Piatak et al. 2006). All mines 

that are not in use are supposed to be decommissioned 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 fisheries along the Northeast 

Atlantic coast by harvest weight exist for herring (39%), lobster (26%), blue mussel (6%), 

hatchery-origin sea-run Atlantic 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 salmon 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 September and mid October for hatchery-origin 

Atlantic salmon has been permitted on the Penobscot River, the only river with listed Atlantic 

salmon that allows salmon fishing. On the Penobscot, spring salmon fishing has not taken place 
245

since 1999, but may be permitted again in 2008. Poaching 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 Virginia. The headwaters of the Connecticut River originate in 

New England/Acadian forests, and as the river descends, it transitions from boreal forest to 

temperate deciduous forest. As the river flows through the low gradient coastal region, the 

ecoregion transitions to Northeastern Coastal Forest. 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 Northeastern Coastal Forest. The headwaters of 

the Delaware River originate in the Allegheny Highland Forest ecoregion, and then as the river 

descends, it transitions to Appalachian/Blue Ridge Forest and then Northeastern Coastal Forest 

ecoregions. 

Connecticut River/Long Island Sound 

Natural History. The Long Island Sound watershed includes 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 surface area of Long Island Sound is 1,320 mi2

, draining an 

area of over 16,000 mi2. 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 least 50 species use the sound as spawning grounds. 

The Connecticut River drains a watershed of 11,259 mi2 and flows approximately 410 miles to 

Long Island Sound. The river flows from the highlands of New Hampshire and Quebec, and is 

bordered by the Green and White Mountains. The Connecticut River’s bed is composed of 

glacial deposits and granitic bedrock. 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 estuarine 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 non-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, nitrogen, lead, mercury, cadmium, cesium, zinc, 
246

copper, and arsenic. Organic and metal contaminants 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, and lead appear to be decreasing 

while copper is increasing (Turgeon and O'Connor 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 York have signed agreements to 

reduce the total nitrogen input to Long Island Sound by 58.5% before 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 dominant 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 watershed is approximately 179 people per square 

mile (Jackson et al. 2005). Throughout the 20th 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, approximately one billion gallons of 

raw sewage enters the river as a result of combined sewer overflow from Hartford, Connecticut 

alone (CRWC 2006). The river has become much cleaner since the CWA was passed, but 

chromium, copper, nickel, lead, mercury, and zinc, chlordane, DDT, DDE, PCBs, and PAHs are 

found in quantities above the EPA recommended levels in sediments and fish tissue throughout 

the watershed (Jackson et al. 2005). Acid rain also affects rivers in the northeast, as it reduces 

the pH of rivers and causes metals to leach from bedrock at a faster 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 fish passage continues to be compromised. For instance, overall 

passage efficiency at Turner Falls fish ladder is 17%, and has historically 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 temperatures throughout the summer. 

Mining. Dating back thousands of years, there 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 currently degrade the river’s water quality, including 

the country’s first chartered copper mine. Towns such as Plymouth, Vermont were famous for 

mining gold, iron, talc, soapstone, marble, asbestos, 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 threaten aquatic life. In other cases, the mines are 

abandoned or failing and need to be cleaned. Such is the case with Elizabeth Mine, an old 

copper mine perched above the Connecticut River that leaches heavy metals into the river. As a 
247

result, Elizabeth Mine has been declared a superfund site. There is little to no mining in Long 

Island Sound and the concept is generally frowned 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 commercial fishery here, although 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, striped bass, bluefish, northern pike, largemouth and 

smallmouth bass, perch, catfish, and other fish. 

Long Island Sound fisheries provide an estimated 5.5 million dollars to the Connecticut 

economy. The primary fisheries target oysters, lobsters, scallops, blue crabs, flounder, striped 

bass, and bluefish. Recently, due to DO deficiencies, the western portion of Long Island Sound 

has seen major declines in fish and shellfish populations. Despite these 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 suffering from an unknown disease 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 for larger predatory fish such as striped bass. 

Directed harvest of shortnose sturgeon is prohibited by the ESA. However, shortnose sturgeon 

are likely taken incidentally in fisheries in the Connecticut River and Long Island Sound. Moser 

and Ross (1993) found that captures of shortnose sturgeon in commercial 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 approximately 315 miles to the ocean, with a 

watershed of 13,365 mi2. The river flows from the Adirondack Mountains, draining most of 

eastern New York State, to the ocean where 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 River’s bed is composed of metamorphosed plutonic 

rock in the Adirondack Mountains, then transitions 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 igneous rocks. The average precipitation is 

approximately 36 inches per year. At the mouth, 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 Newburgh 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 over 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 and 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 population density in Manhattan is over 25,907 people per 

square mile (Jackson et al. 2005). 

Throughout the 20th century, power plants, municipalities, pulp and paper mills, and corporations 

such as IBM, General Motors, and General Electric 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, with 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 low gradient. High tide in this stretch causes the river to flow 

backwards due to the low gradient and this prevents stratification. The brackish tidal estuary 

portion of the Hudson is nutrient rich with hard water. Two hundred miles of the Hudson River, 

from Hudson Falls to New York City, were designated as a superfund site due to the amount of 

pollution. There are still elevated amounts of cadmium, copper, nickel, chromium, lead, mercury, 

and zinc, DDT, PCBs, and PAHs are found in quantities above the EPA recommended levels in 

sediments and fish tissue throughout the watershed (Wall et al. 1998). 

Hydromodification Projects. The mainstem Hudson River has 14 dams and there are dams 

near the mouths of many tributaries, but the lower 154 miles of tidally influenced river is 

undammed. Several flood control dams on tributaries 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 unobstructed for the lower 154 miles, resulting in the 

healthiest population of ESA-listed 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 behind 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 minerals during different 

generations. Initially bought as an iron mine, McIntyre sat dormant for 75 years before titanium 

was discovered there, at which point National Lead purchased it and mined there until 1982 

when NL Industries abandoned the mine. 

Commercial and Recreational Fishing. The Hudson River commercial fishery historically 

caught fish, blue crabs, and oysters. Now, the only fish that is caught commercially in the 

Hudson is American shad. Historically, Atlantic sturgeon, striped bass, American eel, and white 

perch were productive commercial fisheries. The striped bass fishery closed in 1976 due to 

PCBs in the river and fish tissue. Atlantic sturgeon were fished until the mid 1990s. Blue crabs 

are still fished in the estuary all the way to Troy, NY with recent catches over 88,185 pounds per 
249

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 approximately 329 miles to the ocean, with a 

watershed of 12,757 mi2

. The river originates in the Catskill 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 Delaware River’s geology is sandstone with shale conglomerate 

in the upper watershed transitioning to sandstone, shale, and limestone in the middle watershed 

and igneous and metamorphic rock in the lower watershed. The average precipitation is 

approximately 43 inches per year. At the mouth, the average discharge is 9.6 billion gallons 

each day, or 14,903 cubic feet per second, and although it is only the 42nd largest river by 

discharge, Philadelphia is home to the largest 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 home 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 organochlorine frequency in rivers 

ranked the Delaware at Trenton and the Schuylkill, the largest tributary to the Delaware, as the 

8th and 1st worst, respectively in the nation (Jackson et al. 2005). While there aren’t many point 

sources of pollution since the Clean Water Act was enacted, historically, power plants, 

municipalities, pulp and paper mills, and industries such as the Philadelphia Shipyard, 

Bethlehem Steel, New Jersey Zinc Company, contributed large quantities of pollutants to the 

Hudson. Approximately 95% of PCBs are introduced to the river through combined sewage 

overflows from treatment plants. Even 35 years 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 levels 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 naval 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 stretch 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 watershed, 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 mining, 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 mortality rates are recorded in gillnet fisheries. 

Recreational fishermen target weakfish, striped bass, 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 (near 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 Coastal Plain but the watershed includes parts of the 

Piedmont Province and the Appalachian Province. The tributaries provide a mixture of waters 

with a broad geochemical range to the Bay with its own mixture of minerals, nutrients and 

sediments depending on the geology of the place where 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, major rivers include the 

Susquehanna, Potomac, and the James River, which we describe in greater detail below. 

Susquehanna River. Ranked as the 18th 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 mi2 (Kammerer 1990, Jackson et al. 2005). The river flows north to 

south from New York, through Pennsylvania, and 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/riffle 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 approximately 383 miles long and has a watershed of 

14,670 mi2

. The river’s headwaters begin in the Allegheny 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 tributaries 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 below the Washington, D.C. Maryland border, just 
251

below the Little Falls of the Potomac River. Ninety-five fish species live 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 mi2

. 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 state 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 average 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 the 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 usage 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, Pennsylvania and Havre de Grace, Maryland. The 

human population in the watershed 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 cities. The three main events that had the 

greatest effect on the river were logging, dam building, and mining. 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 transport 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 mined. 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 river’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 waste that is dumped directly into the river. In 

Binghampton, New York, there are 10 sewer outfalls, 70 in Scranton, Pennsylvania, 65 in 

Harrisburg, Pennsylvania, and the number of outfalls 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, Virginia. 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, zinc, and pesticides (Beyer and Day 2004). 

The Potomac River watershed usage is 32% agriculture, 58% forested, 5% developed, 4% water, 

1% wetland, and 1% barren (Jackson et al. 2005). Major towns in the Potomac River watershed 

are Washington, D.C.; Arlington and Alexandria, Virginia; and Hagerstown, Maryland. The 

human population in the watershed 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 are still elevated amounts of cadmium, chromium, copper, lead, 

dioxin, PCBs, and chlordane, which may have resulted in recent highly publicized reports of 

male fish producing eggs. 

The James River watershed usage is 23% agriculture, 71% forested, and 6% urban (VDCR 

2006). Major towns in the James River watershed are Charlottesville, Richmond, Petersburg, 

and Hampton Roads, Virginia. The human population 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 were fined and forced to clean up large areas of 

extreme toxicity. Even 35 years after the Clean Water Act, there are still elevated amounts of 

zinc, copper, cadmium, nickel, chromium, lead, arsenic, dioxin, PCBs, and pesticides. 

Hydromodification Projects. There are many dams along the Potomac River and its tributaries, 

but only three impoundments are 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 appear to be responding positively. For instance, 

between 1928 and 1972, no shad passed Conowingo Dam, 10 miles upstream of the mouth of the 

Susquehanna River, but since fish began coming 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 Boscher 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 all of the old mines in the watershed. Abandoned 

coal mines leach sulfuric acid as a result of natural 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, bluefish, oyster, red drum, spot, striped bass, summer 

flounder, weakfish, menhaden, and white perch (CFEPTAP 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 the 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 states around the bay. Species like Spanish 

mackerel, king mackerel, red drum, and summer flounder have ranges that extend beyond the 

bay and are managed under multiple regional management plans. Some species such as 

American shad are allowed to be fished by some states (Virginia and 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 ultimately drain to the Atlantic Ocean between the states 

of North Carolina and Florida. This region includes 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 (southern most tip of Florida’s 

panhandle). The hot continental division is characterized by it’s winter deciduous forest 

dominated by tall broadleaf trees, soils rich in humus and moderately leached (Inceptisols, 

Ultisols, and Alfisols), and rainfall totals that 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 with particularly high humidity during summer 

months, and warm mild winters. Soils are strongly leached and rich in oxides of iron and 

aluminum (Bailey 1995). The subtropical ecoregion is forested, largely by second 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 north, decreasing slightly in 

the central portion of the region, and increasing to 64 inches in southern Florida. The savanna 

ecoregion has a tropical wet-dry climate, controlled by moist warm topical air masses and 

supports flora and fauna that is adapted to fluctuating water levels (Bailey 1995). 

In the sections that follow we describe several basins and estuaries to characterize the general 

ecology and natural history of the area, and past and current human activities 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 climate ensures variability 

in biological productivity and hydrology. Major 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 and has a higher gradient 

and a streambed largely characterized by cobble, gravel and bedrock. 

The southern rivers are characterized by larger portions of low gradient reaches, and streambeds 

that are composed of greater amounts of sand and fine sediments—are often high in suspended 

solids, and have neutral to slightly acidic waters with high concentrations of dissolved organic 

carbon. Rivers emanating entirely within the Coastal Plain are acidic, low alkalinity, blackwater 

systems with dissolved organic carbon concentrations often up to 50 mg/L (Smock and Benke 

2005). We described several river basins in detail to provide additional context for evaluating 
254

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 Estuarine Complex, the largest lagoonal 

estuarine system in the United States, includes seven sounds including Currituck Sound, 

Albemarle Sound, Pamlico Sound and others (EPA 2006b). The Estuarine Complex is separated 

from the Atlantic Ocean by the Outer Banks, a long barrier peninsula, and is characterized by 

shallow waters, wind-driven tides that result in variable patterns of water circulation and salinity. 

Estuarine habitats include salt marshes, hardwood swamp forests, and bald cypress swamps. 

The Albemarle-Pamlico watershed encompasses four physiographic regions—the Valley and 

Ridge, Blue Ridge, Piedmont and Costal Plain Provinces. The 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 gradient 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 Complex 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 approximately 410 miles long and drains a watershed of 

9,580 mi2

. 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-rubble bedrock system. The middle Roanoke River 

is primarily course sand and gravel. The lower section of the 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 day, 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 Roanoke is also home to nine endangered fish 

species, two amphibians, and seven mussels, including several important anadromous fish 

species. 

The Neuse River is 248 miles long and has a watershed of 6,235 mi2 (Smock and Benke 2005). 

The Neuse River watershed is also located entirely within the state of North Carolina, flowing 

through the same habitat as the Cape Fear River, but ultimately entering Pamlico Sound. The 

river originates in weathered crystalline rocks of the piedmont and crosses sandstone, shale, and 

limestone before entering Pamlico Sound (Turekian et al. 1967). The average precipitation is 

approximately 48 inches. At the mouth, the average discharge is 3.4 billion gallons each day, or 

5,297 cubic feet per second (USGS 2005). 

Human Activities and their Impacts 
255

Land Use. Land use in the Roanoke River is dominated by forest (68%) and the basin contains 

some of the largest intact, least disturbed bottomland forest floodplains along the eastern coast. 

Only 3% of the basin qualifies as urban land uses, and 25% is used for agriculture (Smock and 

Benke 2005). The only major town in the Roanoke 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% agriculture, 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 counties in the Southeast (EPA 2006b). Much of the estuarine 

complex is protected by large amounts of state and 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 Albemarle-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 indicate 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, crushed 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 nutrients to the estuary and a contributor to the 

harmful algal blooms (HABs) in summer, although according to McMahon and Woodside 1997 

(EPA 2006a) nearly one-third of the total nitrogen inputs and one-fourth of the total phosphorus 

input to the estuary are from atmospheric sources. Primary agricultural 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 sampled within the estuary also showed elevated concentrations 

of total PAHs and total PCBs—10% of the sampled stations exceeded risk-based EPA Advisory 

Guidance values (EPA 2006b). Water quality studies in the mid-1990s showed the Neuse Basin 

contained the highest nitrogen and phosphorus yields, while the Chowan Basin had the lowest 

yields (Spruill and Survey 1998). 

The Neuse River entered the national spotlight 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 the 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 reported 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 precise cause is not clear, 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 the 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 urban 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 fishery 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. Roughly 100 species are fished commercially or 

recreationally in the region. The Neuse River supports 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 estuarine areas for feeding. 

Commercial fisheries along the South Carolina coast 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 Ocean including the Cape Fear, Great Pee-Dee, 

Altamaha, and the St. Johns Rivers (see Table 30 for a description of several basins within this 

region). Rainfall is abundant in the region and temperatures are generally warm throughout the 

year. Northern rivers originate in the Blue Ridge Mountains or the Piedmont Plateau, but all the 

rivers described in this section have sizeable reaches 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 highest 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 rapids 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 surface horizon, and a loamy or clay subsurface. 

The acidic characteristics, slow flowing water with poor flushing and high organic and mineral 

inputs gives these waters their characteristic “blackwater” (or “brownwater” for those that 

originate in the Piemont Plateau) appearance. The Satilla River is a blackwater river that has a 

naturally low pH (between 4 and 6) and white sandbars--due to the low pH it also has naturally 

lower productivity than other rivers that originate within the mountains or the Plateau. 

Table 30. General Information about Rivers of the Southeast United States (NCDENR 1999, Smock and 

Benke 2005). 

Watershed Length 

(mi.) 

Basin 

Size 

(mi2

) 

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 industry, and to a lesser 

extent timber and paper production, although more than half of most basins remain forested. 

Basin population density is highly variable throughout the region with the greatest 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 contrast, there are only 29 people per square mile in the 

Saltilla River watershed in Georgia (Smock and Benke 2005). See Table 31 for a summary of 

land uses and population densities in several area basins across the region (data from (Smock and 

Benke 2005). 

The largest population centers in the region include Miami and Jacksonville, Florida, and 

Savannah, Georgia. Major towns include Greensboro, Chapel Hill, Fayetteville, South Carolina, 

and Wilmington, North Carolina in the Cape Fear River watershed; Winston-Salem, North 

Carolina and Georgetown, Florence, and Sumter, South Carolina in the Great Pee-Dee River 

Watershed; Charlotte, Hickory, and Gastonia, North Carolina and Greenville and Columbia, 

South Carolina in the Santee-Cooper River watershed; Savannah and Augusta, Georgia, in the 

Savannah River watershed; Louisville, Statesboro, and Savannah, Georgia, in the Ogeechee 
258

River watershed; Athens, and Atlanta, Georgia, 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 pollution, waterfowl, urban runoff, marinas, 

agriculture, and industries including textile manufacturing, 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 (Bricker et al. 2007). Nitrate concentrations (as nitrogen) 

tend to be higher in stream draining basins with agricultural and mixed land uses (Berndt et al.

1998). Based on studies in Georgia, however, nitrate loads did not vary with growing season of 

crops (periods of heaviest fertilizer application), but were influenced 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 Several 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 have 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 agricultural land and a 38% increase in urban 

development. 

Mining. Mining occurs throughout the region. South Carolina is ranked 25th in the states in 

terms of mineral value and 13th among the eastern 26 states, and produces 1% of the total 

nonfuel mineral production value in the United States. There are currently 13 minerals being 

extracted from 485 active mines in South Carolina 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 the 
259

total nonfuel mineral production value in the United States. North Carolina’s leading nonfuel 

minerals in 2004 were crushed stone, phosphate rock, and construction sand and gravel. Georgia 

produces 24% of the clay in the nation; other leading nonfuel minerals 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 other 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 watershed. Mines through Georgia are also major producers of barite 

and crude mica, iron oxide, and feldspar. There is a proposed titanium mine near the mouth of 

the Satilla River. Unfortunately, mines release some toxic materials and negatively impact fish, 

as fish living around dredge tailings have elevated levels of mercury and selenium. 

Hydromodification Projects. Several of the rivers within the 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 projects in this area, such as levees and 

channelization projects. There are three locks and dams along the 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 blackwater 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 utilized 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 Santee River. Lake Moultrie is formed 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 meandering course, but above the fall line, there are 

three large dams that turn the piedmont section 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 above 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 largest 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 commercial 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 the annual bycatch. Shrimpers are responsible for 

50% of all bycatch in Georgia waters and often interact with sea turtles. There are 

approximately 1.15 million recreational 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, South Dakota, Tennessee, the 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, Wisconsin, Wyoming, Mexico, and two Canadian provinces. Almost 
260

2/3 of the continental United States drains to the Gulf of Mexico through the Mississippi River 

Basin. 

While the Mississippi River is the most notable basin that drains to the Gulf 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 and estuaries that enter the Gulf of Mexico to 

characterize the general ecology and 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 drainages in this region, several ecoregions are 

encompassed in this region including the subtropical, the tropical/subtropical steppe, hot 

continental and mountain segments, temperate steppe, and the prairie ecoregions (Bailey 1995). 

Most of the region is within the subtropical ecoregion (division) and is characterized by a humid 

subtropical climate with particularly high humidity during summer months and warm mild 

winters. Soils are strongly leached and rich in oxides of iron and aluminum (Bailey 1995). The 

region is forested, largely by second 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 north, decreasing slightly in the central portion of the region, and 

increasing to 64 inches in southern Florida. 

The geology of the eastern Gulf Coast is primarily sedimentary rocks of both siliclastic (sand, 

silt, clay) and carbonate (limestone and dolomite) types. 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 parts of the region exhibit calcarious 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 decline in total rainfall moving east to west, which strongly affects 

vegetation patterns, river discharge (see Table 32 – rivers are listed in their general east to west 

pattern). 

Table 32. Select Rivers in the Gulf Coast Region (Kammerer 1990, Brown et al. 2005, Dahm et al. 2005, 

Ward and Ward 2005) 

Watershed 

Length 

(mi. 

[approx.) 

Basin Size 

(mi2

) 

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 0 
261

Watershed 

Length 

(mi. 

[approx.) 

Basin Size 

(mi2

) 

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=Colorado 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 the basins east of the Mississippi, whereas 

grass/shrub and rangeland uses dominate in basins west of the Mississippi. The Mississippi also 

appears to be a divide between the less developed eastern 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 population within the San Antonio 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/mi2

, and in other areas of the basin density 

is as little as 16 people/mi2

 (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 per square mile in the Neches River watershed 

(Dahm et al. 2005). 

Major threats to the southwestern basins also include wastewater effluent, water extraction, nonpoint source pollution, nonnative species, existing impoundments, and proposals for dams 

(Dahm et al. 2005), and new reservoirs are proposed for 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 impacted rivers due to water quality and quantity 

concerns. The basin suffers from elevated levels 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 is 
262

very polluted, with excessive amounts of nutrients, pesticide, fecal coliform bacteria, PAHs, and 

oils. The lower Mississippi River is degraded by excess fecal coliform bacteria, PCBs, 

chlordane, turbidity, siltation, nutrients, 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 and lifestock—hog, cattle, sheep, goats), and dairy 

production are significant source of nutrients, fecal coliform, and pesticides. Other basins are 

severely impacted by altered sediment regimes. The Choctawhatchee River watershed has 

highly erodable soils, heavy rains, and intermittent 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 Brazos River are a problem 

resulting from numerous dams interrupting sediment transport within 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 toxic 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, chemical plants, hydroelectric, iron and steel 

manufacturing, and coal plants. 

Pollution of this nature can reduce productivity and health of the fish populations within the 

basin, and at times can lead to fish kills. Since 1998, there have been at least 16 fish kills, at 

least one of which was the result of elevated ammonia 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 easily 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 47 
263

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 Rivers 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 the 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 Mississippi River) have been heavily hydromodified—

fragmented by hydroelectric power plants and navigational dams, channels have been deepened, 

straightened, and contained within levees. For instance, there are 13 dams on the mainstem 

Chattahoochee and three on the Flint River, but there are no major 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 major dams throughout the watershed. 

There are more than a 132 dams on the Brazos River—as a result of the dams there has been a 

reduction in sediment transport to reaches below 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 native fishes and other animals. According to 

Dahm et al. (2005), although development is not prevalent 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 from 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, trawls, paired trawls, and cast nets. Recreational 

fishermen are allowed to use hand lines, rod and reels, spears, and cast nets. This gear poses a 

risk to gulf sturgeon as a potential bycatch species. Gulf of Mexico fishing regulations require 

special gear to release turtles and smalltooth sawfish. 

Ballast Water Bioregions NEP-VI (US/Mexico Border to Los Angeles) and NEP-V 

(California north of Los Angeles) 

The basins described in this section empty into the coastal waters of California and are 

encompassed by the state of California and parts of Oregon. Select watersheds described herein 
264

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 are 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 discussed herein, include: 

(a) the Cascade-Sierra Nevada Mountains province, which extends beyond this region as we 

have defined it and continue north into British Columbia, (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 the Mediterranean Division. Three major vegetation types are 

encompassed by this region: the temperate coniferous forest, the Mediterranean shrub and 

savannah, and the temperate grasslands/savannah/shrub. The area, once dominated by native 

grasses, is naturally prone to fires set by lightening during the dry season (Bailey 1995). 

This region is the most geologically young and tectonically 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 hydrology and biology of the individual basins (Carter and Resh 

2005). The region also covers the Klamath Mountains and the Sierra Nevada. 

The climate is defined by hot dry summers and wet, mild winters, with precipitation generally 

decreasing in southern latitudes although precipitation is strongly influences by topograph and 

generally increases with elevation. Annual precipitation 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 result 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 predictable within a basin, but the 

constancy is low across years (Carter and Resh 2005). 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 Southwest Coast Region (Carter and Resh 2005) 

Watershed 

Length 

(mi. 

[approx.) 

Basin Size 

(mi2

) 

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) 54 
265

Watershed 

Length 

(mi. 

[approx.) 

Basin Size 

(mi2

) 

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, originates within the Sierra Nevada 

near the middle of California and flows in a northwesterly direction through the southern portion 

of the Central Valley. The alluvial fan of the 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, Russian, and Santa Margarita Rivers. The Santa Ana 

Watershed encompasses portions of San Bernardino, Los Angeles, Riverside, and Orange 

counties. About 50% of coastal subbasin of the 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 densely populated portion of the basin is near 

the city of Santa Ana where density reaches 20,000 people per square mile (Burton et al. 1998, 

Belitz et al. 2004). The basin is home to nearly 5 million 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 heavily 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 Geological Survey under the National 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 sources--sources that influence peak flows—include 

stormwater runoff from urban, agricultural, and undeveloped lands (Belitz et al. 2004). 

Concentrations of nitrates and pesticides 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 the country, the primary 

influencing factor was the level of agriculture and the application of manure and pesticides 

within the basin. In the Santa Ana basin the elevated nitrogen is attributed largely to the 

wastewater treatment plants, where downstream reaches consistently exceeding 3 mg/L total 

nitrogen. Samples of total nitrogen taken upstream of the wastewater treatment plants were 

commonly below 2 mg/L (Belitz et al. 2004). Other contaminants detected at high levels 

included volatile organic compounds (VOCs; including chlorform, which sometimes exceeded 

water quality standards), pesticides (including diuron, diazinon, carbaryl, chlophyrifos, lindane, 

malathion, and chlorothalonil), and trace elements (including lead, zinc, arsenic). 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 Several 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 from the San Joaquin River was 

diverted for agriculture, and 1.5% diverted for livestock (Carter and Resh 2005). During the 

same period, Fresno, Kern, Tulare, and Kings Counties ranked top in the nation for nitrogen 

fertilizer use. Nitrogen fertilizer use increased 500% and phosphorus use increased 285% in the 

San Joaquin River basin in a 40 year 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 nonfuel 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 construction sand and gravel, cement, boron and 

crushed stone. California is the only state to produce boron, rare-earth metals and asbestos 

(NMA 2007). 

The State contains some 1,500 abandoned mines and 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 biological productivity within a basin, can result in 

fish kills at high levels and at low levels contributes to sublethal effects including reduced 

feeding, overall activity levels, and growth. The Sacramento Basin and the San Francisco Bay 

watershed is one of the most heavily impacted basins within the state 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 19th 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 concentrations 
267

were about 5 times lower than concentrations detected within the Bay today (Conaway et al.

2003, EPA 2006a). 

Hydromodification Projects. Several of the rivers within the area have been modified by dams, 

water diversions and drainage systems for agriculture and drinking water, and some of the most 

drastic channelization projects within the nation. In all, there are about 1,400 dams within the 

State of California, more than 5,000 miles of levees, 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 water diversion projects meet these 

demands—the Federal Central Valley Project and 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 and holds nearly 6 million acre-feet of water, 

delivering about 3 million acre feet. Together these 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 are impounded at least once and most have multiple 

dams or diversions. The Stanislaus River, a tributary 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 regime are altered, and such changes have had 

profound influences on the biological community within the basin—while the modifications 

generally result in a reduction of suitable habitat for native species, these changes frequently 

result in a concomitant increase of suitable habitat for nonnative species. The Friant Dam on the 

San Joaquin River is attributed with the extirpation of spring-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, squid, red sea urchin, and Dungeness crab (CDFG 

2007). Red abalone are also harvested off of the shores of California. Illegal poaching of 

abalone, including endangered white abalone continues 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 established along the west coast 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 runs 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, Idaho, and includes parts of Nevada, Montana, 

Wyoming, and British Columbia. The region is ecologically diverse, encompassing northern 

marine lowland forests, mountain forests, alpine meadows and Northern desert habitat. In this 
268

section we focus on three primary areas that characterize 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 defined by Bailey (1995), and encompassed within 

this region are the Marine and Marine Mountains Divisions, portions of the Temperate Dessert, 

and Temperate Steppe and Temperate Steppe Mountains. Puget Sound and the coastal drainages 

are contained within the Marine 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 fourth 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 includes 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 Snake, 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 select 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 19th largest river in 

the nation in terms of average annual 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 saltwater 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 

(mi2

) 

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 = Cascade-Sierra Mountains, PB = Pacific Border 

The river and estuary were once home to more than 200 distinct runs of Pacific salmon and 

steelhead, and represented adaptation 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 sturgeon, 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 most abundant non-native fish is the American 

shad, which was introduced to the basin in the late 1800s (Wydoski 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 areas), 39% is in private 

land ownership (most of which occurs in river valleys and plateaus), and the 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 substantially by humans causing dramatic changes 

and declines in many native fish populations. In general the basin supports a variety of mixed 

uses. Predominant human uses include logging, agriculture, ranching, hydroelectric power 

generation, mining, fishing and a variety of recreational activities, and urban uses. 

The decline of salmon runs in the Columbia is attributed to loss of habitat, blocked migratory 

corridors, altered river flows and pollution, over harvest, and competition from hatchery fish. 

Critical ecological connectivity (mainstem to tributaries and riparian floodplains) has been 

disconnected by dams and associated activities such as floodplain deforestation and urbanization. 

The most productive floodplains of the watershed 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 within 120 m was forested, 30% was agricultural fields and 16% 

was urban or suburban lands. In the Flathead River aquatic invasive plants such as pondweed, 

hornwort, watermilfoil, waterweed, cattail and duckweed grow in the floodplain wetlands and 

shallow lakes and in the Yakima River non-native grasses and other plant are commonly found 

along the lower reaches of the river (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 million acres of croplands within the 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 agriculturally productive areas in the United States (Fuhrer et al. 2004). 
270

Croplands within the Yakima Basin account for about 16% of the total basin area of which 77% 

is irrigated. 

Agriculture and ranching increased steadily but slowly within the Columbia River basin from the 

mid to late 1800. By the early 1900s, agricultural opportunities began increasing at a much more 

rapid pace with creation of more irrigation canals and the passage of the Reclamation Act of 

1902 (NRC 2004). Today, agriculture represents the largest water use within the basin. More 

than 105,000 acre feet per day (more than 90 percent) is used for agricultural purposes. 

Agriculture, ranching, and the related 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 1800s and early 1900s with 

deliberate burning to increase grass production (NRC 2004). Several measures are in use to 

reduce the impacts of grazing including restricting grazing in degraded areas, reduced grazing 

allotments, and lower stocking rates. Today agricultural impacts to water quality within the 

basin are second to large scale influences of hydromodification projects for both power 

generation and irrigation. Water quality impacts from agricultural activities include alteration of 

the natural temperature regime, and insecticide 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 Columbia River, many of which have been in place for decades. Water 

volumes, crop rotation patterns, crop-type, and location of within the basin are some of the 

variables that influence the distribution 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 January 2000, detected 25 pesticide compounds 

(Ebbert and Embry 2001). Another 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 occurred most often in the mixtures, along with 

azinphos-methyl, the most heavily applied pesticide, and atrazine, one of the most mobile 

pesticides in water (Fuhrer et al. 2004). However, the most frequently 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 chemicals continue to pose a serious problem to 

water quality and fish communities despite their cancellation in the 1970s and 1980s (Hinck et 

al. 2004). 

Fish and macroinvertebrate communities 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, specifically 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 the river. Portland’s population exceeds 500,000 

people, whereas the next largest cities, Spokane, Salem, Eugene, and Boise, have more than 

100,000 people (Hinck et al. 2004). Overall, however the population 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, and 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 treatment plants and 1% is from the chemical 

and allied products industry. Non-point source discharges (urban stormwater runoff) account for 

more of the total pollutant loading to the lower basin for most organics and over half of the 

metals. Although rural non-point sources contributions 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 case 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 metals, 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 chlorinated dioxins and furans (Hinck et al. 2004). 

Hydromodification Projects. More than 400 dams exist in the basin ranging from mega dams 

that store large amounts of water to small diversion 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 projects of which 18 dams are located on mainstem 

Columbia River and its major tributary, the Snake River. The Federal Columbia River Power 

System encompasses the operations of 14 major dams 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 operates 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 industrial water supply, and irrigation benefits. 

The Bureau of Reclamation has operated irrigation 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 water delivered by two structures, the Columbia River Project (Grand 

Coulee Dam) and the Yakima Project. Grand Coulee Dam delivers water for the irrigation of 
272

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 about 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, and ending with construction of Libby Dam in 

1973 and Lower Granite Dam in 1975. 

Development of the Pacific Northwest regional hydroelectric power system, dating to the early 

twentieth century, has had profound effects on the ecosystems of the Columbia River Basin (ISG 

1996). These effects have been especially adverse to the survival of anadromous salmonids. 

The construction of the Federal power system modified migratory habitat of adult and juvenile 

salmonids, and in many cases presented a complete barrier to habitat access. Both upstream and 

downstream migrating fish are impeded by the dams, and a substantial number of juvenile 

salmonids are killed and injured during downstream migrations. Physical injury and direct 

mortality occurs as juveniles pass through turbines, bypasses, and spillways. Indirect effects of 

passage through all routes may include disorientation, stress, delays in passage, and exposure to 

high concentrations of dissolved gases, warm water, and increased 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 large dams (NWPPC 1986). Construction of Grand 

Coulee Dam blocked 1,000 miles of habitat from migrating salmon and steelhead (Wydoski and 

Whitney 1979a). The mainstem habitats of the 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 disconnected from the main channel, and the amount of 

large woody debris in the mainstem has been reduced. Remaining areas are affected by flow 

fluctuations associated with reservoir management for power generation, flood control and 

irrigation. Overbank flow events, important 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 propagation programs for salmon production 

within the Columbia River Basin, many of which were instituted under Federal law to ameliorate 

the effects of lost natural production of salmon within the basin from the dams on fishing. The 

hatcheries are operated by Federal, state, and 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 minimally 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 spring 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 hatcheries (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 propagation on the total production of Pacific salmon and steelhead has 

been extensive (Hard et al. 1992). Hatchery practices, among other factors, are a contributing 

factor to the 90% reduction in natural coho salmon runs in the lower Columbia River of the past 

30 years (Flagg et al. 1995). Past hatchery and stocking practices have resulted in the 

transplantation of salmon and steelhead from nonnative basins, and the impacts of these practices 

are largely unknown. Adverse effects of these practices 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; increased competition for food, habitat, or mates; 

increased predation; altered migration; and displacement 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 whether or how much artificial propagation during 

the recovery process will compromise the distinctiveness of natural 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 hatcheries—particularly federally-funded 

hatcheries— are affecting the recovery of listed salmon and steelhead in the Basin. Eventually, 

the project team would create a management approach that allows tribal, state and Federal 

managers to effectively manage Columbia River Basin hatcheries to meet conservation and 

harvest goals consistent with their respective legal 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 focused 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 from 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% coming 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 Whitney (1979a) one of the 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 phosphate 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 sites are designated as superfund 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 Columbia River Basin of which nearly 200 pose a 

potential hazard to the environment (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 have 
274

bioaccumulated metals—the exposure and bioaccumulation of these metals in native fishes in the 

basin are suspected of reducing their 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 occupied by Dalles Dam since 1957. Salmon fishing increased with 

better fishing methods and preservation 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 and 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 million 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 salmon and steelhead annually significantly 

reduced population productivity (Mann et al. 2005). The largest harvest 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 harvest consists largely of coho and fall Chinook 

salmon. Most ocean catches are made north of Cape Falcon, Oregon. Over the past five years, 

the number of spring and fall salmon commercially harvested in tribal fisheries has averaged 

between 25,000 and 110,000 fish (Mann 2004). Recreational 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 heights 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 with the highest peak, 

Mount Rainer towering over the region at 14,410 feet above sea level. As the second largest 

estuary in the United States, Puget Sound has about 1330 miles of shoreline, extends from the 

mouth of the Strait of Juan de Fuca east, including 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 major geographic marine basins: Hood Canal, South 

Sound, Whidbey Basin, and the Main Basin. The Main Basin has been further subdivided into 

two sub-basins: Admiralty Inlet and Central Basin. Each of the above basins forms a depression 
275

on the sea floor in which a shallower ledge or sill separates the relatively deep water from the 

adjacent basin. The waters of Puget Sound function as a partially mixed, two-layer system, with 

relatively fresh water flowing seaward at the surface and salty oceanic water entering at depth. 

The main ledge of Puget Sound is located at the north end of Admiralty Inlet where the water 

shoals to a depth of about 200 feet at its shallowest point (King County 2001). The deepest point 

in Puget Sound is found in the Central Basin and is over 920 feet. Approximately 43% of the 

Puget Sound’s tideland is located 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 environment include eelgrass meadows, kelp forest, 

mud flats, tidal marshes, subestuaries (tidally influenced portions of river and stream mouths), 

sand spits, beaches and backshore, banks and bluffs, and marine riparian vegetation. These 

habitats provide critical functions such as primary food production, support habitat for 

invertebrates and juvenile and adult fishes, and provide foraging and refuge opportunities for 

birds and other wildlife. 

The Puget Sound ecoregion is a glaciated area consisting 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 mostly Tertiary sedimentary bedrock formations. 

The land and vegetation surrounding Puget Sound waters 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 (Franklin 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 result of steep topography, soil, and microclimate (sun exposure, 

temperature, and precipitation). Dominant vegetation types include from the Puget lowland 

region – the lowland forest, the mid-montane 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 the treeline (Kruckeberg 1991). 

The Puget Sound region has a Mediterranean-like climate, with warm, dry summers, and mild 

wet winters (Franklin and Dyrness 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, such that it will sometimes persist throughout 

much of the summer months. Average annual rainfall in the north Cascades at Mount Baker 

Lodge is about 110 inches, and at Paradise Station 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, the Puyallup/Green River, and the Lake 

Washington/Cedar River watershed. Major rivers from the Olympic Mountains include the 
276

Hamma Hamma, the Duckabush, the Quilcene, and 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, coastal cutthroat trout, bull trout, and Dolly 

Varden (Wydoski and Whitney 1979a, Kruckeberg 1991). Important commercial fishes include 

the five Pacific salmon species and several rockfish species. A number of introduced species 

occur within the region including brown trout, brook trout, Atlantic salmon, bass, tunicates (sea 

squirts), and a saltmarsh grass (spartina). Estimates suggest that more than 90 species have been 

intentionally or accidentally introduced in the region (Ruckelshaus 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), urban areas (industrial and residential use), and rural areas (low 

density residential with some agricultural activity). In the 1930s, all of Western Washington 

contained about 15.5 million acres of “harvestable” 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 Lowland; in this area forest 

cover declined by 11.1% between 1991 and 1999 (Ruckelshaus and McClure 2007). Projected 

land cover changes (Ruckelshaus and McClure 2007) indicate that trends 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, Cedar-Sammamish Basin, Green-Duwamish, and 

Puyallup River watersheds), and the area is expected 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 have been heavily modified, primarily from 

agricultural land conversion and urban development (NRC 1996b). Although most estuarine 
277

wetland losses result from conversions to agricultural land by ditching, draining, or diking, these 

wetlands are also experiencing increasing effects from industrial and urban causes. 

The most extreme case of river delta conversion 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 Washington directly with Puget Sound. After several private and smaller 

attempts, by the early 1900s locks were built achieving this engineering feat. The resultant 

outcome was that the Black River, which formerly 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 migrated sporadically between the Puyallup and the 

Green/Duwamish basins, was permanently diverted into the Puyallup River basin in 1914 with 

the construction of concrete diversion at river mile 8.5, resulting in a permanent increase of the 

Puyallup River flows by about 50% and a doubling 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 lower river dredging in the early 1900s reducing the 

lower nine miles of the river to 5 miles in length. 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 the 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 especially vulnerable to conversion with losses of 50-90% common 

for individual estuaries. 

More than 100 years of industrial pollution and urban development have affected water quality 

and sediments in Puget Sound. Many different kinds of activities and substances release 

contamination into Puget Sound and the contributing waters. Positive changes in water quality 

in the region, however, are also evident. One of the most notable improvements was the 

elimination of sewage effluent to Lake Washington in the mid 1960s, which significantly 

reduced problems within the lake from phosphorus pollution and triggered a concomitant 

reduction in the cyanobacteria (Ruckelshaus and McClure 2007). 

Even so, as the population and industry has risen 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 impaired due to contaminants, such as toxics, 

pathogens, and low dissolved oxygen or high temperatures, 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 the past 20 years the body burden in harbor seals 

has increased dramatically from 50 ppb to more than 1,000 ppb. Primary pollutants of concern 

in Puget Sound include heavy metals, organic 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, Saratoga Passage, and Sinclair Inlet (DOE 

2002). Hypoxic dissolved oxygen concentration (<3 mg/L) were found at several (11 out of 54) 

stations. Dissolved oxygen concentrations less than 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, with 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 Department of Ecology (DOE) began a program to monitor marine 

sediment conditions called the Puget Sound Assessment and Monitoring Program (PSAMP). 

The PSAMP is a multi-agency partnership administered by the Puget Sound Action Team. From 

1989-1995 the Marine Sediment Monitoring Program 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 sediment quality at ambient (i.e. away from point 

sources of contamination) stations through the Sound. Eighty-six stations were established 

throughout Puget Sound, Hood Canal, the Strait of Georgia, and the Strait of Juan de Fuca. 

Stations were grouped in two categories: core stations sampled annually, and rotating stations 

sampled once every three years alternating between North, Central and South Puget Sound 

regions. At each station, replicate sediment samples were collected for the analysis of chemical 

contaminants, sediment variables, and benthic community structure. 

Overall, contaminant concentrations at monitoring stations were generally low and below state 

sediment quality standards. Metals and semi-volatile organic compounds were most frequently 

detected. The highest metal and organic contamination was found in locations associated with 

urban and industrial centers. Low metal concentrations were also detected in some rural areas 

and in deep depositional environments. Contaminant 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 Dyes Inlet, with concentrations above standards 

for each of the seven years monitored. 

By 2000, annual monitoring of sediments at ten historical PSAMP stations showed mixed trends 

in recent years for some chemicals found in sediments (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, concentrations of a number of priority pollutant and metals also 

decreased significantly. Individual PAH levels decreased at the Point Pully station, but increased 

significantly at the Bellingham 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 in 
279

anthropogenic input of contaminants to the estuarine system over this 12-year study period. 

Also, changes in grain size and benthic infaunal community composition seen at the Strait of 

Georgia station were probably linked to increased precipitation and subsequent 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 chemicals, 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 toxicity, and adversely altered benthos, 

occurred in samples that represented about 1% of the total area (5,700 acres) (DOE 2003). These 

conditions occurred in samples collected within urbanized bays and industrial waterways, 

especially near the urban centers of Everett, Seattle, Tacoma, and Bremerton, where degraded 

conditions had been reported in previous studies. Sediments with high quality (as indicated by 

no elevated chemical concentrations, no significant responses in the toxicity 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 analyses 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 web, 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 found in ship hull paint, 

at 50% of the stations. PAHs were common while phthlalate esters, PCBs, DDTs and 

dibenzofurans appeared at fewer 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 degradation in critical nearshore habitat may 

disproportionately affect important fish, shellfish and aquatic plant species (DOE 1997-2003 

posters). 

The USGS assessed water quality of streams, rivers and groundwater in the Puget Sound Basin 

as part of the National Water-Quality Assessment (NAWQA) Program between 1996 and 1998. 

This assessment focused on the quality of surface and ground waters and biological indicators 

such as fish status, algal status and invertebrate status in relation 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 than half of the pesticide compounds (26 of 47 

analyzed) were detected. The study found that large 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 were detected in streams and rivers, with different mixtures of 

chemicals linked to agricultural and urban settings including atrazine, prometon, simazine and 
280

tebuthiuron, carbaryl, diazinon, and malathion (Bortleson and Ebbert 2000). Commonly 

detected volatile organic compound in the agricultural 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 were 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 withdrawal projects or 

small impoundments that can impede migrating salmon. 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 area rivers and streams, and the volume and quality 

of water delivered to Puget Sound waters. Several 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 River 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, prevent 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 discussed 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 expected to open up more than 70 miles of high 

quality salmon habitat (Wunderlich et al. 1994). Estimates suggestion that nearly 400,000 

salmon could begin using the basin within 30 years 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, mostly from shoreline armoring associated with the 

Burlington Northern Railroad tracks (Ruckelshaus and McClure 2007). Levee development 

within the rivers and their deltas has isolated significant 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, and in 2004 the state was ranked 

13th nationally in total nonfuel mineral production value and 17th 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 between 1940 and 1970 (Palmisano et al. 1993a). Today, construction sand 

and gravel, Portland cement and crushed stone are the predominant materials mined. Where 

sand and gravel is mined from riverbeds (gravel bars and floodplains) it may result in changes in 

channel elevations and patterns, instream sediment loads, and seriously alter instream habitat. In 
281

some cases, instream or floodplain mining has resulted 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 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. Many of the same species are sought by Tribal 

fisheries, and by charter, and recreational anglers. Nets and trolling 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 marine 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 originating in the Klamath Mountains, the 

Oregon Coast Mountains and the Olympic Mountains--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 moderately sized streams dot the coastline. Many of the 

basins in this region are relatively small—the 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 most 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, and 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 ranges from about 60 inches to more than 180 

inches, much of which falls as rain, and supports a rich temperate forest. Vegetation is 

characterized by giant coniferous forests of Sitka 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 rivers, 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 (outside of the Columbia River estuary) include 

Coos Bay, Tillamook Bay and the Nehalem River Estuary in Oregon, and Grays Harbor, and 

Willapa Bay in Washington. 

Human Activities and their Impacts 

Land Use. The rugged topography of the western Olympic 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 coastal region is forested (Palmisano et al. 1993a). 

Approximately 92% of the Nehalem River basin is forested, with only 4% considered 

agricultural (Maser and Johnson 1999). Similarly, in the Umpqua River basin about 86% is 
282

forested land, 5% agriculture and 0.5% are considered urban lands—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, Tillamook 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 private entities. In the Nehalem 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 rivers 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 streams of Washington. 

In the past, temporary splash dams were constructed throughout the region to transport logs out 

of mountainous reaches. The general practice involved building a temporary dam in the creek 

adjacent to the area being logged, the pond was filled with logs and when the dam broke the 

floodwater would carry the logs to downstream reaches where they could be rafted and moved to 

market or downstream mills. Thousands of splash dams were constructed across the Northwest 

in the late 1800s and early 1900s. While the dams typically only temporarily blocked salmon 

habitat, in some cases they remained long enough to wipe out entire runs, the effects of the 

channel scouring and loss of channel complexity resulted in the long term loss of salmon habitat 

(NRC 1996a). 

Mining. Oregon is ranked 35th nationally in total nonfuel mineral production value in 2004, 

while Washington was ranked 13th nationally in total nonfuel mineral production value 2004 and 

17th 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 may result in changes in channel elevations and 

patterns, instream sediment loads, and seriously alter instream habitat. In some cases, instream 

or floodplain mining has resulted 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 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. Many of the same species are sought by Tribal 

fisheries, and by charter, and recreational anglers. Nets and trolling 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 marine 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 storms 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 season, 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 nesting 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; surprisingly, few nests were lost (Jaime Pena, Gladys Porter Zoo, pers. 

comm.). Tropical storm Hermine arrived 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 directly 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. Anthropogenic sources of 

ambient noise include: transportation and shipping traffic, dredging, construction activities, 

geophysical surveys, and sonars. In general, it has been asserted that ocean background noise 

levels have doubled every decade for the last six 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 frequencies from 20 to 300 Hz (Richardson et al. 1995b, 

Andrew et al. 2002, Hildebrand 2009). Background noise has increased significantly in the past 

50 years as a result of increasing vessel traffic, and particularly shipping, with increases of as 

much as 12 dB in low frequency ranges and 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 commercial vessels has tripled, 

carrying an estimated six times as much cargo (requiring larger, more powerful, and 

consequently louder vessels) (Hildebrand 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 surface-active communication (splashes; carry relatively little 

information) when low-frequency background noise increased due to increased sea state. Sonars 

and small vessels also contribute significantly to mid-frequency ranges (Hildebrand 2009). 

The northeastern US hosts some of the busiest commercial shipping lanes in the world, including 

those leading into Boston, Providence, Newark, and New York (MARAD 2011). In addition to 

vessel traffic, marine construction activities occur in the Cape Cod area (liquefied natural gas 
284

terminal construction, pile driving, offshore wind farm construction, dredging, cable laying, 

drilling, and others) that contributes to local and regional 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 other 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 Francisco. Important Alaskan ports include Kodiak, 

Alaska’s largest commercial fishing port, and Valdez, the southern terminus of the 1,300 km 

trans-Alaska pipeline. Additional minor ports are located throughout the region and include: 

Anchorage, Cordova, Homer, Kodiak, Nikiski, Seward, Whittier, and Yakutat. 

The Puget Sound and nearby waters experience very high levels of vessel traffic from both 

commercial and recreational sources, producing the potential for ship strike, high ambient noise 

levels, and behavioral harassment of southern resident killer whales. Commercially, a quarter 

million vessels move within the Puget Sound region annually, with the Ports of Seattle and 

Tacoma combining to be the third largest port in the US (www.washingtonports.org). These 

vessels include tankers, tugs, cargo containers, ferries, and a variety of other vessel types. 

Several cruise ships are also based out of Seattle. Recreationally, 244 marinas, nearly 40,000 

moorage slips, and 331 boat launches are located within the Pacific Northwest, servicing 180,000 

registered recreational vessels and countless vessels not requiring registration (WSDE 2006). 

Haro Strait, one of the regions primary shipping lanes, is frequently 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 Diego (DoT 2005). These ports service a wide 

variety of vessels, including cargo, 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, accounting 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 shipping 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 California have proposed more stringent ballast water discharge 

standards than those proposed by the USCG’s ballast 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 practicability and periodic reviews 

to determine whether it is practicable to implement a more stringent ballast water standard. 

In-water construction activities (e.g., pile driving associated with shoreline projects) in both 

inland waters as well as coastal waters in the action area can produce sound levels sufficient to 

disturb proposed and listed species under some conditions. Pressure levels from 190-220 dB re 1 

μPa were reported for piles of different sizes in a number of studies (NMFS 2006b). The 

majority of the sound energy associated with pile 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 μPa 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 known to be underway or underdevelopment that 

have the potential to impact North Atlantic right whales. A 130-turbine wind farm is proposed 

or reviewed for construction off Long Island, New York and another 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 likely 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 used their primary mechanism for gaining information about their 

environment. Although neither area co-occurs locally with high North 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 southeastern/northeastern US breeding and foraging areas. In addition, a 

liquefied natural gas terminal is planned for construction off Gloucester, Massachusetts. This 

project would involve similar stressors, but located in prime North Atlantic right whale foraging 

habitat. 

Several measures have been adopted to reduce the sound pressure levels associated with in-water 

construction activities or prevent 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 peak pressures that are about 17 dB lower than 

those generated by impact hammers (Nedwell and 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 turtles during construction activities; and maintenance of a buffer 

zone around the project area, within which sound-producing activities would be halted when 

marine mammals enter the zone (NMFS 2008e). 

The state of marine and estuarine wetlands are a particular concern because of the reliance that 

listed anadromous fish have on them (Quinn 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 consequences 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 thought to have been caused by the complete obstruction of the gut 

with plastic marine debris (Lambertsen 1990). The stomach contents of two sperm whales that 

stranded separately in California included extensive amounts of discarded fishing netting (NMFS 

2009). A fifth individual from the Pacific was found to contain nylon netting in its stomach 

when it washed ashore in 2004 (NMFS 2009). Further incidents may occur but remain 
286

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 death or serious injury (Lutcavage et al. 1997a, Laist et al.

1999). Gulko and Eckert (2003) estimated that between one-third and one-half of all sea turtles 

ingest plastic at some point in their lives; this figure is supported by data from Lazar and Gracan 

(Lazar and Gračan 2010), who found 35% of loggerheads had plastic in their gut. A Brazilian 

study found that 60% of stranded green sea turtles had ingested marine 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 curiosity or due to confusion with prey items; for 

example, plastic bags can resemble jellyfish (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 maturity and increasing predation risk (McCauley 

and Bjorndal 1999). Studies of shore cleanups have found that marine debris washing up along 

the northern Gulf of Mexico 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 and 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 already declining (such as loggerhead and leatherback sea turtle 

populations in the action area). 

Invasive species have had major impacts on aquatic, estuarine, and marine ecosystems. As 

information presented here would be largely 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 Mexico 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 and exploitation. These 

activities increase vessel traffic, produce loud sounds 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 humpback whales in southeastern 

Alaska to noise from a single air gun or to playback of recorded sounds of oil drilling, production 

platforms and aircraft. Whales showed no overall 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. These 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: Northstar and Endicott. Multiple marine geophysical 

(seismic) projects are planned for the Beaufort and Chukchi Seas. 

The northern Gulf of Mexico is the location of massive industrial activity associated with oil and 

gas extraction and processing. Over 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 both detrimental and beneficial for sea turtles. These structures appreciably increase the 

amount of hard substrate in the marine environment, providing shelter and foraging opportunities 

for species like loggerhead 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 shock 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 meters of the structure and overt behavioral 

(potentially physiological) impacts for individuals further out (Duronslet et al. 1986, Klima et al.

1988). Although observers and procedures are in place to mitigate impacts to sea turtles (i.e., not 

blasting when sea turtles are present), not all sea turtles are observed all the time and low-level 

sea turtle injury and mortality still occurs (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 abandonment 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 releases below). Overall, these activities provide both positive and 

negative effects at the individual 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 refining activity in the region. Routine discharges into the 

northern Gulf of Mexico (not including oil spills) include roughly 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. Although this is only 10% of the amount discharged in a major oil 

spill, such as the Exxon Valdez spill (roughly 1 million barrels), this 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 oil discharged annually in the northern Gulf of 

Mexico, making non-spilled oil normally one of the leading 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 naturally 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 northern Gulf of Mexico 

daily (MMS 2007b); worldwide, it is estimated that 900,000 barrels of oil are released into the 

environment as a result of oil and gas activities (Epstein and (Eds.). 2002). Even if a small 

fraction of the annual oil and gas extraction is released 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 and Mississippi Delta areas 

(and the extensive activities taking place at these facilities), these locations have the greatest 

probability of experiencing oil spills. Oil released into the marine environment contains 

aromatic organic chemicals known to be toxic to a variety of marine life; these chemicals tend to 

dissolve into the air to a greater or lesser extent, depending upon oil type and composition 
288

(Yender et al. 2002). Solubility of toxic components is generally low, but does vary and can be 

relatively high (0.5-167 parts per billion; (Yender et al. 2002)). Use of dispersants can increase 

oil dispersion, raising the levels of toxic constituents in the water column, but speeding chemical 

degradation overall (Yender et al. 2002). The remaining oil becomes tar, which forms floating 

balls that can be transported thousands 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 prey 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 the 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 herbivorous marine iguanas when the Galapagos Islands 

experienced an oil spill, contributing to a >60% decline in local populations the following year. 

The potential exists for green sea turtles to experience 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 oiling 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 potentially causing death (NOAA 

2003). Dispersants reduce the formation of tar balls. Although the effects of dispersant 

chemicals on sea turtles is unknown, testing on other organisms have found currently used 

dispersants to be less toxic than those used in the past (NOAA 2003). It is possible that 

dispersants can interfere with surfactants in the lungs (surfactants prevent the small spaces in the 

lungs from adhering together due to surface tension, facilitating large surface areas for gas 

exchange), as well as interfere with digestion, excretion, and salt gland 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, cessation 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 beaches 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 Ivan in 2004 on the Gulf Coast included pipeline 

damage causing 16,000 barrels of oil to be released 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 storage facilities, releasing 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 environment for nine months, releasing 2.8-7.5 million 

barrels of oil. Oil from this release eventually reached the Texas coast, including the Kemp’s 

ridley sea turtle nesting beach at Rancho Nuevo, from where 9,000 hatchlings were airlifted and 

released offshore (NOAA 2003). Over 7,600 m3

 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 recovered during the oil spill event; 

although cause of deaths were not determined, oiling was suspected to play a part (NOAA 2003). 

Also in 1979, the oil tanker Burmah Agate collided with another vessel near Galveston, Texas, 

causing an oil spill and fire that ultimately released 65,000 barrels 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 semisubmersible drilling platform 

Deepwater Horizon roughly 80 km southeast of the Mississippi Delta (NOAA 2010a). The 

platform had 17,500 barrels of fuel aboard, which likely burned, escaped, or sank with the 

platform (NOAA 2010a). However, once the platform 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 oiled regions of Texas, Louisiana, Mississippi, Alabama, and 

Florida and widespread oil slicks throughout the northern Gulf of Mexico that closed more than 

one-third of the Gulf of Mexico Exclusive Economic Zone to fishing due to contamination 

concerns. Apart from the widespread surface slick, massive undersea oil plumes formed, 

possibly through the widespread use 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 animals’ health status continue. 

Perhaps the most famous shipwreck of all time occurred 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. The spill was the worst in US history until the 

Deepwater Horizon event in 2010. The Alaska Department of Environmental Conservation 

estimated that 149 km of shoreline was heavily oiled 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 level of helicopter activity over Gulf of Mexico 

waters. It is estimated that between roughly 900,000 and 1.5 million helicopter take-offs and 

landings are undertaken in association with oil and gas activities in the Gulf of Mexico annually 

(NRC 1990a, USN 2008). This likely includes numerous overflights of sea turtles, an activity 

which has been observed to startle and at least 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 significance of specific pollutants 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 (including the fertilizers, 

pesticides, herbicides, and other contaminants that are applied to it) and discharges into the Gulf 

of Mexico near the action area (MMS 1998). Agricultural discharges, as well as discharges from 

large urban centers (ex.: Houston and New Orleans) contribute contaminants as well as coliform 

bacteria to Gulf of Mexico habitats (Garbarino et al. 1995). These contaminants can be carried 

long distances from terrestrial or nearshore sources and ultimately accumulate in offshore pelagic 

environments (USCOP 2004). The ultimate impacts 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 hypoxic “dead zone” (USCOP 2004). This hypoxic event 

occurs annually from as early as February to as late as October, spanning roughly 12,700 km2

(although in 2005 the “dead zone” grew to a record size of 22,000 km2) from the Mississippi 

River Delta to Galveston, Texas (MMS 1998, Rabalais et al. 2002, LUMCON 2005, USGS 

2010). Although sea turtles do not extract 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 “dead zone” decreases biodiversity, alters marine food webs, and 

destroys habitat (Craig et al. 2001, Rabalais et al. 2002). High nitrogen loads 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 resident killer whales (Ross et al. 2000c, CBD 2001a, Krahn et al. 2002b, 
291

Koski 2007b, NMFS 2008e). These pollutants have the potential to cause immune, endocrine, 

and reproductive system effects in killer whales; concern has been expressed by researchers that 

these pollutants may be a factor in the decline of southern resident 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 primarily 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 environment (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 move refined oil products 

to destinations along the 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, the 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 spills of at least 2,222 barrels; 

the largest was over 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 concern are the chronic small-scale 

discharges of oil, which together can be substantial (Clark 1997). These 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 the 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 whales and other marine 

mammals is poorly documented (NMFS 2008e). In addition, the long-term effects of repeated 

ingestion of sub-lethal quantities of petroleum hydrocarbons on marine mammals are not well 

understood. As a result, the magnitude of the risks posed by oil discharges in the proposed 

action area is difficult to precisely quantify or estimate. 

Commercial Fisheries 

Commercial fisheries are a major factor in the decline of several listed species in this Opinion. 
292

Directed harvesting of whales as been the greatest single reason several whale species are listed 

as threatened or endangered. However, harvests of listed whales have been largely eliminated 

and the species dealt with in this Opinion species appear to be generally recovering. The 

significance of whale reductions compared to historical levels is provided under Status of Listed 

Resources. Salmonids have also experienced significant fishing pressure, which is discussed 

below. Sea turtle species also continue to suffer from commercial harvesting in their nesting 

areas. While there is an international moratorium on harvesting whales, Japan harvests up to 101 

sei whales annually (IWC 2008). 

Although protected under the ESA, listed salmonids are killed in commercial harvests along the 

US west coast. As salmonid DPSs/ESUs are readily differentiated genetically, but not 

morphologically, the extent of impact that harvesting 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 the 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 included 1,377 metric tons of Chinook, 5,352 metric 

tons of chum, 1,139 metric tons of Coho, 36 metric tons of sockeye, 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 Washington, 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 mortality, there are additional consequences of fishing. Trawls 

can disturb the ocean floor, causing physical damage 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 also prey to Steller sea lions. Significant concern 

exists that the level of fishing results in significant 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 metric tons of arrowtooth flounder, 8,000-

8,500 metric tons of rockfish, and ~5,500 metric tons of skates were allowed to be harvested 

annually (NMFS 2006a). 

The waters surrounding Washington State 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 regulated, there is the 

potential for overfishing, especially considering the selective nature of southern resident killer 

whale diets (Fraser River populations of Chinook salmon). Lost gill nets, purse seines, and 

longlines may foul and disrupt bottom 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 few 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 females on nesting beaches, and directed hunting of sea 

turtles in foraging areas. These factors have led 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 (Lagueux 1998, Bräutigam and Eckert 2006a). While 

these threats have been largely eliminated in Florida due to successful conservation measures, 

the hunting of juvenile and adult turtles continues both legally and illegally in many foraging 

areas where green sea turtles originating from Florida are known to occur (Fleming 2001, 

Chacon 2002). The killing of nesting hawksbill females continues to threaten 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 hawksbill shell (Milliken and Tokunaga 

1987), and in particular with the 90% reduction in the annual take 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 making it difficult to accurately 

determine the frequency of mortalities. Entanglement may also make whales more vulnerable to 

additional dangers, such as predation and ship strikes, by restricting agility and swimming speed. 

Between 1998 and 2005, observers identified 12 humpback 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 gillnet fishery during the 1980s (Heyning and Lewis. 

1990). From 1996-2000, 22 humpback whales of the Central North 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 individuals showed some kind of scarring from fishing gear 

entanglement (Neilson et al. 2005). 

Fisheries interactions are a particularly problem for North Atlantic right whales (Figure 25). 

Aside from the potential of entrapment and entanglement, 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 vulnerable to additional dangers, such as predation 

and ship strikes, by restricting agility and swimming speed. Along the Atlantic 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 could confirm, North Atlantic right whales 

were injured in five of the entanglements and killed in four entanglements. 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 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. 1948b, Pike 

and Macaskie. 1969, Haley 1970, Olesiuk et al. 1990c, Baird 2001c). This practice has largely 

abated in the past few decades and 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 eventually cut into tissue, leading 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 ingest 
296

flashers and hooks used in commercial 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 fishing 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 and, frequently, the lack there of. 

NMFS (2002) estimated that 62,000 loggerhead sea turtles have been killed 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 waters, mortality still occurs in Gulf of Mexico waters. 

In addition to commercial bycatch, recreational hook-and-line interaction also occurs. Cannon 

and Flanagan (1996) reported that 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 rehabilitated turtles, five died during rehabilitation, and 

96 were reported as released by fishermen. 

Naval Activities 

Naval activity, notable sonar use during training exercises, has gained notoriety for its 

coincidence with marine mammal strandings. However, other activities (also during training 

exercises in designated naval operating areas and training ranges) also have the potential to 

adversely impact marine mammals and sea turtles. 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 activities in several ranges. 

 Marianas Island Range Complex, where humpback whales breed and possibly give birth, 

 The Southern California 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 North Atlantic right and humpback whales 

forage, 

 The Virginia Capes, Cherry Point, and Jacksonville-Charleston Operating Areas, which 

are situated consecutively along the migratory corridor for North Atlantic right whales 

(the later operating area overlaps 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 distribution of North Atlantic right and humpback whales, but 

presence within these areas is rare or undocumented. 

Naval activities to which individuals could be exposed include, among 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 the total sound level and are designed to 

operate quietly, these ships are large and equipped with high-output sonar equipment such as 

ANISQS-53C tactical sonar, which produces signals at source levels of 235 dB re 1 μParms 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. However, an event that occurred in the Strait of Juan de Fuca and 

Haro Strait on May 5, 2003 demonstrates the potential for naval activities to impact southern 

resident killer whales. The US Navy guided missile destroyer U.S.S. Shoup passed through the 

strait operating its mid-frequency sonar during 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 sonar, 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 information 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 dead 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 explosives were used during this exercise 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 disruptive to southern resident killer whales. AHDs used at salmon 

aquaculture farms emit "loud" signals intended to displace harbor seals and sea lions and thereby 

reduce depredation (Petras 2003, NMFS 2008e). However, these signals can also cause strong 

avoidance responses in cetaceans (Olesiuk et al. 2002). Morton and Symonds (2002) describe 
298

one AHD model that broadcasts a 10 kHz signal at 194 dB re 1 μPa at 1 m. A large majority of 

these occur in Arctic waters during exploration 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 the Ballard locks in Seattle, 

where NMFS uses it to deter sea lions (NMFS 2008e). 

Vessel Approaches – Commercial and Private Marine Mammal Watching 

Although considered by many to be a non-consumptive use of marine mammals with economic, 

recreational, educational and scientific benefits, marine mammal watching is not without 

potential negative impacts. Whale watching has 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 that preferred habitats may be abandoned if disturbance levels are 

too high. In the Notice of Availability of Revised Whale Watch Guidelines for Vessel 

Operations in the Northeastern United States (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, increasing the potential for harassment, injury, 

or even the death of these animals. In addition to whale watching vessels, large cruise vessels 

also operate in waters off the coast of Alaska, 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 waters likely influence how they 

react to approaches by vessels in the future (Herman 1979). 

Several studies have specifically examined the effects of whale watching on marine mammals, 

and investigators have observed a variety of short-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 direction, respiration rate, dive time, feeding behavior, and social 

behavior (NMFS 2006b). Responses appear to be dependent on factors 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 watch vessels has also been noted to cause newborn 

calves to separate briefly from their mothers' sides, which leads to greater energy expenditures 

by the calves (NMFS 2006b). Although numerous short-term behavioral responses to whale 

watching vessels are documented, little information is available on whether long-term negative 

effects result from whale watching (NMFS 2006b). 

It is difficult to precisely quantify or estimate the magnitude of the risks posed to marine 

mammals in general and southern resident killer 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). Commercial whale watching in Washington State has 

increased dramatically from small scale operations 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 Juan 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 advantage of whale watching opportunities (MMMP 2002b). Weather 

conditions in the Pacific Ocean in winter limit whale 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 southern resident killer whales on a daily basis (Koski 2006a). In 

Washington State, southern resident killer whales are the primary target species, particularly 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 traffic and sound emitted by it. Whale 

watching activities have the potential to affect southern resident killer whales in the action area, 

resulting in possible disturbance or displacement. Increasing anthropogenic sound levels in the 

Puget Sound region have been associated 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 success and/or inhibiting 

foraging all together (Bain et al. 2006a, Williams et al. 2006). Based on a study in Johnstone 

Strait, British Columbia, northern resident killer whales decreased feeding behaviors 

significantly and increased time engaging in behaviors which required 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 released or escaped. Of the 

individuals captured and displayed or killed, 70% (47 or 48 individuals) were southern resident 

killer whales, including 17 immature males, 10 immature females, nine mature females, and 

seven or eight mature males; 15 individuals were from K pod, five from L, and one from J 

(Olesiuk et al. 1990c, Baird 2001c, NMFS 2006b). The selective removal of younger animals 

and males produced a skewed age and sex composition in the southern resident killer whale DPS, 

which probably affected its ability to recover (Olesiuk et al. 1990c). 

Ship-Strikes 

Ship-strike is a significant concern for the recovery of listed baleen whales and sea turtles to a 

lesser degree. We believe the vast majority of ship-strike mortalities go unnoticed, and that 

actual mortality is higher than currently documented. More humpback whales are killed in 

collisions with ships than any other whale species 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 and 
300

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 quantify 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 information on interactions between ships and commercial 

vessels outside of US waters. With the information 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 right whales in the region. 

We believe the vast majority of ship-strike mortalities 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 traffic 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 injured in ship strikes within the US EEZ and have 

virtually no information on interactions between ships and commercial 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 whale deaths, accounting for 

approximately 35% of all known mortalities, 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 right 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 occurs. 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 especially deleterious to the ability of the North Atlantic right 

whale population to recover. For instance, in 2005, mortalities included six 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 individuals over the short 

term (Kraus et al. 2005). Voluntary recommendations for 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 concentrations 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 April 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 southern breeding areas. Nearly a dozen shipping 
301

lanes transect through coastal waters 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 higher than currently documented. 

A total of six instances have been documented of northern and southern resident killer whales 

being struck by vessels since the 1990s, including lethal 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 potential to be highlysignificant (Work et al. 2010). All sea turtles must surface to breath and several species are 

known to bask at the surface for long periods, including loggerhead sea turtles. Although sea 

turtles can move rapidly, sea turtles apparently are not well able to move 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 Mexico, 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 vessels 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). Commercial vessel operations are also 

extensive. Vessels servicing the offshore oil and gas industry 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 Orleans and Houston)(MMS 2007a, USN 2008). 

Extensive shrimping and other fishery effort is also expended in the area. Overall, ship strike is 

likely highly underestimated as a source of injury or mortality to sea turtles 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 action area. Tables 38 identifies the 

cumulative number of takes for each listed marine mammal and sea turtle species, as well as 

smalltooth sawfish, rockfishes, sturgeon, and eulachon in the action area authorized in scientific 

research permits. Due to the large numbers of research permits authorizing salmonid takes, we 

did not quantitatively assess the numbers of actions on listed salmon and steelhead. We did, 

however, address these in a qualitative manner. Cetacean takes include approach, biopsy, 

suction cup and implantable tagging, breath sampling, acoustic playbacks, and/or ultrasound. 

Pinniped takes consist of approach, capture, handling, restraint, measurements, anaesthesia or 

drugging, temporary captivity, external 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 sampling, captive experiments, laproscopy, 

imaging, and/or antibiotic injections. Smalltooth sawfish may be captured via a variety of 

means, measured, tagged, tissue sampled, and/or ultrasounded. Research actions on sturgeon 

species include capture, handling, restraint, anaesthesia, laproscopy, lavage, boroscopy, fin, 

operculum, or barbel clipping, PIT, floy, sonic, or satellite tagging, gonad sampling, 

prophylactic, and/or mortality. Rockfishes may undergo capture, handling, restraint, anaestheia, 
302

floy or sonic tagging, and mortality. 

Behavioral disruption of Steller sea lions occurs as a result of human disturbance (FOC 2008). 

Research efforts to collect scats, count and weigh pups, and other human activities on or near 

rookeries can lead to stampedes into the water. Mortality can occur directly due to pup 

trampling, separating from mothers, or drowning. If disturbance is too frequent, haulouts may be 

completely abandoned. Although habituation to some 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-North 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 indirect effects to listed species 

and critical habitat that are reasonably likely to occur as a result of activities authorized, funded, 

or carried out by the ballast water management program including the recently proposed 

numerical standard. As discussed 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 associated with ballast water intake 

and discharge are capture of listed species during uptake of ballast water and the introduction of 

non-native species through ballast water discharge at the proposed standard. NMFS does not 

expect any effects from three methods of ballast water maintenance identified in the FR (33 CFR 

151.1510 and 33 CFR 151.2025): (1) retaining ballast water on board, (2) discharging ballast 

water to on-shore or vessel/barge based holding facilities, or (3) using public drinking water, 

meeting the requirements of the Safe Drinking Water Act (42 U.S.C. §§ 300f-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 unload cargo, the ships take on ballast water to offset 

the cargo weight being removed. When these ships take on or discharge ballast, water enters or 

leaves the ship through two to six 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 would 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 impingement, entrainment, or capture in ballast tanks 

or pipes. 

Impingement occurs when organisms that are sufficiently 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 drawn 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 stress pathologies (for example, 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 during ballast intake, these animals are too large to 
304

get impinged or entrained on pipes or screens leading into the sea chest and strong enough to 

swim out of the area. Therefore, marine mammals 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 turtles, especially recently hatched turtles, could be 

sucked into sea chests during ballast water intake activities. Recently hatched sea turtles leave 

their nesting beaches and swim to offshore rearing areas near debris collecting at convergence 

zones. It is unlikely that a hatchling 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 pipes during ballast water intake. Therefore, sea 

turtles may be affected but are not likely to be adversely affected by ballast water intake. 

In the event that juvenile Atlantic or Pacific salmonids or shortnose, 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 ballast intake system would intake less than 250 

cubic meters per hour (or about 0.069 m3

/s). Flow rates through sea chests 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 1.1m2. 

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 possible that these species wouldn’t avoid the intake 

current, however, adults of both species 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 juveniles or smolts. In both cases, interactions 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 interaction 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 (very small juveniles are considered 

to be less than one meter, Wiley and Simpfendorfer 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 ballast water intake because their habitat does not 

overlap with shipping ports. Rockfish adults 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, spend 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, larvae moving downstream would be 

vulnerable to ballast water intake for approximately three weeks until they are large enough to 

move to deep water areas in the 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 the 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 spring/early summer that some eulachon larvae could 

be removed from the system by ballast uptake. 

There are several invertebrates 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 offspring that will successfully establish in offshore 

rocky habitats would be unlikely to have any interaction with vessels intaking ballast water. 

Elkhorn and staghorn coral typically 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 larvae. Neither would be expected in 

ports, which are generally dredged environments with sandy bottoms. However, these corals live 

near port environments and their offspring could be taken up during ballast water intake if 

currents have carried their offspring 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 habitat. It is possible that the planktonic coral could have 

continued drifting to suitable habitat but the chances of a coral plankton being removed from a 

port environment that would have otherwise successfully settled is discountable. 

Johnson’s sea grass grows in bays, inlets, and lagoons. The grass typically occupies areas with 

turbid water and high tidal currents. In some cases, it may grow near ports or areas where ships 

could take up ballast water; however, as a rooted plant, there is essentially no chance it would be 

uprooted and removed from the environment. Furthermore, 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, ballast water intake would not be expected to remove any seeds 

from the environment. 

Ballast Water Discharge 

An often overlooked ballast water discharge issue is the presence of chemical pollutants in 

ballast water taken from locations 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 effects of such exposures will be analyzed in 

EPA’s biological evaluation for the vessel discharge permit. However, the effects of any 
306

pollutant generated in association with ballast 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 water discharge is the establishment of non-native 

species and how those non-native species may affect listed species and their critical habitat by 

altering the basic functional components of ecosystems. Several authors 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 population, 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 species 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 concern in ballast water: (1) planktonic life in 

suspension in the ballast tanks and ballasted cargo holds, (2) fouling organisms that settle on the 

lower walls of ballast tanks and ballasted cargo holds, (3) infaunal 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 ballasted cargo holds, and (5) fish and fish eggs in 

ballast tanks and ballasted cargo holds. The proposed 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 discharge water, which would contain the suspended 

organisms, all potential introductions associated with ballast water 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, McCarthy and Crowder 2000, Carlton 2001). 

Microorganisms, such as phytoplankton, particularly 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 produce toxic compounds. While there was 

previously not a scientific, enforceable, or practical discharge standard for microorganisms, the 

USCG is now able to test ballast water and determine 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 introduction, it has spread 
307

rapidly through rivers, bays, and lakes in the Great 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 richness, non-native species are unlikely to be able to establish in a 

new environment because the native flora and fauna are better able to compete 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 invasive population increases the 

probability that a successful establishment will occur (Kolar and Lodge 2001). The structure and 

function of aquatic and coastal ecosystems are adversely affected by increased stress from 

human activities, which have altered these systems for a long period of time. The scale, 

intensity, and rate of human activities and associated impacts have significantly increased in the 

past century from factors such as population growth, higher levels of consumption, and 

technological advances, making these systems 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 habitat, 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 species are more likely to establish than species from other habitats 

(Nehring 2006a). Estuaries and coastlines are particularly susceptible 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 capable of asexual reproduction may also be 

better able to establish in new habitats due to the lack of allee affects, 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 mediated invasion process. The 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 interaction between that species and the vessel. Second, 

if the potential invasive species is present at the port with the ship, it must then be taken up 

during ballast water intake, but also survive being taken up into the ballast tanks. Now in the 

ballast tanks, that individual 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 discharged ballast water and survive the discharge 
308

process. Fifth, upon release, the potentially invasive species must survive in the new 

environment. Sixth, the newly introduced species must establish a reproductive population, 

producing viable offspring. And finally, after that species is able to reproduce, it must survive 

and reproduce effectively enough to establish a long-term population in the area. 

While the likelihood that any given individual of a particular species might go through all the 

steps previously described and become successfully established in the long-term is quite low, the 

likelihood that a vessel will release live organisms from ballast water (stage five) is very high, 

especially given the large volumes of ballast water discharged annually in US ports. It is 

possible that listed species or critical habitat exposed to recently released organisms could 

exhibit a response, but the potential affects to listed species and their critical habitat increase 

with the duration of exposure (stage six and seven) and the wider the exposed area. Therefore, 

exposure to listed species is considered to have occurred when reproduction occurs in US waters 

and viable offspring are created (stage six). 

The contention in the final PEIS is non-indigenous species “management is best accomplished 

by preventing introduction and establishment, which 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 discharge or if no ballast water is discharged into the environment, 

there is no invasive species effect to the receiving environment. However, most vessels will 

have initial organism densities greater than the organism densities proposed by the USCG so that 

vessels meeting any of the alternatives will achieve 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 Geller 1993) (Dunston and Bax 2003), the level of inoculation is 

critical to the risks of non-native species establishing viable reproductive populations (Bax 

2003). 

For transoceanic introductions, vessel-mediated transport is virtually the only vector, but 

organisms may arrive in the US via several vectors, including the hull, anchor system, sea chest, 

internal seawater piping system, and the bilges (Keller and Lodge 2007) (Padilla and Williams 

2004, Coutts and Dodgshun 2007). Furthermore, non-native 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 that ballast water discharge is likely to be a primary source of 

non-native species associated with shipping. The reductions of organisms released via ballast 

water discharge is dependent on the complex interplay of a number of factors: the types of 

organisms; their reproductive mode; their life stage; their nutritional status; 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 densities (concentrations) of organisms transported 

via ballast water. Sampling studies show that organism 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 densities) (Cohen 1998, (Carlton 2001). These studies 

rarely focus on the bacterial and viral components of ballast, but these organisms may exist in 

even greater densities. The greater the volume discharged the more non-native species that can 

be introduced to the local environment. Additionally, 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 106

 total zooplankton remain possible, 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 discharge are also additive to 

non-native species already established (stage 6 and 7) or recently 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 (stage 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 had 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 newly introduced crabs selectively preyed on the 

native mussels, thereby freeing the non-native clams (which had been living in the area for at 

least 50 years without having a population explosion) from competition and facilitating their 

spread in the ecosystem (Grosholz 2005). 

When ballast water is discharged, there are generally 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 generally accepted that the most important vector responsible for 

transporting and introducing non-native species is ballast water discharge (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 invasive 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 aquarium 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% exclusively by ballast water, and the other 30% by 

both. 

Invasive Species History 

 

Pre-Ballast Water Regulations
310

Invasion corridors have been open since ships first sailed from one port to another; however, 

non-native species generally did not establish in the US until early in the 19th 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 resulting invasions have occurred at an increasing rate through the 

1980s, 1990s, and 2000s. 

Identifying the precise time of an invasion is often 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 have been open for decades with vessels moving 

between ports and discharging ballast water without a non-native species establishing, then for 

no explainable reason, a species like the zebra mussel becomes a problematic invader. While it 

may be impossible to know for sure why the zebra mussels established in the 1980s instead of 

the 1960s, increased habitat alteration has made many areas more susceptible to colonization by 

non-native species (Brothers and Spingarn 1992, Carlton 1996, Lake 2004, Didham et al. 2007). 

Furthermore, some non-native species that are currently 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, and 99% of the biomass there (Cohen and Carleton 

1995, Cohen and Carlton 1998). This story was similar throughout the US as habitat alteration 

and economic globalization led to increasing reports of ecological 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 control the rate of invasions. 

Ballast Water Exchange 

Ballast water exchange, when ocean water from outside the EEZ is pumped through the ballast 

tanks to exchange the ballast water taken up from nearshore port environments and replaced with 

less species-rich water from the mid-ocean, was made mandatory for the great lakes in 1993. It 

was then made voluntary for ships entering all US waters in 2001. In 2004, BWE was made 

mandatory as a ballast water treatment method intended to protect US waters from invasive 

species and their impacts. The effectiveness of BWE varies dependent upon vessel type 

(design), exchange method, ballasting system configuration, 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). Other studies demonstrate that 

the volumetric efficiency of BWE ranges from 50 to 90% (Zhang and Dickman 1999, USCG 

2001, Battelle 2003). Recent work by Ruiz et al. (in NOAA 2007) has shown that BWE can 
311

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 organisms in ballast tanks resulted 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 bound) to one-tenth (lower bound) (USCG 2003). 

As noted above, without any treatment, the mean ballast water discharge 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 because 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 will no longer be an authorized treatment 

method and would only be used by vessels 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 geographic locations is very high. In the Port of 

Hampton Roads in the Chesapeake Bay, Drake et al. (2007) estimated that 1020 microorganisms, 

6.8x1019 viruses, and 3.9x1018 bacteria cells are discharged via ballast water annually. Lawrence 

(2008) estimated 3 x 1022 viruses to be transported annually. 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 other 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, and 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 discharge of 1,000,000 large organisms, under the worst possible 

conditions, the numerical standard would release that many large organisms, but on average, the 

number of large organisms discharged would be reduced to 43,500 (Minton et al. 2005). And 

while research on the number of smaller organisms discharged under BWE hasn’t been 

conducted, the largest expected discharge of small organisms is 100,000,000,000 and the mean 
312

number of small organisms that would be discharged is 4,350,000,000 under the numerical 

standard. 

BWE is more effective than numerical standards for freshwater when ballast water is discharged 

due to the salinity differences (Gray et al. 2007). Despite the loss of the lone potential benefit of 

BWE, this numerical standard will reduce the number 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-native population establishing. Furthermore, NMFS must consider 

whether more stringent standards would provide additional protection for listed species and their 

critical habitat. NMFS uses analyzed the probable exposure associated with a standard 1000 

times more stringent than 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 analysis (PVA) to relate initial population size and 

extinction probability to determine the probability of successful establishment of new non-native 

populations. A PVA is normally used to determine the size of a native, viable population 

necessary to ensure a high probability of survival. In the context of this final PEIS, the concern 

is the initial non-native species population size must 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, theoretical ecology suggests that stochastic events 

and allee effects are more likely to cause the extinction of small populations, regardless of 

whether they are native or non-native. Therefore, NMFS is unaware of a better model for 

predicting the likely changes in invasion risk 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 growth rate and the other to population variability around mean growth. In 

conjunction with an estimate of the proportional reduction in initial population size resulting 

from the alternatives, this model allows for a comparison 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 standard 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 the scientific knowledge 

of invasive species ecology when this analysis first began in 2001 (Ludwig 1999). 

The main premise of the model is that a decrease 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). Introduction of individuals in separate events can 

increase the genetic diversity available for an invasive species, allowing for more rapid evolution 
313

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 introductions are occurring throughout the United States 

at an increasing rate as a result of increased ship traffic (Grosholz 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 ballast 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% reduction 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 model 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 are more ecologically realistic than evaluating a 

single species scenario. Under the multiple species 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 overestimates 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 and 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 proposed 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 propagule pressure. The two studies 

determined there are no valid models to address the questions they were tasked with answering. 

Both of these reviews are complete and address 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 specific 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 probabilities of any particular species, 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 the global probability analysis provided by Drake 

and Lodge (2004) and Drake et al. 2005, NMFS estimated the probability of an invasion using 

their model as well as a validation technique using historic information from a well studied 

location. Applying the best scientific information available, the validation method requires 

knowledge of historic invasions for specific ports and a conversion from no ballast water 

treatment to the current ballast water standard to estimate the current invasion rate. This method 

presents two shortcomings: (1) there is limited information 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 water discharge volumes or vessel 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 available estimate of the impacts of implementing 

the numerical standard. Because the ballast water standard establishes a numerical limit of 

organisms of a particular size allowed 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 anticipated annual invasion rate. 

The PVA designed by the USCG provides information on the generally anticipated results of the 

numerical standard relative to the 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 number of invasive species introduced. Also, 

Dunstan and Bax (2008) show BWE reduces the likelihood of an invasive 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 endangered species and their 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 because 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) argue that while it is difficult to determine whether a newly 

discovered invasive species was recently established or had been present 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 their greatest rate in 

over 200 years of introductions (Cohen and Carlton 1998). The NMFS analysis uses the 

estimated number of invasive species introduced 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 the 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 probable estimate or upper end of a range. NMFS 

acknowledges there is insufficient knowledge to get accurate numbers, but believes the estimated 

rates of invasion presented below represent the most conservative estimate.

In San Francisco Bay research first began on invasive species in 1851. At that time 
315

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 species that established 

in San Francisco Bay increased to 2.3 each year between 1961 and 1995 (Callaway and Josselyn 

1992, Cohen and Carlton 1998). However, there are several sources of non-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 water (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 probability of an established population of nonnative species by 43 to 48%. This correlates with annual ballast-mediated non-native species 

introductions being reduced from 0.968 per year to approximately 0.504 (0.968*0.52) to 0.553 

(0.968*0.57) per year. Then, using the multispecies 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 approximately one new species every three years. 

The calculated rate of 0.318 to 0.348 invasions per year is comparable to the global multi species 

model used by NMFS to estimate invasion risk at the continental scale (Drake and Lodge (2004). 

Drake and Lodge (2004)(, p.578) used a generalized linear regression with Poisson errors 

developed at a global scale to show “that shipping 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 National Ballast Information Clearinghouse, which is 

a record of all reported ballast water discharges. However, reporting rates vary between 80 and 

90% of the actual value. NMFS applied a correction factor assuming the conservative 80% 

reporting rate. This rate is expected to increase 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 Francisco Bay the corrected average from 2005 to 

2010 was 6.469x106 tonnes of ballast water were discharged under BWE management. Using 

Drake and Lodge’s (2004) equation would result in approximately 0.548 species per year being 

introduced. Then, using the multispecies analysis 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 (without the correction factor, the estimate is 0.32 

species introduced per year). 

While this test is not an exhaustive 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 invasions at individual ports as well as throughout 

the US. 

Despite setting a ballast water standard, listed species 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 discharge remains at the same level and there are no 

additional 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 logical location to use as a case study because of 

the vast research history as well as invasion history there. Furthermore, the San Francisco Bay is 

home to six listed species and five critical habitats in two different functional groups under 

NMFS jurisdiction. NMFS also decided to analyze Puget Sound and Chesapeake Bay to assess 

the impacts of invasive species on listed species and their critical habitat. By choosing those 

locations, several listed species, their critical habitats, or reasonable surrogates were not 

considered, so we analyzed 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 estuary in the US (Herborg et al. 2007). San Francisco Bay is also home 

to central valley spring-run Chinook salmon, Sacramento River winter-run Chinook salmon, 

green sturgeon, and California central valley steelhead (Table 39). Additionally, while they do 

not reside within the bay, black and white abalone 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 estuary. Many of these species were established 

before any ballast water treatment was utilized. And as noted, there is considerable uncertainty 

as to when a species establishes due to the delay between it being discharged and it being 

discovered. 

Table 39: Listed species present 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 discharge into the San Francisco Bay bioregion has increased between 

2005 and 2010 with 4.43x106

, 4.6x106, 4.49x106, 5.9x106, 5.7x106, and 5.9x106 tonnes 

discharged annually from around the globe, respectively. These numbers were corrected for the 

unreported discharges. If we accept the USCG’s PVA 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 under the proposed numeric ballast discharge standard. Following 

this trend in discharge volumes; 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 each year. In other words, if the rate of 

0.345 new species per year in 2010 is correct, given the upward trend in ballast discharge 
317

provided by the National Ballast Information Clearinghouse, 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 increases the standard in the event 

better technology is practicable, their goal is a standard that is 1000 times more stringent than the 

proposed numeric standard. In that case, the anticipated rate of invasions would fall to about 

0.09 individuals per year or about one new non-native species every 11 years. Although we can 

estimate the rate of invasion within a given system, we cannot predict the species that would 

constitute future invasions. 

Puget Sound. Puget Sound has a number of ports, including 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 present 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 Elston (1997), Ruiz and Hines (1997), WDFW (1998), and Cohen et al.

(1998) determined there were between 31 and 67 non-native species reproducing in Puget Sound. 

Cohen (2004) later re-sampled the sound and identified 76 non-native species. 

Ballast water discharge into Puget Sound has increased between 2005 and 2010 with corrected 

volumes of 4.99x106

, 5.16x106, 4.92x106, 6.12x106, 5.25x106, and 5.8x106 tonnes discharged 

annually from around the globe, respectively. The mean corrected discharge rate over this time 

was 5.37x106

. 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, over the past 

six years, an invasion rate of approximately 0.455 non-native species per year (8.47x10-8 

x 

5.37x106

 x 0.63) would be expected or approximately one new species in Puget Sound every two 

years. And as we did for San Francisco Bay, observing the increasing trend in discharge 

volumes, there is an increase of approximately 161,000 tonnes per year each year beyond 2010, 

increasing the risk of an invasion by approximately 0.014 new non-native species each year. In 

other words, if the rate of 0.455 new species per year over the past six years is correct, given the 

upward trend in ballast discharge provided by the National Ballast Information Clearinghouse, 

the potential for an invasive species to establish 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 that 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 Atlantic 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 present 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 Chesapeake 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 there 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 efforts to improve water quality are allowing nonnative species to capitalize on the nutrient rich waters and greatly disturbed habitat in the bay. 

Ballast water discharge into Chesapeake Bay has increased between 2005 and 2010 with 

corrected volumes of 6.72x106

, 8.0x106, 10.76x106, 16.63x106, 14.45x106 , and 21.06x106 tonnes 

discharged annually from around the globe, respectively. This is the greatest proportional 

increase in ballast water discharge in the US during this time period. Again, if NMFS adopts the 

USCG’s PVA from the final PEIS, we would expect a 37% reduction in the number of new nonnative species under the numerical standard and an 85% reduction 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-native species per year (8.47x10-8 x 12.94x106

 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 tonnes 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 words, if the rate of 1.1 new species per year is 

correct, given the upward trend in ballast discharge provided by the National Ballast Information 

Clearinghouse, the potential for an invasive species to establish in 2012 would be approximately 
319

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 anticipated rate of invasions would fall 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 predict the species that would constitute future 

invasions. 

Smalltooth Sawfish, Johnson’s Sea Grass, Elkhorn and Staghorn Coral 

Smalltooth sawfish, Johnson’s sea grass, and elkhorn 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 between Florida and some Caribbean Islands. Johnson’s sea grass 

is found in isolated bays, inlets, and lagoons along 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 therefore discharges from Mobile Bay and west were considered 

unlikely to affect these species. Despite excluding some major ports, this area is still exposed to 

discharges in Tampa, Miami, and the Florida Keys. 

In the waters of Florida alone, 23 non-indigenous aquatic plant species have become established 

(McCann et al. 1996). Approximately 50 non-indigenous aquatic species have established in the 

state, but not all from ballast water discharge (Courtenay 1997). 

Ballast water discharge into coastal Florida and Caribbean waters has remained steady between 

2006 and 2010 with corrected volumes of 14.43x106

, 15.19x106, 15.52x106, 14.26x106, and 

14.08x106 tonnes discharged annually from around the globe, respectively. If NMFS accepts the 

USCG’s PVA from the final PEIS, we would expect a 37% reduction in the number of new nonnative species under the numerical standard and an 85% reduction under the ballast management 

program’s published goal. Therefore, over the last 5 years, an average invasion rate of 

approximately 0.784 non-native species per year (8.47x10-8 x 14.698x106 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, observing 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 the downward trend in ballast discharge provided 

by the National Ballast Information Clearinghouse, the potential for an invasive species to 

establish in 2012 would be approximately 0.772 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 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 means that the US imports and exports goods to the rest of the 

world. There are three general types of aquatic ecosystems in the US: freshwater, estuarine, and 

saltwater, and all of those systems receive ballast 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 ships are regulated by this FR. Additionally, approximately 6,000 
320

new ships are built each year around the world and the 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 approximately 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 discharges to NMFS resources, NMFS analyzed the total 

coastal and estuarine ballast water discharges in the US over the past six years (excluded the 

Mississippi River Basin and Great Lakes ports). Ballast water discharge into US waters has 

increased between 2005 and 2010 with corrected volumes of 123.6x106

, 155.5x106, 173.0x106, 

192.0x106

, 215.1x106 , and 236.5x106 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 approximately 9.7 non-native species per year (8.47x10-8 x 

182.6x106 x 0.63) would be expected in US waters every year. And as we did for local port case 

studies, observing the increasing trend in discharge volumes, there is an increase of 

approximately 22,582,000 tonnes per year each year 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 upward trend in ballast discharge provided by the 

National Ballast Information Clearinghouse, the potential for an invasive species to establish in 

2012 would be approximately 13.52 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 anticipated rate of invasions would fall to about 3.005 

individuals per year nationwide. 

Response Analysis 

Consequences of exposing listed species and designated critical habitat 

The preceding section of this Opinion presented the evidence that leads us to conclude that 

endangered or threatened species and proposed or designated critical habitat under NMFS 

jurisdiction are likely to co-occur with discharges of ballast water on, over, or near U. S. waters. 

In this section, we summarize information on the probable physical, 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 intention is to identify the range of representative responses we 

would expect proposed or listed species and their critical habitats to exhibit given exposure to 

these stressors. We do provide probability estimates 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 the 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 regions 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 species 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 threatened and endangered freshwater species in the US as a whole, 

but particularly so in the western 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 surrounding regions. Mitten 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 Chesapeake and San Francisco Bays (Leichsenring and 

Lawrence 2011). Similarly, the Asian kelp expanded 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 aquatic environment has been poorly explored and 

we likely know very little about the true frequency 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 tolerances to physical conditions, and ability to 

form resting stages and, as such, can be considered cosmopolitan and not invaders. A similar 

cosmopolitan nature has been suggested for bryozoans, barnacles, polychaete worms, hydroids, 

and molluscsthat reside on marine debris (Barnes 2002). Despite these hypotheses, NMFS 

would expect that if this assertion were true, the occurrence of novel diseases would not occur on 

a geographic basis or, in the present situation, in association with relatively high levels of ballast 

water discharge, which a majority of the available literature contradicts. Furthermore, although a 

pathogen may already exist in waters receiving ballast 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 adverse effects to native fish populations. Viral 

hemorrhagic septicemia virus was initially introduced to North America from western Europe via 

ballast water discharge in the 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 widely in the Northern Hemisphere. 

Pathogens also have adverse effects to invertebrate communities. Molluscs such as black and 

white abalone seem to be particularly sensitive to pathogens. Various species of the genus 

Vibrio, known to cause cholera in humans, white pox and 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 sustained 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 North 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 face severe declines in 

part due to morbidity and mortality 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 species, including listed fish. The Asian parasitic 

copepod, Neoergasilus japonicas, began parasitizing endangered cichlid 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 among the naïve natives (Josefsson and Andersson 

2011). 

Like fish, invertebrates are highly susceptible to introduced parasites and there are a number or 

examples of adverse effects. A parasitic worm, Terebrasabella heterouncinata, attacked red 

abalone in California aquaculture subsequent to the parasites’ introduction, significantly 

reducing growth rate. The pathogen escaped into the marine environment from waste water and 

shells discharged from these facilities (Cohen 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 waters or re-introduce it to other parts of the US. 

The aquaculture impacts likely highlight the potential 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 significant risk to black abalone’s 

survival and recovery is a disease called withering 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 salmon 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, directly compete with native salmon, introduce new diseases to naïve 

populations, and compete for limited resources. The introduction of brook trout into freshwater 

systems of the western US, particularly the Columbia River Basin, has directly resulted in the 

displacement of native trout in these locations through competitive interaction and hybridization 

(Dunham et al. 2002, Feyrer et al. 2003). Mills et al. (2004) used laboratory experiments to 

show competitive interactions between 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 invertebrates being introduced that directly compete with 

native invertebrates, and sometimes listed species. Stat and Gates (2008) found a non-endemic 

symbiotic dinoflagellate of the same genus used by elkhorn and staghorn coral were present in a 

Hawaiian Acroporid coral; genetic analyses suggest the symbiotic dinoflagellate was transported 

to and established in the isolated Northwestern Hawaiian Islands National Monument via ballast 

water. 

Prey 

Ballast water potentially introduces species that become prey for native biota in the region. For 

instance, Hawaiian green sea turtles historically fed upon native algal species, but have recently 

included the non-native algal species Acanthophora spicifera, Hypnea musciformis, and 

Gracilaria salicornia in their diets. The contribution of these three alga to green sea turtle diets 

in the area are now so significant that they compose 40% of the diet among individuals studied 

(Russell and Balazs 2009). Another two invasive algal species comprised an additional 18% of 

the diet (Russell and Balazs 2009). 

Predators 

Ballast water discharges sometimes 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 estimated to constitute 100 million tons of biomass in 

the Black Sea in 1994. This rapid increase in population caused a decline of native species due 

to predation (mostly copepods and cladocerans), the extinction of some native pontellid 

copepods, declines in benthic zooplankton, 70% declines in inshore fish larvae (also prey to the 

invasive species), benthic anoxia, mucus within the water column, and reductions in biodiversity 

as a whole (Vinogradov et al. 1989, Tsikhon-Lukanina and Reznichenko 1991, Zeitsev and 

Ozturk 2001, Moncheva and Kamburska 2002). This species preyed on native larval fish, while 

also effecting several areas within the native food web. If a species having similar ecological 

impacts was introduced to the US, there is the chance larval stages of listed fish could be 

consumed by an invasive predator. 
324

Direct effects to designated critical habitat and indirect effects to listed 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 and lower survival rates 

(r-selection). This unbalancing of food webs makes subsequent introductions more likely as 

resource utilization shifts, increasing resource availability, and exploitation success by nonnative species (Barko and Smart 1981, Byrnes et al. 2007). Such shifts in the base of food webs 

fundamentally alters predator-prey dynamics up and across food chains (Moncheva and 

Kamburska 2002). The number of extinction events seems to be roughly correlated with the 

number of invasive establishments within 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 instrumental in the environments in which 

they invade, changing metabolic and trophic interactions of predator-prey relationships as well as 

altering physical and chemical processes within an environment (Colwell 1996, Castilla et al.

2004, Hecky et al. 2004, Wallentinus and Nyberg 2007). However, while a large majority of 

available information indicates that invasive species alter the ecosystems 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 expansion of other invasive species (Daehler and 

Strong 1996). Grosholz and Ruiz (1996) studied the impacts of green crabs at three different 

sites, finding that ecological impacts were similar at all three sites, although habitat use and 

expansion rate exhibited considerable variation. 

Pathogens, Parasites, and Toxins 

Primarily, pathogens and species with toxic effects have direct effects to listed species, however 

there are instances when these species 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 critical habitat PCEs or indirectly affect listed 

species. Pathogens are in some cases very specific to hosts, but when a species similar to a listed 

species is introduced, eventually that parasite that specific to the non-native species can shift to 

also affecting similar native populations. In these cases, the effects may be directly adverse to 

listed species or indirect to food resources as identified in a species’ critical habitat. And in other 

cases, parasites can have direct effects to PCEs of designated critical habitat or indirectly affect 

listed species. 

Toxins. Red tide dinoflagellates have been introduced via ballast water discharges and have the 

potential to undergo extreme seasonal population fluctuations. During bloom conditions, high 
325

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 paralytic shellfish poisoning and harmful algal 

blooms in the region (Moore et al. 2011). Another species potentially introduced by ballast 

water, Pseudonitzschia australis, resulted in bird mortality 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 availability 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 Pacific Coast where abalone of the same genera as 

listed white and black abalone resulting in 50% mortality or higher (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 the abundance of native oysters (Chew 1990, Ciguarria and Elston 

1997, Minchin and Rosenthal 2002). 

The introduction of American signal 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 disease 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 discharge (Christison and 

Baker 2007). A commonly-occurring nematode parasite of Asian eels infected 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 predatory effects caused by invasive species. 

European green crabs have invaded both the east and west Coasts of the US, having trophic scale 
326

effects to both environments. In Massachusetts Bay, green crabs prey upon native mussels and 

oysters, altering community structure (Lafferty 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 least 20 invertebrate species (Grosholz and 

Ruiz 1995), with some species showing five to ten fold declines in abundance within the first 

three years of the invader’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 increase 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 endemic gelatinous zooplankton 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 the native and also by 

facilitating predation by fishes, while the shrimp itself preyed upon the amphipod heavily in 

areas where it had become dense (Bailey et al. 2006). Decapods as a whole reduce the presence 

of macrophytes, snails, and other benthic animals (Strayer 2010). Also, P. macrodacylus may 

act as a new prey for bethivore generalists, 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 capable invaders and have proven 

to be disruptive to food webs. The most common 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 increase bioturbation and aerate the sediment 

(Nyberg 2007). Gastropods can alter the biogeochemical cycle through excretion of biogenic 

silicate in the faeces and pseudofaeces (Ragueneau et al. 2005). Molluscan invasions can also 

provide substrate for epibionts, shelter for benthic 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 jurisdiction or directly affecting the habitat they 

rely on. The compound tunicate, Botrylloides sandiegensis, was released near Woods Hole, 

Massachussetts and has outcompeted other encrusting organism in the coastal environment of 

southern New England (Lafferty and Kuris 1996). The invasive snail Batillaria attramentaria

outcompetes a native species, Cerithidea californica, along northern California via a more 

efficient ability to convert food into body tissue and possibly greater dispersal 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 provide an opportunity to study the 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 amount 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 alter the nutrient composition 

and phytoplankton availability in the habitats they 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 reefs (Nehring 2006b). A mollusk of the genus 

Rapanus has been blamed for the decreased abundance 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 forming 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 benefitted from this altered habitat (Crooks 1998, Grosholz 

2002). However, larger bivalves and eelgrass appear to be inhibited by the mat covering the mud 

flats (Crooks 1998). In addition to altering habitat and shifting species distributions and 

abundances, this Asian mussel also caused increased sediment and organic accumulations 

converting sand flats into mud flats (Crooks 1998). 

Maybe the most well known example of a ballast-mediated invasive 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 freshwater systems in the eastern US (Benson 1995, 

Strayer 1999, Pimentel et al. 2004). Shortly after zebra mussels were introduced, their close 

relative the quagga mussel was introduced via ballast water discharge 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 other 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 Detroit 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 similar, 

although not as dramatic findings in reduced diversity of native mussel species. Michigan 

Technological University (2010) reported that zebra mussel numbers 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 (Benson 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, after having zebra mussels grow on them, fail to 

open and close their shells and thus are unable 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 nutrition 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, which is a primary component of many predatory 

salmonid diets including the economically 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 lower 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 effectively filter feed is increased water clarity, altered 

benthic and water chemistry, as well as blue-green 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 possible within the ecosystem 

(Vanderploeg et al. 2001, Ciborowski 2007). These impacts 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 individuals 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 filtering food from the water 

column, 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 adversely 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 and 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 displaced 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 species of coralline algae, serpulid 

polychaetes, bryozoans, and hydroids to exclusively occupy large patches of intertidal and 

subtidal rocky habitat along New Zealand ((Carlton 1999). 

Vegetation. Invasive plants can cause widespread habitat alteration, including native plant 

displacement, changes in benthic and pelagic animal communities, altered sediment deposition, 

altered sediment characteristics, and shifts in chemical processes such as nutrient cycling (Grout

et al. 1997, Wigand et al. 1997, Ruiz et al. 1999). Introduced seaweeds alter habitat by 

colonizing previously unvegetated areas, while algae form extensive mats that exclude most 

native taxa, dramatically reducing habitat complexity and the ecosystem services provided by it 

(Wallentinus and Nyberg 2007). Invasive algae can alter native habitats through a variety of 

impacts, including trapping sediment, reducing the number of suspended particles that reach the 

benthos for benthic suspension and deposit feeders, 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 exclude other algal species during 

some seasons and alter the diversity of other species present, including 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 impair 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 enhance local biodiversity (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). Particularly significant for juvenile salmon is that 

some of their invertebrate prey are adapted to take advantage of seasonal spikes in available 

detritus, which invasives like purple loosestrife can 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 impactful 

to habitats it invades due to its ability to reduce the overall biodiversity 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%, increased sedimentation, and 

changed diurnal dissolved oxygen and pH cycles on the surfaces it established on. Some species 

may benefit from this, as it creates cover for several filter-feeding 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 and 
330

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 species in the San Juan Islands can be outcompeted 

(reductions of 50-75%) and green sea urchin abundance can decline significantly with the loss of 

this primary food, potentially destabilizing kelp ecosystems that entire biological communities 

are dependent upon (Britton-Simmons 2004) (Ambrose & Nelson 1982). Invasive freshwater 

plants in Sweden have reduced available light, 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 shade 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 negatively) dissolved oxygen levels. These changes can affect other 

aspects of the ecosystem including decreased biodiversity, impeding the recovery of endangered 

waterbirds, and disproportionally providing habitat for invasive species over native taxa (Allen 

1998). 

Other invasive plant species can be beneficial for native taxa 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 disturbances, improve water quality, and transform 

beaches into marshes (Posey 1988, Allen 1998, Ronnback 1999, Bruno 2000, Bruno and 

Kennedy 2000, Gribsholt and Kristensen 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 (Carpenter 1987). In Britain, S. anglica has invaded eelgrass habitat, 

reducing the abundance of eelgrass and thus food availability for 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 plants (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 available habitat for benthic invertebrates (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 environments (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 column 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 flows can be decreased in these areas (Ranwell 1964, 

1967, Christie 1974, Carpenter 1987, Daehler and Strong 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 example, 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 introduction 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 water 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 resources 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 predator in some invasive 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 nutrient 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 ballast water discharge and has become a major 

predator of native benthic invertebrates and competes for available prey and habitat with native 

taxa, such as yellow perch and preys upon the 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 habitat 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 affected by human alterations, as identified in the baseline section, 

such as hydromodification, mining, dredging, drilling, and construction. However, invasive 

species, facilitated by human commerce, have the ability to directly 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). This invader, first introduced by 

ballast water from northern Europe to Lake Ontario in the early 1980s may eventually occupy 

thousands of lakes throughout North America (Johannsson et al. 1991, Grout et al. 1997, Kerfoot
332

et al. 2011). Bythotrephes is an important contributor to its native habitat, including as prey to 

salmon; however, in the Great Lakes, they reduce 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 caused 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 resources 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 additional contaminants to accumulate in the 

underlying prey base (Kerfoot et al. 2011, Rennie et al. 2011). Other macroinvertebrate 

predators and fishes are also likely adversely impacted by this disruption 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 appear to be stable and persistent (Yan et al. 2008). Through 

these mechanisms, Bythotrephes alone represents a significant threat to the biodiversity within 

temperate North American aquatic environments (Grout et al. 1997). 

Dreissena mussels introduced to the Baltic Sea altered 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 transfer them to the sediment in their excrement. 

In many cases, this decreases the nutrient outflow from some embayments and causes an 

accumulation of organic material, leading to increased 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 (Wallentinus 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 extensively used by native invertebrates and having 

double the species richness in these habitats than surrounding unaltered habitat (Castilla et al.

2004, Wallentinus and Nyberg 2007). 

Other invertebrates can also have major impacts on the ecosystems they invade. The invasive 

bryozoan Membranipora membranacea established on kelp fronds in the Gulf of Maine, making 

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 biomass 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 denitrification process causing changes to the native habitat (Panov et 

al. 2003, Wallentinus and Nyberg 2007). These changes have adversely affected the prey 

species for native predators. 
333

Asian crabs (Hemigraspus sanguineus) compete for space (burrows) with native fiddler crabs 

along Connecticut (Wallentinus and 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 Pacific Northwest after establishing there via ballast 

water discharge (Kipp et al. 2010). Oriental shrimp (Palaemon macrodactylus) were likely 

introduced into New York State waters via ballast 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 significant 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 the sedimentation process; because individuals 

cumulatively engage in so much grazing, some bottom habitats have become dominated by hardbottom instead of soft bottom as they formerly were (Bertness 1984, Wallentinus and Nyberg 

2007)(Carlton 1999). Significant declines in soft-sediment habitats and fringing salt marshes are 

attributed at least partially to the invasion of this species, possibly 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 introductions 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 least 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 fish 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 local clapper rails by trapping the bird’s beak in its 

shell if the rail inserts it into the open shell; rails cannot pull their beaks out and die of starvation 

(Cohen and Carleton 1995). Adult rails also lose toes by the mussel closing on them as the rails 

walk through the marsh, although the result for 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 significant 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 green sunfish are all introduced species to the 

area, although largely through means other than ballast water discharge (Cohen and Carleton 
334

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 either directly through 

predation or indirectly through competition for prey and/or breeding 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 declines 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 fish 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 delta 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 influential within the bay system, occurring in 

such high abundance so as to transfer the phytoplankton biomass in the water column to the 

benthos that is then exploited by benthic invertebrates and bottom-feeding fishes, instead of 

pelagic species (Alpine and Cloern 1992, Cohen and Carleton 1995, Grosholz 2002). Mya spp. 

alone is estimated to be capable of filtering 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 capacities (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 (Cohen 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 fish species. Alpine and Cloern (1992) 

provided an even larger estimated decline from 106 g C/m2

 to 39 g C/m2. 

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 species including the Asiatic clam, polyzoan, 

and hydroid (Corbicula, Urnatella, and Cordylophora spp., respectively) have led to increased 

sedimentation in canals, reducing flow rates and 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 and indirect effects to listed species described 

above impact the ecosystem to some extent. Additionally, there are species that may not impact 

other species in the area, but they 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 introduction, this small animal has altered mud banks 
335

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 over a century ago (Cohen and Carleton 1995). 

Several invasive plants also impact San 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 more protective habitat from predators, although 

competition between the native and invasive cordgrass 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 rates, reduce light levels by its 

increased canopy density, and loss of shorebird feeding habitat through colonization 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 introduced species in the Chesapeake Bay. A 

comprehensive review of the impacts of invasive 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 invasive species had not been studied 

for their impacts. Of the 39 species with significant ecological impacts, 69% did so through 

competition with native species, 38% altered habitat, 44% served as prey, 15% were predators of 

native species, 21% engaged in extensive herbivory, 8% produced hybrids with native taxa, and 

8% were parasitic (Ruiz et al. 1999). Plants and fish were the largest taxonomic 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 studied, it appears the best documentation 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 marinus, 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 altered 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 also 
336

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 environment 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, subsequently increasing aquatic vegetation 50%, and 

ultimately increasing abundance of fish populations, 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 listed species also generally have 

impacts to habitat. For instance, the description of indirect effects caused by the Asiatic clam 

also have impacts to habitat in the Chesapeake Bay. The reed, Phragmites auatralis, also 

outcompetes local plants and has 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 species, 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 not recovered from an eradication program in the 

1930s (Ruiz et al. 1999). 

As has been documented in other regions, there appears to be a correlation between habitat 

disturbance and the susceptibility to invasions. Jewett et al. (2005) experimentally found that 

low dissolved oxygen levels (common in the Chesapeake 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 situation tends to lead to 

low light levels, which make some invasive species, such as the 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 information documenting biological, physical, chemical, 

or ecological responses of species or habitats 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 available supporting direct effects 

that ballast-mediated invasive species have on species listed under NMFS’ jurisdiction. The only 

direct connection we found supported a beneficial connection of invasive algae providing a 

significant new food source for Hawaiian green sea turtles (Russell and Balazs 2009). By 

assessing the adverse effects of disease, 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 parasites can have significant 

and catastrophic effects, such as that which has occurred to Atlantic salmon in Norway as a 

result of an invasive fluke being introduced, white pox disease in Acroporid corals, withering 

syndrome in California abalone, and harmful algal 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 mussels, silverside minnows, and numerous 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 beneficial to harm of an unpredictable number of 

listed individuals. 

Competition is another means by which invasive species are likely to directly affect listed 

species. Particularly susceptible to direct effects of competition are the listed invertebrates 

(abalone and corals) because filter 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 ballast-mediated invasive species establish in an 

area occupied in the same time and place as listed 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 destroy or adversely modify critical habitat for ESAlisted species by examining any change 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 habitat. Instead, this analysis focuses on statutory provisions of the 

ESA, including those in Section 3 that define “critical habitat” and “conservation,” those in 

Section 4 that describe the designation process, and those in Section 7 setting forth the 

substantive protections and procedural aspects of consultation. 

The Response analysis documents indirect effects to listed species via habitat and ecosystem 

alteration as the major mechanism by which ballast-mediated invasive species can adversely 

affect listed species. While this is not the result of every invasion, it is a probable outcome over 

time. In many cases, this also constitutes adverse affects to critical habitat (specifically of 

eulachon, salmonids, corals, abalone, Cook Inlet beluga 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 have been designated, are described in the Status of 

Listed Resources section of this Opinion. The PCEs identify those physical or biological 

features that are essential to the conservation of the species that may require special management 

considerations or protections. The species addressed in this Opinion have similar life history 

characteristics and therefore, many of the same PCEs. 

The PCEs for gulf and green sturgeon focus on life history stages, rearing, 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, and adult life stages; 

2. Water flow necessary for normal behavior, growth, and survival of all life stages; 

3. Water quality, including temperature, salinity, 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 meters deep for both upstream and downstream 

holding of adult or sub-adult fish, with adequate 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) necessary for normal behavior, growth, 

and viability of all life stages. 

These PCEs for salmonids and Pacific eulachon include sites essential to support one or more 

life stages (sites for spawning, rearing, migration and foraging) and contain physical or 

biological features essential to the conservation of the ESU/DPS, such as: 

1. Freshwater spawning sites with water quantity and quality conditions and substrate 

supporting spawning, incubation and larval development; 
339

2. Freshwater rearing sites with water quantity and floodplain connectivity to form and 

maintain physical habitat conditions and support juvenile growth and mobility; water 

quality and forage supporting juvenile development; 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 obstruction, along with water quantity and quality 

conditions and natural cover such as submerged 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 physiological 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 fishes, supporting growth and maturation; 

5. Nearshore marine areas free of obstruction with water quality and quantity conditions and 

forage, including aquatic invertebrates and fishes, supporting growth and maturation; and 

natural cover such as submerged and overhanging large wood, 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, supporting growth and maturation. 

These PCEs for southern resident killer whales include sites essential to support one or more life 

stages and contain physical or biological features essential to the conservation of the DPS, such 

as: 

1. Water quality to support growth and development; 

2. Prey species of sufficient quantity, quality, and availability to support individual growth, 

reproduction, and development, as well as overall population growth; and 

3. Passage conditions to allow for migration, resting, and foraging. 

The PCE for leatherback sea turtles essential to support population fitness and the conservation 

of the DPS, is the occurrence of prey species, primarily scyphomedusae of the order 

Semaeostomeae (Chrysaora, Aurelia, Phacellophora, and Cyanea), of sufficient condition, 

distribution, diversity, abundance and density necessary to support 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 species of Pacific salmon (Chinook, sockeye, 

chum, and coho), Pacific eulachon, Pacific cod, walleye pollock, saffron cod, and 

yellowfin sole; 

3. Waters free of toxins or other agents of a type and amount harmful to Cook Inlet beluga 

whales; 

4. Unrestricted passage within or between the critical habitat areas; and 
340

5. Waters with in-water noise below levels resulting in the abandonment of critical habitat 

areas by Cook Inlet beluga whales. 

These PCEs for Hawaiian monk seal are essential to ensuring the health, well-being, and 

continued viability of their populations: 

1. Pupping and major hauling beaches including 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 Hawaiian 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 new habitat. Benthic invertebrates, such as 

mussels, polychaetes, and hydroids become dominant filter feeders, greatly reducing the amount 

of organic energy that is available to native taxa in the water column. This transfer of energy 

from the water column into the benthos fundamentally alters the ecology of the host habitat, 

resulting in less food/prey available for other filter feeders. At the same time, this may increase 

resource availability for sturgeon species. Adverse effects of this include reduced body 

condition, growth, survival, and/or reproduction of native pelagic 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 greater 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 benthic invertebrates also cause 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 explosions of bacteria that exploit 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 through shallow water habitats and 

juvenile salmonid rearing areas, leading to increased 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 are susceptible to introduced parasites and 

diseases, sometimes with devastating levels of mortality 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, survival, and reproductive resources 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 introduced dinoflagellates can not only produce 

toxins that adversely affect listed individuals directly, but also produce massive die-offs of fish 

and invertebrate species resulting in less available prey. Estuarine plants, such as Spartina spp. 

and mangroves can rapidly alter the physical environment of an estuary, reducing wave energy, 
341

increasing sedimentation, and leading to estuaries filling 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 sedimentation or particle movement into the water column, 

reducing habitat complexity and, ultimately, the species that can survive within a habitat 

(potentially prey species of listed organisms or the listed species themselves). 

Future Invasive Species Impacts to Listed Species and Critical Habitat 

We have established that, regardless of the mechanism by which invasive species have been 

introduced, they can have a wide range of effects on the environments in which they establish. 

More importantly, ballast-mediated invasive species have a similar affect on the environment 

and the species within the environment as do all invasive 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 beneficial, providing food for listed sea turtles (Russell and Balazs 

2009), increasing biodiversity (Nyberg 2007), and establishing new habitat used by native 

species (Paulo da Cunha Lana 1991, Daehler 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 catastrophic 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 fundamentally 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 species 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 invasive species establishes, it is 

not appropriate to assume there will be an adverse effect to the receiving environment or the 

listed resources within it. For this reason, we determined what the 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 sources of information to evaluate this question: 

(1) evaluations of the impacts of invasive species 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 impacts of invasive species in San Francisco Bay (McNickle et al. 2001). 

While none of these studies identified only ballast-mediated invasive species, the Response 

Analysis in this Opinion shows that ballast-mediated invasive species elicit similar responses as 

other invasive species. Therefore, NMFS expects that any invasive species that are introduced as 
342

a consequence of the USCG’s Ballast Water Management Program will have impacts similar to 

other introduced species with effects ranging from beneficial to harmful. 

US Technology Office of Technology and Assessment. This report (Anttila et al. 1998) 

summarized information on non-indigenous species (primarily terrestrial and freshwater) in the 

United States, as well as what proportion of them were ecologically or economically harmful, 

beneficial, neutral, or unknown. This non-exhaustive but reasonably comprehensive report used 

a panel of experts to identify non-indigenous species and the impacts 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, rangelands, or forests; seriously 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 these categories at least 

represent species that are competitors, predators, diseases, or habitat modifiers of native species 

or habitats and would be impactful to the environments into which they were introduced. 

Overall, 16% were considered harmful, 33% had unknown effects, 26% were considered 

beneficial, and 25% had both beneficial and harmful effects. The survey included purposefully 

introduced species of agriculture, hunting, and commercial use, which should upwardly bias the 

data in the beneficial and beneficial and harmful categories. This is supported by one of the 

report’s findings that only 12% of intentional introductions were considered harmful versus 44% 

of unintentional introductions. Because this assessment includes terrestrial invasions and 

intentional introductions, NMFS believes the probability 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 conducted 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 information existed to support a strong possibility of 

being invasive were included in these 196 while hundreds of others were excluded. The 

definition of whether a species was impactful was based exclusively upon available literature 

that qualitatively or quantitatively ascribed a particular species as being impactful. Additional 

information on impact for some of these species 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 the Bay itself. Impacts were assigned to nine 

categories: competition, habitat change, food-prey, predation, herbivory, hybridization, 

parasitism, toxicity, or bioturbation which caused “significant and measurable changes in the 

abundance or distribution of resident species or habitats.” This definition, unfortunately, 

excluded some effects that we would consider significant, such as hybridization or introduced 

predators for which information was lacking on population-level effects to native prey; this 

likely causes a slight negative bias in the number of invasives we would consider having an 

adverse affect. We also expect that the total number of invasives is actually higher than the 196 

included, as an unknown subset of those that were 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 the 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 impact but significant information was lacking. 

Cryptogenic species represented six (15%) of impactful invasives. Based upon this, cryptogenic 

species compose a relatively small proportion of impactful 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 their impacts. The ecological categories in which 

impacts occurred were: competition—69%; habitat change—38%; food/prey—44%; predation—

15%; herbivory—21%; hybridization—8%; parasitism—8%; toxicity—0%; and bioturbation—

3%. These proportions of impact 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 Francisco Bay/Delta and the number of these that we found in our 

information searches that we found to impact the area. The number of invasives, was adopted 

from Cohen and Carlton (1998). These authors assembled data from the 16th century to the time 

of publication from published, unpublished, and their personal research on the introduced aquatic 

organisms of San Francisco Bay/Delta, including some cryptogenic species. The authors 

identified 234 exotic species as well as 125 cryptogenic species. We expect that some of the 

cryptogenic species identified by Cohen and Carlton (1998) are actually 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 five were considered exotic 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 exotic 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 representative of the number of invasive species 

present. Furthermore, this assumption would lead to an overestimation of the proportion of 

invasive species adversely impacting San Francisco Bay/Delta. 

The next step is to identify the number of invasive 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 species monitoring programs. Based upon our nonexhaustive review in the San Francisco Bay case study, we identified 21 invasive species that 

have adversely affected native species and/or habitats in San Francisco Bay/Delta. We identified 

another 22 species negatively impacting this area from Cohen and Moyle (2004), producing a 

total of 43 adversely affecting invasive species out of 234 (18.4%). 

Calculated Rate of Harm. The Congressional study found 15.3% of terrestrial and aquatic 

invasive species were harmful, while the Chesapeake 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 the data in some direction depending 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 mean 
344

probability of harm (19.9%) identified 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. This method reflects the most probable rate of 

invasions likely to cause harm to listed species and their critical habitat. Although each analysis 

has specific caveats that make each more or less valuable 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 that could be affected 

by an invasive population of organisms. In the Effects Analysis, NMFS further identified the 

probable rate of exposure for these 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 and generally occur throughout the year. Although listed species 

would 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 have indirect effects to 

habitat (such as alteration of physical conditions) and that impact 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 probability of an establishment. 

Through the Response Analysis, we found that likely responses of listed species to future 

invasions would be expected to be similar to past invasions, which can be diverse (including 

competition for prey/nutrition and space, predation, disease, genetic hybridization, and habitat 

alteration) and not necessarily negative (such 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 several 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 rates being between 18.8% and 

19.9% in the two aquatic analyses we conducted, NMFS 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 species would have. By using the mean rate of 

harmful impacts, NMFS can assess the adverse impacts of this action in the most probable way. 

However, because listed species and critical habitats often occur over multiple bioregions, to 

determine the effects to listed species and critical habitat for the conclusions of this Opinion, we 

must summarize the information from this section in the next section. While the probability of 

an invasive species establishing in a specific bioregion may be negligible, the cumulative effect 

of those probabilities to listed species and critical habitats may be significant. 

Furthermore, in order to determine the probability of an invasive species establishing through 

time, NMFS estimated when these numerical standards will be required based on the history of 

ballast water regulations. NANPCA was initially passed in 1990. After six years, that law was 

updated and NISA was passed. After eight years of working under the initial framework 

established by NISA, the USCG issued a final rule 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 numerical standard. Given the history of regulation change, 

it is not unreasonable to anticipate another regulation change in seven to eight years. Although 
345

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 risk of underestimating exposure at the proposed 

numeric standard if we limited our review to eight years. However, given the history of this 

program, NMFS anticipates within a duration of two historic cycles (16 years) there will be 

program differences not reflected in the assumptions of this Opinion (Table 42). We therefore 

do not consider effects beyond this timeframe. We are aware that effects of invaders now and 

within the next decade and a half will extend beyond this period, but have no information 

available to use to determine what, where, when, and how those effects will impact listed 

resources and their critical habitats. As such, any analysis of these 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 establishing invasive species have the potential to impact southern 

DPS spotted seal. 

Between 2005 and 2010 in NA-S1, the corrected average ballast water discharge volume was 

approximately 32,000 tonnes. The reported volume was corrected to account for incomplete 

reporting. If this five year discharge rate is similar for the next 16 years, NMFS would expect 

the probability for a ballast-mediated invasion in this bioregion 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 probability 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 bioregion, 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 8.54x106

 tonnes. The reported volume was corrected to account for incomplete 

reporting. If this five year discharge rate is similar for the next 16 years, NMFS would expect 

the probability for an invasion in this bioregion to be approximately 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 invasion 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 year 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 bioregion, newly establishing invasive species have 

the potential to impact Cook Inlet DPS beluga whale, southern DPS spotted seal, and western 

DPS Steller sea lions and their critical habitat. 

Between 2005 and 2010 in NEP-II, the corrected average ballast water discharge volume was 

approximately 603,000 tonnes. The reported volume was corrected to account for incomplete 

reporting. If this five year discharge rate is similar for the next 16 years, NMFS would expect 

the probability for an invasion in this bioregion to be approximately 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 invasion 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 year 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 Puget Sound. Within this bioregion, new non-native 

species have the potential to impact SONCC coho salmon and their critical habitat, lower 

Columbia River coho salmon, Oregon coast coho salmon and their critical habitat, Pacific 

eulachon and their critical habitat, lower Columbia River Chinook salmon and their critical 

habitat, upper Columbia River spring-run Chinook salmon and their critical habitat, Snake River 

fall run Chinook salmon and their critical habitat, Snake River spring/summer run Chinook 

salmon and their critical habitat, upper Willamette 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 green sturgeon and their critical habitat, Snake 

River sockeye salmon and their critical habitat, Ozette Lake sockeye salmon and their critical 

habitat, Puget Sound steelhead, Snake River basin steelhead and their critical habitat, upper 

Columbia River steelhead and their critical habitat, middle Columbia River steelhead and their 

critical habitat, lower Columbia River steelhead and their critical habitat, upper Willamette River 

steelhead and their critical habitat, 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 Steller 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 1.81x106 tonnes. The reported volume was corrected to account for incomplete 

reporting. If this five year discharge rate is similar for the next 16 years, NMFS would expect 

the probability for an invasion in this bioregion to be approximately 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 invasion 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 year but 0.2607 over 16 years. 

Ballast Water Bioregion NEP-V. This bioregion covers northern 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, central valley spring-run Chinook salmon and their 

critical habitat, Sacramento River winter-run Chinook salmon and their critical habitat, central 

California coast coho salmon and their critical habitat, southern Oregon/northern California coast 

(SONCC) coho salmon and their critical habitat, green sturgeon and their critical habitat, Pacific 

eulachon and their critical habitat, central California coast steelhead and their critical habitat, 

northern California steelhead and their critical habitat, south-central California coast steelhead 

and their critical habitat, California central valley steelhead and their critical habitat, leatherback 

sea turtles and their critical habitat, north Pacific Ocean DPS loggerhead sea turtles, eastern DPS 

Steller sea lions and their critical habitat, and humpback whales. 

Between 2005 and 2010 in NEP-V, the corrected average ballast water discharge volume was 

approximately 6.54x106

 tonnes. The reported volume was corrected to account for incomplete 

reporting. If this five year discharge rate is similar for the next 16 years, NMFS would expect 

the probability for an invasion in this bioregion to be approximately 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 invasion 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 year but 0.6731 over 16 years. 

Ballast Water Bioregion NEP-VI. 

Bioregion NEP-VI includes all ports from southern 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 critical habitat, eastern DPS Steller sea lions and 

their critical habitat, green sea turtles, leatherback 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 8.89x106

 tonnes. The reported volume was corrected to account for incomplete 

reporting. If this five year discharge rate is similar for the next 16 years, NMFS would expect 

the probability for an invasion in this bioregion to be approximately 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 invasion 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 year but 0.7858 over 16 years. 

Ballast Water Bioregion CAR-I. Bioregion CAR-I includes all US ports in the entire Gulf of 

Mexico and around the southern tip of Florida. Within this bioregion, newly establishing 

invasive species have the potential to impact gulf sturgeon and their critical habitat, green sea 

turtle, hawksbill sea turtle, Kemp’s ridley sea turtle, leatherback sea turtles, Northwest Atlantic 

Ocean DPS loggerhead sea turtles, elkhorn coral and their critical habitat, staghorn coral and 

their critical habitat, and Johnson’s sea grass and their critical habitat. 

Between 2005 and 2010 in CAR-I, the corrected average ballast water discharge volume was 

approximately 94.01x106

 tonnes. The reported volume was corrected to account for incomplete 

reporting. If this five year discharge rate is similar for the next 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 invasion 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 but 1.0 over 16 years. 

Ballast Water Bioregion CAR-IV. CAR-IV includes all ports around the US Virgin Islands 

and Puerto Rico. Within this bioregion, newly establishing invasive species have the potential to 

impact green turtles and critical habitat, Hawksbill turtle and their critical habitat, leatherback 

sea turtles and their critical habitat, Northwest Atlantic Ocean DPS loggerhead sea turtles, olive 

ridley sea turtles, smalltooth sawfish, elkhorn coral and their 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 10.19x106 tonnes. The reported volume was corrected to account for incomplete 

reporting. If this five year discharge rate is similar for the next 16 years, NMFS would expect 

the probability for an invasion in this bioregion to be approximately 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 invasion 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 year but 0.8314 over 16 years. 

Ballast Water Bioregion NA-ET2. NA-ET2 includes all ports from Maine to Massechusetts. 

Within this bioregion, newly establishing invasive species have the potential to impact Gulf of 

Maine, Atlantic salmon, Gulf of Maine DPS of Atlantic 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 sturgeon, Kemp’s ridley sea turtle, 

leatherback sea turtles, and Northwest Atlantic Ocean DPS loggerhead sea turtles. 

Between 2005 and 2010 in NA-ET2, the corrected average ballast water discharge volume was 

approximately 579,000 tonnes. The reported volume was corrected to account for incomplete 

reporting. If this five year discharge rate is similar for the next 16 years, NMFS would expect 

the probability for an invasion in this bioregion to be approximately 0.0309 per year under the 
349

numerical standard. Assuming a 16 year timeframe before another rulemaking introduces a new 

management plan, the probability of an invasion 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 year but 0.0918 over 16 years. 

Ballast Water Bioregion NA-ET3. NA-ET3 includes all ports south of Massechusetts to the 

southern edge of the Chesapeake Bay. Within this bioregion, newly establishing invasive 

species have the potential to impact Gulf of Maine DPS of Atlantic 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 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 average ballast water discharge volume was 

approximately 25.56x106

 tonnes. The reported volume was corrected to account for incomplete 

reporting. If this five year discharge rate is similar for the next 16 years, NMFS would expect 

approximately 1.364 new ballast-mediated invasive 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 invasion 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 year but 0.9926 over 16 years. 

Ballast Water Bioregion CAR-VII. Bioregion CAR-VII includes all ports between northern 

Florida and North Carolina. Within this bioregion, newly establishing invasive species have the 

potential to impact Gulf of Maine DPS of Atlantic 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 sturgeon, green sea turtles, hawksbill sea turtles, 

Kemp’s ridley sea turtle, leatherback sea turtles, Northwest Atlantic Ocean DPS loggerhead sea 

turtles, 

Between 2005 and 2010 in CAR-VII, the corrected average ballast water discharge volume was 

approximately 3.65x106

 tonnes. The reported volume was corrected to account for incomplete 

reporting. If this five year discharge rate is similar for the next 16 years, NMFS would expect 

the probability for an invasion in this bioregion to be approximately 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 invasion 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 year but 0.4593 over 16 years. 

Ballast Water Bioregion SP-IX. Bioregion SP-IX includes all ports of American Somoa. 

Within this bioregion, newly establishing invasive species have the potential 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-IX, the corrected average ballast water discharge volume was 

approximately 5,679 tonnes. The reported volume was corrected to account for incomplete 
350

reporting. If this five year discharge rate is similar for the next 16 years, NMFS would expect 

the probability for an invasion in this bioregion to be approximately 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 invasion 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 year but 0.0007 over 16 years. 

Ballast Water Bioregion SP-XII. Bioregion SP-XII includes all the ports of Guam. Within this 

bioregion, newly establishing invasive species have the potential 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 average ballast water discharge volume was 

approximately 62,433 tonnes. The reported volume was corrected to account for incomplete 

reporting. If this five year discharge rate is similar for the next 16 years, NMFS would expect 

the probability for an invasion in this bioregion to be approximately 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 invasion 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 year but 0.0096 over 16 years. 

Ballast Water Bioregion SP-XXI. Bioregion SP-XXI is an isolated 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 turtles, and humpback whales. 

Between 2005 and 2010 in SP-XXI, the corrected average ballast water discharge volume was 

approximately 507,000 tonnes. The reported volume was corrected to account for incomplete 

reporting. If this five year discharge rate is similar for the next 16 years, NMFS would expect 

the probability for an invasion in this bioregion to be approximately 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 invasion 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 year 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 182.6x106

 tonnes. The 

reported volume was corrected to account for incomplete reporting. If this five year discharge 

rate is similar for the next 16 years, NMFS would expect approximately 9.7 new ballastmediated invasive species per year under the numerical standard. Therefore, new invasive 

species are still expected to establish 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 has been functioning for over 20 years. While this is the first 

time the USCG has chosen to have NMFS and USFWS consult on the potential impacts of its 

program to listed species and their 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 ballast water treatment technique. However, there 

were several conditions that would allow vessels to discharge ballast water at a port without 

having conducted BWE. When the USCG evaluated the voluntary program in a 2002 report to 

Congress, the report concluded that low participation 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 published 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 broadened 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 keeping 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, including 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 operate the treatment device, and inspect the treatment 

equipment to determine if it is being maintained. This cursory inspection leads to more thorough 
352

inspections if it is apparent the equipment is not being used. Vessels that have been cited for 

violations previously will receive 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 the vessels. The USCG is currently in the midst 

of a prosecution for presenting false information in the ballast 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 monitoring program in that every vessel 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 instances of criminal actions 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 requesting a penalty of 

$19,000. Finally, USCG inspectors issued 268 deficiencies for oil water separator equipment in 

calendar year 2011 with the large majority of those being rectified during the course of the 

inspection. The oil water separator program accounted for a small fraction (0.70%) of the total 

vessel related deficiencies identified on vessels 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 consultation with NMFS, which will result in NMFS 

determining the effect of the program on listed species 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 listed 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 authorized activity under the USCG’s ballast water 

management program that are likely to have adverse effects to listed species and their critical 

habitat without any further mitigation being available 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 discharge. If no ballast water 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 invasive species from entering US 

waters; however they are tasked with doing so practicably. Therefore, the USCG is required to 

consider economic and environmental impacts of any method to prevent an invasion. It is not 
353

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 appears the action being proposed by the USCG may 

adversely affect listed species and their critical habitat, the actual questions NMFS must address 

are whether their proposal is the 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 number of small actions with relatively minor 

effects that all fall under a single national level 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 without determining whether those small projects 

could cumulately have large impacts. For instance, if the USCG requested consultation with 

NMFS on each individual vessel’s ballast water management, while cumbersome under the ESA, 

the cumulative effects analysis should identify the impacts of all of those smaller actions, which 

may appear to have an individually negligible impact to listed species or their critical habitat. 

Therefore, the analytical process of assessing the function of a program is different than 

analyzing the effects of an individual or collection of projects and in many ways, requires more 

of a long-term approach. Analyzing the USCG’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 their 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 species or adversely modify their 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 from the NISA, they are routinely assessing the impacts of 

invasive species. To assess invasive species impacts, the USCG gathers information on ballast 

water through reporting, record keeping, and monitoring. In addition to program monitoring 

activities, the USCG also reviews invasive species literature and continues, as mandated by 

NISA, to review ballast water treatment performance to reduce individual and cumulative 

impacts from ballast water discharge. While the proposed action is the best numerical standard 

that is technologically achievable at this time, the USCG has implemented practicability and 

periodic reviews with the goal of identifying technology that can practicably achieve a more 

stringent ballast water standard in order to further 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 USCG 
354

has proposed the most stringent discharge standard that is technologically achievable. The 

second, the total volume of ballast water discharged, cannot be regulated by the USCG because 

NISA does not give them that authority. 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, California has proposed the most stringent 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 Permit 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 standards evaluate whether a specific technology can be installed on all 

types of vessels or whether it would only fit some vessel types and designs, whereas, the Vessel 

General Permit does not take into account the achievability 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; however, if ballast water is to be discharged, 

vessels in those states must meet a more stringent 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 all cases, evaluations of current technology were conducted but for instance 

in California, a technology is considered 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 potentially 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 standards 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 state extended the deadline 

to meet the New York standard, meaning until the date that technology is capable of meeting a 

standard 1000 times more stringent than the USCG’s, New York effectively has no ballast water 

standard. 
355

Based on sampling conducted by foreign nations, it appears there is technology available that can 

consistently meet the standard proposed by the USCG. Furthermore, because it appears 

technology is not yet capable of achieving a more stringent standard, the USCG monitors the 

total ballast water discharge through their reporting and record keeping requirements. 

Additionally, the USCG has established periodic reviews every three years with the intention of 

making the ballast water standard more stringent if practical technology is discovered. 

USCG’s Ability to Estimate Stressors Resulting from the Ballast Water Discharges 

Many publications have attempted to predict which 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 determined 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 specific locations those species may establish viable populations. 

Because it is not possible to predict which species 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 effects of future invasions. 

Given the unpredictability of invasive species establishment, the USCG has proposed a 

numerical standard that will limit, to the maximum extent practicable, the probability of an 

invasion occurring. Furthermore, the USCG acknowledges that the total volume discharged is 

critical, but the agency has no authority to regulate 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 reviews to evaluate the practicability 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 with the Ballast Water Management Program 

The USCG has established monitoring, reporting, and record keeping requirements of all vessels 

to ensure that all vessels and vessel operators have complied with the conditions, restrictions, and 

mitigation measures required by the proposed action. The records are checked when monitoring 

takes place; the USCG boards every vessel that enters US waters at least once each year. As 

previously described, a vessel is required to report ballast water discharges prior to making them 

and maintain a record for at least 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 vessels 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 identified, 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 documented and violating vessels 
356

fined. While it is more difficult to prove a vessel is not complying with BWE, the USCG has 

brought two cases related to record keeping and BWE. Under the proposed numerical ballast 

water standard, the USCG expects monitoring water and sediment samples to lead to increased 

compliance and easier enforcement. 

USCG’s Ability to Determine Exposures of Listed 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 Information Clearinghouse. Because every vessel is 

boarded at least once a year and 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 reliability, and the volume of ballast water 

discharged annually. Furthermore, annual compliance monitoring ensures that vessels are 

equipped with ballast water treatment technology and that those technologies 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 various 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 water discharged and therefore can only reduce the 

number of organisms per discharge to the maximum extent practicable. Furthermore, they 

currently monitor the volume of ballast water discharged by each ship, while also conducting 

practicability reviews and periodic reviews to identify new technology capable of achieving a 

more stringent standard. In these ways, the USCG is able to ensure the exposure of listed 

resources to ballast water discharge 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 species 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 determined that the likely effects of new listed 

species will be similar to the effects seen by the 500 non-native species currently inhabiting the 

waters of the US. Furthermore, there is generally a delay between when a species is introduced 

and when its population is first identified as being invasive, the point at which adverse effects 

could potentially occur. This is because, as the final PEIS shows, there is a high likelihood that 

most species that are introduced will go extinct rather 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 resulting 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 effectiveness monitoring that will ensure ballast water 

treatment technology has been installed in all vessels and is effectively meeting the proposed 

numerical standard. That coupled with monitoring 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 effects caused by those invasions. All of these 

monitoring efforts will allow the USCG to undertake 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 USCG’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 invasion and increasing the enforceability of ballast 

water regulations, there is still a potential risk to listed species. 

The USCG proposed the most stringent practicable ballast water management solution to protect 

the environment. Furthermore, the USCG program includes a practicability review and later 

periodic reviews that allow the USCG to increase the stringency of the standard as soon as 

technology is developed that can practicably achieve a greater stringency. Additionally, the 

USCG included three alternative discharge techniques 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 decide not to discharge ballast water due to the 

increased risk of enforcement and cost of installing treatment technologies; however, most 

vessels are unable to operate safely without taking on and discharging 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 water, 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 vessels 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 require 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 discharges 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 critical habitat. The USCG has minimized to the 

maximum extent practicable the likely adverse effects associated with this action. As noted in 

the approach to the assessment section, minimizing 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 expected as a result of this action. 

There are two primary concerns with ballast water discharge: (1) the number of organisms per 

volume of water and (2) the total volume of ballast water discharged. 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 discharged. Therefore, the USCG has selected a 

numerical standard that is currently the most practicably achievable. 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 compliance with these mitigation measures will be 

reviewed by the reports as well as required record keeping on board vessels for 24 months. 

Vessel operators would be required to report and maintain records of their discharges, but there 

is the potential for dishonesty. To make sure vessel operators are in fact treating ballast water 

and discharging it when they say they will, the USCG boards all vessels at least once a year 

determined by a decision matrix. During these exercises, the USCG determines that the vessel is 

in compliance with the proposed FR and has treatment technology installed. Furthermore, the 

USCG takes samples of the ballast water and ballast sediment to ensure the treatment technology 

is meeting pre-installation treatment success and being maintained to continue to meet the 

numerical standard. If the vessel is out of compliance, the USCG is able to take enforcement 

action against the vessel, as has been demonstrated previously with criminal prosecutions for 

falsified BWE recordkeeping and during compliance 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 practicability review to be carried out by 2016 that 

will assess whether there are new technologies available to increase the stringency of the 

standard. After 2016, the USCG will conduct periodic 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 more stringent. In the event of 

a technology that can allow vessels to practicably achieve a more stringent standard, the USCG 

will determine whether that technology will require an economic impact analysis. If the new 

technology is not economically significant, the new practicable numeric standard could be 

implemented in less than a year and if it is economically significant, then it would be 

implemented within three years. 

Because the USCG was not able to avoid all possible harmful exposure associated with ballast 

water discharges, NMFS must assess whether and how quickly the USCG could implement new 

protective measures in the event that the proposed numerical standard is unable to be achieved. 

The USCG is going to assess the potential of all new technologies that may achieve the 
359

numerical standard under a tiered consultation process due to the quantity and variety of 

treatment technologies, as well as the anticipation that new technologies will continually be 

developed until the technologies are able to achieve more stringent standards. Essentially, once a 

technology is developed and can practicably increase the stringency of the ballast water standard, 

all vessels operators will be required to equip their vessels with that technology until a new 

technology is developed that can achieve that same level of stringency. Similarly, if a 

technology is approved but through monitoring is revealed to consistently not be able to achieve 

the numerical standard, the USCG would reinitiate consultation on that technology or quickly 

determine that particular technology is no longer approved for use and suggest one of the other 

technologies that is available that can meet the 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 State, tribal, local, or private actions that are 

reasonably certain to occur in the action area considered in this Opinion. Future Federal actions 

that are unrelated to the proposed action are not considered in this section because they require 

separate consultation pursuant to section 7 of the ESA. 

Government and private actions may include changes in land and water use patterns, including 

ownership and intensity, any of which could affect listed species or their habitat. It is difficult, 

and perhaps speculative, to analyze the effects of such actions, considering the broad geographic 

landscape covered by this BO, the geographic and political variation in the action area, extensive 

private land holdings, the uncertainties associated with State and 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 minimize these effects to listed species, including 

regulations regarding construction best management practices, storm water control, and 

treatment of wastewater. 

Although quantifying an incremental change in survival for the species considered in this 

consultation due to the cumulative effects is not possible, 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 turtles, Hawaiian monk seals, and Cook Inlet beluga 

whales, the proposed action, the environmental baseline, the action area, and cumulative effects, 

it is NMFS’ biological opinion that the program, may affect but is not likely to adversely affect 

these threatened and endangered species or their critical habitat. 
360

NMFS reviewed the current status of California coastal Chinook salmon, central valley Chinook 

salmon, 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 River Chinook salmon, Sacramento River winterrun Chinook salmon, Hood Canal summer run chum salmon, Columbia River chum salmon, 

lower Columbia River coho salmon, Oregon Coast coho salmon, SONCC coho salmon, central 

California coast 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-central California coast 

steelhead, upper Columbia River steelhead, upper Willamette River steelhead, northern 

California steelhead, southern California steelhead, central California coast steelhead, green 

sturgeon, gulf sturgeon, Atlantic sturgeon, shortnose sturgeon, smalltooth sawfish, yelloweye 

rockfish, bocaccio, canary rockfish, Pacific eulachon, elkhorn coral, staghorn coral, white 

abalone, black abalone, Johnson’s sea grass, Steller sea lions, spotted 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 humpback whales, the environmental baseline for the 

action area, the effects of the proposed action, and the cumulative effects. Because the program 

has adaptive management capabilities, feedback loops, monitoring plans, record keeping and 

maintaining requirements, and enforcement capabilities, it is NMFS’ Opinion that the USCG 

ballast water management program, is not likely to jeopardize the continued existence of these 

endangered or threatened species. 

NMFS also reviewed the current status of California coastal Chinook salmon, central valley 

Chinook salmon, lower Columbia River Chinook salmon, Puget Sound Chinook salmon, Snake 

River fall-run Chinook salmon, Snake River spring/summer-run Chinook salmon, upper 

Columbia River spring-run Chinook salmon, Upper Willamette River Chinook salmon, 

Sacramento River winter-run Chinook salmon, Hood 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-central California coast steelhead, upper Columbia 

River steelhead, upper Willamette River steelhead, northern California steelhead, southern 

California steelhead, central California coast steelhead, Pacific eulachon, green sturgeon, gulf 

sturgeon, smalltooth sawfish, Steller sea lions, southern resident killer whales, green sea turtles, 

hawksbill sea turtles, leatherback sea turtles, loggerhead sea turtle, elkhorn coral, staghorn coral, 

and black abalone, the environmental baseline for the action area, the effects of the proposed 

action, and the cumulative effects of this program. Because the program has adaptive 

management capabilities, feedback loops, monitoring plans, record keeping 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, respectively, without special exemption. Take is defined 

as to harass, harm, pursue, hunt, shoot, wound, kill, trap, capture or collect, or to attempt to 
361

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, including breeding, feeding, or sheltering. Harass is 

defined by the USFWS as intentional or negligent actions that create the likelihood of injury to 

listed species to such an extent as to significantly disrupt normal behavior patterns which 

include, but are not limited to, breeding, feeding, or sheltering. Incidental take is defined as take 

that is incidental to, and not the purpose of, the carrying out of an otherwise lawful activity. 

Under the terms of section 7(b)(4) and section 7(o)(2), taking that is incidental to and not 

intended as part of the agency action is not considered to be prohibited 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 limitations of scientific knowledge of the invasion process, predicting 

invasions, invasive species, or the expected effects is impossible at this time. Because there is no 

means of predicting specific invaders, the number or species of listed species that will be 

affected, the proportion of the population that will be taken, or the amount 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 population and habitat scale effects. 

As was explained in the exposure analysis, using 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 water 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 establishing each year. The volume of water in the 

coastal, estuarine, and freshwater regions of the US is so vast that it is not economically possible 

to conduct annual environmental assessments and 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 available science and measuring invasions by the volume of water 

discharged geographically based on the proposed numerical standard for ballast water 

management. Therefore NMFS will provide take in terms of the number of harmful invasive 

species establishing each year using volumes of ballast water discharged as a surrogate for the 

take that is likely to affect listed species. While the USCG does not control the volume of water 

discharged, they do regulate the number of organisms per volume of ballast water that is then 

discharged and therefore NMFS believes this is an appropriate proxy 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, treatment, 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 the probable rate of harmful invasions. The incidental take expected 

annually under this program is identified using the volume of ballast water discharged as a proxy 

for listed species and critical habitat affected. As the periodic reviews allow the USCG to 
362

establish more stringent ballast water standards, the amount and extent of take of listed species is 

expected to decrease. Assuming technology will advance in the future to allow for a new, 

practicable standard to be implemented, the impacts to the listed species and their critical habitat 

will be reduced, much as the reductions of those 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 NMFS 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 Opinion, but not to exceed 

the amount in this Opinion. 

Not all newly established invasive species are harmful to listed species or their critical habitat. 

In fact, some may be beneficial, while most are innocuous. Past research has concluded that the 

proportion of invasive species that 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 limited to the introduction, but 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 94x106 tonnes per year to Bioregion 

NEP-V. 

Central Valley Chinook salmon 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 three in the next 16 

years, which would correlate to a volumetric ballast discharge of 94x106 tonnes per year to 

Bioregion NEP-V. 

Lower Columbia River Chinook salmon 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 16 years, which would correlate to a volumetric ballast discharge of 94x106 per year tonnes 

to Bioregion NEP-IV. 

Snake River fall-run Chinook salmon 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 16 years, which would correlate to a volumetric ballast discharge of 94x106 tonnes per year 

to Bioregion NEP-IV. 

Snake River spring/summer-run Chinook salmon 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 16 years, which would correlate to a volumetric ballast discharge of 94x106 tonnes per 

year to Bioregion NEP-IV. 

Upper Columbia River Chinook salmon 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 
363

next 16 years, which would correlate to a volumetric ballast discharge of 94x106 tonnes per year 

to Bioregion NEP-IV. 

Upper Willamette River Chinook salmon 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 16 years, which would correlate to a volumetric ballast discharge of 94x106 tonnes per year 

to Bioregion NEP-IV. 

Puget Sound Chinook salmon 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 16 

years, which would correlate to a volumetric ballast discharge of 94x106 tonnes to Bioregion 

NEP-IV. 

Sacramento winter-run Chinook salmon 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 three in the 

next 16 years, which would correlate to a volumetric ballast discharge of 94x106 tonnes per year 

to Bioregion NEP-V. 

Columbia River chum salmon 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 16 

years, which would correlate to a volumetric ballast discharge of 94x106 tonnes per year to 

Bioregion NEP-IV. 

Hood Canal summer-run chum salmon 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 16 years, which would correlate to a volumetric ballast discharge of 94x106 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 volumetric ballast discharge of 94x106 tonnes per year to Bioregion NEPIV. 

SONCC coho salmon 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 three in the next 16 years, 

which would correlate to a volumetric ballast discharge of 94x106 tonnes per year to Bioregions 

NEP-IV and NEP-V. 

Central California coast coho salmon 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 three in the next 

16 years, which would correlate to a volumetric ballast discharge of 94x106 tonnes per year to 

Bioregion NEP-V. 

Oregon Coast coho salmon 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 16 years, 
364

which would correlate to a volumetric ballast discharge of 94x106 tonnes per year to Bioregion 

NEP-IV. 

Snake River sockeye salmon 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 16 

years, which would correlate to a volumetric ballast discharge of 94x106 tonnes per year to 

Bioregion NEP-IV. 

Ozette Lake sockeye salmon 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 16 

years, which would correlate to a volumetric ballast discharge of 94x106 tonnes per year to 

Bioregion NEP-IV. 

California central valley 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 three in the next 

16 years, which would correlate to a volumetric ballast discharge of 94x106 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 species per year and no more than one in the next 16 

years, which would correlate to a volumetric ballast discharge of 94x106 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 species per year and no more than one in the next 16 

years, which would correlate to a volumetric ballast discharge of 94x106 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 per year and no more than three in the next 16 

years, which would correlate to a volumetric ballast discharge of 94x106 tonnes per year to 

Bioregion NEP-V. 

Snake River Basin 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 16 

years, which would correlate to a volumetric ballast discharge of 94x106 tonnes per year to 

Bioregion NEP-IV. 

South-central California coast 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 three in the 

next 16 years, which would correlate to a volumetric ballast discharge of 94x106 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 16 
365

years, which would correlate to a volumetric ballast discharge of 94x106 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 species per year and no more than one in the next 16 

years, which would correlate to a volumetric ballast discharge of 94x106 tonnes per year to 

Bioregion NEP-IV. 

Upper Willamette River 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 16 

years, which would correlate to a volumetric ballast discharge of 94x106 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 94x106 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 species per year and no more than three in the next 

16 years, which would correlate to a volumetric ballast discharge of 94x106 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 volumetric ballast discharge of 94x106 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 94x106 tonnes per year to Bioregions NA-ET2, NA-ET3, and 

CAR-VII. 

Gulf of Maine 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 volumetric ballast discharge of 94x106 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 volumetric ballast discharge of 94x106 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 volumetric ballast discharge of 94x106 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 94x106 tonnes per year to Bioregions NA-ET2, 

NA-ET3, and CAR-VII. 

South Atlantic 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 volumetric ballast discharge of 94x106 tonnes per year to Bioregions NAET2, NA-ET3, and CAR-VII. 

Green sturgeon 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 volumetric ballast discharge of 94x106 tonnes per year to Bioregion NEP-V. 

Gulf sturgeon and their critical habitat are likely to be harmed by no more than two ballastmediated harmful invasive species per year and no more than 16 in the next 16 years, which 

would correlate to a volumetric ballast discharge of 188x106 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 94x106 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 years, which would correlate to a volumetric ballast 

discharge of 94x106 tonnes per year to Bioregion NEP-IV. 

Canary 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 94x106 tonnes per year to Bioregion NEP-IV. 

Smalltooth sawfish 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 volumetric ballast discharge of 94x106 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 volumetric ballast discharge of 94x106 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 volumetric ballast discharge of 94x106 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 the next 16 years, which would correlate to a 

volumetric ballast discharge of 94x106 tonnes per year to Bioregion NEP-V. 

Black abalone 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 volumetric ballast discharge of 94x106 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 per year and no more than five in the next 16 

years, which would correlate to a volumetric ballast discharge of 94x106 tonnes per year to 

Bioregions NEP-IV and NEP-V. 

Western DPS Steller sea lions 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 four in the next 16 

years, which would correlate to a volumetric ballast discharge of 94x106 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 the next 16 years, which would correlate to a 

volumetric ballast discharge of 94x106 tonnes per year to Bioregions NEP-I, NEP-II, and NAS1. 

Southern resident killer whales 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 16 

years, which would correlate to a volumetric ballast discharge of 94x106 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 94x106 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 and no more than 23 in the next 16 years, which 

would correlate to a volumetric ballast discharge of 188x106 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 188x106 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 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 
368

volumetric ballast discharge of 188x106 tonnes per year to Bioregions NA-ET3, CAR-VII, 

CAR-IV, and CAR-I. 

Olive ridley sea turtles are likely to be harmed 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 188x106 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 volumetric ballast discharge of 94x106 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 and no more than 23 in the next 16 years, which 

would correlate to a volumetric ballast discharge of 188x106 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 188x106 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 invasive species introduced using the volume of 

ballast water discharged as a surrogate, then provides take coverage for harmful invasions by 

calculating the ratio of harmful invasive species identified from several hundred previous 

invasive species that have been studied. Under a new, more stringent ballast water standard, the 

number of non-native species per volume of ballast water would be reduced. This will reduce 

the probability of a harmful invasive species establishing, and therefore 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 standard would not require 

reinitiation of consultation, but rather it would require NMFS to amend 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 required to provide the amended Opinion with a 

new incidental take statement to the USCG. 
369

Reasonable and Prudent Measures 

Reasonable and Prudent Measures (RPMs) are non-discretionary measures to minimize take that 

may or may not already be part of the description of the proposed action. They must be 

implemented as binding conditions for the exemption in section 7(o)(2) to apply. The USCG has 

the continuing duty to regulate the activities covered in this incidental take statement. If the 

USCG fails to adhere to the terms and conditions 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 resulting from implementation of this action. 

The USCG shall monitor and report to NMFS Office 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 summarizing national patterns of ballast water delivery and 

management. 

2. Provide results of the practicability review. 

3. Provide results of any periodic programmatic reviews which may be undertaken as 

mandated under NISA/NANPCA. 

4. Provide an annual report detailing the number of ships entering US waters, the number of 

inspections, ballast water violations found, and corrective actions taken to fix the 

violation. 

CONSERVATION RECOMMENDATIONS

Section 7(a)(1) of the Act directs Federal agencies to utilize their authorities 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 action 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 chains, 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 programs with the goal of creating more stringent 

ballast water standards by developing new technologies and testing procedures. 

3. Habitat Improvement Measures. The USCG should implement 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-native species to mitigate the impacts 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-based) discharge facilities capable 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 eventually eliminate ballast water discharges and 

invasive species introductions 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 operate 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 under 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 actions minimizing or avoiding adverse effects or 

benefiting listed species or their 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, reinitiation of formal consultation is required where 

discretionary Federal agency involvement or control over the action has been retained (or is 

authorized by law) and if: (1) the amount or extent of take is exceeded; (2) new information 

reveals effects of the agency action that may affect listed species or critical habitat in a manner 
371

or to an extent not considered in this opinion; (3) the agency action is subsequently modified in a 

manner that causes an effect to the listed species or critical habitat not considered in this opinion; 

or (4) a new species is listed or critical habitat designated that may be affected by the action. In 

the event USCG would authorize, fund, or carry out ballast water discharge related activities that 

may affect resources in a manner or to an extent not considered in this Opinion, USCG must 

reinitiate consultation to compensate for information that was not available for consideration 

during this consultation. 

Literature Cited 

(FEMAT) F. E. M. A. T. (1993) Forest Ecosystem Management: An Ecological, Economic and 

Social Assessment. Report of the Forest Ecosystem Management Assessment Team, July 1993. 

Abernathy K. (1999) Foraging ecology of Hawaiian monk seals at French Frigate Shoals, 

Hawaii. p. 65. University of Minnesota, St. Paul, MN. 

Abrego D., Ulstrup K. E., Willis B. L. & Oppen M. J. H. v. (2010) Species–specific interactions 

between algal endosymbionts and coral hosts define their bleaching response to heat and light 

stress. Proceedings of the Royal Society of London Part B 275, 2273-2282. 

Abrego D., Van Oppen M. J. H. & Willis B. L. (2009) Onset of algal endosymbiont specificity 

varies among closely related species of Acropora corals during early ontogeny. Molecular 

Ecology 18, 3532-3543. 

ACC. (2010) Alabama’s debris history. Alabama Coastal Cleanup. 

Acevedo-Whitehouse K. & Duffus. A. L. J. (2009) Effects of environmental change on wildlife 

health. Philosophical Transactions of the Royal Society of London B Biological Sciences 364, 

3429-3438. 

Ackerman R. A. (1997) The nest environment, and the embryonic development of sea turtles. In: 

The Biology of Sea Turtles (eds P. L. Lutz and J. A. Musick) pp. 83-106. CRC Press, Boca 

Raton, Florida. 

Acropora Biological Review Team. (2005) Atlantic Acropora Status Review Document. Report 

to National Marine Fisheries Service, Southeast Regional Office, Miami, Florida. 

Adams P. B., Grimes C., Hightower J. E., Lindley S. T., Moser M. L. & Parsley M. J. (2007) 

Population status of North American green sturgeon, Acipenser medirostris. Environmental 

Biology of Fishes 79, 339-356. 

Adams P. B., Grimes C. B., Hightower J. E., Lindley S. T. & Moser M. L. (2002) Status Review 

for the North American green sturgeon. NOAA, National Marine Fisheries Service, Southwest 

Fisheries Science Center, Santa Cruz, California. 

Adams S. R., Hoover J. J. & Killgore K. J. (1999) Swimming endurance of juvenile pallid 

sturgeon, Scaphirhynchus albus. Copeia 1999, 802-807. 

Adams W. F. & Wilson C. (1995) The status of the smalltooth sawfish, Pristis pectinata Latham 

1794 (Pristiformes: Pristidae) in the United States. Chondros 6, 1-5. 

Addison D. S. (1997) Sea turtle nesting on Cay Sal, Bahamas, recorded June 2-4, 1996. Bahamas 

Journal of Science 5, 34-35. 

Addison D. S. & Morford B. (1996) Sea turtle nesting activity on the Cay Sal Bank, Bahamas. 

Bahamas Journal of Science 3, 31-36. 

Adey W. H. (1977) Shallow water Holocene bioherms of the Caribbean Sea and West Indies. In: 

The 3rd International Coral Reef Symposium pp. xxi-xxiii. 

Adey W. H. (1978) Coral reef morphogenesis: A multidimensional model. Science 202, 831-837. 
372

Aguilar R., Mas J. & Pastor X. (1995) Impact of Spanish swordfish longline fisheries on the 

loggerhead sea turtle Caretta caretta population in the western Mediterranean. In: Proceedings 

of the Twelfth Annual Workshop on Sea Turtle Biology and Conservation (eds J. I. Richardson 

and T. H. Richardson). U.S. Department of Commerce, Jekyll Island, Georgia. 

Aguirre A. A., Balazs G. H., Zimmerman B. & Galey F. D. (1994) Organic contaminants and 

trace metals in the tissues of green turtles (Chelonia mydas) afflicted with fibropapillomas in the 

Hawaiian Islands. Marine Pollution Bulletin 28, 109-114. 

Aguirre A. A., Spraker T. R., Balazs G. H. & Zimmerman B. (1998) Spirochidiasis and 

fibropapillomatosis in green turtles of the Hawaiian Islands. journal of wildlife diseases 34, 91-

98. 

Aguirremacedo M., Vidalmartinez V., Aherrerasilveira J., Valdeslozano D., Herrerarodriguez M. 

& Olveranovoa M. (2008) Ballast water as a vector of coral pathogens in the Gulf of Mexico: 

The case of the Cayo Arcas coral reef. Marine Pollution Bulletin 56, 1570-1577. 

Ainley D. G., Anderson D. W. & Kelly P. R. (1981) Feeding ecology of marine cormorants in 

southwestern North America. Condor 83, 120-131. 

Ainsworth T. D. & Hoegh-Guldberg O. (2009) Bacterial communities closely associated with 

coral tissues vary under experimental and natural reef conditions and thermal stress. Aquatic 

Biology 4, 289-296. 

Ainsworth T. D., Hoegh-Guldberg O., Heron S. F., Skirving W. J. & Leggat W. (2008) Early 

cellular changes are indicators of pre-bleaching thermal stress in the coral host. Journal of 

Experimental Marine Biology and Ecology 364, 63-71. 

Airamé S., Gaines S. & Caldow C. (2003) Ecological linkages: Marine and estuarine ecosystems 

of central and northern California. p. 164. NOAA, National Ocean Service, Silver Spring, 

Maryland. 

Al-Bahry S., Mahmoud I., Elshafie A., Al-Harthy A., Al-Ghafri S., Al-Amri I. & Alkindi A. 

(2009) Bacterial flora and antibiotic resistance from eggs of green turtles Chelonia mydas: An 

indication of polluted effluents. Marine Pollution Bulletin 58, 720-725. 

Alava J. J., Keller J. M., Kucklick J. R., Wyneken J., Crowder L. & Scott G. I. (2006) 

Loggerhead sea turtle (Caretta caretta) egg yolk concentrations of persistent organic pollutants 

and lipid increase during the last stage of embryonic development. Science of the Total 

Environment 367, 170-181. 

Alcaraz C., Bisazza A. & Garcia-Berthou E. (2008) Salinity mediates the competitive 

interactions between invasive mosquitofish and an endangered fish. Oecologia 155, 205-213. 

Alcorn D. J. & Henderson J. R. (1984) Resumption of nursing in "weaned" Hawaiian monk seal 

pups. 'Elepaio 45, 11-12. 

Allen B. M. & Angliss R. P. (2010) Alaska Marine Mammal Stock Assessments, 2009. In: 

NOAA Technical Memorandum NMFS-AFSC-206 p. 276. U.S. Department of Commerce. 

Allen J. A. (1998) Mangroves as Alien Species: The Case of Hawaii. Global Ecology and 

Biogeography 7, 61-71. 

Allen K. R. & Smith. T. D. (1978) A note on the relation between pregnancy rate, age at maturity 

and adult and juvenile mortality rates. Report of the International Whaling Commission 28, 477-

478. 

Allen M. J. & Smith G. B. (1988) Atlas and zoogeography of common fishes in the Bering Sea 

and Northeastern Pacific. 
373

Allen P. & Cech J. (2007) Age/size effects on juvenile green sturgeon, Acipenser medirostris , 

oxygen consumption, growth, and osmoregulation in saline environments. Environmental 

Biology of Fishes 79, 211-229. 

Allen P. J., Hobbs J. A., Cech J. J., Van Eenennaam J. P. & Doroshov S. I. (2009) Using Trace 

Elements in Pectoral Fin Rays to Assess Life History Movements in Sturgeon: Estimating Age at 

Initial Seawater Entry in Klamath River Green Sturgeon. Transactions of the American Fisheries 

Society 138, 240-250. 

Allen P. J., Hodge B., Werner I. & Cech J. J. J. (2006) Effects of ontogeny, season, and 

temperature on the swimming performance of juvenile green sturgeon (Acipenser medirostris). 

Canadian Journal of Fisheries and Aquatic Sciences 63, 1360-1369. 

Allen P. J. & J.J. Cech, Jr. (2007) Age/size effects on juvenile green sturgeon, Acipenser 

medirostris, oxygen consumption, growth, and osmoregulation in saline environments. 

Environmental Biology of Fishes 79, 211-229. 

Almy C. C. & Carrión-Torres C. (1963) Shallow-water stony corals of Puerto Rico. Caribbean 

Journal of Science 3, 133-162. 

Alpine A. E. & Cloern J. E. (1992) Trophic interactions and direct physical effects control 

phytoplankton biomass and production in an estuary. Limnology and Oceanography 37, 946-955. 

American Bureau of Shipping. 2010. Guide for ballast water exchange. Houston, Texas. 

Amos A. F. (1989) Recent strandings of sea turtles, cetaceans and birds in the vicinity of 

Mustang Island, Texas. In: Proceedings of the First International Symposium on Kemp's Ridley 

Sea Turtle Biology, Conservation and Management (eds C. W. C. Jr. and A. M. Landry) p. 51. 

Anan Y., Kunito T., Watanabe I., Sakai H. & Tanabe S. (2001) Trace element accumulation in 

hawksbill turtles (Eretmochelys imbricata) and green turtles (Chelonia mydas) from Yaeyama 

Islands, Japan. Environmental Toxicology and Chemistry 20, 2802-2814. 

Anders P., Richards D. & Powell M. S. (2002) The first endangered white sturgeon population: 

Repercussions in an altered large river-floodplain ecosystem. In: Biology, management, and 

protection of North American sturgeon, Symposium 28. (ed V. W. Webster) pp. 67-82. American 

Fisheries Society, Bethesda, Maryland. 

Anderson J. J. (2000) A vitality-based model relating stressors and environmental properties to 

organism survival. Ecological Monographs 70, 445-470. 

Andreassen P. M. R., Martinussen M. B., Hvidsten N. A. & O. Stefansson S. (2001) Feeding and 

prey selection of wild Atlantic salmon post smolts. Journal of Fish Biology 58, 1667-1679. 

Andrew R. K., Howe B. M. & Mercer J. A. (2002) Ocean ambient sound: Comparing the 1960s 

with the 1990s for a receiver off the California coast. Journal of the Acoustical Society of 

America 3, 65-70. 

Andrews A. H., Burton E. J., Kerr L. A., Cailliet G. M., Coale K. H., Lundstrom C. C. & Brown 

T. A. (2005) Bomb radiocarbon and lead-radium disequilibria in otoliths of bocacio rockfish 

(Sebstes paucispinis): a determination of age and longevity for a difficult-to-age fish. Marine and 

Freshwater Research 56, 517-528. 

Andrews A. H., Kerr L. A., Cailliet G. M., Brown T. A., Lundstrom C. C. & Stanley R. D. 

(2007) Age validation of canary rockfish (Sebastes pinniger) using two independent otolith 

techniques: lead-radium and bomb radiocarbon dating. Marine and Freshwater Research 58, 

531-541. 

Andrews H. & Shanker K. (2002) A significant population of leatherback turtles in the Indian 

Ocean. Kachhapa 6, 19. 
374

Andrews H. V., Krishnan S. & Biswas P. (2002) Leatherback nesting in the Andaman & Nicobar 

Islands. Kachhapa 6, 15-18. 

Andrews J. D. (1984) Epizootiology of diseases of oysters (Crassostrea virginica), and parasites 

of associated organisms in eastern North America. Helgoland Marine Research 37, 149-166. 

Andrews R. D., Calkins D. G., Davis R. W. & Norcross B. L. (2001) Foraging behavior and 

energetics of adult female Steller sea lions. In: Steller sea lion decline: is it food II? (eds D. 

DeMaster and S. Atkinson), Fairbanks, Alaska. 

Andriolo A., Kinas P. G., Engel M. H., Albuquerque Martins C. C. & Rufino A. M. (2010) 

Humpback whales within the Brazilian breeding ground: distribution and population size 

estimate. Endangered Species Research 11, 233-243. 

Angliss R. P. & Allen A. L. (2007) Draft Alaska marine mammal stock assessments 2008. 

National Marine Mammal Laboratory, Alaska Fisheries Science Center, Seattle, Washington. 

Angliss R. P., DeMaster D. P. & Lopez A. L. (2001) Alaska marine mammal stock assessments, 

2001. In: NOAA Technical Memorandum U.S. Department of Commerce. 

Angliss R. P. & Lodge. K. L. (2004) Alaska marine mammal stock assessments, 2003. NOAA 

Technical Memorandum NMFS-AFSC-144. 230pp. 

Angliss R. P. & Outlaw R. B. (2005) Alaska marine mammal stock assessments, 2005. U.S. 

Department of Commerce, NOAA Technical Memorandum NMFSAFSC-161, 250 p. 

Angliss R. P. & Outlaw R. B. (2007a) Alaska Marine Mammal Stock Assessments, 2006. U.S. 

Department of Commerce, NOAA Technical Memorandum NMFS-AFSC-168, 244 p. 

Angliss R. P. & Outlaw R. B. (2007b) Alaska marine mammal stock assessments, 2006. In: 

NOAA Technical Memorandum U.S. Department of Commerce. 

Angliss R. P. & Outlaw R. B. (2008) Alaska marine mammal stock assessments, 2007. In: NOAA 

Technical Memorandum U.S. Department of Commerce. 

Anguiano-Beltrán C., Searcy-Bernal R. & Lizárraga-Partida M. L. (1998) Pathogenic effects of 

Vibrio alginolyticus on larvae and poslarvae of the red abalone Haliotis rufescens. Diseases of 

Aquatic Organisms 33, 119-122. 

Anonmyous. (2010a) References, Index. In: Marine Mammal Ecology and Conservation: A 

Handbook of Techniques (eds I. L. Boyd, W. D. Bowen and S. J. Iverson) pp. 360-450. Oxford 

University Press. 

Anonmyous. (2010b) Title pages, preface, table of contents, list of abbreviations, list of 

contributors. In: Marine Mammal Ecology and Conservation: A Handbook of Techniques (eds I. 

L. Boyd, W. D. Bowen and S. J. Iverson) pp. i-xxiv. Oxford University Press. 

Anonmyous. (2009a) Blue whales re-establishing former migration patterns. Marine Pollution 

Bulletin 58, 949. 

Anonmyous. (2009b) Right whale sedation enables disentanglement effort. Marine Pollution 

Bulletin 58, 640-641. 

Anthony K. R. N., Connolly S. R. & Hoegh-Guldberg O. (2007) Bleaching, energetics, and coral 

mortality risk: Effects of temperature, light, and sediment regime. Limnology and Oceanography

52, 716-726. 

Anthony K. R. N., Kline D. I., Diaz-Pulido G., Dove S. & Hoegh-Guldberg O. (2008) Ocean 

acidification causes bleaching and productivity loss in coral reef builders. Proceedings of the 

National Academy of Sciences 105, 17442-17446. 

Antonelis G. A., Baker J. D., Johanos T. C., Braun R. C. & Harting A. L. (2006) Hawaiian monk 

seal (Monachus schauinslandi): status and conservation issues. Atoll Research Bulletin 543, 75-

101. 
375

Antonelis Jr. G. A. & Fiscus C. H. (1980) The pinnipeds of the California Current. CalCOFI 

Report 21, 68-78. 

Anttila C. K., Daehler C. C., Rank N. E. & Strong D. R. (1998) Greater male fitness of a rare 

invader (Spartina alterniflora, Poaceae) threatens a common native (Spartina foliosa) with 

hybridization. American Journal of Botany 85, 1597-1601. 

Armstrong J. D., Huntingford F. A. & Herbert N. A. (1999) Individual space use strategies of 

wild juvenile Atlantic salmon. Journal of Fish Biology 55, 1201-1212. 

Armstrong J. L. & Hightower J. E. (2002) Potential for restoration of the Roanoke River 

population of Atlantic sturgeon. Journal of Applied Ichthyology 18, 475-480 

Arndt D. S., Baringer M. O., Johnson M. R., Alexander L. V., Diamond H. J., Fogt R. L., Levy J. 

M., Richter-Menge J., Thorne P. W., Vincent L. A., Watkins A. B. & Willett K. M. (2010) State 

of the climate in 2009. Bulletin of the American Meteorological Society 91, S1-S224. 

Arthur K., Limpus C., Balazs G., Capper A., Udy J., Shaw G., Keuper-Bennett U. & Bennett P. 

(2008) The exposure of green turtles (Chelonia mydas) to tumour promoting compounds 

produced by the cyanobacterium Lyngbya majuscula and their potential role in the aetiology of 

fibropapillomatosis. Harmful Algae 7, 114-125. 

Ashe E., Noren D. P. & Williams. R. (2010) Animal behaviour and marine protected areas: 

Incorporating behavioural data into the selection of marine protected areas for an endangered 

killer whale population. Animal Conservation 13, 196-203. 

ASMFC. (1998) American Shad and Atlantic Sturgeon Stock Assessment Peer Review: Terms 

of Reference and Advisory Report. p. 33. Atlantic States Marine Fisheries Commission, 

Washington D.C. 

Asrar F. F. (1999) Decline of marine turtle nesting populations in Pakistan. Marine Turtle 

Newsletter 83, 13-14. 

ASSRT. (2007) Status Review of Atlantic sturgeon (Acipenser oxyrinchus oxyrinchus). p. 174. 

Report to National Marine Fisheries Service, Northeast Regional Office by Atlantic Sturgeon 

Status Review Team. 

Atkinson A., Siegel V., Pakhomov E. & Rothery P. (2004) Long-term decline in krill stock and 

increase in salps within the Southern Ocean. Nature 432, 100-103. 

Atkinson S., DeMaster D. P. & Calkins D. G. (2008) Anthropogenic causes of the western 

Steller sea lion Eumetopias jubatus population decline and their threat to recovery. Mammal 

Review 38, 1-18. 

Au W. W. L. & Green. M. (2000) Acoustic interaction of humpback whales and whale-watching 

boats. Marine Environmental Research 49, 469-481. 

Ault J. S. (1985) Species profiles: Life histories and environmental requirements of coastal 

fisheries and invertebrates (Pacific Southwest)-- black, green, and red abalones. In: Biological 

Report 82(11.21). U.S. Fish and Wildlife Service, US Army Corps of Engineers. 

Ault J. S. & Martini J. D. (1987) Movement, and dispersion of the red abalone, Haliotis 

rufescens, in northern California. California Fish and Game 73, 196-213. 

Avens L. & Goshe L. R. (2007) Skeletochronological analysis of age and growth for leatherback 

sea turtles in the western North Atlantic. In: 27th Annual Symposium on Sea Turtle Biology and 

Conservation (eds M. Frick, A. Panagopoulou, A. F. Rees and K. Williams) p. 223, Myrtle 

Beach, South Carolina. 

Ayuso R. A., Foleyand N. K., Robinson Jr. G. R., Colvin A. S., Lipfert G. & Reeve A. S. (2006) 

Tracing lead isotopic compositions of common arsenical pesticides in a coastal Maine watershed 
376

containing arsenic-enriched ground water. In: Contaminated Soils, Sediments and Water. 

Proceedings of the 21st Annual International Conference on Soils Sediments and Water p. 25. 

Babcock R. & Keesing H. (1999) Fertilization biology of the abalone Haliotis laevigata: 

Laboratory and field studies. Canadian Journal of Fisheries and Aquatic Science 56, 1668-1678. 

Baggett L. S. & Bright T. J. (1985) Coral recruitment at the East Flower Garden Reef. 

Proceeding of the 5th International Coral Reef Congress 4, 379-384. 

Bahn R. A. & Peterson D. L. (2009) Sturgeon bycatch in the Altamaha River shad fishery. In: 

American Fisheries Society Nashville, Tennessee. 

Bailey A. M. (1952) The Hawaiian monk seal. Museum Pictorial, Denver Museum of Natural 

History 7. 

Bailey K. M. & Houde E. D. (1989) Predation on eggs and larvae of marine fishes and the 

recruitment problem. Advances in Marine Biology 25, 1-83. 

Bailey K. M., Piatt J. F., Royer T. C., Macklin S. A., Reed R. K., Shima M., Francis R. C., 

Hollowed A. B., Somerton D. A., Brodeur R. D., Ingraham W. J., Anderson P. J. & Wooster W. 

S. (1995) ENSO events in the northern Gulf of Alaska, and effects on selected marine fisheries. 

CalCOFI Reports 36, 78-96. 

Bailey R. G. (1995) Description of the ecoregions of the United States. p. 108. USDA Forest 

Service, Washington D.C. 

Bailey R. J. E., Dick J. T. A., Elwood R. W. & MacNeil C. (2006) Predatory interactions 

between the invasive amphipod Gammarus tigrinus and the native opossum shrimp Mysis relicta. 

Journal of the North American Benthological Society 25, pp. 

Bain D. (1990) Examining the validity of inferences drawn from photo-identification data, with 

special reference to studies of the killer whale (Orcinus orca) in British Columbia. Report of the 

International Whaling Commission, Special Issue 12, 93-100. 

Bain D. E. & Dahlheim M. E. (1994) Effects of masking noise on detection thresholds of killer 

whales. In: Marine mammals and the Exxon Valdez (ed T. R. Loughlin) pp. 243-256. Academic 

Press, San Diego, CA. 

Bain D. E., Lusseau D., Williams R. & Smith J. C. (2006a) Vessel traffic disrupts the foraging 

behavior of southern resident killer whales (Orcinus spp.). p. 26. IWC Paper SC/59/ForInfo28. 

Bain D. E., Lusseau D., Williams R. & Smith J. C. (2006b) Vessel traffic disrupts the foraging 

behavior of southern resident killer whales (Orcinus spp.). IWC Paper SC/59/ForInfo28. 26p. 

Bain D. E., Williams R., Smith J. C. & Lusseau D. (2006c) Effects of vessels on behavior of 

Southern Resident killer whales (Orcinus spp.) 2003-2005. NMFS Contract Report No. 

AB133F05SE3965. 66p. 

Bain M. B. (1997) Atlantic and shortnose sturgeons of the Hudson River: Common and 

divergent life history attributes. Environmental Biology of Fishes 48, 347-358. 

Bain M. B., Cornwell E. R., Hope K. M., Eckerlin G. E., Casey R. N., Groocock G. H., Getchell 

R. G., Bowser P. R., Winton J. R., Batts W. N., Cangelosi A. & Casey J. W. (2010) Distribution 

of an Invasive Aquatic Pathogen (Viral Hemorrhagic Septicemia Virus) in the Great Lakes and 

Its Relationship to Shipping. Plos One 5, -. 

Bain M. B., Haley N., Peterson D., Arend K. K., Mills K. E. & Sullivan P. J. (2000) Shortnose 

sturgeon of the Hudson River: an endangered species recovery success. In: 20th Annual Meeting 

of the American Fisheries Society p. 14, St. Louis, MO. 

Bain M. B., Peterson D. L., Arend K. K. & Haley N. (1999) Atlantic Sturgeon Population 

Monitoring for the Hudson River Estuary: Sampling Design and Gear Recommendations. p. 34. 
377

Hudson Rivers Fisheries Unit, New York State Department of Environmental Conservation, 

New Paltz, NY and The Hudson River Foundation, New York, NY. 

Baird A. H., Cumbo V. R., Leggat W. & Rodriguez-Lanetty M. (2007) Fidelity and flexibility in 

coral symbioses. Marine Ecology Progress Series 347, 307-309. 

Baird R. W. (2000) The killer whale: foraging specializations and group hunting. In: Cetacean 

societies: Field studies of dolphins and whales (eds J. Mann, R. C. Connor, P. L. Tyack and H. 

Whitehead) pp. 127-153. University of Chicago Press, Chicago, Illinois. 

Baird R. W. (2001a) Status of killer whales, Orcinus orca, in Canada. Canadian Field-Naturalist

115, 676-701. 

Baird R. W. (2001b) Status of killer whales, Orcinus orca, in Canada. Canadian Field-Naturalist

115, 676-701. 

Baird R. W. (2001c) Status of killer whales, Orcinus orca, in Canada. Canadian Field-Naturalist

115, 676-701. 

Bak R. P. M. (1977) Coral reefs and their zonation in the Netherland Antilles. AAPG Studies of 

Geology 4, 3-16. 

Bak R. P. M. & Criens S. R. (1982) Survival after fragmentation of colonies of Madracis 

mirabilis, Acropora palmata, and A. cervicornis (Scleractinia) and the subsequent impact of a 

coral disease. Proceedings of the 4th International Coral Reef Symposium 1, 221-227. 

Baker C. S. (1985) The behavioral ecology and populations structure of the humpback whale 

(Megaptera novaeangliae) in the central and eastern Pacific. University of Hawaii at Manoa. 

Baker C. S. & Herman L. M. (1987) Alternative population estimates of humpback whales in 

Hawaiian waters. Canadian Journal of Zoology 65, 2818-2821. 

Baker C. S., Herman L. M., Perry A., Lawton W. S., Straley J. M. & Straley J. H. (1985) 

Population characteristics and migration of summer and late-season humpback whales 

(Megaptera novaeangliae) in southeastern Alaska. Marine Mammal Science 1, 304-323. 

Baker J. D. (2008) Variation in the relationship between offspring size and survival provides 

insight into causes of mortality in Hawaiian monk seals. Endang Species Res 5, 55-64. 

Baker J. D., Harting A. L., Wurth T. A. & Johanos T. C. (2011) Dramatic shifts in Hawaiian 

monk seal distribution predicted from divergent regional trends. Marine Mammal Science 27, 78-

93. 

Baker J. D. & Johanos T. C. (2004) Abundance of the Hawaiian monk seal in the main Hawaiian 

Islands. Biological Conservation 116, 103-110. 

Baker J. D., Littnan C. L. & Johnston D. W. (2006) Potential effects of sea level rise on the 

terrestrial habitats of endangered and endemic megafauna in the Northwestern Hawaiian Islands. 

Endangered Species Research 4, 1-10. 

Balazs G. H. (1982) Growth rates of immature green turtles in the Hawaiian Archipelago. In: 

Biology and Conservation of Sea Turtles (ed K. A. Bjorndal) pp. 117-125. Smithsonian 

Institution Press, Washington D.C. 

Balazs G. H. (1983) Recovery records of adult green turtles observed or originally tagged at 

French Frigate Shoals, Northwestern Hawaiian Islands. In: NOAA Technical Memorandum. U.S. 

Department of Commerce. 

Balazs G. H. & Chaloupka M. (2004) Thirty-year recovery trend in the once depleted Hawaiian 

green sea turtle stock. Biological Conservation 117, 491-498. 

Balazs G. H. & Chaloupka M. (2006) Recovery trend over 32 years at the Hawaiian green turtle 

rookery of French Frigate Shoals. Atoll Research Bulletin 543, 147-158. 
378

Balcomb III K. C. & Nichols G., Jr. (1982) Humpback whale censuses in the West Indies. Report 

of the International Whaling Commission 32, 401-406. 

Baldwin R., Collins T., Minton G., Findlay K., Corkeron P., Willson A. & Bressem M.-F. V. 

(2010) Arabian Sea humpback whales: Canaries for the northern Indian Ocean? p. 5. IWC 

Scientific Committee, Agadir, Morocco. 

Baldwin R., Hughes G. R. & Prince R. I. T. (2003) Loggerhead turtles in the Indian Ocean. In: 

Loggerhead Sea Turtles (eds A. B. Bolten and B. E. Witherington) pp. 218-232. Smithsonian 

Institution Press, Washington, D.C. 

Baldwin R. M. (1992) Nesting turtles on Masirah Island: Management issues, options, and 

research requirements. Ministry of Regional Municipalities and Environment, Oman. 

Ban S. (2005) Modelling, and characterization of Steller sea lion haulouts and rookeries using 

oceanographic and shoreline type data. University of British Columbia, Vancouver, British 

Columbia. 

Bannister J. C. (2005) Intersessional working group on Southern Hemisphere humpback whales: 

revised tables by breeding stock (as at 1 May 2005). IWC Paper SC/57/SH11. 15p. 

Baraff L. & Weinrich M. T. (1993) Separation of humpback whale mothers and calves on a 

feeding ground in early autumn. Marine Mammal Science 9, 431-434. 

Baraff L. S., Foy R. J. & Wynne K. M. (2005) Summer distribution and habitat characteristics of 

fin whales and humpback whales in Steller sea lion critical habitat off northeast Kodiak Island, 

2002-2003. p. 241. NOAA. 

Barbieri E. (2009) CONCENTRATION OF HEAVY METALS IN TISSUES OF GREEN 

TURTLES (CHELONIA MYDAS) SAMPLED IN THE CANANEIA ESTUARY, BRAZIL. 

Brazilian Journal of Oceanography 57, 243-248. 

Barendse J., Best P. B., Thornton M., Pomilla C., Carvalho I. & Rosenbaum. H. C. (2010) 

Migration redefined? Seasonality, movements and group composition of humpback whales 

Megaptera novaeangliae off the west coast of South Africa. African Journal of Marine Science

32, 1-22. 

Barko J. W. & Smart R. M. (1981) Comparative influences of light and temperature on the 

growth and metabolism of selected submersed freshwater macrophytes. Ecological Monographs

51, 219-236. 

Barlow J. (1997) Preliminary estimates of cetacean abundance off California, Oregon, and 

Washington based on a 1996 ship survey and comparisons of passing and closing modes. In: 

Administrative Report LJ-97-11. Southwest Fisheries Science Center, National Marine Fisheries 

Service, La Jolla, California. 

Barlow J., Calambokidis J., Falcone E. A., Baker C. S., Burdin A. M., Clapham P. J., Ford J. K. 

B., Gabriele C. M., LeDuc R. G., Mattila D. K., Quinn T. J., Rojas-Bracho L., Straley J. M., 

Taylor B. L., Urban J., Wade R. P., Weller D. W., Witteveen B. H. & Yamaguchi M. (2009) 

Humpback whale abundance in the North Pacific estimated by phographic capture-recapture 

with bias correction from simulation studies. In: 18th Biennial Conference on the Biology of 

Marine Mammals p. 25. 

Barlow J., Forney K. A., Hill P. S., Brownell J., R.L., Carretta J. V., DeMaster D. P., Julian F., 

Lowry M. S., Ragen T. & Reeves R. R. (1997) U.S. Pacific marine mammal stock assessment -

1996. U.S. Department of Commerce, NOAA Technical Memorandum NMFS-SWFSC-248., 

Southwest Fisheries Science Center; La Jolla, California. 

Barnes D. K. (2002) Invasions by marine life on plastic debris. Nature 416. 
379

Barnhart R. A. (1986) Species profiles: Life histories and environmental requirements of coastal 

fishes and invertebrates (Pacific Southwest): Steelhead. In: Technical Report. Humboldt State 

University, Arcada, California. 

Barraclough W. E. (1964) Contribution to the marine life history of the eulachon Thaleichthys 

pacificus. Journal of the Fisheries Research Board of Canada 21, 1333-1337. 

Barrett-Lennard L. G. (2000) Population structure, and mating patterns of killer whales as 

revealed by DNA analysis. University of British Columbia, Vancouver, British Columbia. 

Barrett-Lennard L. G. & Ellis G. M. (2001) Population structure and genetic variability in 

northeastern Pacific killer whales: towards an assessment of population viability. In: Research 

Document 2001/065. Department of Fisheries and Oceans Canada, Nanaimo, British Columbia. 

Barrett B. M., Thompson F. M. & Wick S. N. (1984) Adult anadromous investigations: MayOctober 1983. In: Susitna Hydro Aquatic Studies. Alaska Department of Fish and Game. 

Barron M., Heintz G. R. & Krahn M. M. (2003) Contaminant exposure and effects in pinnipeds: 

Implications for Steller sea lion declines in Alaska. Science of the Total Environment 311, 111-

133. 

Barry R. G. (2005) Alpine climate change and cryospheric responses: An introduction. In: 

Climate and hydrology in mountain areas (eds C. d. Jong, D. Collins and R. Ranzi). John Wiley 

and Sons, Ltd, London, England. 

Barss W. H. (1989) Maturity and reproductive cycle for 35 species from the family Scorpaenidae 

found off Oregon. Oregon Department of Fish and Game, Portland, Oregon. 

Bartoli M., Nizzoli D., Viaroli P., Turolla E., Castaldelli G., Fano E. A. & Rossi R. (2001) 

Impact of Tapes philippinarum farming on nutrient dynamics and benthic respiration in the 

Sacca di Goro. Hydrobiologia 455, 203–212. 

Barton B. T. & Roth J. D. (2008) Implications of intraguild predation for sea turtle nest 

protection. Biological Conservation 181, 2139-2145. 

Bartsch P. (1940) The West American Haliotis. Proceedings of the United States National 

Museum 89, 49-58. 

Baskett M. L., Nisbet R. M., Kappel C. V., Mumby P. J. & Gaines S. D. (2010) Conservation 

management approaches to protecting the capacity for corals to respond to climate change: a 

theoretical comparison. Global Change Biology 16, 1229-1246. 

Bass A. L. (1999) Genetic analysis to elucidate the natural history and behavior of hawksbill 

turtles (Eretmochelys imbricata) in the wider Caribbean: a review and re-analysis. Chelonian 

Conservation and Biology 3, 195-199. 

Bass A. L., Epperly S. P., Braun J., Owens D. W. & Patterson R. M. (1998) Natal origin and sex 

ratios of foraging sea turtles in the Pamlico-Albemarle Estuarine Complex. In: NOAA Technical 

Memorandum. U.S. Department of Commerce, National Oceanic and Atmospheric 

Administration, National Marine Fisheries Service, Southeast Fisheries Science Center, Miami, 

Florida. 

Baum E. T. (1997) Maine Atlantic Salmon: A National Treasure. Atlantic Salmon Unlimited, 

Hermon, Maine. 

Baum E. T. & Meister A. L. (1971) Fecundity of Atlantic salmon (Salmo salar) from two Maine 

Rivers. Journal of the Fisheries Research Board of Canada 28, 764-767. 

Baumgartner M. F., Cole T. V. N., Clapham P. J. & Mate B. R. (2003) North Atlantic right 

whale habitat in the lower Bay of Fundy and on the SW Scotian Shelf during 1999-2001. Marine 

Ecology Progress Series 264, 137-154. 
380

Baums I. B., Miller M. W. & Hellberg M. E. (2005) Regionally isolated populations of an 

imperiled Caribbean coral, Acropora palmata. Molecular Ecology 14, 1377-1390. 

Baussant T., Sanni S., Jonsson G., Skadsheim A. & Borseth J. F. (2001) Bioaccumulation of 

polycyclic aromatic compounds: 1. bioconcentration in two marine species and in semipermeable 

membrane devices during chronic exposure to dispersed crude oil. Environmental Toxicology 

and Chemistry 20, 1175-1184. 

Bax N. (2003) Marine invasive alien species: a threat to global biodiversity. Marine Policy 27, 

313-323. 

Bax N., Williamson A., Aguero M., Gonzalez E. & Geeves W. (2003) Marine invasive alien 

species: a threat to global biodiversity. Marine Policy 27, 313-323. 

Bax N. J. (1983) The early marine migration of juvenile chum salmon (Oncorhynchus keta) 

through Hood Canal - its variability and consequences. Ph.D. thesis. University of 

Washington, Seattle, Washington, 196 p. 

Beacham T. D., Hay D. E. & Le K. D. (2005) Population structure and stock identification of 

eulachon (Thaleichthys pacificus), an anadromous smelt, in the Pacific Northwest. Marine 

Biotechnology 7, 363-372. 

Beakes M. P., Satterthwaite W. H., Collins E. M., Swank D. R., Merz J. E., Titus R. G., Sogard 

S. M. & Mangel M. (2010) Smolt Transformation in Two California Steelhead Populations: 

Effects of Temporal Variability in Growth. Transactions of the American Fisheries Society 139, 

1263-1275. 

Beamesderfer R. C. P. & Webb M. A. H. (2002) Green sturgeon status review information. S.P. 

Cramer and Associates, Gresham, Oregon. 

Beardsley R. C., Epstein A. W., Chen C., Wishner K. F., Macaulay M. C. & Kenney R. D. 

(1996) Spatial variability in zooplankton abundance near feeding right whales in the Great South 

Channel. Deep-Sea Research 43, 1601-1625. 

Beauchamp D. A., Shepard M. F. & Pallley G. O. (1983) Species Profiles: Life Histories and 

Environmental Requirements of Coastal Fishes and Invertebrates (Pacific Northwest) - Chinook 

Salmon. FWS/OBS-82/11.6, TR EL-82-4, October 1983. Prepared for the U.S. Department of 

Interior, Fish and Wildlife Service, and the U.S. Army Corps of Engineers, Coastal Ecology 

Group, Waterways Experiment Station. 23p. 

Beaudreau A. H. & Essington T. E. (2007) Spatial, temporal, and ontogenetic patterns of 

predation on rockfishes by lingcod. Transactions of the American Fisheries Society 136, 1438-

1452. 

Bechdel S., Smith B., Atkinson S., Suydam R., Brix K. & O'Corry-Crowe. G. (2010) Genetic 

consequences of population decline in an endangered species: Results on major 

histocompatibility complex (MHC) diversity in Cook Inlet beluga whales. In: Cook Inlet Belgua 

Whale Science Conference p. 3, Anchorage Downtown Marriott, Anchorage, Alaska. 

Beck, A. D. and D. C. Miller. 1974. Analysis of inner plant passage of estuarine biota. Pages 12-

14 Electric power and the civil engineer. Conference papers, Power Division Specialty 

Conference; Boulder, Colorado. American Society of Civil Engineers; United Engineers Center, 

New York. 

Becker P. (2010) Current-use and legacy persistent pollutants in Cook Inlet beluga whales: 

Results from the analysis of banked tissues from the Alaska Marine Mammal Tissue Archival 

Project (AMMTAP). In: Cook Inlet Belgua Whale Science Conference p. 4, Anchorage 

Downtown Marriott, Anchorage, Alaska. 
381

Becker P. R., Krahn M. M., Mackey E. A., Demiralp R., Schantz M. M., Epstein M. S., Donais 

M. K., Porter B. J., Muir D. C. G. & Wise S. A. (2000a) Concentrations of Polychlorinated 

Biphenyls (PCB’s), chlorinated pesticides and heavy metals and other elements in tissues of 

beluga whale, Delphinapterus leucas, from Cook Inlet, Alaska. Marine Fisheries Review 62, 81-

98. 

Becker P. R., Krahn M. M., Mackey E. A., Demiralp R., Schantz M. M., Epstein M. S., Donais 

M. K., Porter B. J., Muir D. C. G. & Wise. S. A. (2000b) Concentrations of polychlorinated 

biphenyls (PCB's) chlorinated pesticides, and heavy metals and other elements in tissues of 

belugas, Delphinapterus leucas, from Cook Inlet, Alaska. Marine Fisheries Review 62, 81-98. 

Beckmen K. B., Duffy L. K., Zhang X. & Pitcher K. W. (2002) Mercury concentrations in the 

fur of Steller sea lions and northern fur seals from Alaska. Marine Pollution Bulletin 44, 1130-

1135. 

Beland K. (1984) Strategic plan for management of Atlantic salmon in the state of Maine. 

Atlantic Sea Run Salmon Commission, Bangor, Maine. 

Beland K. F. (1996) The relationship between redd counts and Atlantic salmon (Salmo salar) 

parr populations in the Dennys River, Maine. Canadian Journal of Fisheries and Aquatic 

Sciences 53, 513–519. 

Beland K. F., Jordan R. M. & Meister A. L. (1982) Water depth and velocity preferences of 

spawning Atlantic salmon in Maine rivers. North American Journal of Fisheries Management 2, 

11-13. 

Beland P., Guise S. D. & Plante. R. (1992) Toxicology and pathology of St. Lawrence marine 

mammals. Report (unpublished) prepared by St. Lawrence National Institute for Ecotoxicology. 

95 pp. Wildlife Toxicology Fund, World Wildlife Fund, Toronto. Available from World Wildlife 

Fund, 90 Eglington Avenue East, #504 Toronto, Ontario M4P 2Z7, Canada. 

Belitz K., Hamlin S. N., Burton C. A., Kent R., Fay R. G. & Johnson T. (2004) Water Quality in 

the Santa Ana Basin, California, 1999-2001. p. 47. U.S. Department of the Interior, Reston 

Virginia. 

Bell L. A. J., Fa'Anunu U. & Koloa T. (1994) Fisheries resources profiles: Kingdom of Tonga. 

Honiara, Solomon Islands. 

Bellido J. J., Báez J. C., Castillo J. J., Pinto F., Martín J. J., Mons J. L. & Real R. (2010) 

Loggerhead strandings and captures along the southern Spanish Coast: Body size–based 

differences in natural versus anthropogenic injury. Chelonian Conservation and Biology 9, 276-

282. 

Bemis W. E. & Canard B. (1997) Sturgeon rivers: an introduction to acipenseriform 

biogeography and life history. Environmental Biology of Fishes 48, 167-183. 

Ben-Haim Y. & Rosenberg E. (2002) A novel Vibrio sp. pathogen of the coral Pocillopora 

damicornis. Marine Biology 141, 47-55. 

Bennett E. R., Ross P. S., Huff D., Alaee M. & Letcher R. J. (2009) Chlorinated and brominated 

organic contaminants and metabolites in the plasma and diet of a captive killer whale (Orcinus 

orca). Marine Pollution Bulletin 58, 1078-1083. 

Benson A. J., Boydstun, C.P. (1995) Invasion of the zebra mussel into the United States. . In: 

LaRoe, E.T., Farris, G.S., Puckett, C.E., Doran, P.D., Mac, M.J., (Eds.), Our Living Resources: 

A Report to the Nation on the Distribution, Abundance, and Health of U.S. Plants, Animals and 

Ecosystems. 
382

Benson S. R., Croll D. A., Marinovic B. B., Chavez F. P. & Harvey J. T. (2002) Changes in the 

cetacean assemblage of a coastal upwelling ecosystem during El Niño 1997-98 and La Niña 

1999. Progress in Oceanography 54, 279-291. 

Benson S. R., Dutton P. H., Hitipeuw C., Samber B., Bakarbessy J. & Parker D. (2007a) Postnesting migrations of leatherback turtles (Dermochelys coriacea) from Jamursba-Medi, Bird's 

Head Peninsula, Indonesia. Chelonian Conservation and Biology 6, 150-154. 

Benson S. R., Eguchi T., Foley D. G., Forney K. A., Bailey H., Hitipeuw C., Samber B. P., 

Tapilatu R. F., Rei V., Ramohia P., Pita J. & Dutton P. H. (2011) Large-scale movements and 

high-use areas of western Pacific leatherback turtles,Dermochelys coriacea. Ecosphere 2, art84. 

Benson S. R., Forney K. A., Harvey J. T., Carretta J. V. & Dutton P. H. (2007b) Abundance, 

distribution, and habitat of leatherback turtles (Dermochelys coriacea) off California, 1990-2003. 

Fishery Bulletin 105, 337-347. 

Benson S. R., Kisokau K. M., Ambio L., Rei V., Dutton P. H. & Parker D. (2007c) Beach use, 

internesting movement, and migration of leatherback turtles, Dermochelys coriacea, nesting on 

the north coast of Papua New Guinea. Chelonian Conservation and Biology 6, 7-14. 

Berg L. (2001) Yakima Subbasin Summary (Draft). Northwest Power Planning Council, 

Portland, Oregon. 

Bergstrom M. A., Evrard L. M. & Mensinger A. F. (2008) Distribution, abundance, and range of 

the round goby, Apollina melanostoma, in the Duluth-Superior Harbor and St. Louis River 

Estuary, 1998-004. Journal of Great Lakes Research 34, 535-543. 

Berkeley S. A., Hixon M. A., Larson R. J. & Love M. S. (2004) Fisheries Sustainability via 

protection of age structure and spatial distribution of fish populations. Fisheries 29, 23-32. 

Bernardo J. & Plotkin P. T. (2007) An evolutionary perspective on the arribada phenomenon, 

and reproductive behavioral polymorphism of olive ridley sea turtles (Lepidochelys olivacea). In: 

Biology and conservation of Ridley sea turtles (ed P. T. Plotkin) pp. 59-87. Johns Hopkins 

University Press, Baltimore, Maryland. 

Berndt M. P., Hatzell H. H., Crandall C. A., Turtora M., Pittman J. R. & Oaksford E. T. (1998) 

Water Quality in the Georgia-Florida Coastal Plain, Georgia and Florida, 1992-96 

In: U.S. Geological Survey Circular 1151 p. 39. 

Bertness M. D. (1984) Habitat and Community Modification by An Introduced Herbivorous 

Snail. Ecology 65, 370-381. 

Bertness M. D. (1999) The ecology of Atlantic shorelines. Sinauer Associates, Inc., Sunderland, 

Massachusetts. 

Best P. B., Bannister J., Brownell R. L. & Donovan G. (2001a) Right whales: Worldwide status. 

Best P. B., Branadâo A. & Butterworth D. S. (2001b) Demographic parameters of southern right 

whales off South Africa. Journal of Cetacean Research and Management, 161-169. 

Beverly S., Curran D., Musyl M. & Molony. B. (2009) Effects of eliminating shallow hooks 

from tuna longline sets on target and non-target species in the Hawaii-based pelagic tuna fishery. 

Fisheries Research-Amsterdam 5-Apr, 281-288. 

Beyer W. N. & Day D. (2004) Role of manganese oxides in the exposure of mute swans (Cygnus 

olor) to Pb and other elements in the Chesapeake Bay, USA. Environmental Pollution 129, 229-

235. 

Bickham J. W., Patton J. C. & Loughlin T. R. (1996) High variability for control-region 

sequences in a marine mammal: Implications for conservation and biogeography of Steller sea 

lions (Eumetopias jubatus). Journal of Mammalogy 77, 95-108. 
383

Bigelow H. B. & Schroeder W. C. (1953) Sawfishes, guitarfishes, skates and rays. In: Fishes of 

the Western North Atlantic, Part Two. Memoir (eds J. Tee-Van, C. M. Breder, A. E. Parr, W. C. 

Schroeder and L. P. Schultz) pp. 1-514. Sears Foundation for Marine Research. 

Bigg M. A. (1985) Status of Steller sea lion (Eumetopias jubatus) and California sea lion 

(Zalophus californianus) in British Columbia. Canadian Special Publication: Fisheries and 

Aquatic Sciences 77, 1-20. 

Bigg M. A. (1988) Status of the Steller sea lion, Eumetopias jubatus, in Canada. Canadian Field 

Naturalist 102, 315-336. 

Bigg M. A., Olesiuk P. F., Ellis G. M., Ford J. K. B. & K. C. Balcomb I. (1990) Social 

organization and genealogy of resident killer whales (Orcinus orca) in the coastal waters of 

British Columbia and Washington State. Report of the International Whaling Commission, 

Special Issue 12, 383-405. 

Binckley C. A., Spotila J. R., Wilson K. S. & Paladino F. V. (1998) Sex determination and sex 

ratios of Pacific leatherback turtles, Dermochelys coriacea. Copeia 2. 

Birkeland C. (1977) The importance of rate of biomass accumulation in early successional 

successes of benthic communities to the survival of coral recruits. Proceedings of the 3rd 

International Coral Reef Symposium 1, 15-21. 

Bishop D. M., Carstensen R. L. & Bishop G. H. (1989) Report on environmental studies 

concerning the proposed Haines airport reconstruction: [Revision of Section A, Hydrology], 

second phase. Alaska Department of Transportation and Public Facilities, Juneau, Alaska. 

Bisson P., Sullivan A. K. & Nielsen J. L. (1988) Channel hydraulics, habitat use, and body form 

of juvenile coho salmon, steelhead, and cutthroat trout in streams. Transactions of the American 

Fisheries Society 117, 262-273. 

Bjorkstedt E., Spence B. C., Garza J. C., Hankin D. G., Fuller D., Jones W. E., Smith J. J. & 

Macedo R. (2005) An analysis of historical population structure for evolutionarily significant 

units of Chinook salmon, coho salmon, and steelhead in the north-Central California Coast 

Recovery Domain. 

Bjorndal K. A. (1982) The consequences of herbivory for the life history pattern of the 

Caribbean green turtle, Chelonia mydas. In: Biology and Conservation of Sea Turtles (ed K. A. 

Bjorndal) pp. 111-116. Smithsonian Institution Press, Washington D.C. 

Bjorndal K. A. (1997a) Foraging ecology and nutrition of sea turtles. In: The Biology of Sea 

Turtles pp. 199–231. CRC Press, Boca Raton, Florida. 

Bjorndal K. A. (1997b) Foraging ecology, and nutrition of sea turtles. In: The biology of sea 

turtles (eds P. L. Lutz and J. A. Musick) pp. 199-232. CRC Press, Boca Raton, Florida. 

Bjorndal K. A. & Bolten A. B. (2000) Proceedings on a workshop on accessing abundance and 

trends for in-water sea turtle populations. NOAA. 

Bjorndal K. A. & Bolten A. B. (2010) Hawksbill sea turtles in seagrass pastures: success in a 

peripheral habitat. Marine Biology 157, 135-145. 

Bjorndal K. A., Bolten A. B. & Chaloupka M. Y. (2000) Green turtle somatic growth model: 

evidence for density dependence. Ecological Applications 10, 269-282. 

Bjorndal K. A., Bolten A. B. & Chaloupka M. Y. (2003) Survival probability estimates for 

immature green turtles Chelonia mydas in the Bahamas. Marine Ecology Progress Series 252, 

273-281. 

Bjorndal K. A., Bolten A. B. & Chaloupka M. Y. (2005) Evaluating trends in abundance of 

immature green turtles, Chelonia mydas, in the greater Caribbean. Ecological Applications 15, 

304-314. 
384

Bjornn T., Craddock D. & Corley D. (1968) Migration and survival of Redfish Lake, Idaho, 

sockeye salmon, Oncorhynchus nerka. Transactions of the American Fisheries Society 97, 360-

373. 

Black B. A. (2009) Climate-driven synchrony across tree, bivalve, and rockfish growthincrement chronologies of the northeast Pacific. Marine Ecology-Progress Series 378, 37-46. 

Blahm T. H. & McConnell R. J. (1971) Mortality of adult eulachon (Thaleichthys pacificus) 

subjected to sudden increases in water temperature. Northwest Science 45, 178-182. 

Blanchet S., Paez D., Bernatchez L. & Dodson J. J. (2008) An integrated comparison of wild and 

hatchery born Atlantic salmon: implications for conservation plans. Biological Conservation

141, 1989-1999. 

Blecha J. B., Steinbeck J. R. & Sommerville D. C. (1992) Aspects of the biology of the black 

abalone (Haliotis cracherodii) near Diablo Canyon, central California. In: Abalone of the world: 

Biology, fisheries, and culture pp. 225-237. Blackwell Scientific Publications, Inc., Cambridge, 

Massachusetts. 

Bley P. W. & Moring J. R. (1988a) Freshwater and ocean survival of Atlantic salmon and 

steelhead: a synopsis. In: Biological report. Maine Cooperative Fish and Wildlife Research Unit, 

Orono, Maine. 

Bley P. W. & Moring J. R. (1988b) Freshwater and ocean survival of Atlantic salmon and 

steelhead: A synopsis. Maine Cooperative Fish and Wildlife Research Unit. 

Blomberg G., Simenstad C. & Hickey P. (1988) Changes in Duwamish River estuary habitat 

over the past 125 years. Proceedings of the First Annual Meeting on Puget Sound Research 2:, 

437-454. 

Blum J. P. (1988) Assessment of factors affecting sockeye salmon (Oncorhynchus nerka) 

production in Ozette Lake, WA. University of Washington, Seattle, Washington. 

Blumenthal J. M., Austin T. J., Bell C. D. L., Bothwell J. B., Broderick A. C., Ebanks-Petrie G., 

Gibb J. A., Luke K. E., Olynik J. R., Orr M. F., Solomon J. L. & Godley B. J. (2009) Ecology of 

Hawksbill Turtles, Eretmochelys imbricata, on a Western Caribbean Foraging Ground. 

Chelonian Conservation and Biology 8, 1-10. 

Boehlert G. W. (1980) Size composition, age composition, and growth of canary rockfish, 

Sebastes pinniger, and splitnose rockfish, S. diploproa, from the 1977 rockfish survey. Marine 

Fisheries Review 42, 57-63. 

Boehlert G. W. & Kappenman R. F. (1980) Variation of growth with latitude in two species of 

rockfish (Sebastes pinniger and S. diploproa) from the northeast Pacific Ocean. Marine Ecology 

Progress Series 3, 1-10. 

BOEMRE. (2010) Gulf of Mexico region-spills = 50 barrels (2,100 gallons) - 2004 - Hurricane 

Ivan. Bureau of Ocean Energy Management, Regulation and Enforcement Offshore Energy and 

Minerals Management. 

Bogan A. E. (1993) Freshwater bivalve extinctions (Mollusca: Unionoida): A search for causes. 

American Zoologist 33, 599-609. 

Bolen L. E. (1997) Growth dynamics of the seagrass Halophila johnsonii from a subtropical 

estuarine lagoon in southeastern Florida, USA. Florida Atlantic University, Boca Raton, Florida. 

Bolten A. B., K.A. Bjorndal & Martins H. R. (1994) Life history model for the loggerhead sea 

turtle (Caretta caretta) populations in the Atlantic: Potential impacts of a longline fishery. . 

Pp.48-55 In: Balazs, G.J. and S.G. Pooley (eds), Research Plan to Assess Marine Turtle 

Hooking Mortality: Results of an Expert Workshop Held in Honolulu, Hawaii, November 16-18, 
385

1993. U.S. Department of Commerce, NOAA Technical Memorandum NMFS-SEFSC-201, 

pp.48-55. 

Bolten A. B., Wetherall J. A., Balazs G. H. & Pooley S. J. (1996) Status of marine turtles in the 

Pacific Ocean relevant to incidental take in the Hawaii-based pelagic longline fishery. In: NOAA 

Technical Memorandum 

Boness D. J. (1990) Fostering behavior in Hawaiian monk seals: Is there a reproductive cost? 

Behavioral Ecology and Sociobiology 27, 113-122. 

Bonnot P. (1928) The sea lions of California. California Fish and Game 14, 1-16. 

Bonnot P. & Ripley W. E. (1948) The California sea lion census for 1947. California Fish and 

Game 34, 89-92. 

Boomer R. (1995) Letter to R. Gustafson, NMFS, from R. Boomer, Project Leader, Western 

Washington Fishery Resource Office, Fish and Wildlife Service, Quilcene National Fish 

Hatchery Log Books, 1920-1958. In: West coast sockeye salmon administrative record, 

Environmental and Technical Services Division, National Marine Fisheries Service, Portland, 

Oregon. 

BOR. (2007) Dataweb Bureau of Reclamation. 

Boreman J. (1997) Sensitivity of North American sturgeons and paddlefish to fishing mortality. 

Environmental Biology of Fishes 48, 399-405. 

Bortleson G. C., Chrzastowski M. J. & Helgerson A. K. (1980) Historical changes of shoreline 

and wetland at eleven major deltas in the Puget Sound region, Washington. In: U.S. Geological 

Survey, Hydrologic Investigations Atlas HA-617. U.S. Department of Justice and Bureau of 

Indian Affairs, Renton, Virginia. 

Bortleson G. C. & Ebbert J. C. (2000) Occurrence of pesticides in streams and ground water in 

the Puget Sound basin, Washington, and British Columbia, 1996-98. United States Geological 

Survey, Tacoma, Washington. 

Bottom D. L., Jones K. K., Cornwell T. J., Gray A. & Simenstad C. A. (2005) Patterns of 

Chinook salmon migration and residency in the Salmon River estuary (Oregon). Estuarine 

Coastal and Shelf Science 64, 79-93. 

Bouchard S., Moran K., Tiwari M., Wood D., Bolten A., Eliazar P. & Bjorndal K. (1998) Effects 

of exposed pilings on sea turtle nesting activity at Melbourne Beach, Florida. Journal of Coastal 

Research 14, 1343-1347. 

Boudouresque C.-F. (2002) The spread of a non-native marine species, Caulerpa taxifolia. 

Impact on the Mediterranean biodiversity and possible economic consequences. In: di Castro, F. 

and Balaji, V. Tourism, Biodiversity and Information, 75-87. 

Boughton D. A., Adams P. B., Anderson E., Fusaro C., Keller E., Kelley E., Lentsch L., Nielsen 

J., Perry K., Regan H., Smith J., Swift C., Thompson L. & Watson F. (2007) Viability criteria for 

steelhead of the south-central and southern California coast. NOAA-NMFS-SWFSC. 

Boulon Jr. R. H. (1980) Patterns of coral community structure and species diversity of a 

submerged shelf-edge reef off southwestern Puerto Rico. University of Puerto Rico. 

Boulon R. H., Jr. (1994) Growth rates of wild juvenile hawksbill turtles, Eretmochelys imbricata, 

in St. Thomas, United States Virgin Islands. Copeia 1994, 811-814. 

Bourgeois S., Gilot-Fromont E., Viallefont A., Boussamba F. & Deem S. L. (2009) Influence of 

artificial lights, logs and erosion on leatherback sea turtle hatchling orientation at Pongara 

National Park, Gabon. Biological Conservation 142, 85-93. 

Bowen B. W., Abreu-Grobois F. A., Balazs G. H., Kamezaki N., Limpus C. J. & Ferl. R. J. 

(1995) TransPacific migrations of the loggerhead turtle (Caretta caretta) demonstrated with 
386

mitochondrial DNA markers. Proceedings of the National Academy of Sciences USA 92, 3731-

3734. 

Bowen B. W., Bass A. L., Chow S.-M., Bostrom M., Bjorndal K. A., Bolten A. B., Okuyama T., 

Bolker B. M., Epperly S., LaCasella E., Shaver D., Dodd M., Hopkins-Murphy S. R., Musick J. 

A., Swingle M., Rankin-Baransky K., Teas W., Witzell W. N. & Dutton P. H. (2004) Natal 

homing in juvenile loggerhead turtles (Caretta caretta). Molecular Ecology 13, 3797–3808. 

Bowen B. W., Bass A. L., Garcia Rodriguez A., Diez C. E., Van Dam R., Bolten A. B., Bjorndal 

K. A., Miyamoto M. M. & Feri R. J. (1996) Origin of hawksbill turtles in a Caribbean feeding 

area as indicated by genetic markers. Ecological Applications 6, 566-572. 

Bowen B. W., Bass A. L., Soares L. & Toonen R. J. (2005) Conservation implications of 

complex population structure lessons from the loggerhead turtle (Caretta caretta). Molecular 

Ecology 14, 2389-2402. 

Bowen B. W., Grant W. S., Hillis-Starr Z., Shaver D. J., Bjorndal K. A., Bolten A. B. & Bass A. 

L. (2007) Mixed stock analysis reveals the migrations of juvenile hawksbill turtles 

(Eretmochelys imbricata) in the Caribbean Sea. Molecular Ecology 16, 49-60. 

Bowen W. D. & Northridge S. (2010) Morphometrics, age estimation, and growth. In: Marine 

Mammal Ecology and Conservation: A Handbook of Techniques (eds I. L. Boyd, W. D. Bowen 

and S. J. Iverson) pp. 98-118. Oxford University Press. 

Bower S. M. (2006) Synopsis of infectious diseases and parasites of commercially exploited 

shellfish: sabellid polychaete infestation disease in abalone. 

Bower S. M. (2009) Synopsis of infectious diseases and parasites of commercially exploited 

shellfish: withering syndrome of abalone. 

Bowlby C. E., Green G. A. & Bonnell M. L. (1994) Observations of leatherback turtles offshore 

of Washington and Oregon. Northwestern Naturalist 75, 33-35. 

Boyd I. L. (2002) Antarctic marine mammals. In: Encyclopedia of Marine Mammals. W. F. 

Perrin, B. Wursig AND J. G. M. Thewissen (eds.). p.30-36. Academic Press, San Diego, CA. 

1414pgs. 

Boye T. K., Simon M. & Madsen P. T. (2010) Habitat use of humpback whales in 

Godthaabsfjord, West Greenland, with implications for commercial exploitation. Journal of the 

Marine Biological Association of the United Kingdom in press, in press. 

Boyett H. V., Bourne D. G. & Willis B. L. (2007) Elevated temperature and light enhance 

progression and spread of black band disease on staghorn corals of the Great Barrier Reef. 

Marine Biology 151, 1711-1720. 

Boyle M. C., FitzSimmons N. N., Limpus C. J., Kelez S., Velez-Zuazo X. & Waycott M. (2009) 

Evidence for transoceanic migrations by loggerhead sea turtles in the southern Pacific Ocean. 

Proceedings of the Royal Society B-Biological Sciences 276, 1993-1999. 

Braham H. W. (1984) Review of reproduction in the white whale, Delphinapterus leucas, 

narwhal, Monodon monoceros, and Irrawaddy dolphin, Orcaella brevirostris, with comments on 

stock assessment. Report of the International Whaling Commission Special Issue 6, 81-89. 

Reproduction in Whales, Dolphins and Porpoises. Proceedings of the Conference Cetacean 

Reproduction Estimating Parameters For stock Assessment and Management. 

Braje T. J., Erlandson J. M., Rick T. C., Dayton P. K. & Hatch M. B. A. (2009) Fishing from 

past to present: continuity and resilience of red abalone fisheries on the Channel Islands, 

California. Ecological Applications 19, 906-919. 

Brandon R. (1978) Adaptation and evolutionary theory. Studies in the History and Philosophy of 

Science 9, 181-206. 
387

Brashares J. S. (2003) Ecological, behavioral, and life-history correlates of mammal extinctions 

in West Africa. Conservation Biology 17, 733-743. 

Braund S. R. & Huntington. H. (2010) Relationship between the native village of Tyonek, 

Alaska and beluga whales in Cook Inlet, Alaska. In: Cook Inlet Belgua Whale Science 

Conference p. 10, Anchorage Downtown Marriott, Anchorage, Alaska. 

Bräutigam A. & Eckert K. L. (2006a) Turning the tide: Exploitation, trade, and management of 

marine turtles in the Lesser Antilles, Central America, Colombia, and Venezuela. TRAFFIC 

International, Cambridge, United Kingdom. 

Bräutigam A. & Eckert K. L. (2006b) Turning the tide: Exploitation, trade, and management of 

marine turtles in the Lesser Antilles, Central America, Colombia, and Venezuela. TRAFFIC 

International, Cambridge, United Kingdom. 

Brennan J. S., Higgins K. F., Cordell J. R. & Stamatiou V. A. (2004) Juvenile salmon 

composition, timing, distribution, and diet in the marine nearshore waters of central Puget Sound 

in 2001-2002 King County Department of Natural Resources and Parks, Seattle, Washington, 

164 p. 

Breton-Provencher M. (1980) Survey of the beluga population in the Poste-de-la-Baleine region 

(New Quebec). Unpublished paper to the IWC Scientific Committee. 60 pp. Cambridge, June 

(SC/32/SM16). 

Brett, J. R. 1958. Implications and assessments of environmental stress, Pages 69-83 in P. A. 

Larkin, ed. H.R. MacMillan lectures in fisheries: the investigation of fish-power problems. 

Vancouver, British Columbia, Institute of Fisheries, University of British Columbia. 

Brett J. R. (1979) Environmental factors and growth. in W.S. Hoar, D.J. Randall, and J.R. Brett 

(editors) Fish Physiology. Academic Press, New York. 

Brewer S., Watson J., Christensen D. & Brocksmith R. (2007) Hood Canal and Eastern Strait of 

Juan de Fuca summer chum salmon recovery plan. Hood Canal Coordinating Council. 

Bricker S., Longstaff B., Dennison W., Jones A., Boicourt K., Wicks C. & Woerner J. (2007) 

Effects of nutrient enrichment in the nation’s estuaries: A decade of change. National Centers for 

Coastal Ocean Science, Silver Spring, Maryland. 

Bright C. (1999) Invasive species: Pathogens of globalization. Foreign Policy 116, 50-60. 

Brito J. L. (1998) The marine turtle situation in Chile. In: Seventeenth Annual Symposium on Sea 

Turtle Biology and Conservation. (eds S. P. Epperly and J. Braun) pp. 12-15. 

Britton-Simmons K. H. (2004) Direct and indirect effects of the introduced alga Sargassum 

muticum on benthic, subtidal communities of Washington State, USA. Marine Ecology-Progress 

Series 277, 61-78. 

Brock K. A., Reece J. S. & Ehrhart L. M. (2009) The Effects of Artificial Beach Nourishment on 

Marine Turtles: Differences between Loggerhead and Green Turtles. Restoration Ecology 17, 

297-307. 

Broderick A., Frauenstein C. R., Glen F., Hays G. C., Jackson A. L., Pelembe T., Ruxton G. D. 

& Godley B. J. (2006) Are green turtles globally endangered? Global Ecology and Biogeography

15, 21-26. 

Broderick A., Glen C. F., Godley B. J. & Hays G. C. (2002) Estimating the number of green and 

loggerhead turtles nesting annually in the Mediterranean. Oryx 36, 227-235. 

Brodeur R. D., Daly E. A., Benkwitt C. E., Morgan C. A. & Emmett R. L. (2010) Catching the 

prey: Sampling juvenile fish and invertebrate prey fields of juvenile coho and Chinook salmon 

during their early marine residence. Fisheries Research. 
388

Brodeur R. D., Frost B. W., Hare S. R., Francis R. C. & Jr. W. J. I. (1996a) Interannual 

variations in zooplankton biomass in the Gulf of Alaska, and covariation with California Current 

zooplankton biomass. CalCOFI Reports 37, 80-99. 

Brodeur R. D., Frost B. W., Hare S. R., Francis R. C. & W.J. Ingraham J. (1996b) Interannual 

variations in zooplankton biomass in the Gulf of Alaska, and covariation with California Current 

zooplankton biomass. CalCOFI Reports 37, 80-99. 

Brodeur R. D. & Ware D. M. (1992) Long-term variability in zooplankton biomass in the 

subarctic Pacific Ocean. Fisheries Oceanography 1, 32-38. 

Brodie P. F. (1971) A reconsideration of aspects of growth, reproduction, and behavior of the 

white whale (Delphinapterus leucas), with reference to the Cumberland Sound, Baffin Island, 

population. Journal of the Fisheries Research Board of Canada 28, 1309-1318. 

Brodie P. F., Parsons J. L. & Sergeant. D. E. (1981) Present status of the white whale 

(Delphinapterus leucas) in Cumberland Sound, Baffin Island. Report of the International 

Whaling Commission 31, 579-582.-Sc/532/Sm517). 

Brosse L., Dumont P., Lepage M. & Rochard E. (2002) Evaluation of a gastric lavage method for 

sturgeons. North American Journal of Fisheries Management 22, 955 - 960. 

Brothers T. S. & Spingarn A. (1992) Forest fragmentation and alien plant invasion of central 

Indiana old-growth forests. Conservation Biology 6, 91-100. 

Brown A. V., Jackson D. C., Brown K. B. & Pierson W. K. (2005) Lower Mississippi River and 

its tributaries. In: Rivers of North America (eds A. C. Benke and C. E. Cushing) pp. 231-280. 

Elsevier, New York. 

Brown C. J., Fulton E. A., Hobday A. J., Matear R. J., Possingham H. P., Bulman C., 

Christensen V., Forrest R. E., Gehrke P. C., Gribble N. A., Griffiths S. P., Lozano-Montes H., 

Martin J. M., Metcalf S., Okey T. A., Watson R. & Richardson A. J. (2009) Effects of climatedriven primary production change on marine food webs: implications for fisheries and 

conservation. Global Change Biology 16, 1194-1212. 

Brown D. W., Chan S. L., Friedman A. J., Grams K. L., Burrows D. G. & MacLeod Jr. W. D. 

(1985) Bioaccumulation study: Organic compounds in sediment and biota from Grays Harbor 

and reference area. Final report to the U.S. Army Corps of Engineers, Seattle, Washington. 

Brown L. R., Moyle P. B. & Yoshiyama R. M. (1994) Historical decline and current status of 

coho salmon in California. North American Journal of Fisheries Management 14, 237-261. 

Brown M., Brault W. S., Hamilton P. K., Kenney R. D., Knowlton A. R., Marx M. K., Mayo C. 

A., Slay C. K. & Kraus S. D. (2001) Sighting heterogeneity of right whales in the western North 

Atlantic: 1980-1992. Journal of Cetacean Research and Management (Special Issue) 2, 245-250. 

Brown M. E., Maclaine J. & Flagg L. (2006) Androscoggin River anadromous fish restroration 

program. State of Maine Department of marine resources, Augusta, Maine. 

Brown R. F., Riemer S. D. & Wright B. E. (2002) Population status and food habits of Steller sea 

lions in Oregon. In: Contract F0225A-01. Oregon Department of Fish and Wildlife. 

Browning C. L., Rolland R. M. & Kraus S. D. (2009) Estimated calf and perinatal mortality in 

western North Atlantic right whales (Eubalaena glacialis). Marine Mammal Science 26, 648-662. 

Bruckner A. W. (2002) Proceedings of the Caribbean Acropora Workshop: Potential Application 

of the U.S. Endangered Species Act as a Conservation Strategy. In: NOAA Technical 

Memorandum p. 199. U.S. Department of Commerce, Silver Spring, MD. 

Brundage III H. M. & O'Herron Biological and Environmental Consulting. (2006) Final Report 

of Shortnose Sturgeon Population Studies in the Delaware River, January 1999 through March 

2003. p. 13. NOAA Fisheries, NJ Division of Fish and Wildlife. 
389

Bruno J. (2008) Macroalgae in the Keys top-down vs bottom-up. In: Discussion board posting 

on Coral-List 

Bruno J. F. (2000) Facilitation of Cobble Beach Plant Communities through Habitat 

Modification by Spartina alterniflora. Ecology 81, 1179-1192. 

Bruno J. F. & Kennedy C. W. (2000) Patch-Size Dependent Habitat Modification and 

Facilitation on New England Cobble Beaches bySpartina alterniflora. Oecologia 122, 98-108. 

Bugoni L., Krause L. & Petry M. V. (2001) Marine debris and human impacts on sea turtles in 

southern Brazil. Marine Pollution Bulletin 42, 1330-1334. 

Buhle E. R., Holsman K. K., Scheuerell M. D. & Albaugh A. (2009) Using an unplanned 

experiment to evaluate the effects of hatcheries and environmental variation on threatened 

populations of wild salmon. Biological Conservation 142, 2449-2455. 

Bulleri F., Airoldi L., Branca G. M. & Abbiati M. (2005) Positive effects of the introduced green 

alga, Codium fragile ssp. tomentosoides, on recruitment and survival of mussels. Marine Biology

148, 1213-1220. 

Burek K., Dushane J. & Goertz. C. (2010) Morbidity and mortality trends in stranded Cook Inlet 

beluga whales. In: Cook Inlet Belgua Whale Science Conference p. 12, Anchorage Downtown 

Marriott, Anchorage, Alaska. 

Burgner R. L. (1991) The life history of sockeye salmon (Oncorhynchus nerka). In: Life history 

of Pacific salmon (eds C. Groot and L. Margolis) pp. 3-117. University of British Columbia 

Press, Vancouver, British Columbia, Canada. 

Burns J. J. & Seaman G. A. (1986a) Investigations of beluga whales in coastal waters of western 

and northern Alaska. II Biology and ecology. In: OCSEAP Final Rep. pp. 221-357. 

Burns J. J. & Seaman G. A. (1986b) Investigations of belukha whales in coastal waters of 

western and northern Alaska. Part II. Biology and ecology. NOAA Outer Continental Shelf 

Environmental Assessment Program. 

Burreson E. M. & Ford S. E. (2004) A review of recent information on the Haplosporidia, with 

special reference toHaplosporidium nelsoni(MSX disease). Aquatic Living Resources 17, 499-

517. 

Burreson E. M., Stokes N. A. & Friedman C. S. (2000) Increased Virulence in an Introduced 

Pathogen:Haplosporidium nelsoni(MSX) in the Eastern OysterCrassostrea virginica. Journal of 

Aquatic Animal Health 12, 1-8. 

Burton C. A., Izbicki J. A. & Paybins K. S. (1998) Water-quality trends in the Santa Ana River 

at MWD Crossing and below Prado Dam, Riverside County, California. In: Water-Resources 

Investigations U.S. Geological Survey Sacramento, California. 

Busby P. J., Wainwright T. C., Bryant G. J., Lierheimer L. J., Waples R. S., Waknitz F. W. & 

Lagomarsine I. V. (1996a) Status review of steelhead from Washington, Oregon, and California. 

In: NOAA Technical Memorandum. U.S. Department of Commerce, Northwest Fisheries 

Science Center, Seattle, Washington. 

Busby P. J., Wainwright T. C., Bryant G. J., Lierheimer L. J., Waples R. S., Waknitz F. W. & 

Lagomarsine I. V. (1996b) Status review of steelhead from Washington, Oregon, and California. 

U.S. Department of Commerce, National Marine Fisheries Service, Northwest Fisheries Science 

Center, Seattle, Washington, NOAA Technical Memorandum NMFS-NWFSC-27. 

Busby P. J., Wainwright T. C., Bryant G. J., Lierheimer L. J., Waples R. S., Waknitz F. W. & 

Lagomarsino I. V. (1996c) Status review of steelhead from Washington, Oregon, and California. 

In: NOAA Technical Memorandum NMFS-NWFSC-27. U.S. Department of Commerce, National 

Marine Fisheries Service, Northwest Fisheries Science Center, Seattle, Washington. 
390

Busby P. J., Wainwright T. C., Bryant G. J., Lierheimer L. J., Waples R. S., Waknitz F. W. & 

Lagomarsino I. V. (1996d) Status Review of West Coast Steelhead from Washington, Idaho, 

Oregon, and California. U.S. Department of Commerce NOAA Technical Memorandum NMFSNWFSC-27, 275p. 

Butterman W. C. & Hilliard H. E. (2005) Mineral Commodity Profiles: Silver. . In: Open-File 

Report U.S. Department of the Interior, U.S. Geological Survey, Reston, Virginia. 

Byles R. A. (1989a) Distribution, and abundance of Kemp's ridley sea turtle, Lepidochelys 

kempii, in Chesapeake Bay and nearby coastal waters. In: First International Symposium on 

Kemp's Ridley Sea Turtle Biology, Conservation and Management (eds C. W. Caillouet Jr. and 

A. M. Landry Jr.) p. 145. 

Byles R. A. (1989b) Satellite telemetry of Kemp's ridley sea turtle Lepidochelys kempii in the 

Gulf of Mexico. In: Proceedings of the Ninth Annual Workshop on Sea Turtle Conservation and 

Biology. NOAA Technical Memorandum NMFS-SEFC-232 (eds S. A. Eckert, K. L. Eckert and T. 

H. Richardson) pp. 25-26. 

Byles R. A. & Plotkin P. T. (1994) Comparison of the migratory behavior of the congeneric sea 

turtles Lepidochelys olivacea and L. kempii. In: Thirteenth Annual Symposium on Sea Turtle 

Biology and Conservation (eds B. A. Schroeder and B. E. Witherington) p. 39. 

Byles R. A. & Swimmer Y. B. (1994) Post-nesting migration of Eretmochelys imbricata in the 

Yucatan Peninsula. In: Fourteenth Annual Symposium on Sea Turtle Biology and Conservation

(eds K. A. Bjorndal, A. B. Bolten, D. A. Johnson and P. J. Eliazar) p. 202. 

Byrne R., Fish J., Doyle T. K. & Houghton J. D. R. (2009) Tracking leatherback turtles 

(Dermochelys coriacea) during consecutive inter-nesting intervals: Further support for direct 

transmitter attachment. Journal of Experimental Marine Biology and Ecology 377, 68-75. 

Byrnes J. E., Reynolds P. L. & Stachowicz J. J. (2007) Invasions and extinctions reshape coastal 

marine food webs. Plos One 2, e295. 

CACSS. (1988) Restoring the balance. In: Annual Report p. 84 California Advisory Committee 

on Salmon and Steelhead Trout, California Department of Fish and Game, Sacramento, 

California. 

Cailliet G. M., Andrews A. H., Burton E. J., Watters D. L., Kline D. E. & Ferry-Graham L. A. 

(2001) Age determination and validation studies of marine fishes: do deep-dwellers live longer? 

pp. 739-764. 

Cailliet G. M., Burton E. J., Cope J. M. & Kerr L. A. (2000) Biological characteristics of 

nearshore fishes of California: A review of existing knowledge. Pacific States Marine Fisheries 

Commission, California Department of Fish and Game 

Caillouet C. C., Fontaine T., Manzella-Tirpak S. A. & Williams T. D. (1995) Growth of headstarted Kemp's ridley sea turtles (Lepidochelys kempii) following release. Chelonian 

Conservation and Biology 1, 231-234. 

Caillouet Jr. C. W., Fontaine C. T., Manzella-Tirpak S. A. & Williams T. D. (1995) Growth of 

head-started Kemp's ridley sea turtles (Lepidochelys kempii) following release. Chelonian 

Conservation and Biology 1, 231-234. 

Cain D. J., Carter J. L., Fend S. V., Luoma S. N., Alpers C. N. & Taylor H. E. (2000) Metal 

exposure in a benthic macroinvertebrate, Hydropsyche californica, related to mine drainage in 

the Sacramento River. Canadian Journal of Fisheries and Aquatic Sciences 57, 380-390. 

Cairns S. D. (1982) Stony corals (Cnidaria: Hydrozoa, Scleractinia) of Carrie Bow Cay, Belize. 

Smithson Contributions to Marine Science 12, 271-302. 
391

Calambokidis J. (1997) Humpback whales and the California - Costa Rica connection. Whales. 

Journal of the Oceanic Society Fall 1997, 4-10. 

Calambokidis J., Barlow J., Ford J. K. B., Chandler T. E. & Douglas A. B. (2009) Insights into 

the population structure of blue whales in the Eastern North Pacific from recent sightings and 

photographic identification. Marine Mammal Science 25, 816-832. 

Calambokidis J., Falcone E. A., Quinn II T. J., Burdin A. M., Clapham P. J., Ford J. K. B., 

Gabriele C. M., LeDuc R. G., Mattila D. K., Rojas-Bracho L., Straley J. M., Taylor B. L., Urbân 

J., Weller R. D. W., Witteveen B. H., Yamaguchi M., Bendlin A., Camacho D., Flynn K. R., 

Havron A., Huggins J. & Maloney N. (2008) SPLASH: Structure of populations, levels of 

abundance, and status of humpback whales in the North Pacific. In: Final report prepared by 

Cascadia Research for. U.S. Department of Commerce, National Oceanic and Atmospheric 

Administration, National Marine Fisheries Service, Seattle, Washington. 

Calambokidis J., Peard J., Steiger G. H., Cubbage J. C. & Delong. R. L. (1984) Chemical 

contaminants in marine mammals from Washington State. NOAA Technical Memorandum 

NOS-OMS-6. 167pp. 

Calambokidis J., Steiger G. H., Straley J. M., Herman L. M., Cerchio S., Salden D. R., Urban J., 

Jacobsen J. K., von Ziegesar O., Balcomb K. C., Gabriele C. M., Dahlheim M. E., Uchida S., 

Ellis G., Miyamura Y., de Guevara P. L., Yamaguchi M., Sato F., Mizroch S. A., Schlender L., 

Rasmussen K., Barlow J. & Quinn T. J. (2001a) Movements and population structure of 

humpback whales in the North Pacific. Marine Mammal Science 17, 769-794. 

Calambokidis J., Steiger G. H., Straley J. M., Herman L. M., Cerchio S., Salden D. R., Urban J. 

R., Jacobsen J. K., Von Ziegesar O., Balcomb K. C., Gabriele C. M., Dahlheim M. E., Uchida S., 

Ellis G. M., Miyamura Y., Ladrón de Guevara P., Yamaguchi M., Sato F., Mizroch S. A., 

Schlender L., Rasmussen K., Barlow J. & Quinn T. J. I. (2001b) Movements and population 

structure of humpback whales in the North Pacific. Marine Mammal Science 17, 769-794. 

Calambokidis J., Steiger G. H., Straley J. M., Quinn T., Herman L. M., Cerchio S., Salden D. R., 

Yamaguchi M., Sato F., Urban J. R., Jacobson J., von Zeigesar O., Balcomb K. C., Gabriele C. 

M., Dahlheim M. E., Higashi N., Uchida S., Ford J. K. B., Miyamura Y., Ladrón de Guevara P., 

Mizroch S. A., Schlender L. & Rasmussen K. (1997) Abundance and population structure of 

humpback whales in the North Pacific basin. Final Report under contract No. 5ABNF500113. 

NMFS Southwest Fisheries Science Center; La Jolla, California. 

Calkins D. (1989a) Status of belukha whales in Cook Inlet. In: Gulf of Alaska, Cook Inlet, and 

North Aleutian Basin information update meeting (eds L. E. Jarvela and L. K. Thorsteinson) pp. 

109-112. U.S. Department of Commerce, NOAA, Outer Continental Shelf Environmental 

Assessment Program, Anchorage, Alaska. 

Calkins D. G. (1983) Susitna hydroelectric project phase II annual report: Big game studies. In: 

Volume IX Beluga whale. Alaska Department of Fish and Game, Anchorage, Alaska. 

Calkins D. G. (1984) Susitna hydroelectric project final report. In: Volume IX beluga whale pp. 

1-17. Alaska Department of Fish and Game, Anchorage, Alaska. 

Calkins D. G. (1986) Marine Mammals. In The Gulf of Alaska, Physical Environment and 

Biological Resources, D.W. Hood and S.T. Zimmerman (editors), Government Printing Office, 

Washington, D.C. p.527−558. 

Calkins D. G. (1989b) Status of beluga whales in Cook Inlet. In: NOAA, OCSEAP Gulf of 

Alaska, Cook Inlet, and North Aleutian Basin information update meeting pp. 109-112, 

Anchorage, Alaska.
392

Calkins D. G. (1996) Movements and habitat use of female Steller sea lions in Southeastern 

Alaska. In: Steller sea lion recovery investigations in Alaska, 1992-1994 pp. 110-134. Alaska 

Department of Fish and Game Wildlife. 

Calkins D. G., Becker E., Spraker T. R. & Loughlin T. R. (1994) Impacts on Steller sea lions. 

Pp.119-139 In: Loughlin, T.R. (Ed), Marine Mammals and the Exxon Valdez. Academic Press, 

Inc. San Diego, CA. 

Calkins D. G., Becker E. F. & Pitcher K. W. (1998) Reduced body size of female Steller sea 

lions from a declining population in the Gulf of Alaska. Marine Mammal Science 14, 232-244. 

Calkins D. G. & Goodwin E. (1988) Investigation of the declining sea lion population in the Gulf 

of Alaska. In: Unpublished Report. Alaska Department of Fish and Game, Anchorage, Alaska. 

Calkins D. G. & Pitcher K. W. (1982a) Population assessment, ecology and trophic relationships 

of Steller sea lions in the Gulf of Alaska. In: Final Report pp. 445-546. U.S. Department of the 

Interior OCSEAP 

Calkins D. G. & Pitcher K. W. (1982b) Population assessment, ecology, and trophic 

relationships of Steller sea lions in the Gulf of Alaska. Pages 447-546 in: Environmental 

assessment of the Alaskan continental shelf. U.S. Department of Commerce and U.S. Department 

of the Interior, Final Report of the Principal Investigators 19, 1-565. 

Call K. A. & Loughlin T. R. (2005) An ecological classification of Alaskan Steller sea lion 

(Eumetopias jubatus) rookeries: A tool for conservation/management. Fisheries and 

Oceanography 14, 212-222. 

Callaway J. C. & Josselyn M. N. (1992) The introduction and spread of smoth cordgrass 

(Spartina alterniflora) in South San Francisco Bay. Estuaries 15, 218-226. 

Campbell C. L. & Lagueux C. J. (2005) Survival probability estimates for large juvenile and 

adult green turtles (Chelonia mydas) exposed to an artisanal marine turtle fishery in the western 

Caribbean. Herpetologica 61, 91-103. 

Canada F. a. O. (2008) Draft management plan for the Steller sea lion (Eumetopias jubatus) in 

Canada. Species at Risk Act management plan series. Fisheries and Oceans Canada, Ottawa. 

Canning-Clode J. o., Fowler A. E., Byers J. E., Carlton J. T. & Ruiz G. M. (2011) ‘Caribbean 

creep’ chills out: Climate change and marine invasive species. Plos One 6, e29657. 

Cannon A. C. & Flanagan J. P. (1996) Trauma and treatment of Kemp's ridley sea turtles caught 

on hook-and-line by recreational fisherman. In: Sea turtles biology and conservation workshop. 

Cardillo M. (2003) Biological determinants of extinction risk: Why are smaller species less 

vulnerable? Animal Conservation 6, 63-69. 

Cardillo M., Mace G. M., Jones K. E. & Bielby J. (2005) Multiple causes of high extinction risk 

in large mammal species. Science 309, 1239-1241. 

Cardona L., Aguilar A. & Pazos L. (2009) Delayed ontogenic dietary shift and high levels of 

omnivory in green turtles (Chelonia mydas) from the NW coast of Africa. Marine Biology 156, 

1487-1495. 

Cardona L., Campos P., Levy Y., Demetropoulos A. & Margaritoulis D. (2010) Asynchrony 

between dietary and nutritional shifts during the ontogeny of green turtles (Chelonia mydas) in 

the Mediterranean. Journal of Experimental Marine Biology and Ecology in press, in press. 

Carlson H. R. & Straty R. R. (1981) Habitat and nursery grounds of Pacific rockfish, Sebastes 

spp., in rocky coastal areas of southeastern Alaska. Marine Fisheries Review 43, 13-19. 

Carlson J. K., Osborne J. & Schmidt T. W. (2007) Monitoring the recovery of smalltooth 

sawfish, Pristis pectinata, using standardized relative indices of abundance. Biological 

Conservation 136, 195-202. 
393

Carlton J. & Geller J. (1993) Ecological Roulette: The Global Transport of Nonindigenous 

Marine Organisms. Science 261, 78-82. 

Carlton J. T. (1996) Pattern, process, and prediction in marine invasion ecology. Biological 

Conservation 78, 97-106. 

Carlton J. T. (1999) Molluscan invasions in marine and estuarine communities. Malacologia 41, 

439-454. 

Carlton J. T. (2001) Introduced species in U.S. coastal waters: Environmental impacts and 

management priorities. Pew Oceans Commission, Arlington, Virginia. 

Carman S. M., Janssen J., Jude D. J. & Berg M. B. (2006) Diel interactions between prey 

behaviour and feeding in an invasive fish, the round goby, in a North American river. 

Freshwater Biology 51, 742-755. 

Carpenter K. E., Abrar M., Aeby G., Aronson R. B., Banks S., Bruckner A., Chiriboga A., Cortés 

J., Delbeek J. C., DeVantier L., Edgar G. J., Edwards A. J., Fenner D., Guzmán H. M., 

Hoeksema B. W., Hodgson G., Johan O., Licuanan W. Y., Livingstone S. R., Lovell E. R., 

Moore J. A., Obura D. O., Ochavillo D., Polidoro B. A., Precht W. F., Quibilan M. C., Reboton 

C., Richards Z. T., Rogers A. D., Sanciangco J., Sheppard A., Sheppard C., Smith J., Stuart i., 

Turak E., Veron J. E. N., Wallace C., Weil E. & Wood E. (2008) One-third of reef-building 

corals face elevated extinction risk from climate change and local impacts. Science 321, 560-563. 

Carpenter S. R. (1987) Complex interactions in lake communities. Springer-Verlag, New York. 

Carr A. & Caldwell D. K. (1956) The ecology, and migrations of sea turtles: 1. Results of field 

work in Florida, 1955. American Museum Novitates 1793, 1-23. 

Carr A., Carr M. H. & Meylan A. B. (1978) The ecology and migration of sea turtles, 7. the west 

Caribbean turtle colony. Bulletin of the American Museum of Natural History, New York 162, 1-

46. 

Carr A. F. (1986) RIPS, FADS, and little loggerheads. Bioscience 36, 92-100. 

Carr M. H. (1983) Spatial and temporal patterns of recruitment of young-of-the-year rockfishes 

(Genus Sebastes) into a central California kelp forest. In: Moss Landing Marine Laboratories 

San Francisco State University. 

Carretta J. V., Barlow J., Forney K. A., Muto M. M. & Baker J. (2001) U.S. Pacific Marine 

Mammal Stock Assessments: 2001. U.S. Department of Commerce. 

Carretta J. V. & Forney K. A. (1993) Report of the two aerial surveys for marine mammals in 

California coastal waters utilizing a NOAA DeHavilland twin otter aircraft: March 9-April 7, 

1991 and February 8-April 6, 1992. NMFS, SWFSC. 

Carretta J. V., Forney K. A., Lowry M. S., Barlow J., Baker J., Johnston D., Hanson B. & Muto 

M. M. (2008) Draft U.S. Pacific marine mammal stock assessments: 2008. Southwest Fisheries 

Science Center. 

Carretta J. V., Forney K. A., Muto M. M., Barlow J., Baker J., Hanson B. & Lowry M. S. (2005) 

U.S. Pacific marine mammal stock assessments - 2004. In: U.S. Department of Commerce, 

NOAA-TM-NMFS-SWFSC-375. 

Carretta J. V., Forney K. A., Muto M. M., Barlow J., Baker J. & Lowry. M. (2004) U.S. Pacific 

Marine Mammal Stock Assessments: 2003. NOAA Technical Memorandum NMFS-SWFSC358. 295p. 

Carrillo E., Webb G. J. W. & Manolis S. C. (1999) Hawksbill Turtles (Eretmochelys imbricata) 

in Cuba: An Assessment of the Historical Harvest and its Impacts. Chelonian Conservation and 

Biology 3, 264-280. 
394

Carruthers E. H., Schneider D. C. & Neilson J. D. (2009) Estimating the odds of survival and 

identifying mitigation opportunities for common bycatch in pelagic longline fisheries. Biological 

Conservation 142, 2620-2630. 

Carter B. T. G. & Nielsen. E. A. (2010) Exploring ecological changes in Cook Inlet beluga whale 

habitat through traditional and local ecological knowledge. In: Cook Inlet Belgua Whale Science 

Conference p. 16, Anchorage Downtown Marriott, Anchorage, Alaska. 

Carter J. L. & Resh V. H. (2005) Pacific Coast rivers of the coterminous United States. In: Rivers 

of North America (eds A. C. Benke and C. E. Cushing) pp. 541-590. Elsevier Academic Press, 

Burlington, Massachusetts. 

Carvalho I., Brito C., Santos M. E. D. & Rosenbaum H. C. (2011) The waters of Sao Tome: A 

calving ground for West African humpback whales? African Journal of Marine Science 33, 91-

97. 

Casale P., d'Astore P. P. & Argano R. (2009a) Age at size and growth rates of early juvenile 

loggerhead sea turtles (Caretta caretta) in the Mediterranean based on length frequency analysis. 

Herpetological Journal 19, 29-33. 

Casale P., Mazaris A. D., Freggi D., Vallini C. & Argano R. (2009b) Growth rates and age at 

adult size of loggerhead sea turtles (Caretta caretta) in the Mediterranean Sea, estimated through 

capture-mark-recapture records. Scientia Marina 73, 589-595. 

Casale P., Nicolosi P., Freggi D., Turchetto M. & Argano R. (2003) Leatherback turtles 

(Dermochelys coriacea) in Italy and in the Mediterranean basin. Herpetological Journal 13, 135-

139. 

Casillas E., Crockett L., deReynier Y., Glock J., Helvey M., Meyer B., Schmitt C., Yoklavich 

M., Bailey A., Chao B., Johnson B. & Pepperell T. (1998) Essential fish habitat west coast 

National Marine Fisheries Service, Seattle, Washington. 

Castellini M. A. (1999) Assessing heavy metals in populations of marine mammals. In: EPA 

Symposium on Western Ecological Systems. April, 1999, San Francisco. 

Castilla J. C., Lagos N. A. & Cerda M. (2004) Marine ecosystem engineering by the alien 

ascidian Pyura praeputialis on a mid-intertidal rocky shore. Marine Ecology-Progress Series

268, 119–130. 

Caswell H., Fujiwara M. & Brault S. (1999) Declining survival probability threatens the North 

Atlantic right whales. Proceedings of the National Academy of Sciences of the United States of 

America. 96, 3308-3313. 

Catry P., Barbosa C., Paris B., Indjai B., Almeida A., Limoges B., Silva C. & Pereira H. (2009) 

Status, Ecology, and Conservation of Sea Turtles in Guinea-Bissau. Chelonian Conservation and 

Biology 8, 150-160. 

Caurant F., Bustamante P., Bordes M. & Miramand P. (1999) Bioaccumulation of cadmium, 

copper and zinc in some tissues of three species of marine turtles stranded along the French 

Atlantic coasts. Marine Pollution Bulletin 38, 1085-1091. 

Caut S., Guirlet E. & Girondot M. (2009a) Effect of tidal overwash on the embryonic 

development of leatherback turtles in French Guiana. Marine Environmental Research 69, 254-

261. 

Caut S., Guirlet E. & Girondot M. (2009b) Effect of tidal overwash on the embryonic 

development of leatherback turtles in French Guiana. Marine Environmental Research in press, 

in press. 
395

CBD. (2001a) Petition to list the Southern Resident killer whale (Orcinus orca) as an endangered 

species under the Endangered Species Act. p. 108. Center for Biological Diversity, Berkeley, 

California. 

CBD. (2001b) Petition to list the southern resident killer whale (Orcinus orca) as an endangered 

species under the Endangered Species Act. Center for Biological Diversity, Berkeley, California. 

CBFWA. (1990) Review of the history, development, and management of anadromous fish 

production facilities in the Columbia River basin. Columbia Basin Fish and Wildlife Authority, 

Portland, Oregon. 

CDFG. (1993) California fishery data for black abalone 1956-1993. In: Unpublished data. 

CDFG. (1998) Report to the Fish and Game Commission: A status review of the spring-run 

Chinook salmon (Oncorhynchus tshawytscha) in the Sacramento River drainage. California 

Department of Fish and Game, Sacramento, CA. 

CDFG. (2002) California Department of Fish and Game comments to NMFS regarding green 

sturgeon listing. California Department of Fish and Game. 

Cech J. J. J., Doroshov S. I., Moberg G. P., May B. P., Schaffter R. G. & Kohlhorst D. M. (2000) 

Biological assessment of green sturgeon in the Sacramento-San Joaquin watershed (phase 1). 

CALFED Bay-Delta Program. 

Cech Jr. J. J., Doroshov S. I., Moberg G. P., May B. P., Schaffter R. G. & Kohlhorst D. M. 

(2000) Biological assessment of green sturgeon in the Sacramento-San Joaquin watershed (phase 

1). In: Fina Report. CALFED Bay-Delta Program. 

Celik A., Kaska Y., Bag H., Aureggi M., Semiz G., Kartal A. A. & Elci L. (2006) Heavy metal 

monitoring around the nesting environment of green sea turtles in Turkey. Water Air and Soil 

Pollution 169, 67-79. 

Cerchio S., Collins T., Strindberg S., Bennett C. & Rosenbaum H. (2010) Humpback whale 

singing activity off northern Angola: An indication of the migratory cycle, breeding habitat and 

impact of seismic surveys on singer number in Breeding Stock B1. p. 21. IWC Scientific 

Committee, Agadir, Morocco. 

Cerchio S., Jacobsen J. K., Cholewiak D. M. & Falcone E. A. (2005) Reproduction of female 

humpback whales off the Revillagigedo Archipelago during a severe El Niño event. In: Sixteenth 

Biennial Conference on the Biology of Marine Mammals p. 55, San Diego, California. 

CETAP. (1982) A characterization of marine mammals and turtles in the mid- and north Atlantic 

areas of the U.S. outer continental shelf. Cetacean and Turtle Assessment Program, University of 

Rhode Island. Final Report #AA551-CT558-548 to the Bureau of Land Management, 

Washington, DC, 538 pp. 

CFEPTAP. (2004) Fisheries Ecosystem Planning for Chesapeake Bay. Chesapeake Fisheries 

Ecosystem Plan Technical Advisory Panel. 

Chacón-Chaverri D. & Eckert K. L. (2007) Leatherback sea turtle nesting at Gandoca Beach in 

Caribbean Costa Rica: Management recommendations from fifteen years of conservation. 

Chelonian Conservation and Biology 6, 101-110. 

Chacón Chaverri D. (1999) Anidacíon de la tortuga Dermochelys coriacea (Testudines: 

Dermochelyidae) en playa Gandoca, Costa Rica (1990 a 1997). Revista de Biologia Tropical 47, 

225-236. 

Chacon D. (2002) Assessment about the trade of sea turtles and their products in the Central 

American isthmus. p. 247 pages. Red Regional para la Conservación de last Tortugas Marinas 

en Centroamérica, San José, Costa Rica. 
396

Chaloupka M. (2001) Historical trends, seasonality, and spatial synchrony in green sea turtle egg 

production. Biological Conservation 101, 263-279. 

Chaloupka M., Bjorndal K. A., Balazs G. H., Bolten A. B., Ehrhart L. M., Limpus C. J., 

Suganuma H., Troeeng S. & Yamaguchi M. (2008a) Encouraging outlook for recovery of a once 

severely exploited marine megaherbivore. Global Ecology and Biogeography 17, 297-304. 

Chaloupka M., Bjorndal K. A., Balazs G. H., Bolten A. B., Ehrhart L. M., Limpus C. J., 

Suganuma H., Troeሷng S. & Yamaguchi M. (2007) Encouraging outlook for recovery of a once 

severely exploited marine megaherbivore. Global Ecology and Biogeography Dec 2007. 

Available online at http://www.cccturtle.org/pdf/Chaloupka_et_alGEB2007.pdf. Accessed 

12/31/2007. 

Chaloupka M. & Limpus C. (2005) Estimates of sex- and age-class-specific survival 

probabilities for a southern Great Barrier Reef green sea turtle population. Marine Biology 146, 

1251-1261. 

Chaloupka M., Limpus C. & Miller J. (2004) Green turtle somatic growth dynamics in a spatially 

disjunct Great Barrier Reef metapopulation. Coral Reefs 23, 325-335. 

Chaloupka M. Y., Kamezaki N. & Limpus C. (2008b) Is climate change affecting the population 

dynamics of the endangered Pacific loggerhead sea turtle? Journal of Experimental Marine 

Biology and Ecology 356, 136-143. 

Chaloupka M. Y. & Limpus C. J. (1997) Robust statistical modelling of hawksbill sea turtle 

growth rates (southern Great Barrier Reef). Marine Ecology-Progress Series 146, 1-8. 

Chaloupka M. Y. & Musick J. A. (1997) Age, growth, and population dynamics. In: The biology 

of sea turtles (eds P. L. Lutz and J. A. Musick) pp. 233-273 CRC Press, Boca Raton, Florida. 

Chamberland K., Lindroth B. A. & Whitaker B. (2002) Genotoxicity in Androscoggin River 

Smallmouth Bass. Northeastern Naturalist 9, 203-212. 

Chan E. H. (2006) Marine turtles in Malaysia: on the verge of extinction? Aquatic Ecosystems 

Health and Management 9, 175-184. 

Chan E. H. & Liew H. C. (1996) Decline of the leatherback population in Terengganu, Malaysia, 

1956-1995. Chelonian Conservation and Biology 2, 196-203. 

Chan S. K. F., Cheng I. J., Zhou T., Wang H. J., Gu H. X. & Song X. J. (2007) A comprehensive 

overview of the population and conservation status of sea turtles in China. Chelonian 

Conservation and Biology 6, 185-198. 

Chapman D., Hillman G. T., Deppert D., Erho M., Hays S., Peven C., Suzumoto B. & Klinge R. 

(1994) Status of summer/fall Chinook salmon in the mid-Columbia region. Report to Chelan, 

Douglas, and Grant County PUDs, Boise, Idaho. 

Chapman D. & Witty K. (1993) Habitat of weak salmon stocks in the Snake River basin and 

feasible recovery measures. In: Report to the Bonneville Power Administration, DOE/BP-99654-

1, Portland, Oregon. 

Chapman D. W. & Bjornn T. C. (1969) Distribution of salmonids in streams, with special 

reference to food and feeding. In: Symposium on salmon and trout in streams: H.R. MacMillan 

Lectures in Fisheries, (ed T. G. Northcote) pp. 153-176. University of British Columbia, Institute 

of Fisheries, Vancouver, British Columbia, Canada. 

Chappell M. A. (1980) Possible physiological effects of space shuttle sonic booms on marine 

mammals. In: Potential Effects of Space Shuttle Sonic Booms on the Biota and Geology of the 

California Channel Islands: Research Reports (eds J. J. R.Jehl and C. F. Cooper) pp. 195-228. 
397

Chaput G., Allard J., Caron F., Dempson J. B., Mullins C. C. & O’Connell M. F. (1998) Riverspecific target spawning requirements for Atlantic salmon (Salmo salar) based on a generalized 

smolt production model. Canadian Journal of Fisheries and Aquatic Sciences 55, 246-261. 

Chen T. L., Wise S. S., Kraus S., Shaffiey F., Levine K. M., Thompson W. D., Romano T., 

O'Hara T. & John Pierce Wise S. (2009) Particulate hexavalent chromium is cytotoxic and 

genotoxic to the North Atlantic right whale (Eubalaena glacialis) lung and skin fibroblasts. 

Environmental and Molecular Mutagenesis 50, 387-393. 

Chen Z., Li B., Zhong Y. & Chen J. (2004) Local competitive effects of introduced Spartina 

alterniflora on Scirpus mariqueter at Dongtan of Chongming Island, the Yangtze River estuary 

and their potential ecological consequences. Hydrobiologia 528, 99–106. 

Chen Z. & Yang. G. (2010) Novel CHR-2 SINE subfamilies and t-SINEs identified in cetaceans 

using nonradioactive southern blotting. Genes and Genomics 32, 345-352. 

Cheng I. J., Huang C. T., Hung P. Y., Ke B. Z., Kuo C. W. & Fong C. L. (2009) Ten Years of 

Monitoring the Nesting Ecology of the Green Turtle, Chelonia mydas, on Lanyu (Orchid Island), 

Taiwan. Zoological Studies 48, 83-94. 

Cheung W. W. L., Lam V. W. Y., Sarmiento J. L., Kearney K., Watson R., Zeller D. & Pauly D. 

(2010) Large-scale redistribution of maximum fisheries catch potential in the global ocean under 

climate change. Global Change Biology 16, 24-35. 

Chew K. K. (1990) Global bivalve shellfish introductions. World Aquaculture 21, 9-22. 

Chilcote M. A. & Waterfield M. R. (1995) MERRYMEETING BAY: An Assessment of 

Pollution Levels and Sources. 

Childers A. R., Whitledge T. E. & Stockwell D. A. (2005) Seasonal and interannual variability in 

the distribution of nutrients and chlorophyll a across the Gulf of Alaska shelf: 1998-2000. DeepSea Research II 52, 193-216. 

Chinese Mitten Crab Working Group. (2003) National management plan for the Genus Eriocheir 

(mitten crabs). Aquatic Nuisance Species Task Force. 

Christensen I. (1984) Growth and reproduction of killer whales, Orcinus orca, in Norwegian 

coastal waters. Report of the International Whaling Commission, Special Issue 6, 253-258. 

Christie W. J. (1974) Changes in the fish species composition of the Great Lakes. Journal of the 

Fisheries Research Board of Canada 31, 827-854. 

Christison K. W. & Baker G. C. (2007) First record of Pseudodactylogyrus anguillae (Yin & 

Sproston, 1948) (Monogenea) from South Africa. African Zoology 42, 279-285. 

Ciborowski J. (2007) Indicator: Invasion of zebra mussels (Dreissena polymorpha) and quagga 

mussels (Dreissena bugensis). 205-211. 

Ciguarria J. & Elston R. (1997) Independent introduction of Bonamia ostreae, a parasite of 

Ostrea edulis, to Spain. Diseases of Aquatic Organisms 29, 157-158. 

Clapham P., Mikhalev Y., Franklin W., Paton D., Baker C. S., IVashchenko Y. V. & Robert L. 

Brownell J. (2009) Catches of humpback whales, Megaptera novaeangliae, by the Soviet Union 

and other nations in the Southern Ocean, 1947–1973. Marine Fisheries Review 71, 39-43. 

Clapham P. J. (1994) Maturational changes in patterns of association among male and female 

humpback whales. Journal of Zoology 71, 440-443. 

Clapham P. J. (1996) The social and reproductive biology of humpback whales: an ecological 

perspective. Mammal Review 26, 27-49. 

Clapham P. J. (2002) Are ship-strikes mortalities affecting the recovery of the endangered whale 

populations off North America? European Cetacean Society Newsletter (special issue) 40, 13–

15. 
398

Clapham P. J., Baraff L. S., Carlson C. A., Christian M. A., Mattila D. K., Mayo C. A., Murphy 

M. A. & Pittman S. (1993) Seasonal occurrence and annual return of humpback whales, 

Megaptera novaeangliae, in the southern Gulf of Maine. Canadian Journal of Zoology 71, 440-

443. 

Clapham P. J., Barlow J., Bessinger M., Cole T., Mattila D., Pace R., Palka D., Robbins J. & 

Seton R. (2003) Abundance and demographic parameters of humpback whales in the Gulf of 

Maine, and stock definition relative to the Scotian shelf. Journal of Cetacean Research and 

Management 5, 13-22. 

Clapham P. J. & Mayo C. A. (1987) Reproduction and recruitment of individually identified 

humpback whales, Megaptera novaeangliae, observed in Massachusetts Bay, 1979-1985. 

Canadian Journal of Zoology 65, 2853-2863. 

Clapham P. J. & Mayo C. A. (1990) Reproduction of humpback whales (Megaptera 

novaeangliae) observed in the Gulf of Maine. Report of the International Whaling Commission

Special Issue 12, 171-175. 

Clapham P. J., Young S. B. & Brownell Jr. R. L. (1999) Baleen whales: conservation issues and 

the status of the most endangered populations. Mammal Review 29, 35-60. 

Clark G. H. (1929) Sacramento-San Joaquin salmon (Oncorhynchus tschawytscha) fishery of 

California. California Fish and Game Bulletin 17, 73. 

Clark R. B. (1997) Marine Pollution. Clarendon Press, Oxford, UK. 

Clark S. T., Odell D. K. & Lacinak C. T. (2000) Aspects of growth in captive killer whales 

(Orcinus orca). Marine Mammal Science 16, 110-123. 

Clarke A. D., Lewis A., Telmer K. H. & Shrimpton J. M. (2007) Life history and age at maturity 

of an anadromous smelt, the eulachon Thaleichthys pacificus (Richardson). Journal of Fish 

Biology 71, 1479-1493. 

Clarke D., Dickerson C. & Reine K. (2003) Characterization of underwater sounds produced by 

dredges. In: Third Specialty Conference on Dredging and Dredged Material Disposal, Orlando, 

Florida. 

Clavero M. & Garcıa-Berthou E. (2005) Invasive species are a leading cause of animal 

extinctions. Trends in Ecology and Evolution 20, 110. 

Clemens W. A., Hart J. L. & Wilby G. V. (1936) Analysis of stomach contents of fur seals taken 

off the west coast of Vancouver Island in April and May, 1935. Canadian Department of 

Fisheries, Ottawa, Canada. 

Clemens W. A. & Wilby G. V. (1961) Fishes of the Pacific coast of Canada. Fisheries Research 

Board of Canada. 

Clifton K., Cornejo D. O. & Felger R. (1982) Sea turtles of the Pacific coast of Mexico. In: 

Biology and Conservation of Sea Turtles pp. 199-209. Smithsonian Institution Press, 

Washington, D.C. 

Cochran J. K., Hirschberg D. J., Wang J. & Dere C. (1998) Atmospheric deposition of metals to 

coastal waters (Long Island Sound, New York U.S.A.): evidence from saltmarsh deposits. 

Estuarine, coastal and shelf science 46, 503-522. 

Cohen A. (2010) Non-native bacterial and viral pathogens in ballast water: Potential for impacts 

to ESA-listed species under NOAA’s jurisdiction. NMFS. 

Cohen A. & Webb K. (2002) Biological Invasions in Aquatic Ecosystems: Impacts on 

Restoration and Potential for Control. 

Cohen A. N. (2004) An exotic species detection program for Puget Sound. Puget Sound Action 

Team, Olympia. 
399

Cohen A. N. & Carleton J. T. (1995) Nonindigenous aquatic species in a United States estuary: 

A case study of the biological invasions of the San Francisco Bay and delta. United States Fish 

and Wildlife Service and The National Sea Gran College Program. 

Cohen A. N. & Carlton J. T. (1998) Accelerating invasion rate in a highly invaded estuary. 

Science 279, 555-558. 

Cohen A. N. & Moyle P. B. (2004) Summary of data and analyses indicating that exotic species 

have impaired the beneficial uses of certain California waters. State Water Resources Control 

Board. 

Cohen R. R. H., Dresler P. V., Phillips E. J. P. & Cory R. L. (1984) The effect of the Asiatic 

clam, Corbicula fluminea, on phytoplankton of the Potomac River, Maryland. Limnology and 

Oceanography 29, 170-180. 

Colautti R. I., Grigorovich I. A. & MacIsaac H. J. (2006) Propagule Pressure: A Null Model for 

Biological Invasions. Biological Invasions 8, 1023-1037. 

Colavecchia M., Katopodis C., Goosney R., Scruton D. A. & Mckinley R. S. (1998) 

Measurement of burst swimming performance in wild Atlantic salmon (Salmo salar L.) using 

digital telemetry. Regulated Rivers: Research and Management 14, 41-51. 

Cole T. V. N., Hartley D. L. & Merrick R. L. (2005a) Mortality and serious injury 

determinations for large whales stocks along the eastern seaboard of the United States, 1999-

2003. NOAA Northeast Fisheries Science Center 05-08. 

Cole T. V. N., Hartley D. L. & Merrick R. L. (2005b) Mortality and serious injury 

determinations for North Atlantic Ocean large whale stocks 1999-2003. In: Northeast Fisheries 

Science Center Reference Document p. 18. U.S. Department of Commerce, NOAA, National 

Marine Fisheries Service, Northeast Fisheries Science Center, Woods Hole, MA. 

Cole T. V. N., Hartley D. L. & Merrick R. L. (2005c) Mortality and seriously injury 

determinations for North Atlantic Ocean large whale stocks 1999-2003. Northeast Fisheries 

Science Center Reference Document 05-08, U.S. Department of Commerce, NOAA, National 

Marine Fisheries Service Northeast Fisheries Science Center. Woods Hole, MA. 18p. 

Collard S. B. (1990) Leatherback turtles feeding near a watermass boundary in the eastern Gulf 

of Mexico. Marine Turtle Newsletter 50, 12-14. 

Collard S. B. & Ogren L. H. (1990) Dispersal scenarios for pelagic post-hatchling sea turtles. . 

Bulletin of Marine Science 47, 233-243. 

Collins B. D. & Sheikh A. J. (2005) Historical reconstruction, classification, and change analysis 

of Puget Sound tidal marshes. Washington Department of Natural Resources Aquatic Resources 

Division, Olympia, Washington. 

Collins M. R. (2010) Problems associated with assessing the status of populations of Atlantic and 

shortnose sturgeons in the Southeastern United States. Fisheries 35, 340-341. 

Collins M. R., Norwood C. & Rourk A. (2008) Shortnose and Atlantic Sturgeon Age-Growth, 

Status, Diet, and Genetics (2006-0087-009): October 25, 2006 - June 1, 2008 Final Report. p. 

41. South Carolina Department of Natural Resources. 

Collins M. R., Post W. C., Russ D. C. & Smith T. I. J. (2002) Habitat use and movements of 

juvenile shortnose sturgeon in the Savannah River, Georgia-South Carolina. Transactions of the 

American Fisheries Society 131, 975-979. 

Collins M. R., Rogers S. G. & Smith T. I. J. (1996) Bycatch of sturgeons along the southern 

Atlantic coast of the USA. North American Journal of Fisheries Management 16, 24 - 29. 
400

Collins M. R., Rogers S. G., Smith T. I. J. & Moser M. L. (2000a) Primary factors affecting 

sturgeon populations in the southeastern United States: Fishing mortality and degradation of 

essential habitats. Bulletin of Marine Science 66, 917-928. 

Collins M. R., Rogers S. G., Smith T. I. J. & Moser M. L. (2000b) Primary factors affecting 

sturgeon populations in the southeastern United States: fishing mortality and degradation of 

essential habitats. Bulletin of Marine Science 66, 917-928. 

Colway C. & Stevenson D. E. (2007) Confirmed records of two green sturgeon from the Bering 

Sea and Gulf of Alaska. Northwestern Naturalist 88, 188-192. 

Colwell R. R. (1996) Global climate and infectious disease: the cholera paradigm. Science 274, 

2025-2031. 

Commander U.S. Pacific Fleet. (2003) Report on the results of the inquiry into allegations of 

marine mammal impacts surroundiilg the use of active sonar by USS Shoup (DDG 86) in the 

Haro Strait on or about 5 May 2003. U.S. Navy. 

Commander U.S. Pacific Fleet. (2010) Northwest training range complex environmental impact 

statement/overseas environmental impact statement. U.S. Navy, Silverdale, Washington. 

Compagno L. J. V. & Cook S. F. (1995) The exploitation and conservation of freshwater 

elasmobranches: Status of taxa and prospects for the future. Journal of Aquaculture and Aquatic 

Science 7, 62-90. 

Conant T. A., Dutton P. H., Eguchi T., Epperly S. P., Fahy C. C., Godfrey M. H., MacPherson S. 

L., Possardt E. E., Schroeder B. A., Seminoff J. A., Snover M. L., Upite C. M. & Witherington 

B. E. (2009) Loggerhead sea turtle (Caretta caretta) 2009 status review under the U.S. 

Endangered Species Act. . p. 222. 

Conaway C. H., Squire S., Mason R. P. & Flegal A. R. (2003) Mercury speciation in the San 

Francisco Bay estuary. Marine Chemistry 80, 199-225. 

Connell J. H., Schroeter S. C. & Swarbrick S. (1988) Variations at different scales of time and 

space on intertidal shores. In: The marine environment of Santa Barbara and its coastal waters: 

A symposium workshop p. 3. 

Connor E. F., Hafernik J., Levy J., Moore V. L. & Rickman J. K. (2002) Insect conservation in 

an urban biodiversity hotspot: The San Francisco Bay Area. Journal of Insect Conservation 6, 

247-259. 

Connor W. P., Sneva J. G., Tiffan K. F., Steinhorst R. K. & Ross D. (2005a) Two alternative 

juvenile life history types for fall Chinook salmon in the Snake River basin. Transactions of the 

American Fisheries Society 134, 291-304. 

Connor W. P., Sneva J. G., Tiffan K. F., Steinhorst R. K. & Ross D. (2005b) Two Alternative 

Juvenile Life History Types for Fall Chinook Salmon in the Snake River Basin. Transactions of 

the American Fisheries Society 134, 291-304. 

Consiglieri L. D., Braham H. W., Dahlheim M. E., Fiscus C., McGuire P. D., Peterson C. E. & 

Pippenger. D. A. (1982) Seasonal distribution and relative abundance of marine mammals in the 

Gulf of Alaska. Final Report. Outer Continental Shelf Environmental Assessment Program, 

Research Unit 68. p.189-343. 

Constantine R., Garrigue C., Steel D., Jackson J., Burns D., Clapham P., Hauser N., Mattila D., 

Oremus M., Poole M., Robbins J., Thompson K. & Baker C. S. (2010) Abundance of humpback 

whales in Oceania based on fluke photo-identification and DNA profiling. p. 30. IWC Scientific 

Committee, Agadir, Morocco. 
401

Conversi A., Piontkovski S. & Hameed. S. (2001) Seasonal and interannual dynamics of Calanus 

finmarchicus in the Gulf of Maine (Northeastern US shelf) with reference to the North Atlantic 

Oscillation. Deep Sea Research Part Ii, Topical studies in Oceanography 48(41-43)519-530. 

Coombs C. I. (1979) Reef fishes near Depoe Bay, Oregon: movement and the recreational 

fishery. Oregon State University. 

Cordeiro J. R., Olivero A. P. & Sovell J. (2007) Corbicula Fluminea (Bivalvia: Sphaeriacea: 

Corbiculidae) in Colorado. The Southwestern Naturalist 52, 424-430. 

Corkeron P. J. (1995) Humpback whales (Megaptera novaeangliae) in Hervey Bay, Queensland: 

Behaviour and responses to whale-watching vessels. Canadian Journal of Zoology 73, 1290-

1299. 

Corsolini S., Aurigi A. & Focardi S. (2000) Presence of polychlorobiphenyls (PCBs), and 

coplanar congeners in the tissues of the Mediterranean loggerhead turtle Caretta caretta. Marine 

Pollution Bulletin 40, 952-960. 

Corwith H. L. & Wheeler P. A. (2002) El Niño related variations in nutrient and chlorophyll 

distributions off Oregon. Progress in Oceanography 54, 361-380. 

COSEWIC. (2005) COSEWIC Assessment and update status report on the shortnose sturgeon 

Acipenser brevirostrum in Canada. p. vi + 27. COSEWIC, Committee on the Status of 

Endangered Wildlife in Canada, Ottawa, Canada. 

COSEWIC. (in press) COSEWIC assessment and status report on the canary rockfish Sebastes 

pinniger in Canada. Committee on the Status of Endangered Wildlife in Canada, Ottawa, 

Canada. 

Costello C., Drake J. M. & Lodge D. M. (2007) Evaluating an invasive species policy: Ballast 

water exchange in the Great Lakes. Ecological Applications 17, 655-662. 

Coughran D. & Gales N. (2010) An unusual peak in recorded mortalities of humpback whales in 

Western Australia: Normal stochastic variability or a regional indication of carrying capacity? 

p. 7. IWC Scientific Committee, Agadir, Morocco. 

Courtenay W. R. (1997) Nonindigenous fishes. In: Strangers in Paradise (eds D. Simberloff, D. 

C. Schmitz and T. C. Brown) pp. 109-122. Island Press, Washington, D.C. 

Coutts A. & Dodgshun T. (2007) The nature and extent of organisms in vessel sea-chests: A 

protected mechanism for marine bioinvasions. Marine Pollution Bulletin 54, 875-886. 

Cowan E., Pennell J., Salmon M., Wyneken J., Cowan C. & Broadwell A. (2002) Influence of 

filtered roadway lighting on the seaward orientation of hatchling sea turtles. In: Twentieth 

Annual Symposium on Sea Turtle Biology and Conservation (eds A. Mosier, A. Foley and B. 

Brost) pp. 295-298. 

Cox K. W. (1960) Review of the abalone of California. California Fish and Game Bulletin 46, 

381-406. 

Cox K. W. (1962) California Abalones, family Haliotidae. California Fish and Game Bulletin

118, 1-133. 

Coyle K. O. & Pinchuk A. I. (2003) Annual cycle of zooplankton abundance, biomass and 

production on the northern Gulf of Alaska shelf, October 1997 through October 2000. Fisheries 

Oceanography 12, 327-338. 

Coyne M. & Landry Jr. A. M. (2007) Population sex ratios, and its impact on population models. 

In: Biology and conservation of Ridley sea turtles (ed P. T. Plotkin) pp. 191-211. Johns Hopkins 

University Press, Baltimore, MD. 

Coyne M., Landry Jr. A. M., Costa D. T. & Williams B. B. (1995) Habitat preference, and 

feeding ecology of the green sea turtle (Chelonia mydas) in south Texas waters. In: Twelfth 
402

Annual Workshop on Sea Turtle Biology and Conservation (eds J. I. Richardson and T. H. 

Richardson) pp. 21-24. 

Cozzi B., Panin M., Butti C., Podesta M. & Zotti A. (2010) Bone density distribution patterns in 

the rostrum of delphinids and beaked whales: Evidence of family-specific evolutive traits. The 

Anatomical Record 293, 235-242. 

Craig J. K., Crowder L. B., Gray C. D., McDaniel C. J., Henwood T. A. & Hanifen J. G. (2001) 

Ecological effects of hypoxia on fish, sea turtles, and marine mammals in the northwestern Gulf 

of Mexico. In: Coastal and estuarine studies series (eds N. N. Rabalais and R. E. Turner) pp. 

262-292. American Geophysical Union, Washington, D.C. 

Craig M. P. & Ragen T. J. (1999) Body size, survival, and decline of juvenile Hawaiian monk 

seals, Monachus schauinslandi. Marine Mammal Science 15, 24. 

Crawford W., Cherniawsky J., Foreman M. & Chandler P. (1999) El Niño sea level signal along 

the west coast of Canada. In: Proceedings of the 1998 Science Board Symposium on the impacts 

of the 1997/98 El Niño event on the North Pacific Ocean and its marginal seas (eds H. J. 

Freeland, W. T. Peterson and A. Tyler). North Pacific Marine Science Organization. 

Crawley A., Kline D. I., Dunn S., Anthony K. R. N. & Dove S. (2010) The effect of ocean 

acidification on symbiont photorespiration and productivity in Acropora formosa. Global 

Change Biology 16, 851-863. 

Credland A. G. (2010) Book review: Whales' Bones of Germany, Austria, Czech Republic and 

Switzerland. Nicholas Redman. Redman Publishing. xxvi + 205pp. ISBN 978-095458002-5. 

2009. Mariners Mirror 96, 396 

391. 

Crognale M. A., Eckert S. A., Levenson D. H. & Harms C. A. (2008) Leatherback sea turtle 

Dermochelys coriacea visual capacities and potential reduction of bycatch by pelagic longline 

fisheries. Endangered Species Research 5, 249-256. 

Crooks J. A. (1998) Habitat alteration and community-level effects of an exotic mussel, 

Musculista senhousia. Marine Ecology-Progress Series 162, 137-152. 

Crouse D. T. (1999) Population modeling and implications for Caribbean hawksbill sea turtle 

management Chelonian Conservation and Biology 3, 185-188. 

Crouse O. T., Crowder L. B. & Caswell H. (1987) A site based population model for loggerhead 

sea turtles and implications for conservation. Ecology 68, 1412-1423. 

Crozier L., Zabel R. & Hamlet A. (2008) Predicting differential effects of climate change at the 

population level with life-cycle models of spring Chinook salmon. Global Change Biology 14, 

236-249. 

Cruz R. D. (2002) Marine turtle distribution in the Philippines. In: Western Pacific Sea Turtle 

Cooperative Research and Management Workshop. (ed I. Kinan) pp. 57-66. Western Pacific 

Regional Fishery Management Council, Honolulu, Hawaii. 

CRWC. (2006) Defending the Watershed: Quabbin at Risk, Redux. In: Currents and Eddies pp. 

1-2. Connecticut River Watershed Council. 

Cuddington K. & Hastings A. (2004) Invasive engineers. Ecological Modelling 178, 335-347. 

Cuffney T. R., Meador M. R., Porter S. D. & Gurtz M. E. (1997) Distribution of fish, benthic 

invertebrate, and algal communities in relation to physical and chemical conditions, Yakima 

River Basin, Washington 1990. In: Water Resources- Investigation Report. U.S. Geological 

Survey, Raleigh, North Carolina. 

Cullon D. L., Yunker M. B., Alleyne C., Dangerfield N. J., O'Neill S., Whiticar M. J. & Ross. P. 

S. (2009) Persistent organic pollutants in Chinook salmon (Oncorhynchus tshawytscha): 
403

Implications for resident killer whales of British Columbia and adjacent waters. (Orcinus orca). 

Environmental Toxicology and Chemistry 28, 148-161. 

Culver C. S. & Kuris A. M. (2002) The Introduction of the South African Worm: Biology, 

History, and Implications for Management. In: Biological Invasions in Aquatic Ecosystems: 

Impacts on Restoration and Potential for Control (eds A. N. Cohen and S. K. Webb) pp. 3-7. 

Cummings V. (2002) Sea turtle conservation in Guam. In: Western Pacific Sea Turtle 

Cooperative Research and Management Workshop. (ed I. Kinan) pp. 37-38. Western Pacific 

Regional Fishery Management Council, Honolulu, Hawaii. 

Cummings W. C. (1985) Right whales--Eubalaena glacialis, and Eubalaena australis. In: The 

sirenians and baleen whales (eds S. H. Ridgway and R. Harrison) pp. 275-304. Academic Press, 

New York, NY. 

Cunjak R. A., Prowse T. D. & Parrish D. L. (1998) Atlantic salmon (Salmo salar) in winter: “the 

season of parr discontent”? Canadian Journal of Fisheries and Aquatic Science 55, 161-180. 

Curran M. A. J., Ommen T. D. v., Morgan V. I., Phillips K. L. & Palmer A. S. (2003) Ice core 

evidence for Antarctic sea ice decline since the 1950s. Science 302, 1203-1206. 

Curry R. G. & McCartney M. S. (2001) Ocean gyre circulation changes associated with the 

North Atlantic Oscillation. Journal of Physical Oceanography 31, 3374-3400. 

Dadswell M. J. (1979) Biology and population characteristics of the shortnose sturgeon, 

Acipenser brevirostrum LeSueur 1818 (Osteichthyes: Acipenseridae), in the Saint John River 

Estuary, New Brunswick, Canada. Canadian Journal of Zoology 57, 2186-2210. 

Dadswell M. J. (2006) A Review of the Status of Atlantic Sturgeon in Canada, with Comparisons 

to Populations in the United States and Europe. Fisheries 31, 218-229. 

Dadswell M. J. & Rulifson R. A. (1994) Macrotidal estuaries: a region of collision between 

migratory marine animals and tidal power development. Biological Journal of the Linnean 

Society 51, 93-113. 

Dadswell M. J., Taubert B. D., Squiers T. S., Marchette D. & Buckley J. (1984a) Synopsis of 

biological data on shortnose sturgeon, Acipenser brevirostrum LeSueur 1818. In: NOAA 

Technical Report U.S. Department of Commerce, National Oceanic and Atmospheric 

Administration, National Marine Fisheries Service, Silver Spring, Maryland. 

Dadswell M. J., Taubert B. D., Squiers T. S., Marchette D. & Buckley J. (1984b) Synopsis of 

biological data on shortnose sturgeon, Acipenser brevirostrum LeSueur, 1818. In: NOAA 

Technical Report-14. 

Daehler C. C. & Strong D. R. (1996) Status, prediction and prevention of introduced cordgrass 

Spartina spp. invasions in PAcific estuaries, USA. Biological Conservation 78, 51-58. 

Dahlheim M. E., White P. A. & Waite. J. M. (2008) Cetaceans of southeast Alaska: Distribution 

and seasonal occurrence. Journal of Biogeography 36, 410-426. Special Issue Southeast Alaska 

Marine Ecology and Biogeography. 

Dahm C. N., Edwards R. J. & Gelwick F. P. (2005) Gulf coast rivers of the southwestern United 

States. In: Rivers of North America (eds A. C. Benke and C. E. Cushing) pp. 181-228. Elsevier, 

New York. 

Dalla Rosa L., Secchi E. R., Maia Y. G., Zerbini A. N. & Heide-Jørgensen M. P. (2008) 

Movements of satellite-monitored humpback whales on their feeding ground along the Antarctic 

Peninsula. Polar Biology 31, 771-781. 

Danie D., Trial J., Stanley J., Shanks L. & Benson N. (1984) Species profiles: life histories and 

environmental requirements of coastal fish and invertebrates (North Atlantic): Atlantic salmon. 

USFWS. 
404

Daniel D. O. & Schneeweis J. C. (1992) Steller sea lion, Eumetopias jubatus, predation on 

glaucous-winged gulls, Larus glaucescens. Canadian Field Naturalist 106, 268. 

Daniels R. & Floren R. (1998) Poaching pressures on northern California’s abalone fishery. 

Journal of Shellfish Research 17, 859-862. 

Danil K. & Chivers S. J. (2005) Habitat-based spatial and temporal variability of life history 

characteristics of female common dolphins (Delphinus delphis) in the eastern tropical Pacific. In: 

Sixteenth Biennial Conference on the Biology of Marine Mammals p. 67, San Diego, California. 

Danilewicz D., Tavares M., Moreno I. B., Ott P. H. & Trigo. C. C. (2009) Evidence of feeding 

by the humpback whale (Megaptera novaeangliae) in mid-latitude waters of the western South 

Atlantic. Jmba2 - Biodiversity Records-Published Online 3Pgs. 

Dark T. A., Wilkins M. E. & Edwards K. (1983) Bottom trawl survey of canary rockfish 

(Sebastes pinninger), yellowtail rockfish (S. flavidus), bocaccio (S. paucispinis), and chilipepper 

(S. goodei) off Washington-California, 1980. NMFS, NWFSC and Alaska Fisheries Center, 

Seattle, Washington. 

Darling J. D. & Mori K. (1993) Recent observations of humpback whales (Megaptera 

novaeangliae) in Japanese waters, Ogasawara, and Okinawa. Canadian Journal of Zoology 71, 

325-333. 

Darling J. D. & Morowitz H. (1986) Census of Hawaiian humpback whales (Megaptera 

novaeangliae) by individual identification. Canadian Journal of Zoology 64, 105-111. 

Darnerud P. O. (2003) Toxic effects of brominated flame retardants in man and in wildlife. 

Environmental International 29, 841-853. 

Davenport J. & Balazs G. H. (1991) “Fiery bodies” – are pyrosomas an important component of 

the diet of leatherback turtles? The British Herpetological Society Bulletin 31, 33-38. 

Davenport J., Wrench J., McEvoy J. & Carnacho-Ibar V. (1990) Metal and PCB concentrations 

in the "Harlech" leatherback. Marine Turtle Newsletter 48, 1-6. 

Davis G. E. (1982) A century of natural change in coral distribution at the Dry Tortugas: A 

comparison of reef maps from 1881 and 1976. Bulletin of Marine Science 32, 608-623. 

Davis G. E. (1993) Mysterious demise of southern California black abalone Haliotis cracherodii 

Leach, 1814. Journal of Shellfish Research 12, 183-184. 

Davis G. E. (1996) California Abalone. In: Status and trends of the nations biological resources 

pp. 22-23. U.S. Geological Survey, Biological Resources Division, Reston, Virginia. 

Davis G. E., Haaker P. L. & Richards D. V. (1998) The perilous condition of white abalone 

Haliotis sorenseni, Bartsch, 1940. Journal of Shellfish Research 17, 871-875. 

Davis R. A. & Evans C. R. (1982) Offshore distribution and numbers of white whales in the 

eastern Beaufort Sea and Amundsen Gulf, summer 1981. p. 76. SOHIO Alaska Petroleum Co. 

Dome Petroleum Ltd., Anchorage, Alaska 

Calgary, Alberta. 

Davis R. A. & Finley. K. J. (1979) Distribution, migrations, abundance and stock identity of 

eastern Arctic white whales. Unpublished paper to the IWC Small Cetaceans Sub-Committee. 

Cambridge, (SC/31/SM10). 

Davis R. W., Evans W. E. & Würsig B. (2000) Cetaceans, Sea Turtles and Seabirds in the 

Northern Gulf of Mexico: Distribution, Abundance and Habitat Associations. Volume II: 

Technical Report. p. 346. Texas A&M, Galveston. 

Dawes C. J., Lobban C. S. & Tomasko D. A. (1989) A comparison of the physiological ecology 

of the seagrasses Halophila decipiens Ostenfeld and H. johnsonii Eiseman from Florida. Aquatic 

Botany 33, 149-154. 
405

De Weede R. E. (1996) The impact of seaweed introductions on biodiversity. Global 

Biodiversity 6, 2-9. 

de Wit C. A. (2002) An overview of brominated flame retardants in the environment. 

Chemosphere 46, 583-624. 

De’ath G., Lough J. M. & Fabricius K. E. (2009) Declining coral calcification on the Great 

Barrier Reef. Science 323, 116-119. 

Deacutis C. F. & Ribb R. C. (2002) Ballast water and introduced species:Management options 

for Narragansett Bay and Rhode Island. Narragansett Bay Estuary Program, R.I. Department of 

Environmental Management. 

Dean R. J. & Durako M. J. (2007) Physiological integration in the threatened seagrass: Halophila 

johnsonii Eiseman. Bulletin of Marine Science 81, 21-35. 

Dean T. A., Schroeter S. C. & Dixon J. D. (1983) Effects of grazing by two species of sea 

urchins (Strongylocentrotus franciscanus and Lytechinus anamesus) on recruitment and survival 

of two kelp species (Macrocystis pyrifera and Pterygophora californica). . Marine Biology 78, 

301-313. 

Deem S. L., Boussamba F., Nguema A. Z., Sounguet G., Bourgeois S., Cianciolo J. & Formia A. 

(2007) Artificial lights as asignificant cause of morbidity of leatherback sea turtles in Pongara 

National Park, Gabon. Marine Turtle Newsletter 116, 15-17. 

Deem S. L., Norton T. M., Mitchell M., Segars A., Alleman A. R., Cray C., Poppenga R. H., 

Dodd M. & Karesh W. B. (2009) COMPARISON OF BLOOD VALUES IN FORAGING, 

NESTING, AND STRANDED LOGGERHEAD TURTLES (CARETTA CARETTA) ALONG 

THE COAST OF GEORGIA, USA. journal of wildlife diseases 45, 41-56. 

DeLacy A. C., Hitz C. R. & Dryfoos R. L. (1964) Maturation, gestation, and birth of rockfish 

(Sebastodes) from Washington and adjacent waters. In: Fisheries Research Papers pp. 51-67. 

Washington Department of Fisheries. 

Delacy A. C., Miller B. S. & Borton S. F. (1972) Checklist of Puget Sound fishes. University of 

Washington, Seattle, Washington. 

Demko D. B. & Cramer S. P. (2000) Effects of pulse flows on juvenile chinook migration in the 

Stanislaus River. p. 48. Report of S.P. Cramer & Associates, Inc. to South San Joaquin 

Irrigation District, Manteca, California and Oakdale Irrigation District, Oakdale, California. 

DeMott G. E. (1983) Movement of tagged lingcod and rockfishes off Depoe Bay, Oregon. 

Oregon State University. 

Deng X. (2000) Artificial reproduction and early life stages of the green surgeon (Acipenser 

medirostris). University of California, Davis. 

DeVries R. J. (2006) Population dynamics, movements, and spawning habitat of the shortnose 

sturgeon, Acipenser brevirostrum, in the Altamaha River. University of Georgia. 

DFO. (2008) Eulachon integrated fisheries management plan April 1, 2008 to March 31, 2009. 

Department of Fisheries and Oceans Canada. 

Diaz-Fernandez R., Okayama T., Uchiyama T., Carrillo E., Espinosa G., Marquez R., Diez C. E. 

& Koike H. (1999) Genetic sourcing for the hawksbill turtle, Eretmochelys imbricata, in the 

Northern Caribbean Region. Chelonian Conservation and Biology 3, 296-300. 

Didham R., Tylianakis J., Gemmell N., Rand T. & Ewers R. (2007) Interactive effects of habitat 

modification and species invasion on native species decline. Trends in Ecology & Evolution 22, 

489-496. 

Dill H. R. & Bryan W. A. (1912) Report on an expedition to Laysan Island in 1911. U.S. 

Department of Agriculture Biological Survey Bulletin 42, 1-30. 
406

Dimick R. E. & Merryfield F. (1945) The fishes of the Willamette River system in relation to 

pollution. Oregon State College Engineering Experiement Station. Bulletin Series No. 20. 

Dobbs F. C. & Rogerson A. (2005) Ridding ships' ballast water of microorganisms. 

Environmental 

Science & Technology 39, 259A-264A. 

Doblin M. A., Popels L. C., Coyne K. J., Hutchins D. A., Cary S. C. & Dobbs F. C. (2004) 

Transport of the Harmful Bloom Alga Aureococcus anophagefferens by Oceangoing Ships and 

Coastal Boats. Applied and Environmental Microbiology 70, 6495-6500. 

Dodd C. K. (1988a) Synopsis of the biological data on the loggerhead sea turtle: Caretta caretta 

(Linnaeus 1758). Fish and Wildlife Service Biological Report 88, 110. 

Dodd C. K. J. (1988b) Synopsis of the biological data on the loggerhead sea turtle Caretta 

caretta (Linnaeus 1758). USFWS Biological Report 88, 110 pp. 

Dodd Jr. C. K. (1988a) Synopsis of the biological data on the loggerhead sea turtle Caretta 

caretta (Linnaeus 1758). In: Biological Report p. 110 pp. U.S. Fish and Wildlife Service. 

Dodd Jr. C. K. (1988b) Synopsis of the biological data on the loggerhead sea turtle, Caretta 

caretta (Linnaeus 1758). In: U.S. Fish and Wildlife Service Biological Report 

Doggett L. & Sowles J. (1989) Maine's Marine Environment: A Plan for Protection. In: A Report 

to the 114th Legislature p. 55. Department of Environmental Protection, Augusta, Maine. 

Dorn M. W. (2002) Advice on west coast rockfish harvest rates from Bayesian meta-analysis of 

stock-recruit relationships. North American Journal of Fisheries Management 22, 280-300. 

DoT. (2005) Report to congress on the performance of ports and the intermodal system. 

Department of Transportation Maritime Administration, Washington, D. C. 

DoT. (2007a) U.S. public port development expenditure report (FYs 2005 & 2006-2010). 

Department of Transportation, Maritime Administration, Washington, D. C. 

DoT. (2007b) U.S. water transportation statistical snapshot. Department of Transportation, 

Maritime Administration, Washington, D. C. 

Dovel W., Pekovitch A. & Berggren T. (1992) Biology of the shortnose sturgeon (Acipenser 

brevirostrum Lesueur, 1818) in the Hudson River estuary, New York. In: Estuarine Research in 

the 1980s (ed C. L. Smith) pp. 187-216. State University of New York Press, Albany, New York. 

Dovel W. L. & Berggren T. J. (1983) Atlantic sturgeon of the Hudson Estuary, New York. New 

York Fish and Game Journal 30, 140-172. 

Drake A. & Wilson L. (1991) Eulachon, a fish to cure humanity. UBC Museum of Anthropology, 

Museum Note 32. 

Drake J. M. (2005) Risk analysis for species introductions: forecasting population growth of 

Eurasian ruffe (<i>Gymnocephalus cernuus</i>). Canadian Journal of Fisheries and Aquatic 

Sciences 62, 1053-1059. 

Drake J. M. & Bossenbroek J. M. (2004) The potential distribution of zebra mussels in the 

United States. Bioscience 54, 931-941. 

Drake J. M. & Lodge D. M. (2004) Global hot spots of biological invasions: evaluating options 

for ballast-water management. Proceedings of the Royal Society B: Biological Sciences 271, 

575-580. 

Drake L. A., Choi K.-H., Ruiz G. M. & Dobbs F. C. (2001) Global redistribution of 

bacterioplankton and virioplankton communities. Biological Invasions 3, 193-199. 

Drake L. A., Doblin M. A. & Dobbs F. C. (2007) Potential microbial bioinvasions via ships’ 

ballast water, sediment, and biofilm. Marine Pollution Bulletin 55, 333-341. 
407

Drinkwater K. F., Belgrano A., Borja A., Conversi A., Edwards M., Greene C. H., Ottersen G., 

Pershing A. J. & Walker H. (2003) The response of marine ecosystems to climate variability 

associated with the North Atlantic oscillation. Geophysical Monograph 134, 211-234. 

Dubrovsky N. M., Kratzer C. R., Brown L. R., Gronberg J. M. & Burow K. R. (1998) Water 

Quality in the San Joaquin-Tulare Basins, California, 1992-1995. In: U.S. Geological Survey 

circular 1159 p. 43. U.S. Department of the Interior, Reston, Virginia. 

Duffield D. A. & Miller K. W. (1988) Demographic features of killer whales in oceanaria in the 

United States and Canada, 1965-1987. Rit Fiskideildar 11, 297-306. 

Duffy E. J., Beauchamp D. A., Sweeting R. M., Beamish R. J. & Brennan J. S. (2010) 

Ontogenetic diet shifts of juvenile Chinook salmon in nearshore and offshore habitats of Puget 

Sound. Transactions of American Fisheries Society 139, 803-823. 

Dumbauld B. R., Holden D. L. & Langness O. P. (2008) Do sturgeon limit burrowing shrimp 

populations in Pacific Northwest Estuaries? Environmental Bioology of Fishes 83, 283-296. 

Dunham J. B., Adams S. B., Schroeter R. E. & Novinger D. C. (2002) Alien invasions in aquatic 

ecosystems: Toward an understanding of brook trout invasions and potential impacts on inland 

cutthroat trout in wester North America. Reviews in Fish Biology and Fisheries 12, 373-391. 

Dunlop R. A., Cato D. H. & Noad M. J. (2010) Your attention please: increasing ambient noise 

levels elicits a change in communication behaviour in humpback whales (Megaptera 

novaeangliae). Proceedings of the Royal Society of London Series B: Biological Sciences in 

press, in press. 

Dunn Jr W. J. (2002) Merrimack River: 5 year watershed action plan 2002-2007. In: Executive 

Office of Environmental Affairs, Massachusetts Watershed Initiative, Department of 

Environmental Management, Devens, Massachusetts. 

Dunstan P. K. & Bax N. J. (2008) Management of an invasive marine species: defining and 

testing the effectiveness of ballast-water management options using management strategy 

evaluation. ICES Journal of Marine Science 65, 841-850. 

Durako M. J., Kunzelman, J.I., Kenworthy, W.J., Hammerstrom, K.K. (2003) Depth-related 

variability in the photobiology of Halophila johnsonii and Halophila decipiens. Mar. Biol 142, 

1219–1228. 

Duronslet M. J., Caillouet C. W., Manzella S., Indelicato K. W., Fontaine C. T., Revera D. B., 

Williams T. & Boss D. (1986) The effects of an underwater explosion on the sea turtles 

Lepidochelys kempii and Caretta caretta with observations of effects on other marine organisms. 

Southeast Fisheries Center, National Marine Fisheries Service, Galveston, Texas. 

Dustan P. (1985) Community structure of reef-building corals in the Florida Keys: Carysfort 

Reef, Key Largo, and Long Key Reef, Dry Tortugas. Atoll Research Bulletin 288, 1-27. 

Dustan P. & Halas J. C. (1987) Changes in the reef-coral community of Carysfort Reef, Key 

Largo, Florida: 1974 to 1982. Coral Reefs 6, 91-106. 

Dutil J. D. & Coutu J. M. (1988) Early marine life of Atlantic salmon (Salmo salar) post smolts 

in the northern Gulf of St. Lawrence. Fisheries Bulletin 86, 197-212. 

Dutton D. L., Bowen B. W., Owens D. W., Barragan A. & Davis S. K. (1999) Global 

phylogeography of the leatherback turtle (Dermochelys coriacea). Journal of Zoology 248, 397-

409. 

Dutton P. (2005-2006) Building our knowledge of the leatherback stock structure. SWOT Report

1, 10-11. 

Dutton P., Balazs G., Dizon A. & Barragan A. (2000) Genetic stock identification and 

distribution of leatherbacks in the Pacific: potential effects on declining populations. In: 
408

Eighteenth International Sea Turtle Symposium (eds F. A. Abreu-Grobois, R. Briseño-Dueñas, 

R. Márquez and L. Sarti) pp. 38-39. 

Dutton P., Benson S. R. & Eckert S. A. (2006) Identifying origins of leatherback turtles from 

Pacific foraging grounds off central California, U.S.A. . In: 23rd Annual Symposiumon Sea 

Turtle Biology and Conservation (ed N. J. Pilcher) p. 228. NMFS. 

Dutton P. H. (2003a) Molecular ecology of Chelonia mydas in the eastern Pacific Ocean. In: 

Proceedings of the 22nd annual symposium on sea turtle biology and conservation 

 (ed J. A. Seminoff). 

Dutton P. H. (2003b) Molecular ecology of the eastern Pacific green turtle. In: 22nd Annual 

Symposium on Sea Turtle Biology and Conservation. 

Dutton P. H. & Balazs G. H. (In review) Molecular ecology of the green turtle (Chelonia mydas) 

in the Hawaiian Archipelago: evidence for a distinct population. Endangered Species Research. 

Dutton P. H., Balazs G. H. & Dizon A. E. (1998) Genetic stock identification of sea turtles 

caught in the Hawaii-based pelagic longline fishery. In: Seventeenth Annual Symposium on Sea 

Turtle Biology and Conservation (eds S. P. Epperly and J. Braun) pp. 45-46. 

Dutton P. H., Davis S. K., Guerra T. & Owens D. (1996) Molecular phylogeny for marine turtles 

based on sequences of the ND4-leucine tRNA and control regions of mitochondrial DNA. 

Molecular Phylogenetics and Evolution 5, 511-521. 

Dutton P. H., Hitipeuw C., Zein M., Benson S. R., Petro G., Pita J., Rei V., Ambio L. & 

Bakarbessy J. (2007) Status and genetic structure of nesting populations of leatherback turtles 

(Dermochelys coriacea) in the western Pacific. Chelonian Conservation and Biology 6, 47-53. 

Ebbert J. & Embry S. (2001) Pesticides in surface water of the Yakima River Basin, Washington, 

1999-2000 - their occurrence and an assessment of factors affecting concentraitons and loads. In: 

Water Investigations Report. U.S. Geological Survey, Portland Oregon. 

Ebbert J. C., Embrey S. S., Black R. W., Tesoriero A. J. & Haggland A. L. (2000) Water quality 

in the Puget Sound basin, Washington and British Columbia, 1996-98. In: U.S. Geological 

Survey Circular p. 39. U.S. Department of the Interior, Reston, Virginia. 

Echeverria T. W. (1987) Thirty-four species of California rockfishes: maturity and seasonality of 

reproduction. Fisheries Bulletin 85, 229-250. 

Eckert K. L. (1993a) The biology and population status of marine turtles in the North Pacific 

Ocean. Final Report to NOAA, NMFS, SWFSC. Honolulu, HI. 

Eckert K. L. (1993b) The biology and population status of marine turtles in the North Pacific 

Ocean. NOAA, NMFS, SWFSC, Honolulu, Hawaii. 

Eckert S. A. (1997) Distant fisheries implicated in the loss of the world’s largest leatherback 

nesting population. Marine Turtle Newsletter 78, 2-7. 

Eckert S. A. (1998) Perspectives on the use of satellite telemetry and electronic technologies for 

the study of marine turtles, with reference to the first year long tracking of leatherback sea 

turtles. In: 17th Annual Symposium on Sea Turtle Biology and Conservation (eds S. P. Epperly 

and J. Braun) pp. 44-46. 

Eckert S. A. (1999) Data acquisition systems for monitoring sea turtle behavior and physiology. 

In: Research and Management Techniques for the Conservation of Sea Turtles. UCN/SSC 

Marine Turtle Specialist Group Publication No. 4 (eds K. L. Eckert, K. A. Bjorndal, F. A. 

Abreu-Grobois and M. Donnelly) pp. 88-93. 

Eckert S. A. (2002) Distribution of juvenile leatherback sea turtle Dermochelys coriacea

sightings. Marine Ecology Progress Series 230, 289-293. 
409

Eckert S. A. (2006) High-use oceanic areas for Atlantic leatherback sea turtles (Dermochelys 

coriacea) as identified using satellite telemetered location and dive information. Marine Biology

149, 1257-1267. 

Eckert S. A., Bagley D., Kubis S., Ehrhart L. & Johnson C. (2006) Internesting and postnesting 

movements and foraging habitats of leatherback sea turtles (Dermochelys coriacea) nesting in 

Florida. Chelonian Conservation and Biology 5, 239–248. 

Edwards M., Johns D. G., Licandro P., John A. W. G. & Stevens D. P. (2007) Ecological Status 

Report: results from the CPR survey 2005/2006. Plymouth, UK. 

Ehlke R. D. & Keller K. (2003) 2002 Chum salmon spawnign ground surveys on the mainstem 

Columbia River and its Washington tributaries. Prepared for the Pacific States Marine Fisheries 

Commission, Vancouver, Washington by the Bonneville Power Administration, Portland, 

Oregon. 

Ehrhart L. M., Bagley D. A. & Redfoot W. E. (2003) Loggerhead turtles in the Atlantic Ocean: 

Geographic distribution, abundance, and population status. In: Loggerhead Sea Turtles (eds A. 

B. Bolten and B. E. Witherington) pp. 157-174. Smithsonian Books, Washington D.C. 

Ehrhart L. M., Redfoot W. E. & Bagley D. A. (2007) Marine turtles of the central region of the 

Indian River Lagoon System, Florida. Florida Scientist 70, 415-434. 

Eiseman N. J. & McMillan C. (1980) A new species of seagrass, Halophila johnsonii, from the 

Atlantic coast of Florida. Aquatic Botany 9, 15-19. 

Eisenberg J. F. & Frazier J. (1983) A leatherback turtle (Dermochelys coriacea) feeding in the 

wild. Journal of Herpetology 17, 81-82. 

Eldredge L. G. (1991) Annotated checklist of the marine mammals of Micronesia. Micronesica

24, 217-230. 

Eldredge L. G. (2003) The marine reptiles and mammals of Guam. Micronesica 35-36, 653-660. 

Eldridge M. B., Norton E. C., Jarvis B. M. & MacFarlane R. B. (2002) Energetics of early 

development in the viviparous yellowtail rockfish. Journal of Fish Biology 61, 1122-1134. 

Elfes C. T., Vanblaricom G. R., Boyd D., Calambokidis J., Clapham P. J., Pearce R. W., Robbins 

J., Salinas J. C., Straley J. M., Wade P. R. & Krahn. M. M. (2010) Geographic variation of 

persistent organic pollutant levels in humpback whale (Megaptera novaeangliae) feeding areas of 

the North Pacific and North Atlantic. Environmental Toxicology and Chemistry 29, 824-834. 

Elliot J. M. (1991) Tolerance and resistance to thermal stress in juvenile Atlantic salmon, Salmo 

salar. Freshwater Biology 25, 61-70. 

Emery B. M., Washburn L., Love M. S., Hishimoto M. M. & Ohlmann J. C. (2006) Do oil and 

gas platforms off California reduce recruitment of bocaccio (Sebastes paucispinis) to natural 

habitat? An analysis based on trajectories derived from high-frequency radar. Fishery Bulletin

104, 391-400. 

Emmett R. L., Hinton S. A., Stoen S. L. & Monaco M. E. (1991a) Distribution and Abundance 

of Fishes and Invertebrates in West Coast Estuaries Volume II: Species Life History Studies. 

National Oceanic and Atmospheric Administration. 

Emmett R. L., Stone S. L., Hinton S. A. & Monaco M. E. (1991b) Distribution and abundance of 

fishes and invertebrates in West Coast estuaries. In: Estuarine Living Marine Resources 

Program Report 8. National Oceanic and Atmospheric Administration, National Ocean Service, 

Strategic Environmental Assessments Division, Rockville, Maryland. 

Emmett R. L., Stone S. L., Hinton S. A. & Monaco M. E. (1991c) Distribution and abundance of 

fishes and invertebrates in West Coast estuaries. Vol. 11: Species life history summaries. In: 

Estuarine Living Marine Resources Program Report 8. National Oceanic and Atmospheric 
410

Administration, National Ocean Service, Strategic Environmental Assessments Division, 

Rockville, Maryland. 

EPA. (2002) Columbia River Basin fish contaminant Survey 1996-1998. Environmental 

Protection Agency. 

EPA. (2003) Biological Evaluation for the Issuance of Ambient Water Quality Criteria for 

Dissolved Oxygen, Water Clarity and Chlorophyll a for the Chesapeake Bay and Its Tidal 

Tributaries p. 61. U.S. Environmental Protection Office. 

EPA. (2006a) EPA identifies pharmaceutical research facility as source of cyanide-related 

discharge to Wissahickon Creek. In: EPA News Release 6/22/2006. U.S. Environmental 

Protection Agency. 

EPA. (2006b) National estuary program coastal condition report. In: EPA Report 

www.epa.gov/nccr. EPA, Office of Water, and Office of Research and Development, 

Washington, D.C. 

EPA. (2007) Biological evaluation of aquatic life criteria – cyanide. p. 232. U.S. Environmental 

Protection Agency, Office of Water & Office of Science and Technology, Washington D.C. . 

EPIC, CBD & WNC. (2001) Petition to list the North American green sturgeon (Acipenser 

medirostris) as an endangered or threatened species under the endangered species act. In: 

National Marine Fisheries Service. Environmental Protection Information Center, Center for 

Biological Diversity, and Waterkeepers Northern California. 

Epstein P. R. & (Eds.). J. S. (2002) Oil, a life cycle analysis of its health and environmental 

impacts. p. 73. Report published by the Center for Health and the Global Environment, Harvard 

Medical School, Boston, MA. 

Erbe C. (2002a) Underwater noise of whale-watching boats and potential effects on killer whales 

(Orcinus orca), based on an acoustic impact model. Marine Mammal Science 18, 394-418. 

Erbe C. (2002b) Underwater noise of whale-watching boats and potential effects on killer whales 

(Orcinus orca), based on an acoustic impact model. Marine Mammal Science 18, 394-418. 

ERC. (2002) Contaminant analysis of tissues from two shortnose sturgeon (Acipenser 

brevirostrum) collected in the Delaware River. Environmental Research and Consulting, Inc., 

National Marine Fisheries Service, Gloucester, Massachusetts. 

ERC. (2003) Contaminant analysis of tissues from a shortnose sturgeon (Acipenser 

brevirostrum) from the Kennebec River, Maine. Environmental Research and Consulting, Inc., 

National Marine Fisheries Service, Gloucester, Massachusetts. 

Erhart L. M., D.A. Bagley & Redfoot W. E. (2003) Loggerhead Turtles in the Atlantic Ocean : 

Geographic Distribution, Abundance, and Population Status. Pp.157-174 In: Bolten, A.B. and 

B.E. Witherington (eds), Loggerhead Sea Turtles. Smithsonian Books, Washington D.C. 

Erickson A. W. (1978) Population studies of killer whales (Orcinus orca) in the Pacific 

Northwest: A radio-marking and tracking study of killer whales. U.S. Marine Mammal 

Commission, Washington, D.C. 

Erickson D. L. & Hightower J. E. (2007) Oceanic distribution and behavior of green sturgeon. 

American Fisheries Society Symposium 56, 197-211. 

Erickson D. L., North J. A., Hightower J. E., Webb J. & Lauck L. (2001a) Movement and habitat 

use of green sturgeon Acipenser medirostris in the Rogue River, Oregon 

Erickson D. L., North J. A., Hightower J. E., Webb J. & Lauck L. (2001b) Movement and habitat 

use of green sturgeon Acipenser medirostris in the Rogue River, Oregon. In: Unpublished draft. 
411

Erickson D. L. & Webb A. H. (2007a) Spawning periodicity, spawning migration, and size at 

maturity of green sturgeon, Acipenser medirostris, in the Rogue River, Oregon. Environmental 

Biology of Fishes 79, 255-268. 

Erickson D. L. & Webb M. A. H. (2007b) Spawning periodicity, spawning migration, and size at 

maturity of green sturgeon, Acipenser medirostris, in the Rogue River, Oregon. Environmental 

Biology of Fishes 79, 255-268. 

Eschmeyer W. N., Herald E. S. & Hammann H. (1983a) A field guide to Pacific coast fishes of 

North America. Houghton Mifflin Company, Boston, Massachussetts. 

Eschmeyer W. N., Herald E. S. & Hammann H. (1983b) A field guide to Pacific coast fishes of 

North America. Houghton Mifflin Company, Boston, Massachusetts. 

Evermann B. W. & Bean B. A. (1898) Indian River and its fishes. U.S. Commission of Fish and 

Fisheries 22, 227-248. 

EVS Environmental Consultants. (2003) Status, trends and effects of toxic contaminants in the 

Puget Sound environment. Puget Sound Action Team, Olympia, Washington. 

Ezer T. (2010) Hydrodynamic models and satellite remote sensing data help study the beluga 

whale habitat in Cook Inlet. In: Cook Inlet Belgua Whale Science Conference p. 8, Anchorage 

Downtown Marriott, Anchorage, Alaska. 

Ezer T., Hobbs R. & Oey. L.-Y. (2008) On the movement of beluga whales in Cook Inlet, 

Alaska: Simulations of tidal and environmental impacts using a hydrodynamic inundation model. 

Oceanography 21, 186-195. 

Fairchild W. L., Swansburg E. O., Arsenault J. T. & Brown S. B. (1999) Does an association 

between pesticide use and subsequent declines in catch of Atlantic salmon (Salmo salar) 

represent a case of endocrine disruption? Environmental Health Perspectives 107, 349-357. 

Falkner M., Dobroski N., Scianni C., Gehringer D. & Takata L. (2009) 2009 biennial report on 

the California marine invasive species program. California State Legislature. 

Farag A. M., Boese C. J., Woodward D. F. & Bergman H. L. (1994) Physiological changes and 

tissue metal accumulation in rainbow trout exposed to foodborne and waterborne metals. 

Environmental Toxicology and Chemistry 13, 2021-2029. 

Fausch K. D. (1988) Tests of competition between native and introduced salmonids in streams: 

What have we learned? Canadian Journal of Fisheries and Aquatic Sciences 45, 2238-2246. 

Fay C. A., Bartron M., Craig S., Hecht A., Pruden J., Saunders R., Sheehan T. & Trial J. (2006) 

Status Review for Anadromous Atlantic Salmon (Salmo salar) in the United States. National 

Marine Fisheries Service and U.S. Fish and Wildlife Service. 

Fay J. A., Foster D. J. & Stanek R. T. (1984) The use of fish and wildlife resources in Tyonek, 

Alaska. Alaska Department of Fish and Game, Anchorage, Alaska. 

Feder H. M., Turner C. H. & Limbaugh C. (1974) Observations on fishes associated with kelp 

beds in southern California. Fisheries Bulletin 160, 144. 

Feist B. E., Buhle E. R., Arnold P., Davis J. W. & Scholz N. L. (2011) Landscape ecotoxicology 

of coho salmon spawner mortality in urban streams. PlosOne 6. 

Feldkamp S. D., DeLong R. L. & Antonelis G. A. (1991) Effects of El Niño 1983 on the 

foraging patterns of California sea lions (Zalophus californianus) near San Miguel Island, 

California. In: Pinnipeds and El Niño: Responses to environmental stress (eds F. Trillmich and 

K. A. Ono) pp. 146-155. Springer-Verlag, Berlin, Germany. 

Felix F., Caballero S. & Olavarria C. (2009) Genetic diversity and population structure of 

humpback whales (Megaptera novaeangliae) from Ecuador based on mitochondrial DNA 

analyses. Update of SC/59/SH11. p. 12. IWC Scientific Committee. 
412

Felleman F., Heimlich-Boran J. R. & Osborne R. W. (1991) The feeding ecology of killer whales 

(Orcinus orca) in the Pacific Northwest. In: Dolphin Societies, discoveries, and puzzles (eds K. 

Pryor and K. Norris) pp. 113-148. University of California Press, Berkeley, California. 

Fernald A. G., Wigington P. J. & Landers D. H. (2001) Transient storage and hyporheic flow 

along the Willamette River, Oregon: Field measurements and model estimates. Water Resources 

Research 37, 1681-1694. 

Fernandes S. J., Zydlewski G. B., Zydlewski J. D., Wippelhauser G. S. & Kinnison M. T. (2010) 

Seasonal Distribution and Movements of Shortnose Sturgeon and Atlantic Sturgeon in the 

Penobscot River Estuary, Maine. Transactions of the American Fisheries Society 139, 1436-

1449. 

Ferraroli S., Georges J. Y., Gaspar P. & Maho Y. L. (2004) Where leatherback turtles meet 

fisheries. Nature 429, 521-522. 

Feyrer F., Herbold B., Matern S. A. & Moyle P. B. (2003) Dietary shifts in a stressed fish 

assemblage: Consequences of a bivalve invasion in the San Francisco Estuary. Environmental 

Biology of Fishes 67, 277-288. 

FFWCC. (2007a) Florida Statewide Nesting Beach Survey Data–2005 Season. Florida Fish and 

Wildlife Conservation Commission. 

FFWCC. (2007b) Long-term monitoring program reveals a continuing loggerhead decline, 

increases in green turtle and leatherback nesting. Florida Fish and Wildlife Conservation 

Commission, Fish and Wildlife Research Institute. 

FFWCC. (2007c) Shortnose sturgeon population evaluation in the St. Johns River, FL: Has there 

ever been a shortnose sturgeon population in Florida’s St. Johns River? Florida Fish and Wildlife 

Conservation Commission. 

Ficetola G. F. (2008) Impacts of Human Activities and Predators on the Nest Success of the 

Hawksbill Turtle, Eretmochelys imbricata, in the Arabian Gulf. Chelonian Conservation and 

Biology 7, 255-257. 

Field I., Hindell M., Slip D. & Michael K. (2001) Foraging strategies of southern elephant seals 

(Mirounga leonina) in relation to frontal zones and water masses. Antarctic Science 13, 371-379. 

Findlay K. P. & Best. P. B. (1995) Summer incidence of humpback whales on the west coast of 

South Africa. (Megaptera novaeangliae). South African Journal of Marine Science 15, 279-282. 

Finkbeiner E. M., Wallace B. P., Moore J. E., Lewison R. L., Crowder L. B. & Read A. J. (2011) 

Cumulative estimates of sea turtle bycatch and mortality in USA fisheries between 1990 and 

2007. Biological Conservation. 

Fiscus C. H. & Braham H. W. (1976) Baseline characterization: marine mammals. In: 

Environmental Assessment of the Alaskan Continental Shelf. Principal Investigators' Reports for 

the Year Ending March 1976. Volume 1. Marine Mammals pp. 57-119. 

Fisher F. W. (1994) Past and present status of Central Valley Chinook salmon. Conservation 

Biology 8, 870-873. 

Fisher T. & Hinrichsen R. (2006) Abundance-based trend results for Columbia Basin salmon and 

steelhead ESUs. Bonneville Power Administration, Portland, Oregon. 

Fisk D. A. & Harriott V. J. (1990) Spatial and temporal variation in coral recruitment on the 

Great Barrier Reef: Implications for dispersal hypotheses. Marine Biology 107, 485-490. 

Fitzsimmons N. N., Tucker A. D. & Limpus C. J. (1995) Long-term breeding histories of male 

green turtles and fidelity to a breeding ground. Marine Turtle Newsletter 68, 2-4. 
413

Flagg T. A., Waknitz F. W., Maynard D. J., Milner G. B. & Mahnken V. W. (1995) The effect of 

hatcheries on native coho salmon populations in the Lower Coloumbia River. American 

Fisheries Society Symposium 15, 366-375. 

Flather C. H., Knowles M. S. & Kendall I. A. (1998) Threatened and endangered species 

geography. Bioscience 48, 365-376. 

Fleming E. H. (2001) Swimming against the tide; recent surveys of exploitation, trade, and 

management of marine turtles in the Northern Caribbean. In: TRAFFIC North America p. 161. 

Flint M., Morton J. M., Limpus C. J., Patterson-Kane J. C., Murray P. J. & Mills P. C. (2009) 

Development and application of biochemical and haematological reference intervals to identify 

unhealthy green sea turtles (Chelonia mydas). The Veterinary Journal. 

Florida Power and Light Company St. Lucie Plant. (2002) Annual environmental operating 

report 2001. Florida Power and Light Company St. Lucie Plant, Juno Beach, Florida. 

Flournoy P. H., Rogers S. G. & Crawford P. S. (1992) Restoration of shortnose sturgeon in the 

Altamaha River, Georgia. U.S. Fish and Wildlife Service, Atlanta. 

FOC. (2008) Draft management plan for the Steller sea lion (Eumetopias jubatus) in Canada. In: 

Species at Risk Act management plan series. Fisheries and Oceans Canada, Ottawa, Canada. 

Foerster R. E. (1968) The sockeye salmon. Bulletin of the Fisheries Research Board of Canada

162. 

Foley A. M., Schroeder B. A., Redlow A. E., Fick-Child K. J. & Teas W. G. (2005) 

Fibropapillomatosis in stranded green turtles (Chelonia mydas) from the eastern United States 

(1980-98): Trends and associations with environmental factors. Journal of Wildlife Diseases 41, 

29-41. 

Foote A. D., Osborne R. W. & Hoelzel A. R. (2004a) Whale-call response to masking boat noise. 

Nature 428, 910. 

Foote A. D., Osborne R. W. & Hoelzel. A. R. (2004b) Whale-call response to masking boat 

noise. Nature 428, 910. 

Foote C. J. (1990) An experimental comparison of male and female spawning territoriality in a 

Pacific salmon. Behaviour 115, 283-313. 

Forcada J., Trathan P. N., Reid K. & Murphy. E. J. (2005) The effects of global climate 

variability in pup production of Antarctic fur seals. (Arctocephalus gazella). Ecology 86, 2408-

2417. 

Ford J. K. B. (2002) Killer whale Orcinus orca. In: Encyclopedia of marine mammals (eds W. F. 

Perrin, B. Würsig and J. G. M. Thewissen) pp. 669-676. Academic Press, San Diego, California. 

Ford J. K. B. & Ellis G. M. (2005) Prey selection and food sharing by fish-eating ‘resident’ killer 

whales (Orcinus orca) in British Columbia. In: Department of Fisheries and Oceans, Canada, 

Canadian Science Advisory Secretariat 2005/041. 

Ford J. K. B. & Ellis G. M. (2006) Selective foraging by fish-eating killer whales Orcinus orca in 

British Columbia. Marine Ecology Progress Series 316, 185-199. 

Ford J. K. B., Ellis G. M. & Balcomb K. C. (2000) Killer whales: The natural history and 

genealogy of Orcinus orca in British Columbia and Washington State, 2nd Edition. UBC Press, 

Vancouver, British Columbia. 

Ford J. K. B., Ellis G. M., Barrett-Lennard L. G., Morton A. B., Palm R. S. & Balcomb III K. C. 

(1998) Dietary specialization in two sympatric populations of killer whale (Orcinus orca) in 

coastal British Columbia and adjacent waters. Canadian Journal of Zoology 76, 1456-1471. 

Ford J. K. B. & Reeves R. R. (2008) Fight or flight: Antipredator strategies of baleen whales. 

Mammal Review 38, 50-86. 
414

Ford M. J., Cooney T., McElhany P., Sands N., Weitkamp L., Hard J., McClure M., Kope R., 

Myers J., Albaugh A., Barnas K., Teel D., Moran P. & Cowen J. (2010) Status review update for 

Pacific salmon and steelhead listed under the Endangered Species Act: Northwest. Draft. NOAA. 

Ford S. E. (1996) Range extension by the oyster parasite Perkinsus marinus into the Northeastern 

United States: Response to climate change? Journal of Shellfish Research 15, 45-56. 

Ford S. F. & Haskin H. H. (1982) History and epizootiology of Haplosporidium nelsoni (MSX), 

an oyster pathogen in Delaware Bay, 1957-1980. Journal of Invertebrate Pathology 40, 118-141. 

Formia A., Tiwari M., Fretey J. & Billes A. (2003) Sea turtle conservation along the Atlantic 

Coast of Africa. Marine Turtle Newsletter 100, 33-37. 

Foss S. (2011) 2011 triennial report on the California Department of Fish and Game’s Marine 

Invasive Species Program. California Department of Fish and Game, Office of Spill Prevention 

and Response Marine Invasive Species Program. 

Foss S. F., Ode P. R., Sowby M. & Ashe M. (2007) Non-indigenous aquatic organisms in the 

coastal waters of California. California Fisha dn Game 93, 111-129. 

Fossette S., Girard C., Bastian T., Calmettes B., Ferraroli S., Vendeville P., Blanchard F. & 

Georges J.-Y. (2009a) Thermal and trophic habitats of the leatherback turtle during the nesting 

season in French Guiana. Journal of Experimental Marine Biology and Ecology. 

Fossette S., Hobson V. J., Girard C., Calmettes B., Gaspar P., Georges J.-Y. & Hays G. C. 

(2009b) Spatio-temporal foraging patterns of a giant zooplanktivore, the leatherback turtle. 

Journal of Marine Systems in press, in press. 

Foster E. P., Fitzpatrick M. S., Feist G. W., Schreck C. B. & Yates J. (2001) Gonad 

organochlorine concentrations and plasma steroid levels in white sturgeon (Acipenser 

transmontanus) from the Columbia River, USA. Bulletin of Environmental Contamination and 

Toxicology 67, 239–245. 

Foster S. E. & Sprules W. G. (2009) Effects of the Bythotrephes invasion on native predatory 

invertebrates. Limnology and Oceanography 54, 757-769. 

Foti M., Giacopello C., Bottari T., Fisichella V., Rinaldo D. & Mammina C. (2009) Antibiotic 

Resistance of Gram Negatives isolates from loggerhead sea turtles (Caretta caretta) in the central 

Mediterranean Sea. Marine Pollution Bulletin 58, 1363-1366. 

Frair W. R., Ackman G. & Mrosovsky N. (1972) Body temperature of Dermochelys coriacea: 

warm turtle from cold water. Science 177, 791-793. 

Francis R. C. & Hare S. R. (1997) Regime scale climate forcing of salmon populations in the 

northeast Pacific--Some new thoughts and findings. In: Estuarine and ocean survival of 

northeastern Pacific salmon: Proceedings of the workshop (eds R. L. Emmett and M. H. 

Schiewe) pp. 113-127. NOAA. 

Francour P., Ganteaume A. & Poulain M. (1999) Effects of boat anchoring in Posidonia oceanica 

seagrass beds in the Port-Cros National Park (north-western Mediterranean Sea). Aquatic 

Conservation: Marine and Freshwater Ecosystems 9, 391-400. 

Franklin J. F. & Dyrness C. T. (1973) Natural vegetation of Oregon and Washington. Oregon 

State University Press. 

Frantzis A., O. Nikolaou, J. M. Bompar & Cammedda A. (2004) Humpback whale (Megaptera 

novaeangliae) occurrence in the Mediterranean Sea. Journal of Cetacean Research and 

Management 6, 25-28. 

Franzel J. & Nelson K. A. (1981) Stikine River eulachon (Thaleichthys pacificus). U.S. 

Department of Agriculture, Forest Service, Petersburg Ranger District, Petersburg, Alaska. 
415

Fraser W. R. & Hofmann E. E. (2003) A predator's perspective on causal links between climate 

change, physical forcing and ecosystem response. Marine Ecology Progress Series 265, 1-15. 

Frasier T. R., Hamilton P. K., Brown M. W., Kraus S. D. & White. B. N. (2010) Reciprocal 

exchange and subsequent adoption of calves by two North Atlantic right whales (Eubalaena 

glacialis). Aquatic Mammals 36, 115-120. 

Frazer N. B. & Ehrhart L. M. (1985a) Preliminary growth models for green, Chelonia mydas, 

and loggerhead, Caretta caretta, turtles in the wild. Copeia 1985, 73-79. 

Frazer N. B. & Ehrhart L. M. (1985b) Preliminary Growth Models for Green, Chelonia mydas, 

and Loggerhead, Caretta caretta, Turtles in the Wild. Copeia 1985, 73-79. 

Frazer N. B., Limpus C. J. & Greene J. L. (1994) Growth and estimated age at maturity of 

Queensland loggerheads. In: Proceedings of the Fourteenth Annual Symposium on Sea Turtle 

Biology and Conservation (eds K. A. Bjorndal, A. B. Bolten, D. A. Johnson and P. J. Eliazar) pp. 

42-45. U.S. Department of Commerce, Hilton Head, South Carolina. 

Frazier J. G. (2001) General natural history of marine turtles. In: Proceedings: Marine turtle 

conservation in the Wider Caribbean Region: A dialogue for effective regional management, 

Santo Domingo, Dominican Republic. 

Freeland H. (2000) The 1997-98 El Niño: The view from Line-P. CalCOFI Reports 41, 56-61. 

Freeland H. J. (1990) Sea surface temperatures along the coast of British Columbia: Regional 

evidence for a warming trend. Canadian Journal of Fisheries and Aquatic Sciences 47, 346-350. 

French J. R. P. & Jude D. J. (2001) Diets and diet overlap of nonindiginous gobies and small 

benthic native fishes co-inhabiting the St. Clair River, Michigan. Journal of Great Lakes 

Research 27, 300-311. 

French R., Bilton H., Osako M. & Hartt A. (1976) Distribution and origin of sockeye salmon 

(Oncorhynchus nerka) in offshore waters of the North Pacific Ocean. In: International North 

Pacific Fisheries Commission Bulletin 34. 

Fresh K. L. (1997) The role of competition and predation in the decline of Pacific salmon and 

steelhead. pp. 245-275. 

Fresh K. L., Casillas E., Johnson L. L. & Bottom D. L. (2005) Role of the estuary in the recovery 

of Columbia River Basin salmon and steelhead: An evaluation of the effects of selected factors 

on population viability. In: NOAA Technical Memorandum p. 155. U.S. Department of 

Commerce, Seattle, Washington. 

Freshwater D. W. & York R. S. (1999) Determination of genetic diversity in the threatened 

species Halophila johnsonii Eiseman. Report prepared for the Johnson’s Seagrass Recovery 

Team., 8 pp. 

Fretey J. (2001a) Biogeography and conservation of marine turtles of the Atlantic coast of 

Africa. In: CMS Technical Series Publication. UNEP/CMS Secretariat, Bonn, Germany. 

Fretey J. (2001b) Biogeography and conservation of marine turtles of the Atlantic coast of 

Africa. In: CMS Technical Series Publication, No. 6, UNEP/CMS Secretariat, Bonn, Germany. 

Fretey J., Billes A. & Tiwari M. (2007) Leatherback, Dermochelys coriacea, nesting along the 

Atlantic coast of Africa. Chelonian Conservation and Biology 6, 126-129. 

Frid A., Burns J., Baker G. G. & Thorne. R. E. (2009) Predicting synergistic effects of resources 

and predators on foraging decisions by juvenile Steller sea lions. Oecologia 158, 775-786. 

Fried S. M., Laner J. J. & Weston S. C. (1979) Investigation of white whale (Delphinapterus 

leucas) predation upon sockeye salmon (Oncorhynchus nerka) smolts in Nushagak Bay and 

associated rivers: 1979 aerial reconnaissance surveys. In: Project 11-41-6-340. Alaska 

Department of Fish and Game, Dillingham, Alaska. 
416

Friedlander M., Gonen Y., Kashman Y. & Beer S. (1996) Gracilaria conferta and its epiphytes: 

3. Allelopathic inhibition of the red seaweed by Ulva cf lactuca. Journal of Applied Phycology 8, 

21-25. 

Friedman C. S., Andree K. B., Beauchamp K. A., Moore J. D., Robbins T. T., Shields J. D. & 

Hedrick R. P. (2000) ‘Candidatus Xenohaliotis californiensis‘, a newly described pathogen of 

abalone, Haliotis spp., along the west coast of North America. International Journal of 

Systematic and Evolutionary Microbiology 50, 847–855. 

Fritts T. H. (1982) Plastic Bags in the Intestinal Tracts of Leatherback Marine Turtles. 

Herpetological Review 13, 72-73. 

Fritts T. H., Hoffman W. & McGehee M. A. (1983) The distribution and abundance of marine 

turtles in the Gulf of Mexico and nearby Atlantic waters. Journal of Herpetology 17, 327-344. 

Fritts T. H. & McGehee M. A. (1981) Effects of petroleum on the development and survival of 

marine turtles embryos. In: U.S. Fish and Wildlife Service, Contract No. 14-16-00009-80-946, 

FWSIOBS-81-3, Washington, D.C. 

Fritz L., Quilliam M. A., Wright J. L. C., Beale A. M. & Work T. M. (1992) An outbreak of 

domoic acid poisoning attributed to the pennate diatom Pseudonitzschia australis. Journal of 

Phycology 28, 439–442. 

Fuentes M., Hamann M. & Limpus C. J. (2009a) Past, current and future thermal profiles of 

green turtle nesting grounds: Implications from climate change. Journal of Experimental Marine 

Biology and Ecology 383, 56-64. 

Fuentes M. M. P. B. & Abbs D. (2010) Effects of projected changes in tropical cyclone 

frequency on sea turtles. Marine Ecology Progress Series 412, 283-292. 

Fuentes M. M. P. B., Hamann M. & Limpus C. J. (2009b) Past, current and future thermal 

profiles of green turtle nesting grounds: Implications from climate change. Journal of 

Experimental Marine Biology and Ecology in press, in press. 

Fuentes M. M. P. B., Limpus C. J. & Hamann M. (2010) Vulnerability of sea turtle nesting 

grounds to climate change. Global Change Biology in press, in press. 

Fuentes M. M. P. B., Maynard J. A., Guinea M., Bell I. P., Werdell P. J. & Hamann M. (2009c) 

Proxy indicators of sand temperature help project impacts of global warming on sea turtles in 

northern Australia. Endangered Species Research 9, 33-40. 

Fuhrer G. J., Morace J. L., Johnson H. M., Rinella J. F., Ebbert J. C., Embrey S. S., Waite I. R., 

Carpenter K. D., Wise D. R. & Hughes C. A. (2004) Water quality in the Yakima Basin, 

Washington, 1999-2000. In: Circular 1237 p. 44. U.S. Department of the Interior, U.S. 

Geological Survey, Portland, Oregon. 

Fujihara J., Kunito T., Kubota R. & Tanabe S. (2003) Arsenic accumulation in livers of 

pinnipeds, seabirds and sea turtles: Subcellular distribution and interaction between 

arsenobetaine and glycine betaine. Comparative Biochemistry and Physiology C-Toxicology & 

Pharmacology 136, 287-296. 

Fullerton A. H., Lamberti G. A., Lodge D. M. & Goetz F. W. (2000) Potential for resource 

competition between Eurasian ruffe and yellow perch: Growth and RNA responses in laboratory 

experiments. Transactions of the American Fisheries Society 129, 1331-1339. 

Fulton L. A. (1968) Spawning areas and abundance of chinook salmon, Oncorhynchus 

tshawytscha, in the Columbia River Basin-Past and present. In: U.S. Fish and Wildlife Service 

Special Scientific Report on Fisheries 571:26. 

Fulton L. A. (1970) Spawning areas and abundance of steelhead trout and coho, sockeye, and 

chum slamon in the Columbia River basin - past and present. In: NOAA Technical Report. 
417

Funk D. W., Markowitz T. M. & (Eds). R. R. (2005) Baseline studies of beluga whale habitat use 

in Knik Arm, Upper Cook Inlet, Alaska: July 2004 - July 2005. Final report prepared by LGL 

Alaska Research Associates, Inc., Anchorage, AK, for Knik Arm Bridge and Toll Authority 

(KABATA), Anchorage, AK. 232p. 

Futer P. & Nassichuk M. (1983) Metals in eulachons from the Nass River and crabs from Alice 

Arm, B.C. Canadian Reports of Fishereis and Aquatic Sciences. 

Gabriele C. M., Straley J. M. & Neilson J. L. (2007) Age at first calving of female humpback 

whales in southeastern Alaska. Marine Mammal Science 23, 226-239. 

Gaines S. D. & Roughgarden J. (1987) Fish in offshore kelp forests affect recruitment to 

intertidal barnacle populations. Science 235, 479-481. 

Gales N. J., Johnston D., Littnan C. & Boyd I. L. (2010) Ethics in marine mammal science. In: 

Marine Mammal Ecology and Conservation: A Handbook of Techniques (eds I. L. Boyd, W. D. 

Bowen and S. J. Iverson) pp. 1-15. Oxford University Press. 

Gallegos C. L. & Kenworthy W. J. (1996) Seagrass depth limits in the Indian River Lagoon 

(Florida, USA): Application of the optical water quality model. Estuarine, Coastal and Shelf 

Science 42, 267-288. 

Gambell R. (1976) World whale stocks. Mammal Review 6, 41-53. 

Gaos A. R., Abreu-Grobois F. A., Alfaro-Shigueto J., Amorocho D., Arauz R., Baquero A., 

Briseno R., Chacon D., Duenas C., Hasbun C., Liles M., Mariona G., Muccio C., Munoz J. P., 

Nicholas W. J., Pena M., Seminoff J. A., Vasquez M., Urteaga J., Wallace B., Yanez I. L. & 

Zarate P. (2010a) Signs of hope in the eastern Pacific: international collaboration reveals 

encouraging status for severely depleted populations of hawksbill turtle Eretmochelys imbricata. 

Oryx in press, in press. 

Gaos A. R., Abreu-Grobois F. A., Alfaro-Shigueto J., Amorocho D., Arauz R., Baquero A., 

Briseño R., Chacón D., Dueñas C., Hasbún C., Liles M., Mariona G., Muccio C., Muñoz J. P., 

Nichols W. J., Peña M., Seminoff J. A., Vásquez M., Urteaga J., Wallace B., Yañez I. L. & 

Zárate P. (2010b) Signs of hope in the eastern Pacific: international collaboration reveals 

encouraging status for a severely depleted population of hawksbill turtles Eretmochelys 

imbricata. Oryx 44, 595-601. 

Gaos A. R., Lewison R. L., Yanez I. L., Wallace B. P., Liles M. J., Nichols W. J., Baquero A., 

Hasbun C. R., Vasquez M., Urteaga J. & Seminoff J. A. (2011) Shifting the life-history 

paradigm: discovery of novel habitat use by hawksbill turtles. Biology Letters. 

Garbarino J. R., Hayes H. C., Roth D. A., Antweiler R. C., Brinton T. I. & Taylor H. E. (1995) 

Heavy metals in the Mississippi River. In: Contaminants in the Mississippi River (ed R. Meade). 

Garcia-Fernandez A. J., Gomez-Ramirez P., Martinez-Lopez E., Hernandez-Garcia A., MariaMojica P., Romero D., Jimenez P., Castillo J. J. & Bellido J. J. (2009) Heavy metals in tissues 

from loggerhead turtles (Caretta caretta) from the southwestern Mediterranean (Spain). 

Ecotoxicology and Environmental Safety 72, 557-563. 

García-Saís J. R., Appeldoorn R., Battista T., Bauer L., Bruckner A., Caldow C., Carrubba L., 

Corredor J., Diaz E., Lilyestrom C., García-Moliner G., Hernández-Delgado E., Menza C., 

Morell J., Pait A., Sabater J., Weil E., Williams E. & Williams S. (2008) The state of coral reef 

ecosystems of Puerto Rico. In: The State of Coral Reef Ecosystems of the United States and 

Pacific Freely Associated States: 2008 (eds J. E. Waddell and A. M. Clarke). 

Garcon J. S., Grech A., Moloney J. & Hamann M. (2010) Relative Exposure Index: an important 

factor in sea turtle nesting distribution. Aquatic Conservation: Marine and Freshwater 

Ecosystems 20, 140-149. 
418

Gardner S. C., Fitzgerald S. L., Vargas B. A. & Rodriguez L. M. (2006) Heavy metal 

accumulation in four species of sea turtles from the Baja California Peninsula, Mexico. 

Biometals 19, 91-99. 

Gardner S. C., Pier M. D., Wesselman R. & Juarez J. A. (2003) Organochlorine contaminants in 

sea turtles from the Eastern Pacific. Marine Pollution Bulletin 46, 1082-1089. 

Garner C. & Mckee. P. C. (2010) Cook Inlet beluga whale monitoring at Eagle Bay, Fort 

Richardson, Alaska 2008-2009. In: Cook Inlet Belgua Whale Science Conference p. 9, 

Anchorage Downtown Marriott, Anchorage, Alaska. 

Garofalo L., Mingozzi T., Mico A. & Novelletto A. (2009) Loggerhead turtle (Caretta caretta) 

matrilines in the Mediterranean: further evidence of genetic diversity and connectivity. Marine 

Biology 156, 2085-2095. 

Gascon D. & Miller R. A. (1981) Colonization by nearshore fish on small artificial reefs in 

Barkley Sound, British Columbia. Canadian Journal of Zoology 59, 1635-1646. 

Gauthier J. M., Metcalf C. D. & Sears R. (1997a) Chlorinated organic contaminants in blubber 

biopsies from northwestern Atlantic balaenopterid whales summering in the Gulf of St 

Lawrence. Marine Environmental Research 44, 201-223. 

Gauthier J. M., Metcalfe C. D. & Sears R. (1997b) Chlorinated organic contaminants in blubber 

biopsies from Northwestern Atlantic Balaenopterid whales summering in the Gulf of St 

Lawrence. Marine Environmental Research 44, 201-223. 

Geist D. R., Murray C. J., Hanrahan T. P. & Xie Y. (2009) A model of the effects of flow 

fluctuations on fall Chinook salmon spawning habitat availability in the Columbia River. North 

American Journal of Fisheries Management 28, 1911-1927. 

Geister J. (1977) The influence of wave exposure on the ecological zonation of Caribbean coral 

reefs. Proceedings of the 3rd International Coral Reef Symposium 1, 23-29. 

Gendron D. & Urban J. (1993) Evidence of feeding by humpback whales (Megaptera 

novaeangliae) in the Baja California breeding ground, Mexico. Marine Mammal Science 9, 76-

81. 

Genov T., Kotnjek P. & Lipej L. (2009) New rescord of the humpback whale (Megaptera 

novaengliae) in the Adriatic Sea. Annales 19, 25-30. 

Gentry R. L. (1970) Social behavior of the Steller sea lion. University of California at Santa 

Cruz, Santa Cruz, California. 

Gerrodette T. & DeMaster D. P. (1990) Quantitative determination of optimum sustainable 

population level. Marine Mammal Science 6, 1-16. 

Gerrodette T. M., Craig M. P. & Johanos T. C. (1992) Human-assisted fostering of Hawaiian 

monk seal pups. 'Elepaio 52, 43-46. 

Ghiold J. & Smith S. H. (1990) Bleaching, and recovery of deep-water, reef-dwelling 

invertebrates in the Cayman Islands, BWI. Caribbean Journal of Science 26, 52-61. 

Gilbert C. R. (1989a) Species profiles: life histories and environmental requirements of coastal 

fishes and invertebrates (Mid-Atlantic Bight): Atlantic and shortnose sturgeons. In: United States 

Fish and Wildlife Service Biological Report. 

Gilbert C. R. (1989b) Species profiles: life histories and environmental requirements of coastal 

fishes and invertebrates (Mid-Atlantic Bight): Atlantic and shortnose sturgeons. In: U.S. Fish 

and Wildlife Service Biological Report. U.S. Department of the Interior, Fish and Wildlife 

Service and U.S. Army Corps of Engineers, Waterways Experiment Station, Washington, D.C. 

Gill B. J. (1997) Records of turtles, and sea snakes in New Zealand, 1837-1996. New Zealand 

Journal of Marine and Freshwater Research 31, 477-486. 
419

Gilman E. L. (2009) Guidelines to reduce sea turtle mortality in fishing operations. FAO, Rome. 

Gilmartin W. G. (1983) Recovery plan for the Hawaiian monk seal, Monachus schauinslandi. 

National Marine Fisheries Service. 

Gilmore M. D. & Hall B. R. (1976) Life history, growth habits, and constructional roles of 

Acropora cervicornis in the patch reef environment. Journal of Sedimentary Petroleum 46, 519-

522. 

Girard C., Tucker A. D. & Calmettes B. (2009) Post-nesting migrations of loggerhead sea turtles 

in the Gulf of Mexico: dispersal in highly dynamic conditions. Marine Biology 156, 1827-1839. 

Girondot M., Godfrey M. H., Ponge L. & Rivalan P. (2007) Modeling approaches to quantify 

leatherback nesting trends in French Guiana and Suriname. Chelonian Conservation and Biology

6, 37-46. 

Gitschlag G. R. & Herczeg B. A. (1994) Sea turtle observations at explosive removals of energy 

structures. Marine Fisheries Review 56, 1-8. 

Gitschlag G. R., Herczeg B. A. & Barcak T. R. (1997) Observations of sea turtles and other 

marine life at the explosive removal of offshore oil and gas structures in the Gulf of Mexico. 

Gulf Research Reports 9, 247-262. 

Glass A. H., Cole T. V. N. & Garron M. (2009) Mortality and serious injury determinations for 

baleen whale stocks along the United States eastern seaboard and adjacent Canadian Maritimes, 

2003-2007 (second edition). 

Gless J. M., Salmon M. & Wyneken J. (2008) Behavioral responses of juvenile leatherbacks 

Dermochelys coriacea to lights used in the longline fishery. Endangered Species Research 5, 

239-247. 

Glockner-Ferrari D. A. & Ferrari M. J. (1985) Individual identification, behavior, reproduction, 

and distribution of humpback whales, Megaptera novaeangliae, in Hawaii. U.S. Marine Mammal 

Commission, Washington, D.C.; National Technical Information Service, Springfield, Virginia: 

36p. 

Godley B., Richardson J. S., Broderick A. C., Coyne M. S., Glen F. & Hays G. C. (2002) Longterm satellite telemetry of the movements and habitat utilization by green turtles in the 

Mediterranean. Ecography 25, 352-362. 

Godley B. J., Thompson D. R. & Furness R. W. (1999) Do heavy metal concentrations pose a 

threat to marine turtles from the Mediterranean Sea? Marine Pollution Bulletin 38, 497-502. 

Godley B. J., Thompson D. R., Waldron S. & Furness R. W. (1998) The trophic status of marine 

turtles as determined by stable isotope analysis. Marine Ecology Progress Series 166, 277-284. 

Godley B. J. E., Lima H. S. M., Akesson S., Broderick A. C., Glen F., Godfrey M. H., Luschi P. 

& Hays G. C. (2003) Movement patterns of green turtles in Brazilian coastal waters described by 

satellite tracking and flipper tagging. Marine Ecology Progress Series 253, 279-288. 

Goenaga C. & Cintrón G. (1979) Inventory of the Puerto Rican coral reefs. Department of 

Natural Resources, San Juan, Puerto Rico. 

Goetz K. T., Rugh D. J., Read r. J. & Hobbs. R. C. (2007) Habitat use in a marine ecosystem: 

Beluga whales Delphinapterus leucas in Cook Inlet, Alaska. Marine Ecology Progress Series

330, 247-256. 

Goff G. P. & Lien J. (1988) Atlantic leatherback turtles, Dermochelys coriacea, in cold water off 

Newfoundland and Labrador. Canadian Field Naturalist 102, 1-5. 

Goldberg W. M. (1973) The ecology of the coral-octocoral community of the southeast Florida 

coast: Geomorphology, species composition, and zonation. Bulletin of Marine Science 23, 465-

488. 
420

Gómez-Cabrera M. d. C., Ortiz J. C., Loh W. K. W., Ward S. & Hoegh-Guldberg O. (2008) 

Acquisition of symbiotic dinoflagellates (Symbiodinium) by juveniles of the coral Acropora 

longicyathus. Coral Reefs 27, 219-226. 

Good T. P., Waples R. S. & Adams P. (2005a) Updated status of federally listed ESUs of West 

Coast salmon and steelhead. In: NOAA Technical Memorandum pp. 1-598. U.S. Department of 

Commerce, Seattle, WA. 

Good T. P., Waples R. S. & Adams P. (2005b) Updated status of federally listed ESUs of West 

Coast salmon and steelhead. In: NOAA Technical Memorandum NMFS-NWFSC-66. U.S. 

Department of Commerce. 

Good T. P., Waples R. S. & Adams P. (2005c) Updated status of federally listed ESUs of West 

Coast salmon and steelhead. U.S. Department of Commerce, NOAA Technical Memorandum 

NMFS-NWFSC-66. 

Goodyear J. D. (1993) A sonic/radio tag for monitoring dive depths and underwater movements 

of whales. Journal of Wildlife Management 57, 503-513. 

Gordon A. N., Pople A. R. & Ng J. (1998) Trace metal concentrations in livers and kidneys of 

sea turtles from south-eastern Queensland, Australia. Marine and Freshwater Research 49, 409-

414. 

Gordon J. & Moscrop A. (1996) Underwater noise pollution, and its significance for whales, and 

dolphins. In: The conservation of whales and dolphins: Science and practice (eds M. P. 

Simmonds and J. D. Hutchinson) pp. 281-319. John Wiley & Sons, Chichester, United Kingdom. 

Goreau N. I., Goreau T. J. & Hayes R. L. (1981) Settling, survivorship, and spatial aggregation 

in planulae and juveniles of the coral Porites porites (Pallas). Bulletin of Marine Science 31, 424-

435. 

Goreau T. F. (1959) The ecology of Jamaican reef corals: I. Species composition, and zonation. 

Ecology 40, 67-90. 

Goreau T. F. & Goreau N. I. (1973) The ecology of Jamaican reefs. II. Geomorphology, 

zonation, and sedimentary phases. Bulletin of Marine Science 23, 399-464. 

Goreau T. F. & Wells J. W. (1967) The shallow-water Scleractinia of Jamaica: Revised list of 

species and their vertical range. Bulletin of Marine Science 17, 442-453. 

Gorsky D., Trial J., Zydlewski J. & McCleave J. (2009) The effects of smolt stocking strategies 

on migratory path selection of adult Atlantic salmon in the Penobscot River, Maine. North 

American Journal of Fisheries Management 29, 949-957. 

Gotelli N. J. & Ellison A. M. (2004) A Primer of Ecological Statistics. Sinauer Associates, Inc. 

Sunderland, Massachusetts. 510p. 

Grace J. B. & Harrison J. S. (1986) The biology of Canadian weeds. 73. Typha latifuliaL., Typha 

angustifolia L. and Typha xglauca Godr. Canadian Journal of Plant Science 66, 361-379. 

Grant G. S., Craig P. & Balazs G. H. (1997) Notes on juvenile hawksbill and green turtles in 

American Samoa. Pacific Science 51, 48-53. 

Grant G. S. & Ferrell D. (1993) Leatherback turtle, Dermochelys coriacea (Reptilia: 

Dermochelidae): Notes on near-shore feeding behavior and association with cobia. Brimleyana

19, 77-81. 

Grant S. C. H. & Ross. P. S. (2002) Southern resident killer whales at risk: Toxic chemicals in 

the British Columbia and Washington environment. (Orcinus orca). CANADIAN TECHNICAL 

REPORT OF FISHERIES AND AQUATIC SCIENCES 2412:i-xii, 1-111. 
421

Gray C. & Tuominen T. (2001) The Fraser River is getting cleaner: will it continue to improve? 

In: 2001 Puget Sound Research Conference (ed T. Droscher). Puget Sound Action Team, 

Olympia, Washington. 

Gray D. K., Johengen T. H., Reid D. F. & MacIsaac H. J. (2007) Efficacy of open-ocean ballast 

water exchange as a means of preventing invertebrate invasions between freshwater ports. 

Limnology and Oceanography 52, 2386-2397. 

Green A. J. & Figuerola J. (2005) Recent advances in the study of long-distance dispersal of 

aquatic invertebrates via birds. Diversity and Distributions 11, 149-156. 

Green G. A., Brueggerman J. J., Grotefendt R. A., Bowlby C. E., Bonnell M. L. & Balcomb III 

K. C. (1992) Cetacean distribution and abundance off Oregon and Washington, 1989-1990. In: 

Oregon and Washington Marine Mammal and Seabird Surveys. Minerals Management Service 

Contract Report 14-12-0001-30426. 

Green G. A., Grotefendt R. A., Smultea M. A., Bowlby C. E. & Rowlett R. A. (1993) Delphinid 

aerial surveys in Oregon and Washington offshore waters. Final report. National Marine 

Fisheries Service, National Marine Mammal Laboratory, Seattle, Washignton. 

Greene C. & Pershing A. J. (2004) Climate and the conservation biology of North Atlantic right 

whales: the right whale at the wrong time? . Front Ecol Environ 2, 29-34. 

Greene C., Pershing A. J., Kenney R. D. & Jossi J. W. (2003a) Impact of climate variability on 

the recovery of endangered North Atlantic right whales. Oceanography 16, 98-103. 

Greene C. H., Pershing A. J., Kenney R. D. & Jossi J. W. (2003b) Impact of climate variability 

on the recovery of endangered North Atlantic right whales. Oceanography 16, 98-103. 

Greening H. & Holland N. (2003) Johnson's seagrass (Halophila johnsonii) monitoring 

workshop. In: Summary report prepared for the Johnson’s Implementation Recovery Team. 

Greer A. E., Lazell Jr. J. D. & Wright R. M. (1973) Anatomical evidence for counter-current heat 

exchanger in the leatherback turtle (Dermochelys coriacea). Nature 244, 181. 

Gregory S., Ashkenas L., Haggerty P., Oetter D., Wildman K., Hulse D., Branscomb A. & 

VanSickle J. (2002a) Riparian vegetation. In: Willamette River basin planning atlas: 

Trajectories of environmental and ecological change (eds D. Hulse, S. Gregory and J. Baker) pp. 

40-43. Oregon State University Press, Corvallis, Oregon. 

Gregory S., Ashkenas L., Oetter D., Minear P. & Wildman K. (2002b) Historical Willamette 

River channel change. In: Willamette River basin planning atlas: Trajectories of environmental 

and ecological change (eds D. Hulse, S. Gregory and J. Baker) pp. 44-45. Oregon State 

University Press, Corvallis, Oregon. 

Gregory S., Ashkenas L., Oetter D., Minear P., Wildman R., Minear P., Jett S. & Wildman K. 

(2002c) Revetments. In: Willamette River basin planning atlas: Trajectories of environmental 

and ecological change (eds D. Hulse, S. Gregory and J. Baker) pp. 32-33. Oregon State 

University Press, Corvallis, Oregon. 

Gregory S., Wildman R., Ashkenas L., Wildman K. & Haggerty P. (2002d) Fish assemblages. 

In: Willamette River basin planning atlas: Trajectories of environmental and ecological change

(eds D. Hulse, S. Gregory and J. Baker) pp. 44-45. Oregon State University Press, Corvallis, 

Oregon. 

Gregory S. V. & Bisson P. A. (1997) Degradation and loss of anadromous salmonid habitat in 

the Pacific Northwest. In: Pacific salmon and their ecosystems: Status and future options (eds D. 

J. Stouder, P. A. Bisson and R. J. Naiman) pp. 277-314. Chapman & Hall, New York, NY. 

Gribsholt B. & Kristensen E. (2002) Effects of bioturbation and plant roots on salt marsh 

biogeochemistry: a mesocosm study. Marine Ecology-Progress Series 241, 71–87. 
422

Gronlund W., Chan S.-L., McCain B., Clark R., Myers M., Stein J., Brown D., Landahl J., Krahn 

M. & Varanasi U. (1991) Multidisciplinary assessment of pollution at three sites in Long Island 

Sound. Estuaries and Coasts 14, 299-305. 

Groot C. & Margolis L. (1991) Pacific salmon life histories. In: University of British Columbia. 

University of British Columbia Press, Vancouver, Canada. 

Groot C., Margolis L. & (editors). (1991) Pacific salmon life histories. University of British 

Columbia. University of British Columbia Press, Vancouver, Canada, 564 p. 

Grosholz E. (2002) Ecological and evolutionary consequences of coastal invasions. Trends in 

Ecology and Evolution 17, 22-27. 

Grosholz E. D. (2005) Recent biological invasion may hasten invasional meltdown by 

accelerating historical introductions. Proceedings of the National Academy of Sciences 102, 

1088-1091. 

Grosholz E. D. & Ruiz G. (1995) Spread and potential impact of the recently introduced 

European green crab, Carcinus maenas, in central California. Marine Biology 122, 239-247. 

Grosholz E. D. & Ruiz G. M. (1996) Predicting the impact of introduced marine species: lessons 

from the multiple invasions of the European green crab Carcinus maenas. Biological 

Conservation 78, 59-66. 

Grosholz E. D., Ruiz G. M., Dean C. A., Shirley K. A., Maron J. L. & Connors P. G. (2000) The 

impacts of a nonindigenous marine predator in a California Bay. Ecology 81, 1206-1224. 

Gross M. R., Repka J., Robertson C. T., Secor D. H. & Winkle W. V. (2002) Sturgeon 

conservation: insights from elasticity analysis. In: Biology, management, and protection of North 

American sturgeon, Symposium 28 (ed V. W. Webster) pp. 13-30. American Fisheries Society, 

Bethesda, Maryland. 

Group S. S. (1996) Eumetopias jubatus. In: 2006 IUCN red list of threatened species. 

Grout J. A., Levings C. D. & Richardson J. S. (1997) Decomposition rates of purple loosestrife 

(Lythrum salicaria) and Lyngbyei's sedge (Carexlyngbyei) in the Fraser River Estuary. Estuaries

20, 96-102. 

Grunwald C., Maceda L., Waldman J., Stabile J. & Wirgin I. (2008) Conservation of Atlantic 

sturgeon Acipenser oxyrinchus oxyrinchus : delineation of stock structure and distinct population 

segments. Conservation Genetics 9, 1111-1124. 

Grunwald C., Stabile J., Waldman J. R., Gross R. & Wirgin I. (2002) Population genetics of 

shortnose sturgeon Acipenser brevirostrum based on mitochondrial DNA control region 

sequences. Molecular Ecology 11, 1885-1898. 

Guerriero A., Meinesz A., D’Ambrosio M. & Pietra F. (1992) Isolation of toxic and potentially 

toxic sesqui- and monoterpenes from the tropical green seaweed Caulerpa taxifolia which has 

invaded the region of Cap Martin and Monaco. Helvetica Chimica Acta 75, 689–695. 

Guilbard F., Munro J., Dumont P., Hatin D. & Fortin R. (2007) Feeding Ecology of Atlantic 

Sturgeon and Lake Sturgeon Co-Occurring in the St. Lawrence Estuarine Transition Zone. 

American Fisheries Society Symposium 56, 85. 

Guillemot P. J., Larson R. J. & Lenarz W. H. (1985) Seasonal cycles of fat and gonad volume in 

five species of northern California rockfish (Scorpaenidae: Sebastes). Fishery Bulletin 1983. 

Guimarães J. P., Mari R. d. B., Marigo J., Rosas F. C. W. & Watanabe I.-S. (2011) Light and 

scanning electron microscopic study of the tongue in the estuarine dolphin (Sotalia 

guianensisvan Bénéden, 1864). Zoological Science 28, 617-622. 

Guisan A. & Thuiller W. (2005) Predicting species distribution: offering more than simple 

habitat models. Ecology Letters 8, 993-1009. 
423

Gulko D. & Eckert K. L. (2003) Sea turtles: an ecological guide. Mutual Publishing, Honolulu, 

Hawaii. 

Gunderson D. R., Callahan P. & Goiney B. (1980) Maturation and fecundity of four species of 

Sebastes. Marine Fisheries Review 42, 74-79. 

Gunderson D. R. & Sample T. M. (1980) Distribution and abundance of rockfish off 

Washington, Oregon, and California during 1977. Marine Fisheries Review MArch-April, 2-16. 

Gustafson-Greenwood K. I. & Moring J. R. (1991) Gravel compaction and permeabilities in 

redds of Atlantic salmon, Salmo salar L. Aquaculture and Fisheries Management 22, 537-540. 

Gustafson-Marjenan K. I. & Dowse H. B. (1983) Seasonal and diel patterns of emergence from 

the red of Atlantic salmon (Salmo salar) fry. Canadian Journal of Fisheries and Aquatic 

Sciences 40, 813-817. 

Gutierrez J. L., Jones C. G., Strayer D. L. & Iribarne O. O. (2003) Mollusks as ecosystem 

engineers: the role of shell production in aquatic habitats. Oikos 101, 79-90. 

Guzmán Del Proó S. A. (1992) A review of the biology of abalone and its fishery in Mexico. In: 

Abalone of the world: Biology fisheries and culture (eds S. A. Shepherd, M. J. Tegner and S. A. 

Guzmán Del Proó) pp. 341-360. Fishing News Books. 

Guzman del Proo S. A., Pineda B. J., Molina J., Uribe F., Lopez F., Aguilar R., Andrade M., 

Leon G., Marin V. & Castro C. (1980) Analisis de la pesqueria del abulon en Mexico. In: Inst. 

Nal de Pesca Doc. Tec. Int. pp. 1-294. 

Haaker P. L. (1994) Assessment of abalone resources at the Channel Islands. In: The Fourth 

California Islands Symposium: Update on the Status of Resources (eds W. L. Halvorson and G. 

J. Maender). Santa Barbara Museum of Natural History. 

Haaker P. L., Henderson K. C. & Parker D. O. (1986) California abalone. In: California Fish and 

Game Marine Resources Leaflet No. 11. . California Department of Fish and Game, Marine 

Resources Division. 

Haaker P. L., Richards D. V., Friedman C., Davis G. E., Parker D. O. & Togstad H. (1989) 

Abalone withering syndrome and mass mortality of black abalone, in California. In: First 

International Symposium on Abalone Biology, Fisheries, and Culture. November 21- 25, La Paz, 

Mexico. 

Haggerty M., Furfey R., Hollenbach M., Hooper T., Longenbaugh M., Morlock M., Tynan M., 

Plummer M., Trask B. & Wheeler R. (2009a) Recovery plan for Lake Ozette sockeye salmon 

(Oncorhynchus nerka). National Marine Fisheries Service, Salmon Recovery Division, Portland, 

Oregon. 

Haggerty M. J., Ritchie A. C., Shellberg J. G., Crewson M. J. & Jalonen J. (2009b) Lake Ozette 

sockeye limiting factors analysis. Makah Indian Tribe and NOAA, Port Angeles, Washington. 

Hahn K. O. (1989) Handbook of culture of abalone and other marine gastropods. CRC Press, 

Inc., Boca Raton, Florida. 

Hain J. H. W., Carter G. R., Kraus S. D., Mayo C. A. & Winn H. E. (1982) Feeding behavior of 

the humpback whale, Megaptera novaeangliae, in the western North Atlantic. Fishery Bulletin

80, 259-268. 

Hain J. H. W., Ellis S. L., Kenney R. D., Clapham P. J., Gray B. K., Weinrich M. T. & Babb I. 

G. (1995) Apparent bottom feeding by humpback whales on Stellwagen Bank. Marine Mammal 

Science 11, 464-479. 

Haley D. (1970) Views on the killer whale dispute. Pacific Search 5, 1-3. 

Haley D. (1986) Marine mammals. Pacific Search Press, Seattle, Washington. 
424

Haley N. (1998) A gastric lavage technique for characterizing diets of sturgeons. North 

American Journal of Fisheries Management 18, 978 - 981. 

Haley N. J. (1999) Habitat Characteristics and Resource Use Patterns of Sympatric Sturgeons in 

the Hudson River Estuary. p. 140. University of Massachusetts Amherst. 

Hall A. J., Englehard G. H., Brasseur S. M. J. M., Vecchione A., Burton H. R. & Reijnders. P. J. 

H. (2003) The immunocompetence handicap hypothesis in two sexually dimorphic pinniped 

species - is there a sex difference in immunity during early development? Developmental and 

Comparative Immunology 27-Jan, 629-637. 

Hall L. M., Hanisak M. D. & Virnstein R. W. (2006) Fragments of the seagrasses Halodule 

wrightii and Halophila johnsonii as potential recruits in Indian River Lagoon, Florida. Marine 

Ecology Progress Series 310, 109-117. 

Hallegraeff G. & Bolch C. (1992) Transport of diatom and dinoflagellate resting spores in ships’ 

ballast water: implications for plankton biogeography and aquaculture. Journal of Plankton 

Research 14, 1067-1084. 

Hallegraeff G. M. (1998) Transport of toxic dinoflagellates via ships' ballast water: bioeconomic 

risk assessment and efficacy of possible ballast water management strategies. Marine EcologyProgress Series 168, 297-309. 

Hallegraeff G. M. (2010) Ocean climate change, phytoplankton community responses, and 

harmful algal blooms: a formidable predictive challenge. Journal of Phycology 46, 220-235. 

Hallock R. J., Van Woert W. F. & Shapovalov L. (1961) An evaluation of stocking hatcheryreared steelhead rainbow trout (Salmo gairdneri gairdneri) in the Sacramento River System. 

Fish Bulletin 114, 1-74. 

Halstead B. W., Auerbach P. S. & Campbell D. R. (1990) A colour atlas of dangerous marine 

animals. Wolfe Medical Publications Ltd, W.S. Cowell Ltd, Ipswich, England. 

Hamann M., Limpus C., Hughes G., Mortimer J. & Pilcher N. (2006a) Assessment of the 

conservation status of the leatherback turtle in the Indian Ocean and South East Asia, including 

consideration of the impacts of the December 2004 tsunami on turtles and turtle habitats. In: 

IOSEA Marine Turtle MoU Secretariat, Bangkok. 

Hamann M., Limpus C., Hughes G., Mortimer J. & Pilcher N. (2006b) Assessment of the 

conservation status of the leatherback turtle in the Indian Ocean and South East Asia, including 

consideration of the impacts of the December 2004 tsunami on turtles and turtle habitats. IOSEA 

Marine Turtle MoU Secretariat, Bangkok, Thailand. 

Hamer J. P., Lucas I. & McCollin T. (2001) Harmful dinoflagellate resting cysts in ships' ballast 

tank sediments: 

potential for introduction into English and Welsh waters. Phycologia 40, 246-255. 

Hamer J. P., McCollin T. & Lucas I. (2000) Dinofagellate Cysts in Ballast Tank 

Sediments: Between Tank Variability. Marine Pollution Bulletin 40, 731-733. 

Hamilton P. & Cooper. L. A. (2010) Changes in North Atlantic right whale (Eubalaena glacialis) 

cow-calf association times and use of the calving ground: 1993-2005. Marine Mammal Science

26, 896-916. 

Hamilton P. K., Knowlton A. R., Marx M. K. & Kraus S. D. (1998) Age structure and longevity 

in North Atlantic right whales Eubalaena glacialis and their relationship to reproduction. Marine 

Ecology Progress Series 171, 285-292. 

Hammerstom K. K. & Kenworthy W. J. (2003) Investigating the existence of a Halophila 

johnsonii sediment seed bank. In: Report prepared for the Johnson’s Seagrass Implementation 

Team. 
425

Hammond P. S. (2010) Estimating the abundance of marine mammals. In: Marine Mammal 

Ecology and Conservation: A Handbook of Techniques (eds I. L. Boyd, W. D. Bowen and S. J. 

Iverson) pp. 42-67. Oxford University Press. 

Handeland S. O., Imsland A. K. & Stefansson S. O. (2008) The effect of temperature and fish 

size on growth, feed intake, food conversion efficiency and stomach evacuation rate of Atlantic 

salmon post-smolts. Aquaculture 283, 36-42. 

Hanson, C. H., J. R. White, and H. W. Li. 1977. Entrapment and impingement of fishes by power 

plant cooling-water intake structures: an overview. Marine Fisheries Review 39:7-17. 

Hansen L. P. & Pethon P. (1985) The food of Atlantic salmon, Salmo salar L., caught by longline in northern Norwegian waters. Journal of Fish Biology 26, 553-562. 

Hansen L. P. & Quinn T. P. (1998) The marine phase of the Atlantic salmon (Salmo salar) life 

cycle, with comparisons to Pacific salmon. Canadian Journal of Fisheries and Aquatic Sciences 

55(Supplemental 1, 104-118. 

Hanson M. B., Baird R. W., Emmons C., Hempelmann J., Schorr G. S. & Doornik D. V. (2007a) 

Stock identification of prey selected by "southern resident" killer whales in their summer range. 

National Marine Fisheries Service, Northwest Fisheries Science Center, Seattle, and Cascadia 

Research Collective, Olympia, Washington. 

Hanson M. B., Baird R. W., Emmons C., Hempelmann J., Schorr G. S., Sneva J. & VanDoornik 

D. (2007b) Summer diet and prey stock identification of the fish-eating "southern resident" killer 

whale: addressing a key recovery need using fish scales, fecal samples, and genetic techniques. 

In: 17th Biennial Conference on the Biology of Marine Mammals, Capetown, South Africa. 

Hanson M. B., Baird R. W., Ford J. K. B., Hempelmann-Halos J., Doornik D. M. V., Candy J. 

R., Emmons C. K., Schorr G. S., Gisborne B., Ayres K. L., Wasser S. K., Balcomb K. C., 

Balcomb-Bartok K., Sneva J. G. & Ford M. J. (2010a) Species and stock identification of prey 

consumed by endangered southern resident killer whales in their summer range. Endangered 

Species Research 11, 69-82. 

Hanson M. B., Baird R. W., Ford J. K. B., Hempelmann-Halos J., Van Doornik D. M., Candy J. 

R., Emmons C. K., Schorr G. S., Gisborne B., Ayres K. L., Wasser S. K., Balcomb K. C., 

Balcomb-Bartok K., Sneva J. G. & Ford M. J. (2010b) Species and stock identification of prey 

consumed by endangered southern resident killer whales in their summer range. Endangered 

Species Research 11, 69-82. 

Hanson M. B., Baird R. W. & Schorr G. S. (2005) Focal behavioral observations and fisheating 

killer whales: improving our understanding of foraging behavior and prey selection. In: 16th 

Biennial Conference on the Biology of Marine Mammals. 

Haraguchi K., Hisamichi Y. & Endo. T. (2009) Accumulation and mother-to-calf transfer of 

anthropogenic and natural organohalogens in killer whales (Orcinus orca) stranded on the Pacific 

coast of Japan. Science of the Total Environment 407, 2853-2859. 

Hard J., Jones R. P., Jr., Delarm M. R. & Waples R. S. (1992) Pacific salmon and artificial 

propagation under the endangered species act. In: Technical Memorandum U.S. Department of 

Commerce, National Oceanic and Atmospheric Administration, National Marine Fisheries 

Service, Northwest Fisheries Science Center, Seattle, Washington. 

Harden Jones F. R. (1968) Fish migration. Edward Arnold Ltd., London. 

Harding Lawson Associates. (1999) Draft engineering evaluation/cost analysis, Eastland Woolen 

Mill site, Corinna, Maine. In: U.S. Army Corps of Engineers, New England District. Concord, 

Massachusetts. 
426

Harding Lawson Associates. (2000) Draft volume I engineering evaluation report, Eastland 

Woolen Mill site, Corinna, Maine. In: U.S. Army Corps of Engineers, New England District. 

Concord, Massachusetts. 

Hare S. R. & Mantua N. J. (2000) Empirical evidence for North Pacific regime shifts in 1977 and 

1989. Progress in Oceanography 47, 103-145. 

Hare S. R., Mantua N. J. & Francis R. C. (1999) Inverse production regimes: Alaska and west 

coast Pacific salmon. Fisheries Habitat 24, 6-14. 

Harington C., Ross R., Mathewes R., Stewart K. & Beattie O. (2004) A late Pleistocene Steller 

sea lion (Eumetopias jubatus) from Courtenay, British Columbia: its death, associated biota, and 

paleoenvironment. Canadian Journal of Earth Sciences 41, 13. 

Harned D. A. & Meyer D. (1983) Water quality of the Yadkin-Peedee River system, North 

Carolina-Variability, pollution loads, and long-term trends. In: Water quality of North Carolina 

streams p. 71. 

Harriott V. J. (1985) Recruitment patterns of scleractinian corals at Lizard Island, Great Barrier 

Reef. Proceedings of the 5th International Coral Reef Congress 4, 367-372. 

Harrison R. J. & Thurley K. W. (1974) Structure of the epidermis in Tursiops, Delphinus, 

Orcinus and Phocoena. In: Functional Anatomy of Marine Mammals, Vol. 2 (ed R. J. Harrison) 

pp. 45-71. Academic Press. 

Hart J. L. (1973a) Pacific fishes of Canada. Fisheries Research Board of Canada Bulletin 180. 

Hart J. L. (1973b) Pacific fishes of Canada. Bulletin of the Fisheries Research Board of Canada

180. 

Hart J. L. & McHugh J. L. (1944) The smelts (Osmeridae) of British Columbia. Bulletin of the 

Fisheries Research Board of Canada 64. 

Hartman G. F., Scrivener J. C. & Miles M. J. (1996) Impacts of logging in Carnation Creek, a 

high-energy coastal stream in British Columbia, and their implication for restoring fish habitat. 

Canadian Journal of Fisheries and Aquatic Sciences 53, 237-251. 

Hartmann A. R. (1987) Movement of scorpionfishes (Scorpaenidae: Sebastes and Scorpaena) in 

the Southern California Bight. Bulletin of the California Department of Fish and Game 73, 68-

79. 

Harvell D., Altizer S., Cattadori I. M., Harrington L. & Weil E. (2009) Climate change and 

wildlife diseases: When does the host matter the most? Ecology 90, 912-920. 

Harvey C. J. (2005) Effects of El Niño events on energy demand and egg production of rockfish 

(Scorpaenidae: Sebastes): a bioenergetics approach. Fishery Bulletin 103, 71-83. 

Harvey C. J., Tolimieri N. & Levin P. S. (2006) Changes in body size, abundance, and energy 

allocation in rockfish assemblages of the Northeast Pacific. Ecological Applications 16, 1502-

1515. 

Harvey P. H. & May R. M. (1997) Case studies of extinction. . Nature 385, 776-777. 

Harwood J. (2010) Approaches to management. In: Marine Mammal Ecology and Conservation: 

A Handbook of Techniques (eds I. L. Boyd, W. D. Bowen and S. J. Iverson) pp. 325-339. Oxford 

University Press. 

Hashagen K. A., Green G. A. & Adams B. (2009) Observations of humpback whales, Megaptera 

novaeangliae, in the Beaufort Sea, Alaska. Northwestern Naturalist 90, 160-162. 

Hastings R. W., O'Herron II J. C., Schick K. & Lazzari M. A. (1987) Occurrence and distribution 

of shortnose sturgeon, Acipenser brevirostrum, in the upper tidal Delaware River. Estuaries 10, 

337-341. 
427

Hatase H., Kinoshita M., Bando T., Kamezaki N., Sato K., Matsuzawa Y., Goto K., Omuta K., 

Nakashima Y., Takeshita H. & Sakamoto W. (2002a) Population structure of loggerhead turtles, 

Caretta caretta, nesting in Japan: bottlenecks on the Pacific population. Marine Biology 141, 

299-305. 

Hatase H., Matsuzawa Y., Sakamoto W., Baba N. & Miyawaki I. (2002b) Pelagic habitat use of 

an adult Japanese male loggerhead turtle Caretta caretta examined by the Argos satellite system. 

Fisheries Science 68, 945-947. 

Hatase H. & Sakamoto W. (2004) Forage-diving behaviour of adult Japanese female loggerhead 

turtles (Caretta caretta) inferred from Argos location data. Journal of the Marine Biological 

Association of the United Kingdom 84, 855-856. 

Hatase H., Sato K., Yamaguchi M., Takahashi K. & Tsukamoto K. (2006) Individual variation in 

feeding habitat use by adult female green sea turtles (Chelonia mydas): Are they obligately 

neritic herbivores? Oecologia 149, 52-64. 

Hatase H., Takai N., Matsuzawa Y., Sakamoto W., Omuta K., Goto K., Arai N. & Fujiwara T. 

(2002c) Size-related differences in feeding habitat use of adult female loggerhead turtles Caretta 

caretta around Japan determined by stable isotope analyses and satellite telemetry. Marine 

Ecology Progress Series 233, 273-281. 

Hatch L., Clark C., Merrick R., Parijs S. V., Ponirakis D., Schwehr K., Thompson M. & Wiley 

D. (2008) Characterizing the relative contributions of large vessels to total ocean noise fields: A 

case study using the Gerry E. Studds Stellwagen Bank National Marine Sanctuary. 

Environmental Management 42, 735-752. 

Hatten J. R. & Tiffan K. F. (2009) A spatial model to assess the effects of hydropower operations 

on Columbia River fall Chinook salmon spawning habitat. North American Journal of Fisheries 

Management 29, 1379-1405. 

Haupt T. M., Griffiths C. L., Robinson T. B. & Tonin A. F. G. (2010) Oysters as vectors of 

marine aliens, with notes on four introduced species associated with oyster farming in South 

Africa. African Zoology 45, 52-62. 

Haw F. & Buckley R. (1971) Saltwater fishing in Washington. Stanley N. Jones, Seattle, 

Washington. 

Hawkes L. A., Broderick A., Godfrey M. H. & Godley B. J. (2007a) The potential impact of 

climate change on loggerhead sex ratios in the Carolinas - how important are North Carolina's 

males? P.153 in: Frick, M.; A. Panagopoulou; A.F. Rees; K. Williams (compilers), 27th Annual 

Symposium on Sea Turtle Biology and Conservation [abstracts]. 22-28 February 2007, Myrtle 

Beach, South Carolina. 296p. 

Hawkes L. A., Broderick A. C., Godfrey M. H. & Godley B. J. (2007b) Investigating the 

potential impacts of climate change on a marine turtle population. Global Change Biology 13, 1-

10. 

Hay D. E. & Beacham T. D. (2005) Stock identification of eulachon (Thaleichthys pacificus), an 

anadromous smelt in the eastern Pacific. In: ICES 2005 Annual Science Conference, Aberdeen, 

Scotland. 

Hay D. E. & McCarter P. B. (2000) Status of the eulachon Thaleichthys pacificus in Canada. 

Department of Fisheries and Oceans Canada. Canadian Stock Assessment Secretariat, Ottawa, 

Ontario. 

Hay D. E., West K. C. & Anderson A. D. (2003) Indicators and ‘response’ points for 

management of Fraser River eulachon: a comparison and discussion with recommendations. 

Canadian Science Advisory Secretariat. 
428

Hayden A. (1998) Merrymeeting Bay an Environmental Review. p. 23. Friends of 

Merrymeeting Bay. 

Hayes S. A., Bond M. H., Hanson C. V. & Freund E. V. (2008) Steelhead growth in a small 

central California watershed: Upstream and estuarine rearing patterns. Transactions of the 

American Fisheries Society 137, 114-128. 

Hayman R. A., Beamer E. M. & McClure R. E. (1996) FY 1995 Skaig River chinook restoration 

research. In: Report by Skagit System Cooperative p. 54, La Conner, Washington. 

Hays G. C., Houghton J. D. R. & Myers A. E. (2004) Pan-Atlantic leatherback turtle movements. 

Nature 429, 522. 

Hayteas D. L. & Duffield. D. A. (2000) High levels of PCB and p,p'-DDE found in the blubber 

of killer whales (Orcinus orca). Marine Pollution Bulletin 40, 558-561. 

Hayward T. L. (2000) El Niño 1997-98 in the coastal waters of southern California: A timeline 

of events. CalCOFl Reports 41, 98-116. 

Hazard K. (1988) Beluga whale, Delphinapterus leucas. In: Selected marine mammals of Alaska: 

Species accounts with research and management recommendations pp. 195-235. Marine 

Mammal Commission, Washington, D.C. 

Hazel J. (2009) Evaluation of fast-acquisition GPS in stationary tests and fine-scale tracking of 

green turtles. Journal of Experimental Marine Biology and Ecology 374, 58-68. 

Hazel J. & Gyuris E. (2006) Vessel-related mortality of sea turtles in Queensland, Australia. 

Wildlife Research 33, 149-154. 

Hazel J., Lawler I. R., Marsh H. & Robson S. (2007) Vessel speed increases collision risk for the 

green turtle Chelonia mydas. Endangered Species Research 3, 105-113. 

HDLNR. (2002) Application for an individual incidental take permit pursuant to the Endangered 

Species Act of 1973 for listed sea turtles in inshore marine fisheries in the main Hawaiian Islands 

managed by the State of Hawaii. p. 54. State of Hawaii, Division of Aquatic Resources. 

Healey M. C. (1983) Coastwide distribution and ocean migration patterns of stream- and oceantype chinook salmon, Oncorhynchus tshawytscha. Canadian Field-Naturalist: 97, 427-433. 

Healey M. C. (1991a) Life history of chinook salmon (Oncorhynchus tshawytscha). In: Pacific 

salmon life histories (eds C. Groot and L. Margolis) pp. 311-393. UBC Press, University of 

British Columbia, Vancouver. 

Healey M. C. (1991b) Life history of Chinook salmon (Oncorhynchus tshawytscha). pages 311-

394 in Groot, C. and Margolis (editors). Pacific salmon life histories. University of British 

Columbia Press. Vancouver, Canada. 

Hearn W. E. (1987) Interspecific competition and habitat segregation among stream-dwelling 

trout and salmon: A review. Fisheries 12, 21-24. 

Hebdon J. L., Kline P., Taki D. & Flagg T. A. (2004) Evaluating reintroduction strategies for 

Redfish Lake sockeye salmon captive broodstock strategy. In: Propagated fish in resource 

management. American Fisheries Society, Symposium 44 (eds M. J. Nickum, P. M. Mazik, J. G. 

Nickum and D. D. MacKinlay) pp. 401-413. American Fisheries Society, Bethesda, Maryland. 

Hecky R. E., Smith R. E., Barton D. R., Guildford S. J., Taylor W. D., Charlton M. N. & Howell 

T. (2004) The nearshore phosphorus shunt: a consequence of ecosystem engineering by 

dreissenids in the Laurentian Great Lakes. Canadian Journal of Fisheries and Aquatic Sciences

61, 1285-1293. 

Heidelbaugh W. S., Hall L. M., Kenworthy W. J., Whitfield P. E., Virnstein R. W., Morris L. J. 

& Hanisak M. D. (2000) Reciprocal transplanting of the threatened seagrass Halophila johnsonii 
429

(Johnson's seagrass) in the Indian River Lagoon, Florida. In: Seagrasses: Monitoring, ecology, 

physiology, and management pp. 197-210. CRC Press, Boca Raton, Florida. 

Heimlich-Boran J. R. (1986) Fishery correlations with the occurrence of killer whales in Greater 

Puget Sound. In: Behavioral biology of killer whales (eds B. C. Kirkevold and J. S. Lockard) pp. 

113-131. Alan R. Liss, New York, NY. 

Heimlich-Boran J. R. (1988) Behavioral ecology of killer whales (Orcinus orca) in the Pacific 

Northwest. Canadian Journal of Zoology 66, 565-578. 

Heithaus M. R., McLash J. J., Frid A., Dill L. M. & Marshall G. J. (2002) Novel insights into 

green sea turtle behaviour using animal-borne video cameras. Journal of the Marine Biological 

Association of the United Kingdom 82, 1049-1050. 

Hellio C., Simon-Colin C., Clare A. C. & Deslandes E. (2004) Isethionic acid and floridoside 

isolated from the red alga, Grateloupia turuturu, inhibit settlement of Balanus amphitrite cyprid 

larvae. Biofouling 20, 139-145. 

Heppell S. S., Crouse D. T., Crowder L. B., Epperly S. P., Gabriel W., Henwood T., Márquez R. 

& Thompson N. B. (2005) A population model to estimate recovery time, population size, and 

management impacts on Kemp’s ridley sea turtles. Chelonian Conservation and Biology 4, 767-

773. 

Heppell S. S., Crowder L. B., Crouse D. T., Epperly S. P. & Frazer N. B. (2003a) Population 

models for Atlantic loggerheads: Past, present, and future. Chapter 16 In: Bolten, A. and B. 

Witherington (eds), Loggerhead Sea Turtles. Smithsonian Books, Washington, D.C. Pp.255-

273. 

Heppell S. S., Snover M. L. & Crowder L. B. (2003b) Sea turtle population ecology. Chapter 11 

In: Lutz, P.L., J.A. Musick, and J. Wyneken (eds), The Biology of Sea Turtles: Volume II. CRC 

Press. Pp.275-306. 

Herborg L. M., Jerde C. L., Lodge D. M., Ruiz G. M. & MacIsaac H. J. (2007) Predicting 

invasion risk using measures of introduction effort and environmental niche models. Ecological 

Applications 17, 663-674. 

Herman L. M. (1979) Humpback whales in Hawaiian waters: A study in historical ecology. 

Pacific Science 33, 1-15. 

Herman L. M., Pack A. A., Rose K., Craig A., Herman E. Y. K., Hakala S. & Milette A. (2011) 

Resightings of humpback whales in Hawaiian waters over spans of 10–32 years: Site fidelity, sex 

ratios, calving rates, female demographics, and the dynamics of social and behavioral roles of 

individuals. Marine Mammal Science. 

Hermanussen S., Matthews V., Papke O., Limpus C. J. & Gaus C. (2008) Flame retardants 

(PBDEs) in marine turtles, dugongs and seafood from Queensland, Australia. Marine Pollution 

Bulletin 57, 409-418. 

Hernandez R., Buitrago J., Guada H., Hernandez-Hamon H. & Llano M. (2007) Nesting 

distribution and hatching success of the leatherback, Dermochelys coriacea, in relation to human 

pressures at Playa Parguito, Margarita Island, Venezuela. Chelonian Conservation and Biology

6, 79-86. 

Hessen D., Bakkestuen V. & Walseng B. (2011) Ecological niches of Bythotrephes and 

Leptodora: lessons for predicting long-term effects of invasion. Biological Invasions 13, 561–

2572. 

Heublein J. C., Kelly J. T., Crocker C. E., Klimley A. P. & Lindley S. T. (2009) Migration of 

green sturgeon, Acipenser medirostris, in the Sacramento River. Environmental Biology of 

Fishes 84, 245-258. 
430

Hewitt C. & Inglis G. (2004) Baseline evaluation and monitoring of marine invasions in 

Australia and New Zealand. In: ASFB Invasive Species: Fish and Fisheries Workshop, 

Wellington, New Zealand. 

Heyning J. E. & Lewis. T. D. (1990) Entanglements of baleen whales in fishing gear off southern 

California. (Eschrichtius robustus, Balaenoptera acutorostrata, Megaptera novaeangliae). Report 

of the International Whaling Commission 40, 427-431.-Sc/441/Ps414). 

Hickie B., Ross P. & Macdonald. R. (2001) An examination of the history of PCB accumulation 

by North Pacific killer whales. (Orcinus orca). In: Fourteenth Biennial Conference on the 

Biology of Marine Mammals, 28 November-3 December Vancouver Canada. p.98. 

Hicks D. (2005) Lower Rogue watershed assessment. South Coast Watershed Council, Gold 

Beach, Oregon. 

Higdon J. W. & Ferguson S. H. (2009) Loss of Arctic sea ice causing punctuated change in 

sightings of killer whales (Orcinus orca) over the past century. Ecological Applications 19, 1365-

1375. 

Hildebrand H. H. (1963) Hallazgo del area de anidacion de la tortuga marina "lora", 

Lepidochelys kempi (Garman), en la costa occidental del Golfo de Mexico (Rept., Chel.). 

Ciencia, Mexico 22, 105-112. 

Hildebrand H. H. (1983) Random notes on sea turtles in the western Gulf of Mexico. Western 

Gulf of Mexico Sea Turtle Workshop Proceedings, January 13-14, 1983, 34-41. 

Hildebrand J. A. (2009) Anthropogenic and natural sources of ambient noise in the ocean. 

Marine Ecology Progress Series 395, 5-20. 

Hill P. S. & DeMaster D. P. (1998) Alaska marine mammal stock assessments, 1998. In: NOAA 

Technical Memorandum NMFS-AFSC-97. 

Hill P. S., DeMaster D. P. & Small R. J. (1997) Alaska stock assessments - 1996. U.S. 

Department of Commerce, NOAA Technical Memorandum, Alaska Fisheries Science Center; 

Auke Bay, Alaska. . 

Hillis-Starr, Coyne Z. M. & Monaco M. (2000) Buck Island and back: Hawksbill turtles make 

their move. In: Nineteenth Annual Symposium on Sea Turtle Biology and Conservation (eds H. J. 

Kalb and T. Wibbels) p. 159. 

Hilterman M. L. & Goverse E. (2003) Aspects of Nesting and Nest Success of the Leatherback 

Turtle (Dermochelys coriacea) in Suriname, 2002. Guianas Forests and Environmental 

Conservation Project (GFECP). Technical Report, World Wildlife Fund Guianas/Biotopic 

Foundation, Amsterdam, the Netherlands, 31p. 

Himmelman J. H. & Steele D. H. (1971) Foods and predators of the green sea urchin 

Strongylocentrotus droebachiensis in Newfoundland waters. Marine Biology 9, 315-322. 

Hinck J. E., Schmitt C. J., Bartish T. M., Denslow N. D., Blazer V. S., Anderson P. J., Coyle J. 

J., Dethloff G. M. & Tillitt D. E. (2004) Biomonitoring of environmental status and trends 

(BEST) program: environmental contaminants and their effects on fish in the Columbia River 

basin. In: Scientific investigation report 2004-5154. U.S. Department of the Interior, U.S. 

Geological Survey, Columbia Environmental Research Center, Columbia, Missouri. 

Hindell M. A., Crocker D., Mori Y. & Tyack P. (2010) Foraging behaviour. In: Marine Mammal 

Ecology and Conservation: A Handbook of Techniques (eds I. L. Boyd, W. D. Bowen and S. J. 

Iverson) pp. 243-262. Oxford University Press. 

Hines A. H., Alvarez F. & Reed S. A. (1997) Introduced and native populations of the marine 

parasitic castrator. Variation in prevalence of the rhizocephalan Loxothylacus panopaei in 

xanthid crabs. . Bulletin of Marine Science 61, 197-214. 
431

Hirth H. F. (1980) Some aspects of the nesting behavior and reproductive biology of sea turtles. 

Presented at: Symposium on Behavioral and Reproductive Biology of Sea Turtles Tampa, FL 

(USA) 27 Dec 1979. American Zoologist 20, 507-523. 

Hirth H. F. (1997a) Synopsis of the biological data on the green turtle Chelonia mydas (Linnaeus 

1758). Biological Report 91, U.S. Fish and Wildlife Service Report, 120p. 

Hirth H. F. (1997b) Synopsis of the biological data on the green turtle, Chelonia mydas 

(Linnaeus 1758). 

Hitipeuw C., Dutton P. H., Benson S. R., Thebu J. & Bakarbessy J. (2007) Population status and 

internesting movement of leatherback turtles, Dermochelys coriacea, nesting on the northwest 

coast of Papua, Indonesia. Chelonian Conservation and Biology 6, 28-36. 

Hitz C. R. (1962) Seasons of birth of rockfish (Sebastodes) in Oregon coastal waters. 

Transactions of the American Fisheries Society 91, 231-233. 

Hoar W. S. (1976) Smolt transformation: Evaluation, behavior, and physiology. Journal of the 

Fisheries Research Board of Canada 33, 1233-1252. 

Hobbs R. C., Laidre K. L., Vos D. J., Mahoney B. A. & Eagleton. M. (2005) Movements and 

area use of belugas, Delphinapterus leucas, in a subarctic Alaskan estuary. Arctic 58, 331-340. 

Hobbs R. C., Shelden K. E. W., Rugh D. J. & Norman. S. A. (2008) 2008 status review and 

extinction risk assessment of Cook Inlet belugas (Delphinapterus leucas). NMFS, Alaska 

Fisheries Science Center. AFSC Processed Report 2008-02. 116pp. 

Hobbs R. C., Shelden K. E. W., Vos D. J., Goetz K. T. & Rugh D. J. (2006) Status review and 

extinction assessment of Cook Inlet belugas (Delphinapterus leucas). In: Alaska Fisheries 

Science Center, Processed Report 2006-16. 

Hobbs R. C. & Shelden. K. E. W. (2008) Supplemental status review and extinction assessment 

of Cook Inlet belugas (Delphinapterus leucas). NMFS, Alaska Fisheries Science Center. AFSC 

Processed Report 2008-08. xv + 76pp. 

Hobbs R. C., Waite J. M. & Rugh D. J. (2000) Beluga, Delphinapterus leucas, group sizes in 

Cook Inlet, Alaska, based on observer counts and aerial video. Marine Fisheries Review 62, 46-

59. 

Hobday A. J. & Tegner M. J. (2000) Status review of white abalone (Haliotis sorenseni) 

throughout its range in California and Mexico. In: NOAA Technical Memorandum NMFS-SWR035. 

Hobday A. J., Tegner M. J. & Haaker P. L. (2001) Over-exploitation of a broadcast spawning 

marine invertebrate: Decline of the white abalone. Reviews in Fish Biology and Fisheries 10, 

493-514. 

Hodge R. P. & Wing B. L. (2000) Occurrences of marine turtles in Alaska waters: 1960-1998. 

Herpetological Review 31, 148-151. 

Hoegh-Guldberg O. & Bruno J. F. (2010) The Impact of Climate Change on the World's Marine 

Ecosystems. Science 328, 1523-1528. 

Hoffman W. S., Kaufman G. D. & Jule K. (2010) Estimation of survival, recruitment and 

realized growth rates of the East Australia humpback population (BS-1) using temporal 

symmetry models. p. 17. IWC Scientific Committee, Agadir, Morocco. 

Holbrook C. M., Zydlewski J., Gorsky D., Shepard S. L. & Kinnison M. T. (2009) Movements 

of prespawn adult Atlantic salmon near hydroelectric dams in the lower Penobscot River, Maine. 

North American Journal of Fisheries Management 29, 495-505. 

Holmberg E. K., Day D., Pasquale N. & Pattie B. (1967) Research report on the Washington 

trawl fishery 1962-64. Washington Department of Fisheries, Research Division. 
432

Holmes A., Wise S., Goertz C., Dunn J., Gulland F., Gelatt T., Beckmen K., Burek K., Atkinson 

S. & Bozza M. (2008) Metal tissue levels in Steller sea lion (Eumetopias jubatus) pups. Marine 

Pollution Bulletin 56, 1416-1421. 

Holmes E. E. & York A. E. (2003) Using age structure to detect impacts on threatened 

populations: a case study using Steller sea lions. Conservation Biology 17, 1794-1806. 

Holt M. M., Noren D. P., Veirs V., Emmons C. K. & Veirs. S. (2009) Speaking up: Killer whales 

(Orcinus orca) increase their call amplitude in response to vessel noise. Journal of the Acoustical 

Society of America 125, El27-El32. 

Honea J. M., Jorgensen J. C., McClure M. M., Cooney T. D., Engie K., Holzer D. M. & Hilborn 

R. (2009) Evaluating habitat effects on population status: influence of habitat restoration on 

spring-run Chinook salmon. Freshwater Biology 54, 1576-1592. 

Hornell J. (1927) The turtle fisheries of the Seychelles Islands. H.M. Stationery Office, London, 

UK. 

Horrocks J. A. & Scott N. (1991) Nest site location, and nest success in the hawksbill turtle 

Eretmochelys imbricata in Barbados, West Indies. Marine Ecology Progress Series 69, 1-8. 

Horrocks J. A., Vermeer L. A., Krueger B., Coyne M., Schroeder B. A. & Balazs G. H. (2001) 

Migration routes and destination characteristics of post-nesting hawksbill turtles satellite-tracked 

from Barbados, West Indies. Chelonian Conservation and Biology 4, 107-114. 

Hoshino H., Fujita S., Goto Y., Isono T., Ishinazaka T., Burkanov V. N. & Sakurai Y. (2006) 

Organochlorines in Steller sea lions (Eumetopias jubatus) from the western North Pacific. In: Sea 

Lions of the World (eds A. W. Trites, S. K. Atkinson, D. P. DeMaster, L. W. Fritz, T. S. Gelatt, 

L. D. Rea and K. M. Wynne) pp. 1-11. Alaska Sea Grant College Program, Fairbanks, Alaska, 

USA. 

Houston J. J. (1988) Status of green sturgeon, Acipenser medirostris, in Canada. Canadian FieldNaturalist 102, 286-290. 

Howell E. A., Kobayashi D. R., Parker D. M., Balazs G. H. & Polovina J. J. (2008) TurtleWatch: 

a tool to aid in the bycatch reduction of loggerhead turtles Caretta caretta in the Hawaii-based 

pelagic longline fishery. Endangered Species Research 5, 267-278. 

Howell M. D. (2001) Characterization of development in Columbia River prolarval eulachon, 

Thaleichthys pacificus, using selected morphometric characters. Washington Department of Fish 

and Wildlife. 

Howell P., Jones K., Scarnecchia D., LaVoy L., Kendra W. & Ortmann D. (1985a) Stock 

assessment of Columbia River anadromous salmonids. In: Volume II: Steelhead stock 

summaries, stock transfer guidelines, and information needs. Final Report to the Bonneville 

Power Administration for Contract No. DE-A179-84BP12737 Bonneville Power Administration, 

Portland, Oregon. 

Howell P., Jones K., Scarnecchia D., LaVoy L., Kendra W. & Ortmann D. (1985b) Stock 

assessment of Columbia River anadromous salmonids. In: Volume II: Steelhead stock 

summaries, stock transfer guidelines, and information needs. Final Report to the Bonneville 

Power Administration Bonneville Power Administration, Portland, Oregon. 

Howorth P. C. (1978) The abalone book. Naturegraph Publishers, Happycamp, California. 

Hoyt E. (1990) Orca: the whale called killer. Camden House Publishing, North York, Ontario. 

Hoyt E. (2001) Whale watching 2001: Worldwide tourism numbers, expenditures, and 

expanding socioeconomic benefits. A special report from the International Fund for Animal 

Welfare. 157pgs. 
433

Hoyt E. (2002) Whale watching. In: Encyclopedia of Marine Mammals. W. F. Perrin, B. Wursig 

AND J. G. M. Thewissen (eds.). p.1305-1310. Academic Press, San Diego, CA. 1414pgs. 

Huang X., Ke C. & Wang W. X. (2008) Bioaccumulation of silver, cadmium and mercury in the 

abalone Haliotis diversicolor from water and food sources. 

Hugg D. O. (1996) MAPFISH georeferenced mapping database. Freshwater and estuarine fishes 

of North America. (eds D. O. and S. Hugg). Life Science Software, Edgewater, Maryland. 

Hughes G. R., Luschi P., Mencacci R. & Papi F. (1998) The 7000-km oceanic journey of a 

leatherback turtle tracked by satellite. Journal of Experimental Marine Biology and Ecology 229, 

209-217. 

Hughes T. P. & Jackson J. B. C. (1985) Population dynamics and life histories of foliaceous 

corals. Ecological Monographs 55, 141-166. 

Hulin V., Delmas V., Girondot M., Godfrey M. H. & Guillon J. M. (2009) Temperaturedependent sex determination and global change: are some species at greater risk? Oecologia 160, 

493-506. 

Humann P. & Deloach N. (2003) Reef Coral Identification: Florida, Caribbean, Bahamas 

Including Marine Plants, Enlarged 2nd Edition. New World Publications, Inc., Jacksonville, 

Florida. 

Huntington H. P. (2000a) Traditional knowledge of the ecology of beluga whale, Delphinapterus 

leucas, in Cook Inlet, Alaska. Marine Fisheries Review 62, 134-140. 

Huntington H. P. (2000b) Traditional knowledge of the ecology of belugas, Delphinapterus 

leucas, in Cook Inlet, Alaska. Marine Fisheries Review 62, 134-140. 

Hurrell J. W. (1995) Decadal trends in the North Atlantic Oscillation: Regional temperatures and 

precipitation. Science 269, 676-679. 

Hutchinson B. J. & Dutton P. (2007) Modern genetics reveals ancient diversity in the 

loggerhead. In: SWOT report, volume II. State of the World’s Sea Turtles. 

ICBTRT. (2003) Independent populations of chinook, steelhead, and sockeye for listed 

evolutionarily significant units within the Interior Columbia River Domain. Northwest Fisheries 

Science Center, Seattle, Washington. 

ICBTRT. (2005) Interior Columbia Basin TRT: Viability criteria for application to interior 

Columbia Basin salmonid ESUs. Interior Columbia Basin Technical Recovery Team, National 

Marine Fisheries Service. 

ICTRT. (2003a) Independent populations of chinook, steelhead, and sockeye for listed 

evolutionarily significant units within the Interior Columbia River Domain. In: Recovery 

Document. Interior Columbia Basin Technical Recovery Team, National Marine Fisheries 

Service, Seattle, Washington. 

ICTRT. (2003b) Working draft. Independent populations of Chinook, steelhead, and sockeye for 

listed evolutionarily significant units within the Interior Columbia River domain. Interior 

Columbia Technical Recovery Team. 

ICTRT. (2006) Draft Snake River salmon and steelhead recovery plan. National Marine 

Fisheries Service, Northwest Region, Protected Resources Division, Portland, Oregon. 

IDEQ. (2011) Idaho Department of Environmental Quality final 2010 integrated report. Idaho 

Department of Environmental Quality, Boise, Idaho. 

IEPSPWT. (1999) Monitoring, assessment, and research on central valley steelhead: Status of 

knowledge, review of existing programs, and assessment of needs in comprehensive monitoring, 

assessment, and research program plan. In: Technical Application VII-A-11. Interagency 

Ecological Program Steelhead Project Work Team. 
434

IGFA. (1991) World record game fishes. International Game Fish Association. 

Illingworth and Rodkin Inc. (2001) Noise and vibration measurements associated with the pile 

installation demonstration project for the San Francisco-Oakland Bay Bridge east span, final data 

report. 

Illingworth and Rodkin Inc. (2004) Conoco/Phillips 24-inch steel pile installation – Results of 

underwater sound measurements. In: Letter to Ray Neal, Conoco/Phillips Company. 

Ings J. S., Servos M. R. & Vijayan M. M. (2011) Exposure to municipal wastewater effluent 

impacts stress performance in rainbow trout. Aquatic Toxicology 103, 85-91. 

Innis C., Nyaoke A. C., Williams C. R., Dunnigan B., Merigo C., Woodward D. L., Weber E. S. 

& Frasca S. (2009) PATHOLOGIC AND PARASITOLOGIC FINDINGS OF COLDSTUNNED KEMP'S RIDLEY SEA TURTLES (LEPIDOCHELYS KEMPII) STRANDED ON 

CAPE COD, MASSACHUSETTS, 2001-2006. journal of wildlife diseases 45, 594-610. 

Innis C., Tlusty M. F., Perkins C., Holladay S., Merigo C. & Scott Webber III E. (2008) Trace 

metal and organochlorine pesticide concentrations in cold-stunned kuvenile Kemp’s ridley turtles 

(Lepidochelys kempii) from Cape Cod, Massachusetts. Chelonian Conservation and Biology 7, 

230-239. 

IPCC. (2000) Land use, land-use change, and forestry. Cambridge University Press, Cambridge, 

England. 

IPCC. (2001a) Climate change 2001: Impacts, adaptation and vulnerability, contribution of 

working group II to the third assessment report of the intergovernmental panel on climate 

change. Cambridge University Press, Cambridge, England. 

IPCC. (2001b) Climate change 2001: The scientific basis, contribution of working group I to the 

third assessment report of the intergovernmental panel of climate change. Cambridge University 

Press, Cambridge, England. 

IPCC. (2002) Climate change and biodiversity. 

Isaac J. L. (2008) Effects of climate change on life history: Implications for extinction risk in 

mammals. Endangered Species Research. 

ISAB. (2007a) Climate change impacts on Columbia River Basin fish and wildlife. Columbia 

River Basin Indian Tribes and National Marine Fisheries Service, Portland, Oregon. 

ISAB. (2007b) Climate change impacts on Columbia River Basin fish and wildlife. Columbia 

River Basin Indian Tribes and National Marine Fisheries Service, Portland, Oregon. 

ISAB. (2007c) Climate change impacts on Columbia River basin fish and wildlife. In: ISAB, 

Report 2007-2, . Independent Scientific Advisory Board, Portland, Oregon. 

Ischer T., Ireland K. & Booth D. T. (2009) Locomotion performance of green turtle hatchlings 

from the Heron Island Rookery, Great Barrier Reef. Marine Biology 156, 1399-1409. 

ISG. (1996) Return to the river: restoration of salmonid fishes in the Columbia River ecosystem. 

In: Independent Science Group report 96-6. Northwest Power Planning Council, Portland, 

Oregon. 

Israel J. A., Bando K. J., Anderson E. C. & May B. (2009) Polyploid microsatellite data reveal 

stock complexity among estuarine North American green sturgeon (Acipenser medirostris). 

Canadian Journal of Fisheries and Aquatic Sciences 66, 1491-1504. 

Issac J. L. (2009) Effects of climate change on life history: Implications for extinction risk in 

mammals. Endangered Species Research 7, 115-123. 

Iwamoto T., Ishii M., Nakashima Y., Takeshita H. & Itoh A. (1985) Nesting cycles and 

migrations of the loggerhead sea turtle in Miyazaki, Japan. Japanese Journal of Ecology 35, 505-

511. 
435

IWC. (1996) Report of the sub-committee on Southern Hemisphere baleen whales, Annex E 

Report of the International Whaling Commission 46, 117-131. 

IWC. (2004) Scientific committee - Annex K: Report of the standing working group on 

environmental concerns. Sorrento, Italy. 

IWC. (2005) Chair's Report of the 57th Annual Meeting. International Whaling Commission. 

Available online at: http://www.iwcoffice.org/_documents/meetings/ulsan/CRREP57.pdf 

Accessed 7/26/2006. 

IWC. (2007) Whale Population Estimates. International Whaling Commission. Accessed 

02/07/2007 online at: http://www.iwcoffice.org/conservation/estimate.htm. 

IWC. (2008) Catch limits & catches taken. International Whalign Commission. 

Jaap W. C. (1974) Scleractinian growth rate studies. In: Florida Keys Coral Reef Workshop p. 

17. Florida Department of Natural Research Coastal Coordinating Council. 

Jaap W. C. (1979) Observation on zooxanthellae expulsion at Middle Sambo Reef, Florida Keys. 

Bulletin of Marine Science 29, 414-422. 

Jaap W. C. (1984) The ecology of the south Florida coral reefs: A community profile. In: US 

Fish and Wildlife Service. FWS/OBS 82/08. 

Jaap W. C., Lyons W. G., Dustan P. & Halas J. C. (1989) Stony coral (Scleractinia, and 

Milleporina) community structure at Bird Key Reef, Ft. In: Florida Marine Research 

Publication 46:31. Jefferson National Monument, Dry Tortugas, Florida. 

Jackson J. K., Huryn A. D., Strayer D. L., Courtemanch D. L. & Sweeney B. W. (2005) Atlantic 

coast rivers of the northeastern United States. In: Rivers of North America (eds A. C. Benke and 

C. E. Cushing) pp. 21-70. Elsevier, New York. 

Jackson N. D., Garvey J. E. & Colombo R. E. (2007) Comparing aging precision of calcified 

structures in shovelnose sturgeon. Journal of Applied Ichthyology 23, 525-528. 

Jacobs R., Larson G., Meyer J., Currence N., Hinton J., Adkinson M., Burgner R., Geiger H. & 

Lestelle L. (1996) The sockeye salmon Oncorhynchus nerka population in Lake Ozette, 

Washington, USA. In: National Park Service Technical Report NPS/CCSOSU/NRTR-96/04. U.S. 

Department of the Interior. 

Jacobsen J. K. (1986) The behavior of Orcinus orca in the Johnstone Strait, British Columbia. In: 

Behavioral Biology of Killer Whales. Barbara C. Kirkevold and Joan S. Lockard, eds. p.135-

185. 

Jager H. I., Chandler J. A., Lepla K. B. & Van Winkle W. (2001) A theoretical study of river 

fragmentation by dams and its effects on white sturgeon populations. Environmental Biology of 

Fishes 60, 347-361. 

Jagoe C. H. & Haines T. A. (1990) Morphometric effects of low pH and limed water on the gills 

of Atlantic salmon (Salmo salar). Canadian Journal of Fisheries and Aquatic Sciences 47, 2451-

2460. 

James M. C., Eckert S. A. & Myers R. A. (2005) Migratory and reproductive movements of male 

leatherback turtles (Dermochelys coriacea). Marine Biology 147, 845-853. 

James M. C., Ottensmeyer C. A., Eckert S. A. & Myers R. A. (2006) Changes in the diel diving 

patterns accompany shifts between northern foraging and southward migration in leatherback 

turtles. . Canadian Journal of Zoology 84, 754-765. 

James M. C., Sherrill-Mix S. A. & Myers R. A. (2007) Population characteristics and seasonal 

migrations of leatherback sea turtles at high latitudes. Marine Ecology Progress Series 337, 245-

254. 
436

Janssen J. & Jude D. J. (2001) Recruitment failure of mottled sculpin Cottus Bairdi in Calumet 

Harbor, southern Lake Michigan, induced by the newly introduced round goby Neogobius 

melanostromus. Journal of Great Lakes Research 27, 319-328. 

Jasny M., Reynolds J., Horowitz C. & Wetzler A. (2005) Sounding the depths II: The rising toll 

of sonar, shipping and industrial ocean noise on marine life. Natural Resources Defense Council, 

New York, New York. 

Jassby A. D., Cloern J. E. & Cole B. E. (2002) Annual primary production: Patterns and 

mechanisms of change in a nutrient-rich tidal 

ecosystem. Limnology and Oceanography 47, 698-712. 

JCRMS. (2009) 2010 joint staff report concerning stock status and fisheries for sturgeon and 

smelt. Oregon Department of Fish and Wildlife and Washington Department of Fish and 

Wildlife. 

Jeffries S. (1984) Marine mammals of the Columbia River Estuary. Washington State 

Department of Game, Olympia, Washington. 

Jehl J., Joseph R. & Cooper C. F. (1980) Introduction. In: Potential Effects of Space Shuttle 

Sonic Booms on the Biota and Geology of the California Channel Islands: Research Reports (eds 

J. J. R.Jehl and C. F. Cooper) pp. 1-3. 

Jennings M. R. (1996) Past occurrence of eulachon, Thaleichthys pacificus, in streams tributary 

to Humboldt Bay, California. California Fish and Game 82, 147-148. 

Jensen A. S. & Silber G. K. (2003) Large whale ship strike database. In: NOAA Technical 

Memorandum NMFS-OPR-25. 

Jewett E. B., Hines A. H. & Ruiz G. M. (2005) Epifaunal disturbance by periodic low levels of 

dissolved oxygen: native vs. invasive species response. Marine Ecology-Progress Series 304, 31-

44. 

Jewitt-Smith J., McMillan C., Kenworthy W. J. & Bird K. T. (1997) Flowering and genetic 

banding patterns of Halophila johnsonii and conspecifics. Aquatic Botany 59, 323-331. 

Johannsson O. E., Mills E. L. & O'Gorman R. (1991) Changes in the Nearshore and Offshore 

Zooplankton Communities in Lake Ontario: 1981-88. Canadian Journal of Fisheries and 

Aquatic Sciences 48, 1546-1557. 

Johanos T. C., Becker B. L., Baker J. D., Ragen T. J., Gilmartin W. G. & Gerrodette T. (2010) 

Impacts of sex ratio reduction on male aggression in the Critically Endangered Hawaiian monk 

seal Monachus schauinslandi. Endangered Species Research 11, 123-132. 

Johanos T. C., Becker B. L. & Ragen T. J. (1994) Annual reproductive cycle of the female 

Hawaiian monk seal (Monachus schauinslandi). Marine Mammal Science 10, 13-30. 

Johanos T. C. B. L. B. & Ragen T. J. (1994) Annual reproductive cycle of the female Hawaiian 

monk seal (Monachus schauinslandi). Marine Mammal Science 10, 13-30. 

Johnson A. & Newman A. (1983) Water quality in the gap-to-gap reach of the Yakima River, 

June - October 1982. (ed W. D. o. Ecology) p. 52, Olympia, Washington. 

Johnson A. M. & Kridler E. (1983) Interisland movement of Hawaiian monk seals. 'Elepaio 44, 

43-45. 

Johnson B. W. & Johnson P. A. (1978) The Hawaiian monk seal on Laysan Island: 1977. Final 

report to U.S. Marine Mammal Commission in fulfillment of contract MM7AC009, Report No. 

MMC-77/05. p. 38. U.S. Department of Commerce, National Technical Information Service, 

Springfield, VA, PB-285-428. 

Johnson C. K., Tinker M. T., Estes J. A., Conrad P. A., Staedler M., Miller M. A., Jessup D. A. 

& Mazet J. A. K. (2009) Prey choice and habitat use drive sea otter pathogen exposure in a 
437

resource-limited coastal system. Proceedings of the National Academy of Sciences of the United 

States of America 106, 2242-2247. 

Johnson D. W. (2006) Predation, habitat complexity, and variation in density-dependent 

mortality of temperate reef fishes. Ecology 87, 1179-1188. 

Johnson J. H., Dropkin D. S., Warkentine B. E., Rachlin J. W. & Andrews W. D. (1997) Food 

habits of Atlantic sturgeon off the central New Jersey coast. Transactions of the American 

Fisheries Society 126, 166-170. 

Johnson J. H. & Wolman A. A. (1984) The humpback whale, Megaptera novaeangliae. Marine 

Fisheries Review 46, 30-37. 

Johnson L. E. & Carlton J. T. (1996) Post-establishment spread in large-scale invasions: 

Dispersal mechanisms of the zebra mussel Dreissena Polymorpha. Ecology 77, 1686-1690. 

Johnson M. L. (1964) San Lorenzo River System, Santa Cruz County. In: CDFG Region 3, 

unpublished file report. Administrative Record for coho salmon in Scott and Waddell Creeks. . 

National Marine Fisheries Service, Southwest Regional Office, Long Beach, California. 

Johnson S. A. & Ehrhart L. M. (1994) Nest-site fidelity of the Florida green turtle. In: 

Proceedings of the Thirteenth Annual Symposium on Sea Turtle Biology and Conservation (eds 

B. A. Schroeder and B. E. Witherington) p. 83. 

Johnson S. W., Murphy M. L. & Csepp D. J. (2003) Distribution, habitat, and behavior of 

rockfishes, Sebastes spp., in nearshore waters of southeastern Alaska: observations from a 

remotely operated vehicle. Environmental Biology of Fishes 66, 259-270. 

JOI/USSSP C. G. E. a. (2003) The interplay of collisional tectonics and late Cenozoic glacial 

climate in Alaska and the northeastern Pacific Ocean. Continental Dynamics Program-Earth 

Sciences Division-National Science Foundation and Joint Oceanographic Institutions/U.S. 

Science Support Program, Austin, Texas. 

Jones R. E. (1981) Food habits of smaller marine mammals from northern California. 

Proceedings of the California Academy of Science 42, 409-433. 

Jordan R. M. & Beland K. F. (1981) Atlantic salmon spawning and evaluation of natural 

spawning success. Atlantic Sea Run Salmon Commission, Augusta, Maine. 

Josefsson M. & Andersson B. (2011) The environmental consequences of alien species in the 

Swedish Lakes Mälaren, Hjälmaren,Vänern and Vättern. Ambio 30, 514-521. 

Joy J. (2002) Upper Yakima River basin suspended sediment and organochlorine pesticide total 

maximum daily load evaluation. Washington Department of Ecology, Olympia, Washington. 

Joy J. & Madrone A. (2002) Data summary: upper Yakima River basin suspended sediment and 

organochlorine pesticide TMDL evaluation. p. 32. Washington Department of Ecology, 

Olympia, Washington. 

Jurasz C. M. & Jurasz V. (1979) Feeding modes of the humpback whale, Megaptera 

novaeangliae, in southeast Alaska. Scientific Reports of the Whales Research Institute, Tokyo 31, 

69-83. 

Kagan J. S., Hak J. C., Csuti B., Kiilsgaard C. W. & Gaines E. P. (1999) Oregon gap analysis 

project final report:a geographic approach to planning for biological diversity. Oregon Natural 

Heritage Program, Portland, Oregon. 

Kahl S. (2001) A public forum for information and perspectives on dioxin in the Maine 

environment. Waterlines 6, 1-10. 

Kahn A. E. & Durako M. J. (2009) Photosynthetic tolerances to desiccation of the co-occurring 

seagrasses Halophila johnsonii and Halophila decipiens. Aquatic Botany 90, 195-198. 
438

Kahnle A. W., Hattala K. A. & McKown K. A. (2007) Status of Atlantic Sturgeon of the Hudson 

River Estuary, New York, USA. American Fisheries Society Symposium 56, 347-363. 

Kahnle A. W., Hattala K. A., McKown K. A., Shirey C. A., Collins M. R., Squiers J. T. S. & 

Savoy T. (1998a) Stock status of Atlantic sturgeon of Atlantic coast estuaries. Draft III. In: 

Report for the Atlantic States Marine Fisheries Commission. 

Kahnle A. W., Hattala K. A., McKown K. A., Shirey C. A., Collins M. R., T.S. Squiers J. & 

Savoy T. (1998b) Atlantic Sturgeon Stock Assessment: Peer Review Report. p. 123. Atlantic 

States Marine Fisheries Commission, Washington, D.C. . 

Kamezaki N. (1989) The nesting sites of sea turtles in the Ryukyu Archipelago and Taiwan. In: 

Current herpetology in East Asia (eds M. Matsui, T. Hikida and R. C. Goris) pp. 342-348. 

Herpetological Society of Japan, Kyoto. 

Kamezaki N., Matsuzawa Y., Abe O., Asakawa H., Fujii T., Goto K., Hagino S., Hayami M., 

Ishii M., Iwamoto T., Kamata T., Kato H., Kodama J., Kondo Y., Miyawaki I., Mizobuchi K., 

Nakamura Y., Nakashima Y., Naruse H., Omuta K., Samejima M., Suganuma H., Takeshita H., 

Tanaka T., Toji T., Uematsu M., Yamamoto A., Yamato T. & Wakabayashi I. (2003) 

Loggerhead turtles nesting in Japan. In: Loggerhead sea turtles (eds A. B. Bolten and B. E. 

Witherington) pp. 210-217. Smithsonian Books, Washington D.C. 

Kamezaki N., Oki K., Mizuno K., Toji T. & Doi O. (2002) First nesting record of the leatherback 

turtle, Oermochelys coriacea, in Japan. Current Herpetology 21, 95-97. 

Kammerer J. C. (1990) Largest rivers in the United States. In: Water Fact Sheet p. 2. U.S. 

Department of the Interior, U.S. Geological Survey, Reston, Virginia. 

Kanda N., Goto M. & Pastene L. A. (2006) Genetic characteristics of western North Pacific sei 

whales, Balaenoptera borealis, as revealed by microsatellites. Marine Biotechnology 8, 86-93. 

Kaniewska P., Campbell P. R., Fine M. & Hoegh-Guldberg O. (2009) Phototropic growth in a 

reef flat acroporid branching coral species. The Journal of Experimental Biology 212, 662-667. 

Kaplan C. C., White G. C. & Noon B. R. (2008) Neonatal survival of Steller sea lions 

(Eumetopias jubatus). Marine Mammal Science 24, 443-461. 

Kapurisinghe T. (2006) Status and conservation of marine turtles in Sri Lanka. In: Marine 

Turtles of the Indian Subcontinent (eds K. Shanker and B. C. Choudhury) pp. 173-187. 

Universities Press, India. 

Katona S. K. & Beard J. A. (1990) Population size, migrations and feeding aggregations of the 

humpback whale (Megaptera novaeangliae) in the western North Atlantic Ocean. Report of the 

International Whaling Commission, 295-306. 

Katona S. K., Rough V. & Richardson D. T. (1983) A field guide to the whales, porpoises, and 

seals of the Gulf of Maine and eastern Canada. Charles Scribner's Sons, New York, NY. 

Kawamura G., Naohara T., Tanaka Y., Nishi T. & Anraku K. (2009) Near-ultraviolet radiation 

guides the emerged hatchlings of loggerhead turtles Caretta caretta (Linnaeus) from a nesting 

beach to the sea at night. MARINE AND FRESHWATER BEHAVIOUR AND PHYSIOLOGY 42, 

19-30. 

Keefer M. L., Peery C. A. & Caudill C. C. (2008) Migration Timing of Columbia River Spring 

Chinook Salmon: Effects of Temperature, River Discharge, and Ocean Environment. 

Transactions of the American Fisheries Society 137, 1120-1133. 

Keinath J. A., Musick J. A. & Barnard D. E. (1996) Abundance and distribution of sea turtles off 

North Carolina. OCS Study, MMS 95-0024, 156. 
439

Keleher C. J. & Rahel F. J. (1996) Thermal limits to salmonid distributions in the Rocky 

Mountain region and potential habitat loss due to global warming: A geographic information 

system (GIS) approach. Transactions of the American Fisheries Society 125, 1-13. 

Keller J. M., Kannan K., Taniyasu S., Day R. D., Arendt M. D., Segars A. L. & Kucklick J. R. 

(2005) Perfluorinated compounds in the plasma of loggerhead and Kemp’s ridley sea turtles 

from the southeastern coast of the United States. Environmental Science and Technology 39, 

9101-9108. 

Keller J. M., Kucklick J. R., Harms C. A. & McClellan-Green P. D. (2004a) Organochlorine 

contaminants in sea turtles: Correlations between whole blood and fat. Environmental 

Toxicology and Chemistry 23, 726-738. 

Keller J. M., Kucklick J. R. & McClellan-Green P. D. (2004b) Organochlorine contaminants in 

loggerhead sea turtle blood: Extraction techniques and distribution among plasma, and red blood 

cells. Archives of Environmental Contamination and Toxicology 46, 254-264. 

Keller J. M., Kucklick J. R., Stamper M. A., Harms C. A. & McClellan-Green P. D. (2004c) 

Associations between organochlorine contaminant concentrations and clinical health parameters 

in loggerhead sea turtles from North Carolina, USA. Environmental Health Parameters 112, 

1074-1079. 

Keller J. M. & McClellan-Green P. (2004) Effects of organochlorine compounds on cytochrome 

P450 aromatase activity in an immortal sea turtle cell line. Marine Environmental Research 58, 

347-351. 

Keller J. M., McClellan-Green P. D., Kucklick J. R., Keil D. E. & Peden-Adams M. M. (2006) 

Turtle immunity: Comparison of a correlative field study and in vitro exposure experiments. 

Environmental Health Perspectives 114, 70-76. 

Kelly B. C., Ikonomou M. G., Blair J. D., Surridge B., Hoover D., Grace R. & Gobas. F. A. P. C. 

(2009) Perfluoroalkyl contaminants in an arctic marine food web: Trophic magnification and 

wildlife exposure. Environmental Science AND Technology 43, 4037-4043. 

Kemmerich J. (1945) A review of the artificial propagation and transplantation of the sockeye 

salmon of the Puget Sound area in the State of Washington conducted by the federal government 

from 1896 to 1945. In: Report to the Regional Director. U.S. Fish and Wildlife Service. 

Keniry T., Marsden, J.E. (1995) Zebra mussels in southwestern Lake Michigan. . In: LaRoe, 

E.T., Farris, G.S., Puckett, C.E., Doran, P.D., Mac, M.J. (Eds.), Our Living Resources: A Report 

to the Nation on the Distribution, Abundance, and Health of U.S. Plants, Animals and 

Ecosystems., 445-448. 

Kenney R. D. (2001) Anomalous 1992 spring, and summer right whale (Eubalaena glacialis) 

distributions in the Gulf of Maine. Journal of Cetacean and Research Management (special 

issue) 2, 209-223. 

Kenney R. D. (2007) Right whales and climate change: Facing the prospect of a greenhouse 

future. In: The Urban Whale: North Atlantic Right Whales at the Crossroads. S. D. Kraus AND 

R. Rolland (eds.). p.436-459. Harvard University Press, Cambridge, MA. ISBN 0-674-02327-7. 

543pp. 

Kenney R. D., Hyman M. A. M. & Winn H. E. (1985a) Calculation of standing stocks and 

energetic requirements of the cetaceans of the northeast United States outer continental shelf. In: 

NOAA Technical Memorandum NMFS-F/NEC-41. 

Kenney R. D., Hyman M. A. M. & Winn H. E. (1985b) Calculation of standing stocks and 

energetic requirements of the cetaceans of the northeast United States Outer Continental Shelf. 

U.S. Department of Commerce, NOAA Technical Memorandum NMFS-F/NEC-41. 
440

Kenney R. D., Mayo C. A. & Winn H. E. (2001) Migration and foraging strategies at varying 

spatial scales in western North Atlantic right whales: a review of hypotheses. Journal of 

Cetacean Research and Management, 251-260. 

Kenney R. D., Winn H. E. & Macaulay M. C. (1995) Cetaceans in the Great South Channel, 

1979-1989: Right whale (Eubalaena glacialis). Continental Shelf Research 15, 385-414. 

Kent M. L., Whyte J. N. C. & LaTrace C. (1995) Gill lesions and mortality in seawater penreared Atlantic salmon Salmo salar associated with a dense bloom of Skeletonema costatum and 

Thalassiosira species. Diseases of Aquatic Organisms 22, 77-81. 

Kenworthy W. J. (1993) The distribution, abundance, and ecology of Halophila johnsonii 

Eiseman in the lower Indian River, Florida. National Marine Fisheries Service, Silver Spring, 

Maryland. 

Kenworthy W. J. (1997) An updated biological status review and summary of the proceedings of 

a workshop to review the biological status of the seagrass Halophila johnsonii Eiseman. 

Southeast Fisheries Science Center, National Marine Fisheries Service, Beaufort, North Carolina. 

Kenworthy W. J. (2000) The role of sexual reproduction in maintaining populations of Halophila 

decipiens: Implications for the biodiversity and conservation of tropical seagrass ecosystems. 

Pacific Conservation Biology 5, 260-268. 

Kenworthy W. J. (2003) Proceedings of the Johnson's seagrass monitoring workshop. St. 

Petersburg, Florida. 

Kenyon K. W. & Rice D. W. (1959) Life history of the Hawaiian monk seal. Pacific Science 53, 

215-252. 

Kenyon K. W. & Rice D. W. (1961) Abundance and distribution of the Steller sea lion. Journal 

of Mammalogy 42, 223-234. 

Kerfoot W. C., Yousef F., Hobmeier M. M., Maki R. P., Jarnagin S. T. & Churchill J. H. (2011) 

Temperature, recreational fishing and diapause egg connections: dispersal of spiny water fleas 

(Bythotrephes longimanus). Biological Invasions. 

Kerwin J. (1999) Salmon habitat limiting factors report for the Puyallup River Basin (Water 

Resource Inventory Area 10). Washington Conservation Commission, Olympia, Washington. 

Ketten D. (1998) Marine Mammal Auditory Systems: A summary of audiometric and 

anatomical data and its implications for underwater acoustic impacts. 

Khuzin R. S. (1961) The procedure used in age determination and materials relating to 

reproduction of the beluga. Nauchno-Tekh. Byull. TINRO 1, 58-60. 

Kideys A. E., Roohi A., Bagheri S., Finenko G. & Kamburska L. (2005) Impacts of invasive 

ctenophores on the fisheries of the Black Sea and Caspian Sea. Oceanography 18. 

Kieffer M. C. & Kynard B. (1993) Annual movements of shortnose and Atlantic sturgeons in the 

Merrimack River, Massachusetts. Transactions of the American Fisheries Society 122, 1088-

1103. 

Killgore K. J., Morgan R. P. & Rybicki N. B. (1989) Distribution and Abundance of Fishes 

Associated with Submersed Aquatic Plants in the Potomac River. North American Journal of 

Fisheries Management 9, 101-111. 

Kim G. B., Tanabe S., Tatsukawa R., Loughlin T. R. & Shimazaki K. (1996) Characteristics of 

butyltin accumulations and its biomagnification in Steller sea lion (Eumetopias jubatus). 

Environmental Toxicology and Chemistry 15, 2043-2048. 

King County. (2001) Reconnaissance assessment of the state of the nearshore ecosystem: eastern 

shore of Central Puget Sound, including Vashon and Maury Islands (WRIAs 8 and 9). King 

County Department of Natural Resources, Seattle, Washington. 
441

King T. L., Lubinski B. A. & Spidle A. P. (2001) Microsatellite DNA variation in Atlantic 

sturgeon (Acipenser oxyrinchus oxyrinchus) and cross-species amplification in the 

Acipenseridae. Conservation Genetics 2, 103-119. 

Kinghorn A. M. M. H., Outridge P. & Chan H. M. (2008) Teeth as biomonitors of selenium 

concentrations in tissues of beluga whales (Delphinapterus leucas). Science of the Total 

Environment 402, 43-50. 

Kipp R., Bailey S. A., MacIsaac H. J. & Ricciardi A. (2010) Transoceanic ships as vectors for 

nonindigenous freshwater bryozoans. Diversity and Distributions 16, 77-83. 

Kistner D. & Pettigrew N. (2001) A variable turbidity maximum in the Kennebec estuary, 

Maine. Estuaries and Coasts 24, 680-687. 

Kiszka J., Ersts P. J. & Ridoux V. (2010a) Structure of a toothed cetacean community around a 

tropical island (Mayotte, Mozambique Channel). African Journal of Marine Science 32, 543-

551. 

Kiszka J., Vely M. & Breysse. O. (2010b) Preliminary account of cetacean diversity and 

humpback whale (Megaptera novaeangliae) group characteristics around the Union of the 

Comoros (Mozambique Channel). Mammalia 74, 51-56. 

Kitching J. A. & Thain V. M. (1983) The ecological impact of the sea urchin Paracentrotus 

lividus (Lamarck) in Lough Ine, Ireland. Philosophical Transactions of the Royal Society of 

London 300, 513-552. 

Klima E. F., Gitschlag G. R. & Renaud M. L. (1988) Impacts of the explosive removal of 

offshore petroleum platforms on sea turtles and dolphins. Marine Fisheries Review 50, 33-42. 

Klinkhart E. G. (1966) The beluga whale in Alaska. In: Alaska Department of Fish and Game. 

Federal Aid in Wildlife Restoration Project Report, Volume VII. Project W-6-R and W-14-R. 

Knowlton A. R. & Kraus S. D. (2001) Mortality and serious injury of northern right whales 

(Eubalaena glacialis) in the western North Atlantic Ocean. Journal of Cetacean Research and 

Management Special Issue, 193 - 208. 

Knowlton A. R., Kraus S. D. & Kenney R. D. (1994) Reproduction in North Atlantic right 

whales (Eubalaena glacialis). Canadian Journal of Zoology 72, 1297-1305. 

Kobayashi D. R., Polovina J. J., Parker D. M., Kamezaki N., Cheng I.-J., Uchida I., Dutton P. H. 

& Balazs G. H. (2008) Pelagic habitat characteriation of loggerhead sea turtles, Caretta caretta, 

in the North Pacific Ocean (1997-2006): insights from satellite tag tracking and remotely sensed 

data. Journal of Marine Biology and Ecology 356, 96-114. 

Kolar C. S. & Lodge D. M. (2001) Progress in invasion biology: predicting invaders. Trends in 

Ecology & Evolution 16, 199-204. 

Kornicker L. S. & Boyd D. W. (1962) Shallow-water geology and environments of Alacran Reef 

complex, Campeche Bank, Mexico. Bulletin of the American Association of Petroleum Geology

46, 640-673. 

Koski K. (2004) Final Program Report: Soundwatch Public Outreach/Boater Education Project. 

The Whale Museum, Friday Harbor, Washington. 

Koski K. (2006a) 2004-2005 final program report. p. 25. Soundwatch Public Outreach Boater 

Education Project 

Koski K. (2006b) 2004-2005 Final Program Report: Soundwatch Public Outreach/Boater 

Education Project. The Whale Museum, Friday Harbor, Washington. 

Koski K. (2007a) 2006. In: Final Program Report: Soundwatch Public Outreach/Boater 

Education Project. The Whale Museum, Friday Harbor, Washington. 
442

Koski K. (2007b) 2006 final program report. Soundwatch Public Outreach Boater Education 

Project 

Kostow K. (1995) Biennial report on the status of wild fish in Oregon. Oregon Department of 

Fish and Wildlife. 

Kragh Boye T., Simon M. & Madsen. P. T. (2010) Habitat use of humpback whales in 

Godthaabsfjord, West Greenland, with implications for commercial exploitation. Journal of the 

Marine Biological Association of the United Kingdom 90, 1529-1538. 

Krahn M., Ford M. J., Perrin W. F., Wade P. R., Angliss R. P., Hanson M. B., Taylor B. L., 

Ylitalo G. M., Dahlheim M. E., Stein J. E. & Waples R. S. (2004a) 2004 Status review of 

Southern Resident killer whales (Orcinus orca) under the Endangered Species Act. In: NOAA 

Technical Memorandum NMFS-NWFSC-62. U.S. Department of Commerce. 

Krahn M. M., Ford M. J., Perrin W. F., Wade P. R., Angliss R. P., Hanson M. B., Taylor B. L., 

Ylitalo G. M., Dahlheim M., Stein J. E. & Waples. R. S. (2004b) 2004 status review of southern 

resident killer whales (Orcinus orca) under the Endangered Species Act. NOAA Technical 

Memorandum NMFS-NWFSC-62. 73p. 

Krahn M. M., Hanson M. B., Baird R. W., Boyer R. H., Burrows D. G., Emmons C. K., Ford J. 

K. B., Jones L. L., Noren D. P., Ross P. S., Schorr G. S. & Collier. T. K. (2007) Persistent 

organic pollutants and stable isotopes in biopsy samples (2004/2006) from Southern Resident 

killer whales. (Orcinus orca). Marine Pollution Bulletin 54, 1903-1911. 

Krahn M. M., Hanson M. B., Schorr G. S., Emmons C. K., Burrows D. G., Bolton J. L., Baird R. 

W. & Ylitalo G. M. (2009) Effects of age, sex and reproductive status on persistent organic 

pollutant concentrations in ‘‘Southern Resident” killer whales. Marine Pollution Bulletin. 

Krahn M. M., Herman D. P., Ylitalo G. M., Sloan C. A., Burrows D. G., Hobbs R. C., Mahoney 

B. A., Yanagida G. K., Calambokidis J. & Moore. S. E. (2004c) Stratification of lipids, fatty 

acids and organochlorine contaminants in blubber of white whales and killer whales. 

(Delphinapterus leucas, Orcinus orca). Journal of Cetacean Research and Management 6, 175-

189. 

Krahn M. M., Wade P. R., Kalinowski S. T., Dahlheim M. E., Taylor B. L., Hanson M. B., 

Ylitalo G. M., Angliss R. P., Stein J. E. & Waples R. S. (2002a) Status review of Southern 

Resident killer whales (Orcinus orca) under the Endangered Species Act. U.S. Dept. Commer., 

NOAA Tech. Memo. NMFS-NWFSC-54, 133p. 

Krahn M. M., Wade P. R., Kalinowski S. T., Dahlheim M. E., Taylor B. L., Hanson M. B., 

Ylitalo G. M., Angliss R. P., Stein J. E. & Waples. R. S. (2002b) Status review of southern 

resident killer whales (Orcinus orca) under the Endangered Species Act. NOAA Technical 

Memorandum NMFS-NWFSC-54. 159pp. 

Kramer D. E. & O'Connell V. M. (1995) Guide to Northeast Pacific rockfishes. Genera Sebastes 

and Sebastolobus. In: Marine Advisory Bulletin. Alaska Sea Grant. 

Kraus S. D., Brown M. W., Caswell H., Clark C. W., Fujiwara M., Hamilton P. K., Kenney R. 

D., Knowlton A. R., Landry S., Mayo C. A., McLellan W. A., Moore M. J., Nowacek D. P., 

Pabst D. A., Read A. J. & Rolland R. M. (2005) North Atlantic right whales in crisis. In: Science 

pp. 561-562. 

Kraus S. D., Kenney R. D., Knowlton A. R. & Ciano J. N. (1993) Endangered right whales of the 

southwestern North Atlantic. In: OCS Study MMS 930024. Prepared by the New England 

Aquarium. Herndon, Virginia: U.S. Department of the Interior, Minerals Management Service, 

Atlantic OCS Region. 
443

Kraus S. D., Pace R. M. & Frasier T. R. (2007) High investment, low return: The strange case of 

reproduction in Eubalaena glacialis. In: The urban whale: North Atlantic right whales at the 

crossroads (eds S. D. Krauss and R. M. Rolland) pp. 172-199. Harvard University Press, 

Cambridge, Massachusetts. 

Kriete B. (2002) Bioenergetic changes from 1986 to 2001 in the Southern Resident killer whale 

population, Orcinus orca. Orca Relief Citizens’ Alliance, Friday Harbor, Washington. 26p. 

Krigsman L. M. (2000) A review of larval duration for Pacific coast temperate reek fishes; 

including kelp rockfish, Sebastes atrovirens. University of California at Santa Cruz Santa Cruz, 

California. 

Kruckeberg A. R. (1991) The natural history of Puget Sound country. 

Kruse G. O. & Scarnecchia D. L. (2001) Assessment of bioaccumulated metal and 

organochlorine compounds in relation to physiological biomarkers in Kootenai River white 

sturgeon. Journal of Applied Ichthyology 18, 430-438. 

Kruse S. (1991) The interactions between killer whales and boats in Johnstone Strait, B.C. In: 

Dolphin societies: Discoveries and puzzles (eds K. Pryor and K. S. Norris) pp. 149-159. 

University of California Press, Berkeley, California. 

Kuller Z. (1999) Current status and conservation of marine turtles on the Mediterranean coast of 

Israel. Marine Turtle Newsletter 86, 3-5. 

Kynard B. (1997a) Life history, latitudinal patterns and status of shortnose sturgeon, Acipenser 

brevirostrum. Environmental Biology of Fishes 48, 319-334. 

Kynard B. (1997b) Life history, latitudinal patterns, and status of the shortnose sturgeon, 

Acipenser brevirostrum. Environmental Biology of Fishes 48, 319-334. 

Kynard B., Breece M., Atcheson M., Kieffer M. & Mangold M. (2009) Life history and status of 

shortnose sturgeon (Acipenser brevirostrum) in the Potomac River. Journal of Applied 

Ichthyology 25, 34-38. 

Kynard B. & Horgan M. (2002) Ontogenetic behavior and migration of Atlantic sturgeon, 

Acipenser oxyrinchus oxyrinchus, and shortnose sturgeon, A. brevirostrum, with notes on social 

behavior. Environmental Biology of Fishes 63, 137-150. 

Kynard B., Horgan M., Kieffer M. & Seibel D. (2000) Habitats used by shortnose sturgeon in 

two Massachusetts rivers, with notes on estuarine Atlantic sturgeon: A hierarchical approach. 

Transactions of the American Fisheries Society 129, 487-503. 

Lachmuth C. L., Barrett-Lennard L. G., Steyn D. Q. & Milsom W. K. (2011) Estimation of 

southern resident killer whale exposure to exhaust emissions from whale-watching vessels and 

potential adverse health effects and toxicity thresholds. (Orcinus orca). Marine Pollution Bulletin

62, 792-805. 

Lackey R. T. (2003) Pacific Northwest salmon: Forecasting their status in 2100. Reviews in 

Fisheries Science 11, 35-88. 

Lackey R. T. (2009) Challenges to sustaining diadromous fishes through 2100: lessons learned 

from western North America. In Challenges for diadromous fishes in a dynamic global 

environment. In: American Fisheries Society Symposium (eds A. Haro, K. L. Smith, R. A. 

Rulifson, C. M. Moffitt, R. J. Klauda, M. J. Dadswell, R. A. Cunjak, J. A. Cooper, K. L. Beal 

and T. S. Avery), Bethesda, Maryland. 

LADEQ. (2010) Beach sweep and inland waterway cleanup. Louisiana Department of 

Environmental Quality Litter Reduction and Public Action. 
444

Lafferty K. D. & Kuris A. M. (1993) Mass mortality of abalone Haliotis cracherodii on the 

California Channel Islands: Tests of epidemiological hypotheses. Marine Ecology Progress 

Series 96, 239-248. 

Lafferty K. D. & Kuris A. M. (1996) Biological Control of Marine Pests. Ecology 77, 1989-

2000. 

Lagueux C. J. (1998) Marine Turtle fishery of Caribbean Nicaragua: human Use Patterns and 

Harvest Trends. p. 236. University of Florida. 

Lagueux C. J., Campbell C. L. & McCoy W. A. (2003) Nesting, and conservation of the 

hawksbill turtle, Eretmochelys imbricata, in the Pearl Cays, Nicaragua. Chelonian Conservation 

and Biology 4, 588-602. 

Laidig T. E., Chess J. R. & Howard D. F. (2007) Relationship between abundance of juvenile 

rockfishes (Sebastes spp.) and environmental variables documented off northern California and 

potential mechanisms for the covariation. Fishery Bulletin 105, 39-48. 

Laidre K. L., Shelden K. E. W., Rugh D. J. & Mahoney B. A. (2000) Beluga, Delphinapterus 

leucas, distribution, and survey effort in the Gulf of Alaska. Marine Fisheries Review 62, 27-36. 

Laist D. W., Coe J. M. & O’Hara K. J. (1999) Marine debris pollution. In: Conservation and 

management of marine mammals (eds J. Twiss and R. R. Reeves) pp. 342-366. Smithsonian 

Institution Press, Washington, D.C. 

Laist D. W., Knowlton A. R., Mead J. G., Collet A. S. & Podesta M. (2001) Collisions between 

ships and whales. Marine Mammal Science 17, 35-75. 

Lake J. (2004) Invasion success of exotic plants in natural ecosystems: the role of disturbance, 

plant attributes and freedom from herbivores. Biological Conservation 117, 215-226. 

Lake J., Haebler L. R., McKinney R., Lake C. A. & Sadove S. S. (1994) PCBs and other 

chlorinated organic contaminants in tissues of juvenile Kemp's ridley turtles (Lepidochelys 

kempii). Marine Environmental Research 38, 313-327. 

Lal A., Arthur R., Marbà N., Lill A. W. T. & Alcoverro T. (2010) Implications of conserving an 

ecosystem modifier: Increasing green turtle (Chelonia mydas) densities substantially alters 

seagrass meadows. Biological Conservation in press, in press. 

Lamb A. & Edgell P. (1986) Coastal fishes of the Pacific northwest. Harbour Publishing Co. Ltd. 

Lambert W. J., Levin P. S. & Berman J. (1992) Changes in the structure of a New England 

(USA) kelp bed: the effects of an introduced species? Marine Ecology-Progress Series 88, 303-

307. 

Lambertsen R. H. (1992) Crassicaudosis: a parasitic disease threatening the health and 

population recovery of large baleen whales. Rev. Sci. Technol., Off. Int. Epizoot. 11, 1131-1141. 

Lammers M. O., Fisher-Pool P. I., Au W. W. L., Meyer C. G., Wong K. B. & Brainard R. E. 

(2011) Humpback whale Megaptera novaeangliae song reveals wintering activity in the 

northwestern Hawaiian Islands. Marine Ecology Progress Series 423, 261-268. 

Lance M. M. & Jeffries S. J. (2007) Temporal and spatial variability of harbor seal diet in the 

San Juan archipelago. Washington Department of Fish and Wildlife. 

Lander M. E., Loughlin T. R., Logsdon M. G., VanBlaricom G. R. & Fadely B. S. (2010) 

Foraging effort of juvenile Steller sea lions Eumetopias jubatus with respect to heterogeneity of 

sea surface temperature. Endangered Species Research 10, 145-158. 

Landry A. M., Jr. & Costa D. (1999) Status of sea turtle stocks in the Gulf of Mexico with 

emphasis on the Kemp's ridley. In: The Gulf of Mexico large marine ecosystem: Assessment, 

sustainability, and management (eds H. Kumpf, K. Steidinger and K. Sherman) pp. 248-268. 

Blackwell Science, Malden, Massachusetts. 
445

Landry A. M. J., Costa D. T., Kenyon F. L., Hadler M. C., Coyne M. S., Hoopes L. A., Orvik L. 

M., St.John K. E. & VanDenburg K. J. (1996) Population Dynamics and Index Habitat 

Characterization for Kemp's Ridley Sea Turtles in Nearshore Waters of the Northwestern Gulf of 

Mexico. Report of Texas A&M Research Foundation pursuant to NOAA Award No. 

NA57FF0062, 153. 

Landsberg J. H., Balazs G. H., Steidinger K. A., Baden D. G., Work T. M. & Russell D. J. 

(1999) The potential role of natural tumor promoters in marine turtle fibropapillomatosis. 

Journal of Aquatic Animal Health 11, 199-210. 

Lane R. C. & Taylor W. A. (1997) Washington's wetland resources. U.S. Geological Survey. 

Langer O. E., Shepherd B. G. & Vroom P. R. (1977) Biology of the Nass River eulachon 

(Thaleichthys pacificus). 

Larson Z. S. & Belchik M. R. (2000) A preliminary status review of eulachon and Pacific 

lamprey in the Klamath River Basin. Yurok Tribal Fisheries Program, Klamath, California. 

Lassalle G., Crouzet P., Gessner J. & Rochard E. (2010) Global warming impacts and 

conservation responses for the critically endangered European Atlantic sturgeon. Biological 

Conservation. 

Last P. R. & Stevens J. D. (1994) Sharks and rays of Australia. CSIRO Australia, East 

Melbourne, Australia. 

Laufle J. C., Pauley G. B. & Shepard M. F. (1986) Species Profiles: Life Histories and 

Environmental Requirements of Coastal Fishes and Invertebrates (Pacific Northwest) - Coho 

Salmon U.S. Fish and Wildlife Service Biological Report 82(11.48). U.S. Army Corps of 

Engineers, TR EL-82-4. 18p. 

Laurance W. F., Fay J. M., Parnell R. J., Sounguet G., Formia A. & Lee M. D. (2008) Does 

rainforest logging threaten endangered sea turtles? Oryx 42, 245-251. 

Law R. J. & Hellou J. (1999) Contamination of fish and shellfish following oil spill incidents. 

Environmental Geoscience 6, 90-98. 

Lawrence J. E. (2008) Furtive foes: algal viruses as potential invaders. ICES Journal of Marine 

Science 65, 716-722. 

Lazar B. & Gračan R. (2010) Ingestion of marine debris by loggerhead sea turtles, Caretta 

caretta, in the Adriatic Sea. Marine Pollution Bulletin. 

LCFRB. (2004) Lower Columbia Salmon Recovery and Fish and Wildlife Subbasin Plan. 

LCFRB. (2010) Washington lower Columbia salmon recovery & fish and wildlife subbasin plan. 

Lower Columbia Fish Recovery Board, Olympia, Washington. 

LCREP. (2007) Lower Columbia River and estuary ecosystem monitoring: Water quality and 

salmon sampling report. Lower Columbia River Estuary Partnership, Portland, Oregon. 

Le Boeuf B. J. & Crocker D. E. (2005) Ocean climate and seal condition. BMC Biology 3, 9. 

Le Gall J., Bosc Y. P., Chateau D. & Taquet M. (1986) Estimation du nombre de tortues vertes 

femelles adultes Chelonia mydas par saison de ponte á Tromelin et Europa (Océan Indien) 

(1973-1985). Océanographie Tropicale 21, 3-22. 

Lea R. N., McAllister R. D. & VenTresca D. A. (1999) Biological aspects of nearshore 

rockfishes of the genus Sebastes from central California. Fishery Bulletin 177, 1-109. 

Leaper R., Cooke J., Trathan P., Reid K., Rowntree V. & Payne R. (2006) Global climate drives 

southern right whale (Eubalaena australis) population dynamics. Biology Letters 2, 289-292. 

Learmonth J. A., Macleod C. D., Santos M. B., Pierce G. J., Crick H. Q. P. & Robinson R. A. 

(2006) Potential effects of climate change on marine mammals. Oceanography and Marine 

Biology: an Annual Review 44, 431-464. 
446

Leblanc A. M. & Wibbels T. (2009) Effect of Daily Water Treatment on Hatchling Sex Ratios in 

a Turtle With Temperature-Dependent Sex Determination. Journal of Experimental Zoology 

Part a-Ecological Genetics and Physiology 311A, 68-72. 

Lederer A., Massart J. & Janssen J. (2006) Impact of Round Gobies (Neogobius melanostomus) 

on Dreissenids (Dreissena polymorpha and Dreissena bugensis) and the Associated 

Macroinvertebrate Community Across an Invasion Front. Journal of Great Lakes Research 32, 

1-10. 

Lee J. S., Tanabe S., Umino H., Tatsukawa R., Loughlin T. R. & Calkins D. C. (1996) Persistent 

organochlorines in Steller sea lion (Eumetopias jubatus) from the bulk of Alaska and the Bering 

Sea, 1976-1981. Marine Pollution Bulletin 32, 535-544. 

Lee Long W. J., Coles R. G. & McKenzie L. J. (2000) Issues for seagrass conservation 

management in Queensland. Pacific Conservation Biology 5, 321-328. 

Lee Lum L. (2003) An assessment of incidental turtle catch in the gillnet fishery in Trinidad and 

Tobago, West Indies. p. 22. Institute for Marine Affairs, Chaguaramas, Trinidad. 

Lee Y. W. & Sampson D. B. (2009) Dietary variations in three co-occurring rockfish species off 

the Pacific Northwest during anomalous oceanographic events in 1998 and 1999. Fishery 

Bulletin 107, 510-522. 

Leet W., Dewees C., Klingbeil R. & Larson E. (2001) California’s living marine resources: A 

status report. In: California Department of Fish and Game pp. 89-97. 

Leff L. G., Burch J. L. & McArthur J. V. (1990) Spatial distribution, seston removal, and 

potential competititive interactions of the bivalves Corbicula fluminea and Elliptio complanata, 

in a coastal stream. Freshwater Biology 24, 409-416. 

Leichsenring J. & Lawrence J. (2011) Effect of mid-oceanic ballast water exchange on virus-like 

particle abundance during two trans-Pacific voyages. Marine Pollution Bulletin. 

Leighton D. L. (1968) A comparative study of food selection and nutrition in the abalone, 

Haliotis rufescens Swainson and the sea urchin, Strongylocentrotus purpuratus Swainson. 

University of California, San Diego, California. 

Leighton D. L. (1972) Laboratory observations on the early growth of the abalone, Haliotis 

sorenseni, and the effect of temperature on larval development and settling success. Fisheries 

Bulletin 70, 373-381. 

Leighton D. L. (1989) Abalone (Genus Haliotis) mariculture on the North American Pacific 

coast. Fisheries Bulletin 87, 698-702. 

Leighton D. L. (2000) The biology and culture of the California abalones. Dorrance Publishing 

Company, Inc., Pittsburgh, Pennsylvania. 

Lenarz W. H. & Echeverria T. W. (1991) Sexual dimorphism in Sebastes. Environmental 

Biology of Fishes 30, 71-80. 

Leppaekoski E. & Olenin S. (2001) The meltdown of biogeographical peculiarities of the Baltic 

Sea:The interaction of natural and man-made processes. Ambio 30, 202-209. 

Lesage V. & Kingsley. M. C. S. (1998) Updated status of the St. Lawrence River population of 

the beluga, Delphinapterus leucas. Canadian Field-Naturalist 112, 98-113. 

Lesser M. P., Bythell J. C., Gates R. D., Johnstone R. W. & Hoegh-Guldberg O. (2007) Are 

infectious diseases really killing corals? Alternative interpretations of the experimental and 

ecological data. Journal of Experimental Marine Biology and Ecology 346, 36-44. 

Levenson C. & Leapley W. T. (1978) Distribution of humpback whales (Megaptera 

novaeangliae) in the Caribbean determined by a rapid acoustic method. Journal of the Fisheries 

Research Board of Canada 35, 1150-1152. 
447

Levi F. & Francour P. (2004) Behavioural response of Mullus surmuletus to habitat modification 

by the invasive macroalga Caulerpa taxifolia. Journal of Fish Biology 64, 55–64. 

Lewis A. F. J., McGurk M. D. & Galesloot M. G. (2002) Alcan’s Kemano River eulachon 

(Thaleichthys pacificus) monitoring program 1988-1998. Alcan Primary Metal Ltd., Kitimat, 

British Columbia. 

Lewis J. B. (1974) Settlement, and growth factors influencing the continuous distribution of 

some Atlantic reef corals. Proceedings of the 2nd International Coral Reef Symposium 2, 201-

207. 

Lewis J. B. (1977) Suspension feeding in Atlantic reef corals and the importance of suspended 

particulate matter as a food source. Proceedings of the 3rd International Coral Reef Symposium

1, 405-408. 

Lewison R. L., Crowder L. B., Read A. J. & Freeman S. L. (2004) Understanding impacts of 

fisheries bycatch on marine megafauna. TRENDS in Ecology and Evolution 19, 598–604. 

LGL Ltd. (2007) Environmental Assessment of a Marine Geophysical Survey by the R/V Marcus 

G. Langseth off Central America, January–March 2008. Prepared for the Lamont-Doherty Earth 

Observatory, Palisades, NY, and the National Science Foundation, Arlington, VA, by LGL Ltd., 

environmental research associates, Ontario, Canada. LGL Report TA4342-1. 

LHR. (2010) Energy, oil & gas. Louisiana Hurricane Resources. 

Lichatowich J. A. (1999) Salmon without rivers. In: A history of the Pacific salmon crisis. Island 

Press, Washington, D.C. 

Lien J. (1994) Entrapments of large cetaceans in passive inshore fishing gear in Newfoundland 

and Labrador (1979-1990). Reports of the International Whaling Commission Special Issue 15, 

149-157. 

Liew H. C. (2002) Status of marine turtle conservation and research in Malaysia. In: Western 

Pacific Sea Turtle Cooperative Research and Management Workshop (ed I. Kinan) pp. 51-56. 

Western Pacific Regional Fishery Management Council, Honolulu, Hawaii. 

Lighty R. G., Macintyre I. G. & Stuckenrath R. (1982) Acropora palmata reef framework: A 

reliable indicator of sea level in the western Atlantic for the past 10,000 years. Coral Reefs 1, 

125-130. 

Lilly E. L., Kulis D. M., Gentien P. & Anderson D. M. (2002) Paralytic shellfish poisoning 

toxins in France linked to a human-introduced strain of Alexandrium catenella from the western 

Pacific: evidence from DNA and toxin analysis. Journal of Plankton Research 24, 443-452. 

Limpus C. (2002) Conservation, and research of sea turtles in the western Pacific region: an 

overview. In: Western Pacific Sea Turtle Cooperative Research and Management Workshop. (ed 

I. Kinan) pp. 41-50. Western Pacific Regional Fishery Management Council, Honolulu, Hawaii. 

Limpus C. & Chaloupka M. (1997) Nonparametric regression modeling of green sea turtle 

growth rates (southern Great Barrier Reef). Marine Ecology Progress Series 149, 23-34. 

Limpus C. J. (1985) A study of the loggerhead turtle, Caretta caretta, in eastern Australia. 

University of Queensland, Brisbane, Australia. 

Limpus C. J. & Limpus D. J. (2003) Loggerhead turtles in the equatorial, and southern Pacific 

Ocean: A species in decline. In: Loggerhead sea turtles (eds A. B. Bolten and B. E. 

Witherington) pp. 199-209. Smithsonian Books, Washington D.C. 

Limpus C. J. & Miller J. D. (2008) Australian Hawksbill Turtle Population Dynamics Project. 

Queensland Environmental Protection Agency. 
448

Limpus C. J., Miller J. D., Parmenter C. J. & Limpus D. J. (2003) The green turtle, Chelonia 

mydas, population of Raine Island and the northern Great Barrier Reef, 1843-2001. Memoirs of 

the Queensland Museum 49, 349-440. 

Limpus C. J. & Nicholls N. (1988) The Southern Oscillation regulates the annual numbers of 

green turtles (Chelonia mydas) breeding around northern Australia. Australian Journal of 

Wildlife Research 15, 157-161. 

Lindley S. T., Moser M. L., Erickson D. L., Belchik M., Welch D., Rechisky E. L., Kelly J. T., 

Heublein J. C. & Klimley A. P. (2008) Marine migration of North American green sturgeon. 

Transactions of the American Fisheries Society 137, 182-194. 

Lino S. P. P., Gonçalves E. & Cozens J. (2010) The loggerhead sea turtle (Caretta caretta) on Sal 

Island, Cape Verde: nesting activity and beach surveillance in 2009. Arquipelago 27, 59-63. 

Lirman D. (2002) Biology and ecology working group report. In: Proceedings of the Caribbean 

Acropora workshop: Potential application of the U.S. Endangered Species Act as a conservation 

strategy (ed A. W. Bruckner). 

Lleonart M., Handlinger J. & Powell M. (2003) Spionid mudworm infestation of farmed abalone 

(Haliotis spp.). Aquaculture 221, 85-96. 

Lockwood B. L. & Somero G. N. (2011) Invasive and native blue mussels (genus Mytilus) on 

the California coast: The role of physiology in a biological invasion☆. Journal of Experimental 

Marine Biology and Ecology. 

Lodge D. M., Stein R. A., Brown K. M., Covich A. P., Brönmark C., Garvey J. E. & Klosiewski 

S. P. (1998) Predicting impact of freshwater exotic species on native biodiversity: Challenges in 

spatial scaling. Australian Journal of Ecology 23, 53-67. 

Lohoefener R. R., Hoggard W., Mullin K., Roden C. & Rogers C. (1990) Association of sea 

turtles with petroleum platforms in the north-central Gulf of Mexico. OCS Study, MMS 90-0025, 

90 pp. 

Long E. R., Dutch M., Aasen S., Ricci C. & Welch K. (2001) Spatial extent of contamination, 

toxicity and associated biological effects in Puget Sound sediments. In: 2001 Puget Sound 

Research Conference (ed T. Droscher). Puget Sound Action Team, Olympia, Washington. 

López E. & Arauz R. (2003) Nesting records of East Pacific green turtles (Chelonia mydas 

agassizii) in south Pacific Costa Rica, including notes on incidental capture by shrimping and 

longline activities. In: Twenty-second Annual Symposium on Sea Turtle Biology and 

Conservation. (ed J. A. Seminoff) pp. 84-85. 

Loseto L. L. G. A. S. & Ferguson S. H. (2008) Size and biomagnification: How habitat selection 

explains beluga mercury levels. Environmental Science and Technology 42, 3982-3988. 

Loughlin T. R. (1997) Using the phylogeographic method to identify Steller sea lion stocks. In: 

Molecular genetics of marine mammals (eds A. E. Dizon, S. J. Chivers and W. F. Perrin) pp. 

159-171. Society for Marine Mammalogy 

Loughlin T. R. (2002) Steller's Sea Lion Eumetopias jubatus. In: Encyclopedia of Marine 

Mammals (eds W. F. Perrin, B. Würsig and J. G. M. Thewissen) pp. 1181-1185. Academic Press 

San Diego. 

Loughlin T. R., Ballachey B. E. & Wright B. A. (1996) Overview of studies to determine injury 

caused by the Exxon Valdez oil spill to marine mammals. American Fisheries Society 

Symposium 18, 798-808. 

Loughlin T. R., Perez M. A. & Merrick R. L. (1987) Eumetopias jubatus. In: Mammalian 

Species Account No. 283. 
449

Loughlin T. R., Perlov A. S. & Vladimirov V. A. (1992) Range-wide survey and estimation of 

total number of Steller sea lions in 1989. Marine Mammal Science 8, 220-239. 

Loughlin T. R., Rugh D. J. & Fiscus C. H. (1984) Northern sea lion distribution and abundance: 

1956-80. Journal of Wildlife Management 48, 729-740. 

Love M. S. (1978) Aspects of the life history of the olive rockfish (Sebastes serranoides). 

University of California at Santa Barbara, Santa Barbara, California. 

Love M. S., Carr M. H. & Haldorson L. J. (1991) The ecology of substrate-associated juveniles 

of the genus Sebastes. Environmental Biology of Fishes 30, 225-243. 

Love M. S., Caselle J. E. & Herbinson K. (1998a) Declines in nearshore rockfish recruitment and 

populations in the southern California Bight as measured by impingement rates in coastal 

electrical power generating stations. Fishery Bulletin 96, 492-501. 

Love M. S., Caselle J. E. & Van Buskirk W. (1998b) A severe decline in the commercial 

passenger fishing vessel rockfish (Sebastes spp.) catch in the Southern California Bight, 1980-

1996. California Cooperative Oceanic Fisheries Investigations Reports 39, 180-195. 

Love M. S., Schroeder D. M. & Lenarz W. (2005) Distribution of bocaccio (Sebastes 

paucispinis) and cow cod (Sebastes levis) around oil platforms and natural outcrops off 

California with implications for larval production. Bulletin of Marine Science 77, 397-408. 

Love M. S., Schroeder D. M., Lenarz W., MacCall A., Bull A. S. & Thorsteinson L. (2006) 

Potential use of offshore marine structures in rebuilding an overfished rockfish species, bocaccio 

(Sebastes paucispinis). Fishery Bulletin 104, 383-390. 

Love M. S. & Yoklavich M. (2008) Habitat characteristics of juvenile cow cod, Sebastes levis 

(Scorpaenidae), in Southern California. Environmental Biology of Fishes 82, 195-202. 

Love M. S., Yoklavich M. M. & Thorsteinson L. (2002) The rockfishes of the Northeast Pacific. 

University of California Press, Berkeley, California. 

Love M. S. & York A. (2005) A comparison of the fish assemblages associated with an oil/gas 

pipeline and adjacent seafloor in the Santa Barbara Channel, southern California bight. Bulletin 

of Marine Science 77, 101-117. 

Love M. S. & York A. (2006) The relationships between fish assemblages and the amount of 

bottom horizontal beam exposed at California oil platforms: fish habitat preferences at manmade platforms and (by inference) at natural reefs. Fishery Bulletin 104, 542-549. 

Lövei G. L. (1997) Global change through invasion. Nature 388, 627-628. 

Lovell S. J. & Stone S. F. (2005) The economic impact of aquatic invasive species: A review of 

the literature. U.S. Environmental Protection Agency, Washington, D.C. 

Lowry L., O’Corry-Crowe G. & Goodman D. (2006) Delphinapterus leucas (Cook Inlet 

population). In: 2006 IUCN red list of threatened species. 

Lubchenco J., McNutt M., Lehr B., Sogge M., Miller M., Hammond S. & Conner W. (2010) 

Deepwater Horizon/BP oil budget: What happened to the oil? (eds USGS, NMFS and DOI). 

Ludwig D. (1999) Is it meaningful to estimate a probability of extinction? Ecology 80, 298-310. 

LUMCON. (2005) Mapping of dead zone completed. Louisiana Universities Marine 

Consortium, Chauvin, Louisiana. 

Lundqvist H. (1980) Influence of photoperiod on growth of Baltic salmon parr (Salmo salar L.) 

with specific reference to the effect of precocious sexual maturation. Canadian Journal of 

Zoology 58, 940-944. 

Luschi P., Hays G. C. & Papi F. (2003) A review of long-distance movements by marine turtles, 

and the possible role of ocean currents. Oikos 103, 293-302. 
450

Luschi P., Lutjeharms J. R. E., Lambardi P., Mencacci R., Hughes G. R. & Hays G. C. (2006) A 

review of migratory behaviour of sea turtles off southeastern Africa. South African Journal of 

Science 102, 51-58. 

Lusseau D., Williams R., Wilson B., Grellier K., Barton T. R., Hammond P. S. & Thompson P. 

M. (2004) Parallel influence of climate on the behaviour of Pacific killer whales and Atlantic 

bottlenose dolphins. Ecology Letters 7, 1068-1076. 

Lutcavage M. E., Plotkin P., Witherington B. & Lutz P. L. (1997a) Human impacts on sea turtle 

survival. In: The Biology of Sea Turtles pp. 387-409. CRC Press, Boca Raton, Florida. 

Lutcavage M. E., Plotkin P., Witherington B. E. & Lutz P. L. (1997b) Human impacts on sea 

turtle survival. In: Biology of sea turtles (eds P. L. Lutz and J. A. Musick) pp. 387-409. CRC 

Press, New York, New York. 

Lux J., Reina R. & Stokes L. (2003) Nesting activity of leatherback turtles (Dermochelys 

coriacea) in relation to tidal and lunar cycles at Playa Grande, Costa Rica. Pp.215-216 In: J.A. 

Seminoff (compiler), Proceedings of the 22nd Annual Symposium on Sea Turtle Biology and 

Conservation. Department of Commerce, NOAA Technical Memorandum NMFS-SEFSC-503. 

Lyubimova T. G. (1965) Main stages in the life cycle of the rockfish Sebastodes alutus (Gilbert) 

in the Gulf of Alaska. Trudy Vniro 49, 85-111. 

MacCall A., Batchelder H., King J., Mackas D., Mantua N., McFarlane G., Perry I., Schweigert 

J. & Schwing F. (2005) Recent ecosystem changes in the California Current System. In: Report 

of the study group on fisheries and ecosystem responses to recent regime shifts (ed J. King) p. 

162. 

MacCall A. D. (1996) Patterns of low frequency variability in fish populations of the California 

Current. pp. 100-110. California Cooperative Oceanic Fisheries Investigation Report. 

MacCall A. D. (2002a) Status of bocaccio off California in 2002. Pacific Fishery Management 

Council, Portland, Oregon. 

MacCall A. D. (2002b) Use of known-biomass production models to determine productivity of 

west coast groundfish stocks. North American Journal of Fisheries Management 22, 272-279. 

MacCall A. D. (2003) Status of bocaccio off California in 2003. Pacific Fishery Management 

Council, Portland, Oregon. 

MacCall A. D. (2008) Bocaccio rebuilding analysis for 2007. NMFS, SWFSC, Fish Ecology 

Division. 

MacCall A. D. & He X. (2002a) Bocaccio rebuilding analysis for 2002. Pacific Fishery 

Management Council, Portland, Oregon. 

MacCall A. D. & He X. (2002b) Status review of the southern stock of bocaccio (sebastes 

paucispinis). NMFS, SWFSC, Santa Cruz Laboratory, Santa Cruz, California. 

MacDonald I. R., Jr. N. L. G., Ackleson S. G., Amos J. F., Duckworth R., Sassen R. & Brooks J. 

M. (1993) Natural oil slicks in the Gulf of Mexico visible from space. Journal of Geophysical 

Research 98, 16,351-316,364. 

MacGregor J. S. (1970) Fecundity, multiple spawning, and description of the gonads in 

Sebastodes. United States Fish and Wildlife Service, Washington, D. C. . 

Maciolek J. A. (1984) Exotic fishes in Hawaii and other islands of Oceania. . In: Distribution, 

Biology, and Management of Exotic Fishes (eds W. R. Courtenay and J. R. Stauffer) pp. 131– 

161. Johns Hopkins University Press, Baltimore. 

Macleod C. D. (2002) Possible functions of the ultradense bone in the rostrum of Blainville's 

beaked whale (Mesoplodon densirostris). Canadian Journal of Zoology 80, 178-184. 
451

Macleod C. D. (2009) Global climate change, range changes and potential implications for the 

conservation of marine cetaceans: A review and synthesis. Endangered Species Research 7, 125-

136. 

Macleod C. D., Santos M. B., Reid R. J., Scott B. E. & Pierce. G. J. (2007) Linking sandeel 

consumption and the likelihood of starvation in harbour porpoises in the Scottish North Sea: 

Could climate change mean more starving porpoises? Biology Letters 3, 185-188. 

Magalhaes S., Prieto R., Silva M. A., Goncalves J., Afonso-Dias M. & Santos. R. S. (2002) 

Short-term reactions of sperm whales (Physeter macrocephalus) to whale-watching vessels in the 

Azores. Aquatic Mammals 28, 267-274. 

Maggs C. A. & Stegenga H. (1999) Red algal exotics on North Sea coasts. Helgolaender 

Meeresuntersuchungen 52, 243-258. 

Maguire M. (2001) Chetco River watershed assessment. South Coast Watershed Council, Gold 

Beach, Oregon. 

Mahoney B. A., Migura M., Jackson K. & Jenson. A. (2010) Cook Inlet beluga whale strandings. 

In: Cook Inlet Belgua Whale Science Conference p. 11, Anchorage Downtown Marriott, 

Anchorage, Alaska. 

Mahoney B. A. & Shelden K. E. W. (2000) Harvest history of beluga whale, Delphinapterus 

leucas, in Cook Inlet, Alaska. Marine Fisheries Review 62, 124-133. 

Mahoney B. A. & Shelden. K. E. W. (2000) Harvest history of belugas, Delphinapterus leucas, 

in Cook Inlet, Alaska. Marine Fisheries Review 62, 124-133. 

Mahtab M. A., Stanton K. L. & Roma. V. (2005) Environmental impacts of blasting for stone 

quarries near the Bay of Fundy. In: The Changing Bay of Fundy: Beyond 400 Years (eds J. A. 

Percy, A. J. Evans, P. G. Wells and S. J. Rolston) pp. 87-97. Proceedings of the 6th Bay of 

Fundy Workshop, Cornwallis, Nova Scotia, 9/29-10/2, 2004. 

Maine Rivers. (2007) Kennebec Watershed. 

MaineRivers. (2007) Penobscot watershed. 

Maison K. (2006) Do turtles move with the beach? Beach profiling and possible effects of 

development on a leatherback (Dermochelys coriacea) nesting beach in Grenada. In: Twentysixth Annual Symposium on Sea Turtle Biology and Conservation (eds M. Frick, A. 

Panagopoulou, A. F. Rees and K. Williams) p. 145. International Sea Turtle Society, Athens, 

Greece. 

Maison K. A., Kinan-Kelly I. & Frutchey K. P. (2010) Green turtle nesting sites and sea turtle 

legislation throughout Oceania. U.S. Department of Commerce, National Oceanic and 

Atmospheric Administration, National Marine Fisheries Service. 

Makowski C., Seminoff J. A. & Salmon M. (2006) Home range and habitat use of juvenile 

Atlantic green turtles (Chelonia mydas L.) on shallow reef habitats in Palm Beach, Florida, USA. 

Marine Biology 148, 1167-1179. 

Mallela J. & Crabbe M. J. C. (2009) Hurricanes and coral bleaching linked to changes in coral 

recruitment in Tobago. Marine Environmental Research 68, 158-162. 

Malme C. I., Miles P. R., Tyack P., Clark C. W. & Bird. J. E. (1985) Investigation of the 

potential effects of underwater noise from petroleum industry activities on feeding humpback 

whale behavior. Report No. 5851, prepared for Minerals Management Service, Alaska OCS 

Offfice, 949 East 36th Avenue, Anchorage, AK 99508. MMS 85-0019. 205pp. 

Maniscalco J., Parker P. & Atkinson S. (2006) Interseasonal, and interannual measures of 

maternal care among individual Steller sea lions (Eumetopias jubatus). Journal of Mammalogy

87, 304-311. 
452

Maniscalco J. M. & Parker. P. (2009) A case of twinning and the care of two offspring of 

different age in Steller sea lions. Marine Mammal Science 25, 206-213.-Research Note). 

Maniscalco J. M., Springer A. M. & Parker. P. (2010) High natality rates of endangered Steller 

sea lions in Kenai Fjords, Alaska and perceptions of population status in the Gulf of Alaska. 

PLoS ONE 5, 9Pp. 

Mann R. (2004) The Yakima basin regional economy and the contribution of fish and wildlife. 

In: Yakima County Public Services. 

Mann R., Netusil N. R., Casavant K. L., Huppert D. D., Hamilton J. R., Peters L. L., Hanna S. S. 

& Radtke H. (2005) Economic effects from Columbia River basin anadromous salmonid 

production. p. 166. Northwest Power and Conservation Council, Independent Economic 

Analysis Board. 

Mansfield K. L., Saba V. S., Keinath J. A. & Musick J. A. (2009) Satellite tracking reveals a 

dichotomy in migration strategies among juvenile loggerhead turtles in the Northwest Atlantic. 

Marine Biology 156, 2555-2570. 

Mantua N. (2002) Pacific-Decadal Oscillation (PDO). In: Encyclopedia of global environmental 

change. Volume 1: The earth system: Physical and chemical dimensions of global environmental 

change (eds M. C. MacCracken and J. S. Perry) pp. 592-594. John Wiley & Sons, Ltd., 

Chichister, England. 

Mantua N. J. & Hare S. R. (2002) The Pacific Decadal Oscillation. Journal of Oceanography 58, 

35-44. 

Mantua N. J., Hare S. R., Zhang Y., Wallace J. M. & Francis R. C. (1997) A Pacific interdecadal 

climate oscillation with impacts on salmon production. Bulletin of the American Meteorological 

Society 78, 1069-1079. 

Manzer J. I. (1964) Preliminary observations on the vertical distributions of Pacific salmon 

(genus Oncorhynchus) in the Gulf of Alaska. Journal of the Fisheries Research Board of 

Canada 21, 891-903. 

MARAD. (2011) Data and statistics. Maritime Administration Department of Transportation. 

Marcovaldi M. A. & Chaloupka M. (2007) Conservation status of the loggerhead sea turtle in 

Brazil: An encouraging outlook. Endangered Species Research 3, 133-143. 

Margaritoulis D., Argano R., Baran I., Bentivegna F., Bradai M. N., Camiñas J. A., Casale P., De 

Metrio G., Demetropoulos A., Gerosa G., Godley B. J., Haddoud D. A., Houghton J., Laurent L. 

& Lazar B. (2003) Loggerhead turtles in the Mediterranean Sea: Present knowledge and 

conservation perspectives. In: Loggerhead sea turtles (eds A. B. Bolten and B. E. Witherington) 

pp. 175-198 Smithsonian Books, Washington D.C. 

Marins F. O., Novaes R. L. M., Rocha R. M. & Junqueira A. O. R. (2010) Non indigenous 

ascidians in port and natural environments in a tropical Brazilian bay. Zoologia 27, 213-221. 

Marks M., Lapin B. & Randall J. (1994) Phragmites australis (P. communis): Threats, 

management, monitoring. Natural Areas Journal 14, 285-294. 

Marquez-M. R. (1994) Synopsis of biological data on the Kemp's ridley turtle, Lepidochelys 

kempii, (Garman, 1880). NOAA Technical Memorandum NMFS-SEFSC-343, or OCS Study 

MMS 94-0023. 91p. 

Márquez M. R. (1990) Sea turtles of the world. An annotated and illustrated catalogue of sea 

turtle species known to date. FAO Species Catalog, FAO Fisheries Synopsis 11(125):81p. 

Marquez M. R., Villanueva A. & Burchfield P. M. (1989) Nesting population, and production of 

hatchlings of Kemp's ridley sea turtle at Rancho Nuevo, Tamaulipas, Mexico. In: First 
453

International Symposium on Kemp's Ridley Sea Turtle Biology, Conservation, and Management

(eds C. W. Caillouet Jr. and A. M. Landry Jr.) pp. 16-19. 

Marr J. C. (1943) Age, length, and weight studies of three species of Columbia River salmon 

(Oncorhynchus keta, O. gorbuscha, and O. kisutch). Stanford Ichthyology Bulletin 2, 157-197. 

Marsh J. W., Chipman J. K. & Livingstone D. R. (1992) Activation of xenobiotics to reactive 

and mutagenic products by the marine invertebrates Mytilus edulis, Carcinus maenus, and 

Asterias rubens. Aquatic Toxicology 22, 115-128. 

Marshall C. D. & Dehnhardt. G. (2005) Behavioral performance of suction feeding in harbor 

seals (Phoca vitulina). In: Sixteenth Biennial Conference on the Biology of Marine Mammals p. 

180, San Diego, California. 

Marshall C. D., Moss A. L. & Guzman A. (2009) Loggerhead Sea Turtle (Caretta caretta) 

Feeding on Mackerel-Baited Longline Hooks. Integrative and Comparative Biology 49, E266-

E266. 

Martel A. L., Pathy D. A., madill J. B., Renaud C. B., Dean S. L. & Kerr S. J. (2001) Decline 

and regional extirpation of freshwater mussels (Unionidae) in a small river system invaded by 

Dreissena polymorpha: the Rideau River, 1993–2000. Canadian Journal of Zoology 79, 2181-

2191. 

Martinez J. A., et al. (2012) Invasive algal mats degrade coral reef physical habitat quality. 

Estuarine, Coastal and Shelf Science. 

MASGC. (2010) Mississippi coastal cleanup. Mississippi Alabama Sea Grant Consortium. 

Matala A. P., K.Gray A. & Gharrett A. J. (2004) Microsatellite variation indicates population 

genetics structure of bocaccio. North American Journal of Fisheries Management 24, 1189-

1202. 

Matarese A. C., Kendall A. W., Blood D. M. & Vinter B. M. (1989) Laboratory guide to early 

life history stages of northeast Pacific fishes. NOAA. 

Mate B. R. (1973) Population kinetics and related ecology of the northern sea lion, Eumetopias 

jubatus, and the California sea lion, Zalophus californianus, along the Oregon coast. University 

of Oregon. 

Mate B. R. & Baumgartner M. (2001) Summer feeding season movements and fall migration of 

North Atlantic right whales from satellite-monitored radio tags. In: Abstracts, Fourteenth 

Biennial Conference on the Biology of Marine Mammals. 28 November–3 December 2001 p. 

137, Vancouver, British Columbia. 

Mate B. R., Nieukirk S. L. & Kraus S. D. (1997) Satellite-monitored movements of the northern 

right whale. Journal of Wildlife Management 61, 1393-1405. 

Matkin C. O. (1994) An observer's guide to the killer whales of Prince William Sound, Alaska. 

(Orcinus orca). Prince William Sound Books. 

Matkin C. O., Ellis G., Barrett-Lennard L., Yurk H., Saulitis E., Scheel D., Olesiuk P. & Ylitalo 

G. (2003) Photographic and acoustic monitoring of killer whales in Prince William Sound and 

Kenai Fjords. In: Restoration Project 030012 Final Report, Exxon Valdes Oil Spill Restoration 

Project. North Gulf Oceanic Society, Homer, Alaska. 

Matkin C. O., Matkin D. R., Ellis G. M., Saulitis E. & McSweeney D. (1997) Movements of 

resident killer whales in southeastern Alaska and Prince William Sound, Alaska. Marine 

Mammal Science 13, 469-475. 

Matthews G. M. & Waples R. S. (1991) Status review for Snake River spring and summer 

Chinook salmon. In: NOAA Technical Memorandum NMFS F/NWC-200. U.S. Department of 

Commerce. 
454

Matthiopoulos J. & Aarts G. (2010) The spatial analysis of marine mammal abundance. In: 

Marine Mammal Ecology and Conservation: A Handbook of Techniques (eds I. L. Boyd, W. D. 

Bowen and S. J. Iverson) pp. 68-97. Oxford University Press. 

Mattila D., Clapham P. J., Vásquez O. & Bowman R. S. (1994) Occurrence, population 

composition, and habitat use of humpback whales in Samana Bay, Dominican Republic. 

Canadian Journal of Zoology 72, 1898-1907. 

Mattila D. K. & Rowles T. (2010) A review of large whale entanglement. p. 8. IWC Scientific 

Committee, Agadir, Morocco. 

Maynard J. A., Anthony K. R. N., Marshall P. A. & Masiri I. (2008) Major bleaching events can 

lead to increased thermal tolerance in corals. Marine Biology 155, 173-182. 

Mayo C. A. & Marx M. K. (1990) Surface foraging behaviour of the North Atlantic right whale, 

Eubalaena glacialis, and associated zooplankton characteristics. Canadian Journal of Zoology

68, 2214-2220. 

Mayor P. A., Rogers C. S. & Hillis-Starr Z. M. (2006) Distribution and abundance of elkhorn 

coral, Acropora palmata, and prevalence of white-band disease at Buck Island Reef National 

Monument, St. Croix, US Virgin Islands. Coral Reefs 25, 239-242. 

Mazaris A. D., Kallimanis A. S., Sgardelis S. P. & Pantis J. D. (2008) Do long-term changes in 

sea surface temperature at the breeding areas affect the breeding dates and reproduction 

performance of Mediterranean loggerhead turtles? Implications for climate change. Journal of 

Experimental Marine Biology and Ecology. 

Mazaris A. D., Kallimanis A. S., Tzanopoulos J., Sgardelis S. P. & Pantis J. D. (2009a) Sea 

surface temperature variations in core foraging grounds drive nesting trends and phenology of 

loggerhead turtles in the Mediterranean Sea. Journal of Experimental Marine Biology and 

Ecology. 

Mazaris A. D., Matsinos G. & Pantis J. D. (2009b) Evaluating the impacts of coastal squeeze on 

sea turtle nesting. Ocean & Coastal Management 52, 139-145. 

MBC. (1987) Ecology of important fisheries species offshore California. Minerals Management 

Service, Camarillo, California. 

Mcaloose D. & Newton. A. L. (2009) Wildlife cancer: A conservation perspective. Nature 

Reviews Cancer 9, 517-526. 

McCarthy A. L., Heppell S., Royer F., Freitas C. & Dellinger T. (2010) Identification of likely 

foraging habitat of pelagic loggerhead sea turtles (Caretta caretta) in the North Atlantic through 

analysis of telemetry track sinuosity. Progress in Oceanography. 

McCarthy H. P. & Crowder L. B. (2000) An overlooked scale of global transport: phytoplankton 

species richness in 

ships’ ballast water. Biological Invasions 2, 321-322. 

McCarthy I. D. & Houlihan D. F. H. (1997) The effect of temperature on protein metabolism in 

fish: the possible consequences for wild Atlantic salmon stocks in Europe as a result of global 

warming. In: Global warming: Implications for freshwater and marine fish. (eds C. M. Wood 

and D. G. McDonald) pp. 51-77. Cambridge University Press, New York, NY. 

McCauley S. J. & Bjorndal K. A. (1999) Conservation implications of dietary dilution from 

debris ingestion: Sublethal effects in post-hatchling loggerhead sea turtles. Conservation Biology

13, 925-929. 

McClellan C. M., Braun-McNeill J., Avens L., Wallace B. P. & Read A. J. (2010) Stable 

isotopes confirm a foraging dichotomy in juvenile loggerhead sea turtles. Journal of 

Experimental Marine Biology and Ecology 387, 44-51. 
455

McClellan C. M., Read A. J., Price B. A., Cluse W. M. & Godfrey M. H. (2009) Using telemetry 

to mitigate the bycatch of long-lived marine vertebrates. Ecological Applications 19, 1660-1671. 

Mcconnell B., Fedak M., Hooker S. & Patterson T. (2010) Telemetry. In: Marine Mammal 

Ecology and Conservation: A Handbook of Techniques (eds I. L. Boyd, W. D. Bowen and S. J. 

Iverson) pp. 222-242. Oxford University Press. 

McCormick J. H., Broderius S. J. & Fiandt J. T. (1984) Toxicity of ammonia to early life stages 

of the green sunfish Lepomis cyanellus. Environmental Pollution Series A, Ecological and 

Biological 36, 147-163. 

McCormick S. D., Hansen L. P., Quinn T. P. & Saunders R. L. (1998) Movement, migration, and 

smolting of Atlantic salmon (Salmo salar). Canadian Journal of Fisheries and Aquatic Sciences 

55(Supplemental 1, 77-92. 

McCormick S. D., Shrimpton J. M. & Zydlewski J. D. (1997) Temperature effects on 

osmoregulatory physiology of juvenile anadromous fish. In: Global warming: implications for 

freshwater and marine fish (eds C. M. Wood and D. G. McDonald) pp. 279-301. Cambridge 

University Press, Cambridge, England. 

McCullough D. (1999) A review and synthesis of effects of alterations to the water temperature 

regime on freshwater life stages of salmonids, with special referene to Chinook salmon. 

Columbia Intertribal Fisheries Commission, Portland, OR. Prepared for the U.S. Environmental 

Protection Agency Region 10. Published as EPA 910-R-99-010. 

McDonald Dutton D. & Dutton P. H. (1998) Accelerated growth in San Diego Bay green turtles? 

In: Seventeenth Annual Sea Turtle Symposium (eds S. P. Epperly and J. Braun) pp. 175-176. 

McDonald M. A., Mesnick S. L. & Hildebrand J. A. (2006) Biogeographic characterization of 

blue whale song worldwide: using song to identify populations. Journal of Cetacean Research 

and Management 8, 55-65. 

McElhaney P., Chilcote M., Myers J. & Beamesderfer R. (2007) Viability status of Oregon 

salmon and steelhead populations in the Willamette and lower Columbia basins. Draft report by 

NOAA Fisheries and Oregon Department of Fish and Wildlife. 

McElhany P., Chilcote M., Myers J. & Beamesderfer R. (2007) Viability status of Oregon 

salmon and steelhead populations in the Willamette and lower Columbia basins. In: 

Willamette/Lower Columbia Technical Recovery Team Report. NMFS and Oregon Department 

of Fish and Wildlife, Seattle, Washington. 

McEwan D. & Jackson T. A. (1996a) Steelhead restoration and management plan for California. 

California Department of Fish and Game, Sacramento, California. 

McEwan D. & Jackson T. A. (1996b) Steelhead restoration and management plan for California. 

California Department of Fish and Game, Sacramento, California. 

McEwan D. R. (2001) Central valley steelhead. In: Contributions to the biology of the central 

valley salmonids (ed R. L. Brown) pp. 1-43. California Department of Fish and Game, 

Sacramento, California. 

McFarlane G. A. & Beamish R. J. (1992) Climatic influence linking copepod production with 

strong yearclasses in sablefish, Anoplopoma fimbria. Canadian Journal of Fisheries and Aquatic 

Sciences 49, 743-753. 

McFarlane G. A., King J. R. & Beamish R. J. (2000) Have there been recent changes in climate? 

Ask the fish. Progress in Oceanography 47, 147-169. 

Mcguire T. L., Blees M. K. & Kaplan. C. C. (2010a) The use of photo-identification to study 

associations and distribution of beluga whale mothers and calves in upper Cook Inlet, Alaska. In: 
456

Cook Inlet Belgua Whale Science Conference pp. 18-19, Anchorage Downtown Marriott, 

Anchorage, Alaska. 

Mcguire T. L., Kaplan C. C. & Blees. M. K. (2010b) Group structure of Cook Inlet beluga 

whales: Insights from photo-identification. In: Cook Inlet Belgua Whale Science Conference p. 

18, Anchorage Downtown Marriott, Anchorage, Alaska. 

McKenna P. (2007) Fresh danger for fish as killer virus fans out. New Scientist 194, 9-9. 

Mckenzie C., Godley B. J., Furness R. W. & Wells D. E. (1999) Concentrations and patterns of 

organochlorine contaminants in marine turtles from Mediterranean and Atlantic waters. Marine 

Environmental Research 47, 117-135. 

McKenzie F. & Benton M. (1972) Biological inventory of the waters and keys of north-east 

Puerto Rico. Division of Natural Resources, Department of Public Works, San Juan, Puerto Rico. 

McKenzie J. & Wynne K. M. (2008) Spatial and temporal variation in the diet of Steller sea lions 

in the Kodiak Archipelago, 1999 to 2005. Marine Ecology Progress Series 360, 265-283. 

McLaughlin E. & Knight A. (1987) Habitat criteria for Atlantic salmon. U.S. Fish and Wildlife 

Service,, Laconia, New Hampshire. 

McLean J. E., Hay D. E. & Taylor E. B. (1999) Marine population structure in an anadromous 

fish: Life history influences patterns of mitochondrial DNA variation in the eulachon, 

Thaleichthys pacificus. Molecular Ecology 8, S143-S158. 

McLean J. E. & Taylor E. B. (2001) Resolution of population structure in a species with high 

gene flow: microsatellite variation in the eulachon (Osmeridae: Thaleichthys pacificus). Marine 

Biology 139, 411-420. 

McLeese J., Johnsson M. J., Huntley F. M., Clarke W. C. & Weisbart M. (1994) Seasonal 

changes in osmoregulation, cortisol, and cortisol receptor activity in the gills of parr/smolt of 

steelhead trout and steelhead/rainbow trout hybrids, Oncorhynchus mykiss. General and 

Comparative Endocrinology 93, 103-113. 

McMahon C. R. & Burton. H. R. (2005) Climate change and seal survival: Evidence for 

environmentally mediated changes in elephant seal, Mirounga leonina, pup survival. 

Proceedings of the Royal Society of London Series B Biological Sciences 272, 923-928. 

McMahon C. R. & Hays G. C. (2006) Thermal niche, large-scale movements and implications of 

climate change for a critically endangered marine vertebrate. Global Change Biology 12, 1330-

1338. 

McMinn A., Hallegraeff G., Thomson P., Jenkinson A. V. & Heijnis H. (1997) Cyst and 

radionucleotide evidence for the recent introduction of the toxic dinoflagellate Gymnodinium 

catenatum into Tasmanian waters. Marine Ecology-Progress Series 161, 65-172.171. 

McNickle G. G., Rennie M. D. & Sprules W. G. (2001) Changes in benthic invertebrate 

communities of South Bay, Lake Huron following invasion by zebra mussels (Dreissena 

polymorpha), and potential effects on lake whitefish (Coregonus clupeaformis) diet and growth. 

Journal of Great Lakes Research 32, 180-193. 

McNickle G. G., Rennie M. D. & Sprules W. G. (2006) Changes in benthic invertebrate 

communities of South Bay, Lake Huron following invasion by zebra mussels (Dreissena 

polymorpha), and potential effects on Lake Whitefish (Coregonus clupeaformis) diet and growth. 

Journal of Great Lakes Research 32, 180-193. 

McPhail J. D. & Lindsey C. C. (1970) Freshwater fishes of northwestern Canada and Alaska. In: 

Fisheries Research Board of Canada, Bulletin 173. , Ottawa, Canada. 

MDEP. (1999) Hazard ranking system package, Eastland Woolen Mill (a.k.a. Corinna Main 

Street), Corinna, Maine. Maine Department of Environmental Protection, Boston, Massachusetts. 
457

MDMR. (2007) Kennebec River diadromous fish restoration project. Maine Department of 

Marine Resources. 

Meador J. P., Stein R. & Varanasi U. (1995) Bioaccumulation of polycyclic aromatic 

hydrocarbons by marine organisms. Reviews of Environmental Contamination and Toxicology

143, 79-165. 

Mearns A. J. (2001) Long-term contaminant trends and patterns in Puget Sound, the Straits of 

Juan de Fuca, and the Pacific Coast. In: 2001 Puget Sound Research Conference (ed T. 

Droscher). Puget Sound Action Team, Olympia, Washington. 

Meehan W. R. & Bjornn T. R. (1991) Salmonid distributions and life histories: Cutthroat trout. 

Influence of forest and rangeland management on salmonid fishes and their habitats. Bethesda, 

American Fisheries Society Special Publication 19, 66-67. 

Meier M. F., Dyurgerov M. B. & McCabe G. J. (2003) The health of glaciers: Recent changes in 

glacier regime. Climate Change 53, 123-135. 

Meister A. L. (1958) The Atlantic salmon (Salmo salar) of Cove Brook. University of Maine. 

Mellgren R. L. & Mann M. A. (1996) Comparative behavior of hatchling sea turtles. In: 

Fifteenth Annual Symposium on Sea Turtle Biology and Conservation (eds J. A. Keinath, D. E. 

Barnard, J. A. Musick and B. A. Bell) pp. 202-204. 

Mellgren R. L., Mann M. A., Bushong M. E., Harkins S. R. & Krumke V. K. (1994) Habitat 

selection in three species of captive sea turtle hatchlings. In: Fourteenth Annual Symposium on 

Sea Turtle Biology and Conservation (eds K. A. Bjorndal, A. B. Bolten, D. A. Johnson and P. J. 

Eliazar) pp. 259-260. 

Melsom A., Meyers S. D., Hurlburt H. E., Metzger J. E. & O'Brien J. J. (1999) ENSO effects on 

Gulf of Alaska eddies. Earth Interactions 3, 1-30. 

Merkel T. J. (1957) Food habits of the king salmon Oncorhynchus tshawytscha. California Fish 

and Game 43, 249-270. 

Merrick R. L., Chumbley M. K. & Byrd G. V. (1997) Diet diversity of Steller sea lions 

(Eumetopias jubatus) and their population decline in Alaska: a potential relationship. Canadian 

Journal of Fisheries and Aquatic Sciences 54, 1342-1348. 

Merrick R. L. & Loughlin T. R. (1997) Foraging behavior of adult female and young-of-the year 

Steller sea lions in Alaskan waters. Canadian Journal of Zoology 75, 776-786. 

Merz J. E. (1994) A dietary analysis of juvenile Sacramento squawfish (Ptychocheilus grandis), 

Chinook salmon (Oncorhynchus tshawytcha), and steelhead trout (O. mykiss) in the Lower 

American River, Sacramento County, California. California State University, Sacramento, CA. 

Metcalfe C., Koenig B., Metcalfe T., Paterson G. & Sears R. (2004) Intra- and inter-species 

differences in persistent organic contaminants in the blubber of blue whales and humpback 

whales from the Gulf of St. Lawrence, Canada. Marine Environmental Research 57, 245–260. 

Metcalfe N. B., Huntingford F. A. & Thorpe J. E. (1988) Feeding intensity, growth rates, and the 

establishment of life history patterns in juvenile Atlantic salmon (Salmo salar). Journal of 

Animal Ecology 57, 463-474. 

Methot R. D. & Stewart J. J. (2005) Status of the U.S. canary rockfish resource in 2005. National 

Marine Fisheries Service, NWFSC, Seattle, Washington. 

Meyers J. M., Kope R. G., Bryant G. J., Teel D. J., Lierheimer L. J., Wainwright T. C., Grand W. 

S., Waknitz F. W., Neely K., Lindley S. T. & Waples R. S. (1998) Status review of Chinook 

salmon from Washington, Idaho, Oregon, and California. U.S. Department of Commerce, NOAA 

Technical Memorandum. NMFS-NWFSC-35. 

Meylan A. (1988) Spongivory in hawksbill turtles: A diet of glass. Science 239, 393-395. 
458

Meylan A. & Donnelly M. (1999) Status Justification for Listing the Hawksbill Turtle 

(Eretmochelys imbricata) as Critically Endangered on the 1996 IUCN Red List of Threatened 

Animals. Chelonian Conservation and Biology 3, 200-224. 

Meylan A., Schroeder B. & Mosier A. (1995a) Sea turtle nesting activity in the State of Florida 

1979-1992. Florida Marine Research Publications 52. 

Meylan A. B. (1999) Status of the hawksbill turtle (Eretmochelys imbricata) in the Caribbean 

region. Chelonian Conservation and Biology 3, 177-184. 

Meylan A. B., Bowen B. W. & Avise J. C. (1990) A genetic test of the natal homing versus 

social facilitation models for green turtle migration. Science 248, 724-727. 

Meylan A. B., Schroeder B. A. & Mosier A. (1995b) Sea turtle nesting activity in the State of 

Florida 1979-1992. Florida Department of Environmental Protection (In press Florida Marine 

Research Publications), 63. 

MFM. (2000) Lake Ozette sockeye – hatchery and genetic management plan – Biological 

assessment, Section 7 consultation. In: October 23, 2000. Prepared by Makah Fisheries 

Management for Bureau of Indian Affairs. Makah Indian Tribe. Neah Bay, Washington. 

Miao X., Balazsb G. H., Murakawa S. K. K. & Li Q. X. (2001) Congener-specific profile, and 

toxicity assessment of PCBs in green turtles (Chelonia mydas)from the Hawaiian Islands. The 

Science of the Total Environment 281, 247-253. 

Michigan Technological University. (2010) Quagga mussels make their presence felt. Marine 

Pollution Bulletin 60, 1904-1905. 

Middlebrook R., Anthony K. R. N., Hoegh-Guldberg O. & Dove S. (2010) Heating rate and 

symbiont productivity are key factors determining thermal stress in the reef-building coral 

Acropora formosa. Journal of Experimental Biology 213, 1026-1034. 

Mieog J. C., Olsen J. L., Berkelmans R., Bleuler-Martinez S. A., Willis B. L. & Oppen M. J. H. 

v. (2009) The roles and interactions of symbiont, host and environment in defining coral fitness. 

PLoS ONE 4. 

Mikhalev Y. A. (1997) Humpback whales Megaptera novaeangliae in the Arabian Sea. Marine 

Ecology Progress Series 149, 13-21. 

Mikhalev Y. A. (2000) Whaling in the Arabian Sea by the whaling fleets slava and Sovetskaya 

Ukraina. In: Soviet Whaling Data (1949-1979) pp. 141-181. Center for Russian Environmental 

Policy Marine Mammal Council, Moscow. 

Milette L. L. (1999) Behavior of lactating Steller sea lions (Eumetopias jubatus) during the 

breeding season: A comparison between a declining, and stable population in Alaska. University 

of British Columbia. 

Milette L. L. & Trites A. W. (2003) Maternal attendance patterns of lactating Steller sea lions 

(Eumetopias jubatus) from a stable and a declining population in Alaska. Canadian Journal of 

Zoology 81, 340-348. 

Miller A. C. & Lawrenz-Miller S. E. (1993) Long-term trends in black abalone, Haliotis 

cracherodii Leach 1814, populations along the Palos Verdes Peninsula, California. Journal of 

Shellfish Research 12, 195-200. 

Miller B. S. & Borton S. F. (1980) Geographical distribution of Puget Sound fishes: Maps and 

data source sheets. University of Washington Fisheries Research Institute. 

Miller D. J. & Geibel J. J. (1973) Summary of blue rockfish and lingcod life histories; a reef 

ecology study; and giant kelp, Macrocystis pyrifera, experiments in Monterey Bay, California. 

Fishery Bulletin 137. 
459

Miller D. J. & Lea R. N. (1972) Guide to the coastal marine fishes of California. Fishery Bulletin

157, 1-249. 

Miller J. D., Dobbs K. A., Limpus C. J., Mattocks N. & Landry A. M. (1998) Long-distance 

migrations by the hawksbill turtle, Eretmochelys imbricata, from north-eastern Australian. 

Wildlife Research 25, 89-95. 

Miller M. & Williamson D. H. (2010) Population status and restoration potential of threatened 

Acropora (with emphasis on the Florida Keys). NOAA, NMFS, Southeast Fisheries Science 

Center. 

Miller M. J. (2004) The ecology and functional morphology of North American sturgeon and 

paddlefish. In: Sturgeon and paddlefish of North America (ed G. T. O. LeBreton) pp. 87-102. 

Kluwer Academic Publishers, Netherlands. 

Miller R. J. & Brannon E. L. (1982) The origin and development of life history patterns in 

Pacific salmonids. In: The Salmon and Trout Migratory Behavior Symposium (eds E. L. Brannon 

and E. O. Salo) pp. 296-309. School of Fisheries, University of Washington, Seattle, 

Washington. 

Milliken T. & Tokunaga H. (1987) The Japanese sea turtle trade 1970-1986. A special report 

prepared by TRAFFIC (Japan). p. 171 pages. Center for Environmental Education, Washington 

D.C. 

Mills M. D., Rader R. B. & Belk M. C. (2004) Complex interactions between native and invasive 

fish: the simultaneous effects of multiple negative actions. Oecologia 141, 713-721. 

Mills S. K. & Beatty J. H. (1979) The propensity interpretation of fitness. Philosophy of Science

46, 263-286. 

Milton S. L., Leone-Kabler S., Schulman A. A. & Lutz P. L. (1994) Effects of Hurricane 

Andrew on the sea turtle nesting beaches of South Florida. Bulletin of Marine Science 54, 974-

981. 

Milton S. L. & Lutz P. L. (2003) Physiological and genetic responses to environmental stress. In: 

The Biology of Sea Turtles, Volume II (eds P. L. Lutz, J. A. Musick and J. Wyneken) pp. 163-

197. CRC Press, Boca Raton, Florida. 

Minchin D. & Rosenthal H. (2002) Exotics for stocking and aquaculture, making correct 

decisions. In: Invasive Aquatic Species of Europe (eds E. Leppakoski, S. Gollasch and S. Olenin) 

pp. 206-216. 

Minckley W. L., Hendrickson D. A. & Bond C. E. (1986) Geography of western North American 

freshwater fishes: Description and relationship to intercontinental tectonism. In: The 

zoogeography of North American freshwater fishes (eds C. H. Hocutt and E. O. Wiley) pp. 519-

613. John Wiley and Sons, New York, New York. 

Minobe S. (1997) A 50-70 year climatic oscillation over the North Pacific and North America. 

Geophysical Research Letters 24, 683-686. 

Minobe S. (1999) Resonance in bidecadal and pentadecadal climate oscillations over the North 

Pacific: Role in climatic regime shifts. Geophysical Research Letters 26, 855-858. 

Minton G., Cercho S., Collins T., Ersts P., Findlay K. P., Pomilla C., Bennet D., Meyer M. A., 

Razafindrakoto Y., Kotze P. G. H., Oosthuizen W. H., Leslie M., Andrianarivelo N., Baldwin R., 

Ponnampalam L. & Rosenbaum. H. (2008) Comparison of humpback whale tail fluke catalogues 

from the Sultanate of Oman (Breeding and Feeding area X) with Madagascar (Breeding area C3) 

and the East Africa Mainland (Breeding area C1). Unpublished paper to the IWC Scientific 

Committee. 4 pp. Santiago, Chile, June (SC/60/SH43). 
460

Minton G., Collins M. & Finlay. K. (2003) A note on re-sights of individually identified 

humpback whales (Megaptera novaeangliae) off the coast of Oman. Unpublished paper to the 

IWC Scientific Committee. 7 pp. Berlin, May (SC/55/O10). 

Minton M. S., Verling E., Miller A. W. & Ruiz G. M. (2005) Reducing propagule supply and 

coastal invasions via ships: effects of emerging strategies. Frontiers in Ecology and the 

Environment 3, 304–308. 

Mitchell E. (1975) Trophic relationships and competition for food in northwest Atlantic right 

whales. Proceedings of the Canadian Society of Zoology Annual Meeting 1974, 123-133. 

Mitchell E. & Reeves R. R. (1983) Catch history, abundance and present status of northwest 

Atlantic humpback whales. Report of the International Whaling Commission, 153-212. 

Miyahara K., Nagai S., Itakura S., Yamamoto K., Fujisawa K., Iwamoto T., Yoshimatsu S., 

Matsuoka S., Yuasa A., Makino K., Hori Y., Nagata S., Nagasaki K., Yamaguchi M. & Honjo T. 

(1996) First Record of a Bloom of Thalassiosira diporocyclus in the Eastern Seto Inland Sea. 

Fisheries Science 62, 878-882. 

MMC. (2002) Marine Mammal Commission Annual Report to Congress 2001. p. 260. Marine 

Mammal Commission. 

MMMP. (2002a) Marine Mammal Monitoring Annual Report 2001-2002. Marine Mammal 

Monitoring Project, Victoria, British Columbia. 25p. 

MMMP. (2002b) Marine Mammal Monitoring Annual Report 2001-2002. p. 25. Marine 

Mammal Monitoring Project, Victoria, British Columbia. 

MMS. (1998) Pages III-3 to III-72 in Gulf of Mexico OCS oil and gas lease sales 171, 174, 177, 

and 180—Western Planning Area. In: Final Environmental Impact Statement. Minerals 

Management Service, New Orleans, Louisiana. 

MMS. (2007a) Gulf of Mexico OCS oil and gas lease sale 224, Eastern planning area. Final 

supplemental environmental impact statement. Minerals Management Service. 

MMS. (2007b) Gulf of Mexico OCS oil and gas lease sales: 2007-2012, Western planning area 

sales 204, 207, 210, 215, and 218; Central planning area sales 205, 206, 208, 213, 216, and 222. 

Final environmental impact statement. Minerals Management Service. 

Moffett J. W. (1949) The first four years of king salmon maintenance below Shasta Dam, 

Sacramento River, California. California Department of Fish and Game 35, 77-102. 

Moir H. J., Soulsby C. & Youngson A. (1998) Hydraulic and sedimentary characteristics of 

habitat utilized by Atlantic salmon for spawning in the Girnock Burn, Scotland. Fisheries 

Management and Ecology 5, 241–254. 

Moline M. A., Claustre H., Frazer T. K., Schofields O. & Vernet M. (2004) Alterations of the 

food web along the Antarctic Peninsula in response to a regional warming trend. Global Change 

Biology 10, 1973-1980. 

Molnar J. L., Gamboa R. L., Revenga C. & Spalding M. D. (2008) Assessing the global threat of 

invasive species to marine biodiversity. Frontiers in Ecology and the Environment 6, 485-492. 

Momma H. & Sato R. (1969) The locomotion behavior of the disc abalone, Haliotis discus 

hannai Ino, and the Siebold’s abalone, Haliotis sieboldi Reeve, in the fishing grounds. Tohoku 

Journal of Agriculture Research 20. 

Monagas P., Oros J., Anana J. & Gonzalez-Diaz O. M. (2008) Organochlorine pesticide levels in 

loggerhead turtles (Caretta caretta) stranded in the Canary Islands, Spain. Marine Pollution 

Bulletin 56, 1949-1952. 

Moncheva S. P. & Kamburska L. T. (2002) Plankton stowaways in the Black Sea - Impacts on 

biodiversity and ecosystem health. In: Alien marine organisms introduced by ships in the 
461

Mediterranean and Black seas pp. 47-51. CIESM Workshop Monographs [CIESM Workshop 

Monogr.]. 2002. 

Montie E. W., Letcher R. J., Reddy C. M., Moore M. J., Rubinstein B. & Hahn M. E. (2010) 

Brominated flame retardants and organochlorine contaminants in winter flounder, harp and 

hooded seals, and North Atlantic right whales from the Northwest Atlantic Ocean. Marine 

Pollution Bulletin 60, 1160-1169. 

Monzon-Arguello C., Rico C., Carreras C., Calabuig P., Marco A. & Lopez-Jurado L. F. (2009) 

Variation in spatial distribution of juvenile loggerhead turtles in the eastern Atlantic and western 

Mediterranean Sea. Journal of Experimental Marine Biology and Ecology 373, 79-86. 

Monzon-Arguello C., Rico C., Marco A., Lopez P. & Lopez-Jurado L. F. (2010) Genetic 

characterization of eastern Atlantic hawksbill turtles at a foraging group indicates major 

undiscovered nesting populations in the region. Journal of Experimental Marine Biology and 

Ecology in press, in press. 

Moore B. & Platt D. D. (1997) The river and its watershed. In: Penobscot pp. 23-30. 

Moore J. D., Juhasz C. I., Robbins T. T. & Vilchis L. I. (2009) Green abalone, Haliotis fulgens 

infected with the agent of withering syndrome do not express disease signs under a temperature 

regime permissive for red abalone, Haliotis rufescens. Marine Biology 156, 2325-2330. 

Moore J. D., Robbins T. T. & Friedman C. S. (2000a) The role of a rickettsia-like prokaryote in 

withering syndrome in California red abalone, Haliotis rufescens. Journal of Shellfish Research

19, 525-526. 

Moore M. E., Berejikian B. A. & Tezak E. P. (2010) Early marine survival and behavior of 

steelhead smolts through Hood Canal and the Strait of Juan de Fuca. Transactions of American 

Fisheries Society 139, 49-61. 

Moore M. M. (1980) Factors influencing survival of juvenile steelhead rainbow trout (Salmo 

Gairdnerii gairdnerii) in the Ventura River. M.s. thesis, Humboldt State University, Arcata, 

California, 82 p. 

Moore S. E., Shelden K. E. W., Litzky L. K., Mahoney B. A. & Rugh D. J. (2000b) Beluga, 

Delphinapterus leucas, habitat associations in Cook Inlet, Alaska. Marine Fisheries Review 62, 

60-80. 

Moore S. E., Shelden K. E. W., Litzky L. K., Mahoney B. A. & Rugh. D. J. (2000c) Beluga, 

Delphinapterus leucas, habitat associations in Cook Inlet, Alaska. Marine Fisheries Review 62, 

60-80. 

Moore S. K., Mantua N. J. & Salathé E. P. (2011) Past trends and future scenarios for 

environmental conditions favoring the accumulation of paralytic shellfish toxins in Puget Sound 

shellfish. Harmful Algae 10, 521-529. 

Moreira L. & Bjorndal K. A. (2006) Estimates of green turtle (Chelonia mydas) nests on 

Trindade Island, Brazil, South Atlantic. In: Twenty-third Annual Symposium on Sea Turtle 

Biology and Conservation (ed N. Pilcher) p. 174. 

Morejohn G. V., Harvey J. T. & Krasnow L. T. (1978) The importance of Loligo opalescens in 

the food web of marine vertebrates in Monterey Bay, California. In: Biological, oceanographic, 

and acoustic aspects of the market squid, Loligo opalescens (eds C. W. Recksiek and H. W. 

Frey) pp. 67-98. 

Morreale S. J., Plotkin P. T., Shaver D. J. & Kalb H. J. (2007) Adult migration and habitat 

utilization. In: Biology and conservation of Ridley sea turtles (ed P. T. Plotkin) pp. 213-229. 

Johns Hopkins University Press, Baltimore, Maryland. 
462

Morreale S. J., Standora E. A., Paladino F. V. & Spotila J. R. (1994) Leatherback migrations 

along deepwater bathymetric contours. Pp.109-110 In: Schoeder, B.A. and B.E. Witherington 

(Eds), Proceedings of the 13th Annual Symposium on Sea Turtle Biology and Conservation. 

NOAA Technical Memorandum NMFS-SEFSC-341, Miami, Florida. 

Morris R. H., Abbott D. P. & Haderlie E. C. (1980) Intertidal invertebrates of California. In: 

1980. Stanford University Press, Stanford, California. 

Morrison J., Quick M. & Foreman M. (2002) Climate change in the Fraser River watershed: flow 

and temperature projections. Journal of Hydrology 263. 

Morse D. E., Hooker N., Morse A. N. C. & Jensen R. A. (1988) Control of larval metamorphosis 

and recruitment in sympatric agariciid corals. Journal of Experimental Marine Biology and 

Ecology 116, 192-217. 

Mortimer J. A. (1982) Factors influencing beach selection by nesting sea turtles. In: The biology 

and conservation of sea turtles (ed K. Bjorndal) pp. 45-51. Smithsonian Institution Press, 

Washington, D.C. 

Mortimer J. A., Collie J., Jupiter T., Chapman R., Liljevik A. & Betsy B. (2003) Growth rates of 

immature hawksbills (Eretmochelys imbricata) at Aldabra Atoll, Seychelles (Western Indian 

Ocean). In: Twenty-second Annual Symposium on Sea Turtle Biology and Conservation (ed J. A. 

Seminoff) pp. 247-248. 

Mortimer J. A. & Donnelly M. (in review) 2007 IUCN red list status assessment: hawksbill turtle 

(Eretmochelys imbricata). 

Morton A. B. & Symonds H. K. (2002) Displacement of Orcinus orca (L.) by high amplitude 

sound in British Columbia, Canada. ICES Journal of Marine Science 59, 71-80. 

Moser H. G. (1967) Reproduction and development of Sebastodes paucispinis and comparison 

with other rockfishes off southern California. Copeia 4, 773-797. 

Moser H. G. (1996a) The early life stages of fishes in the California current region. National 

Marine Fisheries Service, La Jolla, California. 

Moser H. G. (1996b) Scorpaenidae: scorpionfishes and rockfishes. In: The early life stages of 

fishes in the California current region (ed H. G. Moser) pp. 733-795. Allen Press, Lawrence, 

Kansas. 

Moser H. G. & Boehlert G. W. (1991) Ecology of pelagic larvae and juveniles of the genus 

Sebastes. Environmental Biology of Fishes 30, 203-224. 

Moser H. G., Charter R. L., Watson W., Ambrose D. A., Butler J. L., Charter J. & Sandknop E. 

M. (2000a) Abundance and distribution of rockfish (Sebastes) larvae in the southern California 

Bight in relation to environmental conditions and fishery exploitation. pp. 132-147. California 

Cooperative Oceanic Fisheries Report. 

Moser H. G., Charter R. L., Watson W., Ambrose D. A., Butler J. L., Charter S. R. & Sandknop 

E. M. (2000b) Abundance and distribution of rockfish (Sebastes) larvae in the Southern 

California Bight in relation to environmental conditions and fishery exploitation. California 

Cooperative Oceanic Fisheries Investigations Reports 41, 132-147. 

Moser M. L. & Lindley S. T. (2007) Use of Washington estuaries by subadult and adult green 

sturgeon. Environmental Biology of Fishes 79, 243-253. 

Moser M. L. & Ross S. W. (1993) Distribution and movements of shortnose sturgeon (Acipenser 

brevirostrum) and other anadromous fishes of the lower Cape Fear River, North Carolina. In: 

Final Report to the U.S. Army Corps of Engineers, Wilmington, North Carolina. 
463

Moser M. L. & Ross S. W. (1995) Habitat Use and Movements of Shortnose and Atlantic 

Sturgeons in the Lower Cape Fear River, North Carolina. Transactions of the American 

Fisheries Society 124, 225. 

Mote P. W. (2003) Trends in snow water equivalent in the Pacific Northwest and their climatic 

causes. Geophysical Research Letters 30. 

Mount J. F. (1995) California rivers: The conflict between fluvial process and land use. 

University of California Press, Berkeley and Los Angeles, California. 

Moyle P. B. (1976a) The decline of anadromous fishes in California. Conservation Biology 8, 

869-870. 

Moyle P. B. (1976b) Inland fishes of California. University of California Press, Berkeley, 

California. 

Moyle P. B. (2002a) Inland fishes of California. p. 502. University of California Press, 

Berkeley, CA. 

Moyle P. B. (2002b) Inland fishes of California. Revised and Expanded. University of California 

Press, Berkeley, California. 

Moyle P. B., Foley P. J. & Yoshiyama R. M. (1992a) Status of green sturgeon, Acipenser 

medirostris, in California. In: Final Report submitted to National Marine Fisheries Service. 

University of California, Davis, California. 

Moyle P. B., Foley P. J. & Yoshiyama R. M. (1992b) Status of Green Sturgeon, Acipenser 

Medirostris, in California. p. 12. National Marine Fisheries Service, Terminal Island, CA. 

Moyle P. B., Yoshiyama R. M., Williams J. E. & Wikramanayake E. D. (1995a) Fish species of 

special concern in California. In: Second edition. Final report to California Department of Fish 

and Game, contract 2128IF. 

Moyle P. B., Yoshiyama R. M., Williams J. E. & Wikramanayake E. D. (1995b) Fish Species of 

Special Concern in California. p. 277. California Department of Fish and Game, Inland 

Fisheries Division. 

Mrosovsky N. (1994) Sex ratios of sea turtles. The Journal of Experimental Zoology 270, 16-27. 

Mrosovsky N., Hopkins-Murphy S. R. & Richardson J. I. (1984) Sex ratio of sea turtles: seasonal 

changes. Science 225, 739-741. 

Mrosovsky N., Ryan G. D. & James M. C. (2009) Leatherback turtles: The menace of plastic. 

Marine Pollution Bulletin 58, 287-289. 

MRWCI. (2007) The voice of the Merrimack. 

MSDEQ. (2000) Pearl Riiver basin status report 2000. Mississippi Department of Evironmental 

Quality. 

Muir W. D. & Williams J. G. (2011) Improving connectivity between freshwater and marine 

environments for salmon migrating through the lower Snake and Columbia River hydropower 

system. Ecological Engineering. 

Muller A., Butterworth D. S. & Johnston S. J. (2010) Preliminary results for a combined 

assessment of all seven Southern Hemisphere humpback whale breeding stocks. p. 28. IWC 

Scientific Committee, Agadir, Morocco. 

Mundy P. R. (1997) The role of harvest management in determining the status and future of 

Pacific salmon populations. In: Pacific salmon and their ecosystems: Status and future options

(eds D. J. Stouder, P. A. Bisson and R. J. Naiman) pp. 315-330. Chapman Hall, New York. 

Mundy P. R. & Cooney R. T. (2005) Physical and biological background. In: The Gulf of Alaska: 

Biology and oceanography (ed P. R. Mundy) pp. 15-23. Alaska Sea Grant College Program, 

University of Alaska, Fairbanks, Alaska. 
464

Mundy P. R. & Olsson P. (2005) Climate and weather. In: The Gulf of Alaska: Biology and 

oceanography (ed P. R. Mundy) pp. 25-34. Alaska Sea Grant College Program, University of 

Alaska, Fairbanks, Alaska. 

Munk K. (2001) Maximum ages of groundfishes in waters off Alaska and British Columbia and 

consideration of age determination. Alaska Fisheries Research Bulletin 8, 12-21. 

Munoz L. T. & Camacho A. J. (1976) Ensayo para la determincion de la primera edad de 

madurez de acuerdo a muestreos biometricos y cortes de concha de H. corrugata en Punta 

Abreojos, B.C. Sur Informe de Desarrollo. In: Mecanografiado. Inst. Nal de Pesca, EIPc-209, 20 

Abril 1976. 

Murakawa S. K. K., Balazs G. H., Ellis D. M., Hau S. & Eames S. M. (2000) Trends in 

fibropapillomatosis among green turtles stranded in the Hawaiian Islands, 1982-98. In: 

Nineteenth Annual Symposium on Sea Turtle Biology and Conservation (eds K. H. J. and T. 

Wibbels). 

Murison L. D. & Gaskin D. E. (1989) The distribution of right whales and zooplankton in the 

Bay of Fundy, Canada. Canadian Journal of Zoology 67, 1411-1420. 

Murphy M. L. & Hall J. D. (1981) Varied effects of clear-cut logging on predators and their 

habitat in small streams of the Cascade Mountains, Oregon. Canadian Journal of Fisheries and 

Aquatic Sciences 38, 137-145. 

Murphy T. M. & Hopkins S. R. (1984) Aerial and ground surveys of marine turtle nesting 

beaches in the southeast region. Final Report to NOAA/NMFS/SEFC, U.S. Department of 

Commerce, 73p. 

Murray N. K. & Fay F. H. (1979) The white whales or belugas, Delphinapterus leucas, of Cook 

Inlet, Alaska. In: International Whaling Commission Scientific Committee. Unpublished 

Document SC/31/SM12. 

Musick J. A., Harbin M. M., Berkeley S. A., Burgess G. H., Eklund A. M., Findley L., Gilmore 

R. G., Golden J. T., Ha D. S., Huntsman G. R., McGovern J. C., Parker S. J., Poss S. G., Sala E., 

Schmidt T. W., Sedbeny G. R., Weeks H. & Wright S. G. (2000a) Marine, estuarine, and 

diadromous fish stocks at risk of extinction in North America (exclusive of Pacific salmonids). 

Fisheries 25, 6-30. 

Musick J. A., Harbin M. M., Berkeley S. A., Burgess G. H., Eklund A. M., Findley L., Gilmore 

R. G., Golden J. T., Ha D. S., Huntsman G. R., McGovern J. C., Parker S. J., Poss S. G., Sala E., 

Schmidt T. W., Sedberry G. R., Weeks H. & Wright S. G. (2000b) Marine, estuarine, and 

diadromous fish stocks at risk of extinction in North America (exclusive of Pacific salmonids). 

Fisheries 25, 6-30. 

Musick J. A. & Limpus C. J. (1997) Habitat utilization, and migration in juvenile sea turtles. In: 

The biology of sea turtles (eds P. L. Lutz and J. A. Musick) pp. 137-163. CRC Press, Boca 

Raton, Florida. 

Myers M. J., Ylitaloc G. M., Krahnc M. M., Boydc D., Calkinsa D., Burkanovd V. & Atkinson 

S. (2008) Organochlorine contaminants in endangered Steller sea lion pups (Eumetopias jubatus) 

from western Alaska and the Russian Far East. Science of the Total Environment 369, 60-69. 

Mysing J. O. & Vanselous T. M. (1989) Status of satellite tracking of Kemp’s ridley sea turtles. 

In: First International Symposium on Kemp’s Ridley Sea Turtle Biology, Conservation, and 

Management. (eds C. W. Caillouet Jr. and A. M. Landry Jr.) pp. 122-115. Texas A&M 

University 
465

Nakamoto R. J., Kisanuki T. T. & Goldsmith G. H. (1995a) Age and growth of Klamath River 

green sturgeon (Acipenser medirostris). p. 27. US Fish and Wildlife Service, Klamath River 

Fishery Resource Office, Yreka, California. 

Nakamoto R. J., Kisanuki T. T. & Goldsmith G. H. (1995b) Age and growth of Klamath River 

green sturgeon (Acipenser medirostris). In: Project #93-FP-13. U.S. Fish and Wildlife Service. 

NCDENR. (1999) Basinwide Assessment Report Cape Fear River Basin. p. 24. North Carolina 

Department of Environmental and Natural Resources, Raleigh, NC. 

NCDEQ. (2004) North Carolina division of water quality annual report of fish kill events 2004. 

Division of Water Quality Environmental Sciences Section, Raleigh, North Carolina. 

NCDWQ. (1998) Yadkin-Pee Dee basinwide water quality management plan. North Carolina 

Division of Water Quality Water Quality Section Planning Branch, Raleigh, North Carolina. 

Nedwell J. & Edwards B. (2002) Measurements of underwater noise in the Arun River during 

piling at County Wharf, Littlehampton. Subacoustech, Ltd. 

Neel J., Hart C., Lynch D., Chan S. & Harris J. (1997) Oil spills in Washington state: a historical 

analysis. Washington Department of Ecology, Olympia, Washington. 

Nehlsen W. J., Williams E. & Lichatowich J. A. (1991) Pacific salmon at the crossroads: Stocks 

at risk from California, Oregon, Idaho, and Washington. Fisheries 16, 4-21. 

Nehring S. (2006a) Four arguments why so many alien species settle into estuaries, with special 

reference to the German river Elbe. Helgoland Marine Research 60, 127-134. 

Nehring S. (2006b) NOBANIS – Invasive alien species fact sheet-Crassostrea gigas. Online 

Database of the North European and Baltic Network on Invasive Alien Species. 

Neigel J. E. & Avise J. C. (1983) Clonal diversity and population structure in a reef-building 

coral, Acropora cervicornis: Self-recognition analysis and demographic interpretation. Evolution

37, 437-454. 

Neilson J., Gabriele C., Straley J., Hills S. & Robbins. J. (2005) Humpback whale entanglement 

rates in southeast Alaska. In: Sixteenth Biennial Conference on the Biology of Marine Mammals 

pp. 203-204, San Diego, California. 

Nell J. (2002) The Australian oyster industry. World Aquaculture 613, 8-10. 

Nelson M., Garron M., Merrick R. L., Pace III R. M. & Cole T. V. N. (2007a) Mortality and 

serious injury determinations for baleen whale stocks along the United States eastern seaboard 

and adjacent Canadian Maritimes, 2001-2005. In: U.S. Department of Commerce. Northeast 

Fisheries Science Center Reference Document 07-05. 

Nelson M., Garron M., Merrick R. L., Pace III R. M. & Cole T. V. N. (2007b) Mortality and 

serious injury determinations for baleen whale stocks along the United States eastern seaboard 

and adjacent Canadian Maritimes, 2001-2005. In: Reference Document 07-05 p. 26. U.S. 

Department of Commerce, NOAA, Northeast Fisheries Science Center. 

Nelson W. G., Brock R., Lee II H., Lamberson J. O. & Cole F. (2007c) Condition of bays and 

estuaries of Hawaii for 2002: A statistical summary. In: Environmental Protection Agency, 

Office of Research and Development, National Health and Environmental Effects Research 

Laboratory, EPA/620-R-07/001, Washington, D.C. . 

Nemoto T. (1957) Foods of baleen whales in the northern Pacific. . Scientific Reports of the 

Whales Research Institute, Tokyo 12, 33-89. 

Nichol L. M. & Shackleton D. M. (1996) Seasonal movements and foraging behaviour of 

northern resident killer whales (Orcinus orca) in relation to the inshore distribution of salmon 

(Oncorhynchus spp.) in British Columbia. Canadian Journal of Zoology 74, 983-991. 
466

Nichols O. C., Kenney R. D. & Brown M. W. (2008) Spatial and temporal distribution of North 

Atlantic right whales (Eubalaena glacialis) in Cape Cod Bay, and implications for management. 

Fishery Bulletin 106, 270-280. 

Nichols W. J. (2005) Following redwood logs, rubber ducks, and drift bottles: Transoceanic 

developmental migrations of loggerhead turtles in the North Pacific Ocean. In: Proceedings of 

the Twenty-First Annual Symposium on Sea Turtle Biology and Conservation (eds M. S. Coyne 

and R. D. Clark) p. 66. 

Nichols W. J., Resendiz A., Seminoff J. A. & Resendiz B. (2000) Transpacific migration of a 

loggerhead turtle monitored by satellite telemetry. Bulletin of Marine Science 67, 937-947. 

Nicieza A. G. & Metcalfe N. B. (1997) Growth compensation in juvenile Atlantic salmon: 

responses to depressed temperature and food availability. Ecology 78, 2385-2400. 

Nickelson T. E., Nicholas J. W., McGie A. M., Lindsay R. B., Bottom D. L., Kaiser R. J. & 

Jacobs S. E. (1992) Status of anadromous salmonids in Oregon coastal basins. In: Oregon 

Department of Fish and Wildlife, Resource Development Section and Ocean Salmon 

Management, Oregon Department of Fish and Wildlife, Portland, Oregon. 

Niklitschek E. J. & Secor D. H. (2005) Modeling spatial and temporal variation of suitable 

nursery habitats for Atlantic sturgeon in the Chesapeake Bay. Estuarine, coastal and shelf 

science 64, 135-148. 

Nishimura W. (1994) Internesting interval and nest frequency for loggerhead turtles on 

Inakahama Beach, Yakushima Island, Japan. Marine Turtle Newsletter 67, 21-22. 

Nishiwaki M. (1972) General biology. In: Mammals of the sea: Biology and medicine (ed S. H. 

Ridgway) pp. 3-204. Thomas, Springfield, IL. 

Nishiwaki M. & Handa C. (1958) Killer whales caught in the coastal waters off Japan for recent 

10 years. Scientific Reports of the Whales Research Institute 13, 85-96. 

Nishiymama T. (1984) Two important factors in oceanic stages of salmon: food and temperature. 

In: The Pacific Salmon Biology Conference pp. 227-258. 

NMA. (2007) 2004 State Mining Statistics. National Mining Association. 

NMFS-AKR. (2005) Final environmental impact statement for essential fish habitat 

identification and conservation in Alaska. National Marine Fisheries Service, Juneau, Alaska. 

NMFS-NWR. (2003) Amendment 17 to the Pacific coast groundfish fishery management plan 

(multi-year management and the specifications and management measures process). National 

Marine Fisheries Service-Northwest Region, Seattle, Washington. 

NMFS-NWR. (2005) Pacific coast groundfish fishery management plan. Essential fish habitat 

designation and minimization of adverse impacts. Final environmental impact statement. 

National Marine Fisheries Service-Northwest Region, Seattle, Washington. 

NMFS. (1987) Marine Mammal Protection Act of 1972. In: Annual Report 1986/87. National 

Marine Fisheries Service. 

NMFS. (1988) Biological Opinion on the removal of oil and gas platforms and related structures 

in the Gulf of Mexico. 

NMFS. (1991) Final recovery plan for the humpback whale (Megaptera novaeangliae). National 

Marine Fisheries Service, Office of Protected Resources, Silver Spring, Maryland. 

NMFS. (1992) Final recovery plan for Steller sea lions Eumetopias jubatus. NMFS Office of 

Protected Resources, Silver Spring, Maryland. 

NMFS. (1996) Status review update for coho salmon from Washignton, Oregon, and California. 

NMFS. (1997a) Status review update for coho salmon from the Oregon and Northern California 

coasts. West coast coho salmon biological review team. 
467

NMFS. (1997b) Status review update for West Coast steelhead from Washington, Idaho, Oregon, 

and California. Memorandum date 7 July 1997 from the biological review team to the National 

Marine Fisheries Service Northwest Regional Office. 

NMFS. (1998a) Final Recovery Plan for the Shortnose Sturgeon Acipenser brevirostrum. (ed 

Prepared by the Shortnose Sturgeon Recovery Team for the National Marine Fisheries Service) 

p. 104, Silver Spring, Maryland. 

NMFS. (1998b) Recovery Plan for the Shortnose Sturgeon (Acipenser brevirostrum). Prepared 

by the Shortnose Sturgeon Recovery Team for the National Marine Fisheries Service, Silver 

Spring, Maryland. 104p. 

NMFS. (1998c) Status review update for sockeye salmon from Ozette Lake and Baker River, 

Washington. West coast sockeye biological review team. Northwest Fisheries Science Center, 

Seattle, Washington. 

NMFS. (1999) Endangered Species Act Section 7 consultation (biological opinion) for the U.S. 

Army Engineer District, Alaska on the proposed construction and operation of the Northstar oil 

and gas project. National Marine Fisheries Service. 

NMFS. (2000) Status review of smalltooth sawfish (Pristis pectinata). Department of Commerce, 

National Marine Fisheries Service, Silver Spring, Maryland. 

NMFS. (2001a) Stock assessments of loggerhead and leatherback sea turtles and an assessment 

of the impact of the pelagic longline fishery on the loggerhead and leatherback sea turtles of the 

western North Atlantic. NOAA Technical Memorandum NMFS-SEFSC-455. 

NMFS. (2001b) Stock assessments of loggerhead and leatherback sea turtles and an assessment 

of the impact of the pelagic longline fishery on the loggerhead and leatherback sea turtles of the 

western North Atlantic. 

NMFS. (2002) Endangered Species Act - Section 7 consultation, biological opinion. Shrimp 

trawling in the southeastern United States under the sea turtle conservation regulations and as 

managed by the fishery management plans for shrimp in the South Atlantic and Gulf of Mexico. 

National Marine Fisheries Service, Southeast Regional Office, St. Petersburg, Florida. 

NMFS. (2004) Biological opinion on the authorization of pelagic fisheries under the fisheries 

management plan for the pelagic. National Marine Fisheries Service, Silver Spring, Maryland. 

NMFS. (2005a) Assessment of acoustic exposures on marine mammals in conjunction with USS 

Shoup active sonar transmissions in the Eastern Strait of Juan de Fuca and Haro Strait, 

Washington 5 May 2003. National Marine Fisheries Service, Office of Protected Resources. 

NMFS. (2005b) Assessment of acoustic exposures on marine mammals in conjunction with USS 

Shoup active sonar transmissions in the eastern Strait of Juan de Fuca and Haro Strait, 

Washington ~ 5 May 2003 ~. NMFS, Office of Protected Resources, Silver Spring, Maryland. 

NMFS. (2005c) Biological Opinion on the Issuance of ESA Section 10(a)(1)(A) Permit No. 1451 

to the National Marine Fisheries Service - Office of Sustainable Fisheries for Research on Sea 

Turtles. National Marine Fisheries Service, Office of Protected Resources, Silver Spring, 

Maryland. 48p. 

NMFS. (2005d) Draft Stock assessment report for Steller sea lions (Eumetopias jubatus): 

Western U.S. Stock. National Marine Fisheries Service.

NMFS. (2005e) Final assessment of NOAA Fisheries' critical habitat analytical review teams for 

12 evolutionarily significant units of West Coast salmon and steelhead. In: CHART Report. 

National Oceanic and Atmospheric Administration, National Marine Fisheries Service, Protected 

Resources Division, Portland, Oregon. 
468

NMFS. (2005f) Final assessment of NOAA Fisheries' critical habitat analytical review teams for 

12 evolutionarily significant units of West Coast salmon and steelhead. National Marine 

Fisheries Service. 

NMFS. (2005g) Proposed conservation plan for southern resident killer whales (Orcinus orca). 

National Marine Fisheries Service, Northwest Region, Seattle, Washington. 

NMFS. (2005b) Final assessment of NOAA Fisheries' critical habitat analytical review teams for 

12 evolutionarily significant units of West Coast salmon and steelhead. available online at 

http://www.nwr.noaa.gov/Salmon-Habitat/Critical-Habitat/2005 Biological -Teams-Report. 

accessed July 2008. 

NMFS. (2006a) Biological assessment of the Alaska groundfish fisheries and NMFS managed 

Endangered Species Act listed marine mammals and sea turtles. National Marine Fisheries 

Service, Juneau, Alaska. 

NMFS. (2006b) Biological opinion on the issuance of section lO(a)(l)(A) permits to conduct 

scientific research on the southern resident killer whale (Orcinus orca) distinct population 

segment and other endangered or threatened species. p. 92. National Marine Fisheries Service, 

Northwest Region, Seattle, Washington. 

NMFS. (2006c) Conservation plan for the Cook Inlet beluga whale (Delphinapterus leucas). 

National Marine Fisheries Service, Juneau, Alaska. 

NMFS. (2006d) Draft Columbia River estuary recovery plan module. Lower Columbia River 

Estuary Partnership under contract to NOAA Fisheries. 

NMFS. (2006e) Endangered and threatened species: final listing determinations for elkhorn coral 

and staghorn coral. National Marine Fisheries Service. 71 FR 26852. 

NMFS. (2006f) Endangered Species Act Section 7 consultation (biological opinion) on oil and 

gas leasing and exploration activities in the Beaufort Sea. National Marine Fisheries Service, 

Silver Spring, Maryland. 

NMFS. (2006g) Recovery plan for smalltooth sawfish (Pristis pectinata). Prepared by the 

Smalltooth Sawfish Recovery Team for the National Marine Fisheries Service, Silver Spring, 

Maryland. 

NMFS. (2006h) Recovery Plan for Smalltooth Sawfish (Pristis pectinata). Prepared by the 

Smalltooth Sawfish Recovery Team for the National Marine Fisheries Service, Silver Spring, 

Maryland. 

NMFS. (2006i) Supplement to the draft Snake River salmon recovery plan for Southeast 

Washington. National Marine Fisheries Service, Northwest Region, Seattle, Washington. 

NMFS. (2006e) Biological Opinion on Permitting Structure Removal Operations on the Gulf of 

Mexico Outer Continental Shelf and the Authorization for Take of Marine Mammals Incidental 

to Structure Removals on the Gulf of Mexico Outer Continental Shelf. National Marine 

Fisheries Service, Silver Spring, Maryland. 131p. 

NMFS. (2006f) National Marine Fisheries Service, Office of Protected Resources website: 

http://www.nmfs.noaa.gov/pr/. 

NMFS. (2007a) Draft Revised Steller Sea Lion Recovery Plan, Eastern and Western Distinct 

Population Segments (Eumetopias jubatus). Prepared by the Steller Sea Lion Recovery Team for 

the National Oceanic and Atmospheric Administration, National Marine Fisheries Service, 

Office of Protected Resources, May 2007. 

NMFS. (2007b) Recovery plan for the Hawaiian Monk Seal (Monachus schauinslandi). Silver 

Spring, Maryland. 
469

NMFS. (2007c) Report to Congress: Pacific coastal salmon recovery fund. National Marine 

Fisheries Service. 

NMFS. (2008a) Biological opinion on the continued authorization of spiny lobster fishing 

managed under the Spiny Lobster Fishery Management Plan of Puerto Rico and the U.S. Virgin 

Islands (SLFMP) (Consultation Number F/SER/2008/09173). NMFS. 

NMFS. (2008b) Draft U.S. Atlantic marine mammal stock assessments 2008. 

NMFS. (2008c) Listing endangered and threatened species: Notification of finding on a petition 

to list Pacific eulachon as an endangered or threatened species under the Endangered Species 

Act. National Marine Fisheries Service. 

NMFS. (2008d) Preliminary scientific conclusions of the review of the status of 5 species of 

rockfish: bocaccio (Sebastes paucispinis), canary rockfish (Sebastes pinniger), yelloweye 

rockfish (Sebastes ruberrimus), greenstriped rockfish (Sebastes elongatus) and redstripe rockfish 

(Sebastes proriger) in Puget Sound, Washington. NMFS, NWFSC, Seattle, Washignton. 

NMFS. (2008e) Recovery Plan for Southern Resident Killer Whales (Orcinus orca). p. 251. 

National Marine Fisheries Service, Northwest Region, Seattle, Washington. 

NMFS. (2008f) Recovery plan for southern resident killer whales (Orcinus orca). National 

Marine Fisheries Service, Northwest Region, Seattle, Washington. 

NMFS. (2008g) Recovery Plan for the Steller Sea Lion (Eumetopias jubatus). Revision. National 

Marine Fisheries Service, Silver Spring, Maryland. 325p. 

NMFS. (2009a) Environmental assessment on issuance of a permit for field research and 

enhancement activities on the endangered Hawaiian monk seal (Monachus schauinslandi) p. 52. 

Office of Protected Resources, Silver Spring, Maryland. 

NMFS. (2009b) Middle Columbia River steelhead distinct population segment ESA recovery 

plan. NMFS, Seattle, Washington. 

NMFS. (2009c) Recovery plan for Lake Ozette sockeye salmon. National Marine Fisheries 

Service. 

NMFS. (2010) Other significant oil spills in the Gulf of Mexico. (ed N. M. F. Service). National 

Marine Fisheries Service, Office of Response and Restoration, Emergency Response Division, 

Silver Sprirng, Maryland. 

NMFS. (2011a) 5-year review: summary and evaluation of California coastal Chinook salmon 

ESU and central California coast coho salmon ESU. National Marine Fisheries Service, Long 

Beach, California. 

NMFS. (2011b) 5-year review: summary and evaluation of central valley spring-run Chinook 

salmon ESU. National Marine Fisheries Service, Long Beach, California. 

NMFS. (2011c) 5-year review: summary and evaluation of Central Valley steelhead DPS 

National Marine Fisheries Service, Long Beach, California. 

NMFS. (2011d) 5-year review: summary and evaluation of lower Columbia River Chinook, 

Columbia River chum, lower Columbia River coho, and lower Columbia River steelhead. 

National Marine Fisheries Service, Portland, Oregon. 

NMFS. (2011e) 5-year review: summary and evaluation of Middle Columbia River steelhead 

National Marine Fisheries Service, Portland, Oregon. 

NMFS. (2011f) 5-year review: summary and evaluation of Puget Sound Chinook, Hood Canal 

summer chum, and Puget Sound steelhead. National Marine Fisheries Service, Portland, Oregon. 

NMFS. (2011g) 5-year review: summary and evaluation of Sacramento River winter-run 

Chinook salmon ESU. National Marine Fisheries Service, Long Beach, California. 
470

NMFS. (2011h) 5-year review: summary and evaluation of Southern Oregon/Northern California 

coast coho salmon ESU. National Marine Fisheries Service, Long Beach, California. 

NMFS. (2011i) 5-year review: summary and evaluation of Upper Columbia River steelhead and 

Upper Columbia River spring-run Chinook National Marine Fisheries Service, Portland, Oregon. 

NMFS. (2011j) 5-year review: summary and evaluation of upper Willamette River steelhead and 

upper Willamette River Chinook. National Marine Fisheries Service, Portland, Oregon. 

NMFS. (2011k) Biennial report to Congress on the recovery program for threatened and 

endangered species October 1, 2008 – September 30, 2010. National Oceanic and Atmospheric 

Administration, National Marine Fisheries Service. 

NMFS. (2011l) Biological opinion on the continued authorization of reef fish fishing under the 

Gulf of Mexico (Gulf) Reef Fish Fishery Management Plan (RFFMP). NMFS. 

NMFS & USFWS. (1991) Recovery Plan for U.S. Population of Atlantic Green Turtle Chelonia 

mydas. National Marine Fisheries Service and U.S. Fish and Wildlife Service, Washington, D.C. 

NMFS & USFWS. (1993) Recovery Plan for the hawksbill turtle in the U.S. Caribbean Sea, 

Atlantic Ocean, and Gulf of Mexico. In: National Marine Fisheries Service and U.S. Fish and 

Wildlife Service, St. Petersburg, Florida. 

NMFS & USFWS. (1998a) Recovery plan for U.S. Pacific populations of the green turtle 

(Chelonia mydas). National Marine Fisheries Service, Silver Spring, Maryland. 

NMFS & USFWS. (2005) Final recovery plan for the Gulf of Maine distinct population segment 

of Atlantic salmon (Salmo salar). NMFS and USFWS, Silver Spring, Maryland and Hadley, 

Massachussetts. 

NMFS & USFWS. (2007a) Green Sea Turtle (Chelonia mydas) 5-Year Review: Summary and 

Evaluation p. 105. National Marine Fisheries Service and U.S. Fish and Wildlife Service, Silver 

Spring, MD. 

NMFS & USFWS. (2007b) Hawksbill Sea Turtle (Eretmochelys imbricata) 5-Year Review: 

Summary and Evaluation p. 93. National Marine Fisheries Service and U.S. Fish and Wildlife 

Service. 

NMFS & USFWS. (2007c) Kemp's Ridley Sea Turtle (Lepidochelys kempii) 5-Year Review: 

Summary and Evaluation. p. 50. National Marine Fisheries Service and U.S. Fish and Wildlife 

Service. 

NMFS & USFWS. (2007d) Kemp's Ridley sea turtle (Lepidochelys kempii) 5-year review: 

Summary and evaluation. U.S. Department of Commerce, National Oceanic and Atmospheric 

Administration, National Marine Fisheries Service, Office of Protected Resources 

U.S. Department of the Interior, U.S. Fish and Wildlife Service, Southeast Region, Jacksonville 

Ecological Services Field Office, JSilver Spring, Maryland 

acksonville, Florida. 

NMFS & USFWS. (2008) DRAFT Recovery Plan for the Northwest Atlantic Population of the 

Loggerhead Sea Turtle (Caretta caretta): Second Revision. p. 306. National Marine fisheries 

Service, U.S. Fish and Wildlife Service, Silver Spring, MD. 

NMFS and USFWS. (1991b) Recovery Plan for U.S. Population of Loggerhead Turtle (Caretta 

caretta). National Marine Fisheries Service, Washington, D.C. 

NMFS and USFWS. (1998b) Recovery plan for U.S. Pacific populations of the loggerhead turtle 

(Caretta caretta). NMFS and USFWS, Silver Spring, Maryland 

Portland, Oregon. 

NMFS and USFWS. (1998d) Recovery Plan for U.S. Pacific Populations of the Loggerhead 

Turtle (Caretta caretta). National Marine Fisheries Service, Silver Spring, MD. 
471

NMFS and USFWS. (2007e) Hawksbill sea turtle (Eretmochelys imbricata) 5-year review: 

Summary and evaluation U.S. Department of Commerce, National Oceanic and Atmospheric 

Administration, National Marine Fisheries Service, Office of Protected Resources 

U.S. Department of the Interior, U.S. Fish and Wildlife Service, Southeast Region, Jacksonville 

Ecological Services Field Office, Silver Spring, Maryland 

Jacksonville, Florida. 

NMFS and USFWS. (2007d) Leatherback Sea Turtle (Dermochelys coriacea) 5-Year Review: 

Summary and Evaluation. U.S. Department of Commerce, National Oceanic and Atmospheric 

Administration, National Marine Fisheries Service, Office of Protected Resources, Silver Spring, 

Maryland; and U.S. Department of the Interior, U.S. Fish and Wildlife Service, Southeast 

Region, Jacksonville Ecological Services Field Office, Jacksonville, Florida. 81p. 

NMFS and USFWS. (2010) Final draft report: Summary report of a meeting of the 

NMFS/USFWS cross-agency working group on joint listing of North Pacific and northwest 

Atlantic loggerhead turtle distinct population segments. NMFS and USFWS, Washington, D.C. 

NMFS/SEFSC. (2001) Stock assessments of loggerhead and leatherback sea turtles and an 

assessment of the impact of the pelagic longline fishery on the loggerhead and leatherback sea 

turtles of the western North Atlantic. U.S. Department of Commerce, NOAA Technical 

Memorandum NMFS-SEFSC-455. 

NMMA. (2007) 2006 Recreational boating statistical abstract. National Marine Manufacturers 

Association, Chicago, Illinois. 

NOAA. (2003) Oil and sea turtles: Biology, planning, and response. (ed G. Shigenaka). National 

Oceanic and Atmospheric Administration, National Ocean Service, Office of Response and 

Restoration. 

NOAA. (2010a) Deepwater Horizon. 

NOAA. (2010b) NOAA's oil spill response: Sea turtle strandings and the Deepwater oil spill. 

(ed N. O. a. A. Administration). 

NOAA. (2010c) NOAA web update July 26, 2010: Deepwater Horizon incident. National 

Oceanic and Atmospheric Administration. 

Nobriga M. L. & Feyrer F. (2008) Diet composition in San Francisco Estuary striped bass: does 

trophic adaptability have its limits? Environmental Biology of Fishes 83, 495-503. 

Noda N., Ichihashi H., Loughlin T. R., Baba N., Kiyota M. & Tatsukawa R. (1995) Distribution 

of heavy metals in muscle, liver, and kidney of northern fur seal (Callorhinus ursinus) caught off 

Sanriku, Japan, and from the Pribilof Islands, Alaska. Environmental Pollution 90, 51-59. 

Noren D. P., Johnson A. H., Rehder D. & Larson. A. (2009) Close approaches by vessels elicit 

surface active behaviors by southern resident killer whales. Endangered Species Research 8, 

179-192. 

Norkko J., Reed D. C., Timmerman K., Norkko A., Gustafsson B. G., Bonsdorff E., Slomp C. P., 

Carstensen J. & Conley D. J. (2011) A welcome can of worms? Hypoxia mitigation by an 

invasive species. Geophysical Research Abstracts 13. 

Norman J. R. & Fraser F. C. (1937) Giant fishes, whales and dolphins. Putman and Co., Ltd., 

London, UK. 

Norman S. A., C. E. Bowlby, M. S. Brancato, J. Calambokidis, D. Duffield P. J. G., T. A. 

Gornall, M. E. Gosho, B. Hanson, J. Hodder, S. J. Jefferies, B. Lagerquist, D. M. Lambourn, B. 

Mate, B. Norberg, R. W. Osborne, J. A. Rash, S. Riemer & Scordino J. (2004) Cetacean 

strandings in Oregon and Washington between 1930 and 2002. Journal of Cetacean Research 

and Management 6, 87-99. 
472

Norrgard J. (1995) Determination of stock composition and natal origin of a juvenile loggerhead 

turtle population (Caretta caretta) in Chesapeake Bay using mitochondrial DNA analysis. 

College of William and Mary, Williamsburg, Virginia. 

Northcote T. G. (1978) Migratory strategies and production in freshwater fishes. In: Ecology of 

Freshwater Fishes (ed S. D. Gerking) pp. 326-329. Blackwell Science, Oxford. 

Nowacek D. P., Johnson M. P. & Tyack P. L. (2004) North Atlantic right whales (Eubalaena 

glacialis) ignore ships but respond to alerting stimuli. Proceedings of the Royal Society of 

London B 271, 227-231. 

Nowak R. M. (1991) Walker’s marine mammals of the world. The Johns Hopkins University 

Press, Baltimore, Maryland. 

Nowak R. M. (2003) Walker's marine mammals of the world. John Hopkins University Press, 

London. 

NPFMC. (1990) Fishery management plan for the salmon fisheries in the EEZ off the coast of 

Alaska. North Pacific Fishery Management Council, Anchorage, Alaska. 

NPS. (2008) Freshwater fishes of Wrangell-St. Elias National Park and Preserve, Alaska. 

National Park Service, U.S. Department of the Interior. 

NRC. (1990a) Decline of the sea turtles: Causes and prevention. National Research Council, 

Washington, D.C. 

NRC. (1990b) Decline of the Sea Turtles: Causes and Prevention. National Academy of 

Sciences, National Academy Press, Washington, D.C. 

NRC. (1990c) Decline of the sea turtles: Causes and prevention. In: (National Research 

Council). National Academy Press, Washington, D.C. 

NRC. (1994) Low-frequency sound and marine mammals, current knowledge and research 

needs. In: (National Research Council). National Academy Press, Washington, D.C. 

NRC. (1996a) Upstream: salmon and society in the Pacific Northwest. National Research 

Council, National Academy Press, Washington, D.C. 

NRC. (1996b) Upstream: salmon and society in the Pacific Northwest. National Research 

Council, National Academy Press, Washington, D.C. . 

NRC. (1996c) Upstream: Salmon and society in the Pacific northwest. In: (National Research 

Council). National Academy Press, Washington, D.C. . 

NRC. (2003) Ocean Noise and Marine Mammals. National Research Council: Committee on 

Potential Impacts of Ambient Noise in the Ocean on Marine Mammals. 

NRC. (2004) Managing the Columbia River. In: Instream flows, water withdrawals, and salmon 

survival. (National Research Council). National Academy Press, Washington D.C. 

NRC. (2005) Marine mammal populations and ocean noise: determining when noise causes 

biologically significant effects. In: (National Research Council). National Academies Press, 

Washington, D.C. 

NWPPC. (1986) Compilation of information on salmon and steelhead losses in the Columbia 

River basin. In: Report to the Northwest Power Planning Council, Portland, Oregon. 

Nyberg C. D. (2007) Introduced marine macroalgae and habitat modifiers - their ecological role 

and significant attributes. p. 61. Goeteborg University, Goeteborg, Sweden. 

O'Connell V. & Carlisle D. W. (1993) Habitat-specific density of adult yelloweye rockfish 

Sebastes ruberrimus in the eastern Gulf of Alaska. Fishery Bulletin 91, 304-309. 

O'Connell V. M. (1987) Reproductive seasons for some Sebastes species in southeastern Alaska. 

O'Corry-Crowe G. (2008) Climate change and the molecular ecology of Arctic marine mammals. 

Ecological Applications 18, S56-S76. Arctic Marine Mammals and Climate Change. 
473

O'Corry-Crowe G., Suydam R., Frey A., Mahoney B., Quakenbush L., Harwood L., Goodwin 

W. & Hobbs. R. (2010) Molecular genetic analysis of population structure, dispersal and gene 

flow of beluga whales, Delphinapterus leucas, in the western Nearctic: New findings on the 

Cook Inlet population. In: Cook Inlet Belgua Whale Science Conference p. 2, Anchorage 

Downtown Marriott, Anchorage, Alaska. 

O'Corry-Crowe G., Suydam R., Orr J., Lowry L., Frost K., Litovka D. & Hobbs. R. (2007) 

Combination of satellite telemetry and molecular genetic analysis in the study of beluga whales 

in the Western Nearctic. In: NOAA Technical Memorandum NMFS-F/SPO-82. vi + 82pp. p.37. 

Report of the National Marine Fisheries Service Workshop on Advancing Electronic Tag 

Technologies and Their Use in Stock Assessments. 

O'Hara K. J., Iudicello S. & Bierce R. (1988) A citizens guide to plastics in the ocean: More than 

a litter problem. Center for Marine Conservation, Washington, D.C. 

O'Herron II J. C., Able K. W. & Hastings R. W. (1993) Movements of shortnose sturgeon 

(Acipenser brevirostrum) in the Delaware River. Estuaries 16, 235-240. 

O’Corry-Crowe G. M., Suydam R. S., Rosenberg A., Frost K. J. & Dizon A. E. (1997) 

Phylogeography, population structure, and dispersal patterns of the beluga whale Delphinapterus 

leucas in the western Arctic revealed by mitochondrial DNA. Molecular Ecology 6, 955-970. 

Odemar M. W. (1964) Southern range extension of the eulachon, Thaleichthys pacificus. 

California Fish and Game 50, 305-307. 

ODFW and NMFS. (2011) Upper Willamette River conservation and recovery plan for Chinook 

salmon and steelhead. Oregon Department of Fish and Wildlife and National Marine Fisheries 

Service, Northwest Region. 

ODFW and WDFW. (2007) Status report: Columbia River fish runs and fisheries, 1938-2002. 

Oregon Department of Fish and Wildlife and Washington Department of Fish and Wildlife. 

Ognetov G. N. (1981) Studies on the ecology and the taxonomy of the white whale 

(Delphinapterus leucas Pall., 1776) inhabiting the Soviet Arctic SC/32?SM1. Report of the 

International Whaling Commission 31, 515-520. 

Ogren L. H. (1989) Distribution of juvenile and subadult Kemp's ridley sea turtles: Preliminary 

results from 1984-1987 surveys. In: First International Symposium on Kemp's Ridley Sea Turtle 

Biology, Conservation, and Management (eds C. W. Caillouet Jr. and A. M. Landry Jr.) pp. 116-

123. 

Ohsumi S. (1979) Interspecies relationshops among some biological parameters in cetaceans and 

estimation of the natural mortality coefficient of the southern hemishpere minke whale. Report of 

the International Whaling Commission 29, 397-406. 

Øien N. (2001) Humpback whales in the Barents and Norwegian Seas. Paper SC/53/NAH21 

presented to the International Whaling Commission Scientific Committee. Available from IWC, 

135 Station Road, Impington, Cambridge, UK. 

Okemwa G. M. & Wamukota A. (2006) An overview of the status of green turtles (Chelonia 

mydas) in Kenya. In: Book of Abstracts. Twenty-sixth Annual Symposium on Sea Turtle Biology 

and Conservation (eds M. Frick, A. Panagopoulou, A. F. Rees and K. Williams) p. 311. 

International Sea Turtle Society, Athens, Greece. 

Okuyama J., Abe O., Nishizawa H., Kobayashi M., Yoseda K. & Arai N. (2009) Ontogeny of the 

dispersal migration of green turtle (Chelonia mydas) hatchlings. Journal of Experimental Marine 

Biology and Ecology. 

Olenin S. & Leppäkoski E. (1999) Non-native animals in the Baltic Sea: alteration of benthic 

habitats in coastal inlets and lagoons. Hydrobiologia 393, 233–243. 
474

Olesiuk P. F., Bigg M. A. & Ellis G. M. (1990a) Life history and population dynamics of 

resident killer whales (Orcinus orca) in the coastal waters of British Columbia and Washington 

stat. In: Individual recognition of cetaceans: use of photo-identification and other techniques to 

estimate population parameters (eds P. S. Hammond, S. A. Mizroch and G. P. Donovan) pp. 

209-243. International Whaling Commission, Cambridge. 

Olesiuk P. F., Bigg M. A. & Ellis G. M. (1990b) Life history and population dynamics of 

resident killer whales (Orcinus orca) in the coastal waters of British Columbia and Washington 

State. Report of the International Whaling Commission Special Issue 12, 209-243. 

Olesiuk P. F., Bigg M. A. & Ellis. G. M. (1990c) Life history and population dynamics of 

resident killer whales (Orcinus orca) in the coastal waters of British Columbia and Washington 

state. Report of the International Whaling Commission Special Issue 12, 209-243. 

Olesiuk P. F., Bigg M. A. & Ellis. G. M. (1990d) Life history and population dynamics of 

resident killer whales (Orcinus orca) in the coastal waters of British Columbia and Washington 

state. Report of the International Whaling Commission Special Issue 12, 209-243.-Individual 

Recognition of Cetaceans Use of Photo-Identification and Other Techniques To Estimate 

Population Parameters). 

Olesiuk P. F., Nichol L. M., Sowden M. J. & Ford J. K. B. (2002) Effect of the sound generated 

by an acoustic harassment device on the relative abundance and distribution of harbor porpoises 

(Phocoena phocoena) in Retreat Passage, British Columbia Marine Mammal Science 18, 843-

862. 

Olson J. M. (1998) Temporal and spatial distribution patterns of sightings of southern 

community and transient orcas in the inland waters of Washington and British Columbia. 

Western Washington University, Bellingham, Washington. 

Oros J., Gonzalez-Diaz O. M. & Monagas P. (2009) High levels of polychlorinated biphenyls in 

tissues of Atlantic turtles stranded in the Canary Islands, Spain. Chemosphere 74, 473-478. 

Orvik L. M. (1997) Trace metal concentration in blood of the Kemp's ridley sea turtle 

(Lepidochelys kempii). Texas A & M University, College Station, Texas. 

Osborne R. W. (1991) Trends in killer whale movements, vessel traffic, and whale watching in 

Haro Strait. In: Puget Sound Research ’91 pp. 672-688. Puget Sound Water Quality Authority, 

Olympia, Washington. 

Osborne R. W. (1999) A historical ecology of Salish Sea “resident” killer whales (Orcinus orca): 

with implications for management. University of Victoria, Victoria, British Columbia. 

Osborne R. W., Koski K. L., Tallmon R. E. & Harrington S. (1999) Soundwatch 1999 final 

report. Soundwatch, Roche Harbor, Washington. 

Pace III R. M. & Merrick R. L. (2008) Northwest Atlantic Ocean habitats important to the 

conservation of North Atlantic right whales (Eubalaena glacialis). In: Northeast Fisheries 

Science Center Reference Document 08-07. 

Pack A. A., Herman L. M., Spitz S. S., Hakala S., Deakos M. H. & Herman. E. Y. K. (2009) 

Male humpback whales in the Hawaiian breeding grounds preferentially associate with larger 

females. Animal Behaviour 77, 653-662. 

Padilla D. K. & Williams S. L. (2004) Beyond ballast water: aquarium and ornamental trades as 

sources of invasive species in aquatic ecosystems. Frontiers in Ecology and the Environment 2, 

131-138. 

Paine R. T. (1974) Ecological determinism in the competition for space. Ecology 65, 339-348. 

Palmisano J. F., Ellis R. H. & Kaczynski V. W. (1993a) The impact of environmental and 

management factors on Washington’s wild anadromous salmon and trout. Washington Forest 
475

Protection Association and Washington Department of Natural Resources, Olympia, 

Washington. 

Palmisano J. F., Ellis R. H. & Kaczynski V. W. (1993b) The impact of environmental and 

management factors on Washington’s wild anadromous salmon and trout. Washington Forest 

Protection Association and Washington Department of Natural Resources, Olympia, 

Washington. 

Palsbøll P. J., Allen J., Bérubé M., Clapham P. J., Feddersen T. P., Hudson R. R., Jørgensen H., 

Katona S., Larsen A. H., Larsen F., Lien J., Mattila D. K., Sigurjónsson J., Sears R., Smith T., 

Sponer R., Stevick P. & Øien N. (1997) Genetic tagging of humpback whales. Nature 388, 767-

769. 

Palsson W. A., Tsou T.-S., Barbman G. G., Buckley R. M., West J. E., Mills M. L., Cheng Y. W. 

& Pacunski R. E. (2008) The biology and assessment of rockfishes in Puget Sound. Washington 

Department of Fish and Wildlife, Olympia, Washignton. 

Panov V. E., Bychenkov D. E., Berezina N. A. & Maximov A. A. (2003) Alien species 

introductions in the eastern Gulf of Finland: current state and possible management options. 

Proceedings of the Estonian Academy of Science, Bioliogy, and Ecology 52, 254–267. 

Parente W. D. & Snyder G. R. (1970) A pictorial record of the hatching and early development 

of the eulachon (Thaleichthys pacificus). Northwest Science 44, 50-57. 

Parker C. & O’Reilly J. (1991) Oxygen depletion in Long Island Sound: A historical perspective. 

Estuaries and Coasts 14, 248-264. 

Parker D. M. & Balazs G. H. (in press) Diet of the oceanic green turtle, Chelonia mydas, in the 

North Pacific. In: Twenty-fifth Annual Symposium on Sea Turtle Biology and Conservation. 

Parker D. M., Cooke W. J. & Balazs G. H. (2005) Diet of oceanic loggerhead sea turtles (Caretta 

caretta) in the central North Pacific. Fishery Bulletin 103, 142-152. 

Parker R. O., Jr., Colby D. R. & Willis T. D. (1983) Estimated amount of reef habitat on a 

portion of the U.S. South Atlantic and Gulf of Mexico continental shelf. Bulletin of Marine 

Science 33, 935-940. 

Parker Stetter S. L. P., Rudstam L. G. & Mills E. L. (2005) Energetic consequences of diet shifts 

in Lake Erie rainbow smelt (Osmerus mordax). Canadian Journal of Fisheries and Aquatic 

Sciences 62, 145-152. 

Parrish F. A. & Abernathy K. (2006) Movements of monk seals relative to ecological depth 

zones in the lower Northwestern Hawaiian Islands. Atoll Research Bulletin 543, 115-130. 

Parrish F. A., Abernathy K., Marshall G. J. & Buhleier B. M. (2002) Hawaiian monk seals 

(Monachus schauinslandi) foraging in deep-water coral beds. Marine Mammal Science 18, 15. 

Parrish F. A., Craig M. P., Ragen T. J., Marshall G. J. & Buhleier B. M. (2000) Identifying 

diurnal foraging habitat of endangered Hawaiian monk seals using a seal-mounted video camera. 

Marine Mammal Science 16, 392. 

Parsons K. M., III K. C. B., Ford J. K. B. & Durban. J. W. (2009) The social dynamics of 

southern resident killer whales and conservation implications for this endangered population. 

(Orcinus orca). Animal Behaviour 77, 963-971. 

Pascual M. A. & Adkison M. D. (1994) The decline of the Steller sea lion in the northeast 

Pacific: Demography harvest or environment. Ecological Applications 4, 393-403. 

Pauley G. B., Bortz B. M. & Shepard M. F. (1986) Species Profiles: Life Histories and 

Environmental Requirements of Coastal Fishes and Invertebrates (Pacific Northwest) - Steelhead 

Trout. U.S. Fish Wildlife Service Biological Report 82(11.62). U.S. Army Corps of Engineers, TR 

EL-82-4. 24p. 
476

Pauley G. B., Bowers K. L. & Thomas G. L. (1988) Species Profiles: Life Histories and 

Environmental Requirements of Coastal Fishes and Invertebrates (Pacific Northwest) - Chum 

Salmon. Biological Report 82(11.81), TR EL-82-4, March 1988. Prepared for the U.S. 

Department of Interior, Fish and Wildlife Service, and the U.S. Army Corps of Engineers, 

Coastal Ecology Group, Waterways Experiment Station. 26p. 

Pauley G. B., Risher R. & Thomas G. L. (1989) Species Profiles: Life Histories and 

Environmental Requirements of Coastal Fishes and Invertebrates (Pacific Northwest) - Sockeye 

Salmon. Biological Report 82 (1 1.116), TR EL-82-4, December 1989. Prepared for the U.S. 

Department of the Interior, Fish and Wildlife Service, and the U.S. Army Corps of Engineers, 

Coastal Ecology Group, Waterways Experiment Station. 32p. 

Paulo da Cunha Lana C. G. (1991) Influence of Spartina alterniflora on structure and temporal 

variability of macrobenthic associations in a tidal flat of Paranagua Bay (southeastern Brazil). 

Marine Ecology Progress Series 73, 231-244. 

Payne P., Nicholas J., O’Brien L. & Powers K. (1986) The distribution of the humpback whale, 

Megaptera novaeangliae, on Georges Bank and in the Gulf of Maine in relation to densities of 

the sand eel, Ammodytes americanus. Fisheries Bulletin 84, 271-277. 

Payne P. M., Wiley D. N., Young S. B., Pittman S., P.J. Clapham & Jossi J. W. (1990) Recent 

fluctuations in the abundance of baleen whales in the southern Gulf of Maine in relation to 

changes in selected prey. Fishery Bulletin 88, 687-696. 

Payne S. A., Johnson B. A. & Otto R. S. (1999) Proximate composition of some north-eastern 

Pacific forage fish species. Fisheries Oceanography 8, 159-177. 

PEARL. (2007) Facilitating the transfer of information on Penobscot River science and research. 

Pearson J. P. & Verts J. P. (1970) Abundance and distribution of harbor seals and northern sea 

lions in Oregon. Murrelet 51, 1-5. 

Peckham S. H., Maldonado-Diaz D., Koch V., Mancini A., Gaos A., Tinker M. T. & Nichols W. 

J. (2008) High mortality of loggerhead turtles due to bycatch, human consumption and 

strandings at Baja California Sur, Mexico, 2003 to 2007. Endangered Species Research 5, 171-

183. 

Pelletier D., Roos D. & Ciccione S. (2003) Oceanic survival and movements of wild and captivereared immature green turtles (Chelonia mydas) in the Indian Ocean. Aquatic Living Resources

16, 35-41. 

Pendleton D. E., Pershing r. J., Brown M. W., Mayo C. A., Kenney R. D., Record N. R. & Cole. 

T. V. N. (2009) Regional-scale mean copepod concentration indicates relative abundance of 

North Atlantic right whales. Marine Ecology Progress Series 378, 211-225. 

Pennington J. T., Mahoney K. L., Kuwahara V. S., Kolber D. D., Calienes R. & Chavez F. P. 

(2006) Primary production in the eastern tropical Pacific: A review. Progress in Oceanography

69, 285–317. 

Perez M. A. (1990) Review of marine mammal and prey population information for Bering Sea 

ecosystem studies. NOAA Technical Memorandum NMFS-NWC-186. 

Perkins J. & Beamish D. (1979) Net entanglements of baleen whales in the inshore fishery of 

Newfoundland. Journal of the Fisheries Research Board of Canada 36, 521-528. 

Perrin W. F. (1982) Report of the sub-committee on small cetaceans. Report of the International 

Whaling Commission 32, 113-125. 

Perrin W. F. & Reilly S. B. (1984) Reproductive parameters of dolphins and small whales of the 

family Delphinidae. Pages 97-134 in: Perrin, W.G., R.L. Brownell, Jr., and D.P. DeMaster, 
477

editors. Reproduction in whales, dolphins, and porpoises. International Whaling Commission 

(Special Issue 6), Cambridge, U.K. 

Perry S. L., DeMaster D. P. & Silber G. K. (1999) The Great Whales: History and Status of Six 

Species Listed as Endangered Under the U.S. Endangered Species Act of 1973. Marine Fisheries 

Review 61, 1-74. 

Perugini M., Giammarino A., Olivieri V., Guccione S., Lai O. R. & Amorena M. (2006) 

Polychlorinated biphenyls and organochlorine pesticide levels in tissues of Caretta caretta from 

the Adriatic Sea. Diseases of Aquatic Organisms 71, 155-161. 

Pess G. R., Montgomery D. R., Beechie T. J. & Holsinger L. (2003) Anthropogenic alterations to 

the biogeography of Puget Sound salmon. Pages 129-154 in: Montgomery, D.R., S. Bolton, D.B. 

Booth, and L. Wall, editors. Restoration of Puget Sound rivers. University of Washington Press, 

Seattle, Washington. 

Petersen S. L., Honig M. B., Ryan P. G., Nel R. & Underhill L. G. (2009) Turtle bycatch in the 

pelagic longline fishery off southern Africa. African Journal of Marine Science 31, 87-96. 

Peterson D. L., Bain M. B. & Haley N. (2000) Evidence of Declining Recruitment of Atlantic 

Sturgeon in the Hudson River. North American Journal of Fisheries Management 20, 231-238. 

Peterson R. H. (1978) Physical characteristics of Atlantic salmon spawning gravel in some New 

Brunswick, Canada streams. Canada Fisheries and Marine Service, Technical Report Number

785, 1-28. 

Petras E. (2003) A review of marine mammal deterrents and their possible applications to limit 

killer whale (Orcinus orca) predation on Steller sea lions (Eumetopias jubatus). In: AFSC 

Processed Report 2003-02 Alaska Fisheries Science Center, National Oceanic and Atmospheric 

Administration, Seattle, Washington. 

PFMC. (2000) Amendment 14 to the Pacific coast salmon plan (1997). Incorporating the 

regulatory impact review/initial regulatory flexibility analysis and final supplemental 

environmental impact statement. p. 420. Pacific Fishery Management Council, Portland, 

Oregon. 

PFMC. (2005) Review of 2004 ocean salmon fisheries. p. 320. Pacific Fishery Management 

Council, Portland, Oregon. 

Phelps H. L. (1994) The Asiatic clam (Corbicula fluminea) invasion and system-level ecological 

change in the Potomac River Estuary near Washington, D. C. Estuaries 17, 614-621. 

Philippart C. J. M., Anadón R., Danovaro R., Dippner J. W., Drinkwater K. F., Hawkins S. J., 

Oguz T., O'Sullivan G. & Reid P. C. (2011) Impacts of climate change on European marine 

ecosystems: Observations, expectations and indicators☆. Journal of Experimental Marine 

Biology and Ecology. 

Phillips A. J., Ralston S., Brodeur R. D., Auth T. D., Emmett R. L., Johnson C. & Wespestad V. 

G. (2007) Recent pre-recruit Pacific hake (Merluccius productus) occurrences in the northern 

California current suggest a northward expansion of their spawning area. CalCOFI Reports 48, 

215-229. 

Phillips C. D., Bickham J. W., Patton J. C. & Gelatt. T. S. (2009) Systematics of Steller sea lions 

(Eumetopias jubatus): Subspecies recognition based on concordance of genetics and 

morphometrics. Occasional Papers of the Museum of Texas Tech University 283, 1-15. 

Phillips J. B. (1960) Canary rockfish. p. 39. California Department of Fish and Game. 

Piatak N. M., Seal II R. R., Sanzolone R. F., Lamothe P. J. & Brown Z. A. (2006) Preliminary 

Results of Sequential Extraction Experiments for Selenium on Mine Waste and Stream 

Sediments from Vermont, Maine, and New Zealand p. 21. U.S. Geological Survey 
478

Picanco C., Carvalho I. & Brito. C. (2009) Occurrence and distribution of cetaceans in Sao Tome 

and Príncipe tropical archipelago and their relation to environmental variables. Journal of the 

Marine Biological Association of the United Kingdom 89, 1071-1076. 

Pike D. A. (2009) Do Green Turtles Modify Their Nesting Seasons in Response to 

Environmental Temperatures? Chelonian Conservation and Biology 8, 43-47. 

Pike D. G., Gunnlaugsson T., Vikingsson G. A., Desportes G. & Mikkelsen B. (2010) Estimates 

of the abundance of humpback whales (Megaptera novaengliae) from the T-NASS Icelandic and 

Faroese ship surveys conducted in 2007. p. 15. IWC Scientific Committee, Agadir, Morocco. 

Pike D. G., Paxton C. G. M., Gunnlaugsson T. & Víkingsson G. A. (2009) Trends in the 

distribution and abundance of cetaceans from aerial surveys in Icelandic coastal waters, 1986-

2001. Nammco Scientific Publications 7, 117-142. 

Pike G. C. & Macaskie. I. B. (1969) Marine mammals of British Columbia. Bulletin of the 

Fisheries Research Board of Canada 171, 1-54. 

Pike G. C. & Maxwell B. E. (1958) The abundance and distribution of the northern sea lion 

(Eumetopias jubata) on the coast of British Columbia. Journal of the Fisheries Research Board 

of Canada 15, 5-17. 

Pilcher N. J. (2000) The green turtle, Chelonia mydas, in the Saudi Arabian Gulf. Chelonian 

Conservation and Biology 3, 730-734. 

Pimentel D., Zuniga R. & Morrison D. (2004) Update on the environmental and economic costs 

associated with alien-invasive species in the United States. Ecological Economics. 

Piner K. R., Wallace J. R., Hamel O. S. & Mikus R. (2006) Evaluation of ageing accuracy of 

bocaccio (Sebastes paucispinis) rockfish using bomb radiocarbon. Fisheries Research 77, 200-

206. 

Pinto De Sa Alves L. C., Andriolo A., Zerbini A. N., Pizzorno J. L. A. & Clapham. P. J. (2009) 

Record of feeding by humpback whales (Megaptera novaeangliae) in tropical waters off Brazil. 

Marine Mammal Science 25, 416-419. 

Pitcher K. W., Burkanov V. N. & Calkin D. G. (2001) Spatial and temporal variation in the 

timing of births of Steller sea lions. LeBoeuf, E.G. Mamaev, R.L. Merrick, and G.W. Pendleton

82, 1047-1053. 

Pitcher K. W. & Calkins D. G. (1981) Reproductive biology of Steller sea lions in the Gulf of 

Alaska. Journal of Mammalogy 62, 599-605. 

Pitcher K. W., Calkins D. G. & Pendleton G. W. (1998) Reproductive performance of female 

Steller sea lions: an energetics-based reproductive strategy? Canadian Journal of Zoology 76, 

2075-2083. 

Pitcher K. W. & Fay F. H. (1982) Feeding by Steller sea lions on harbor seals. Murrelet 63, 70-

71. 

Pitcher T. J. (1986) Functions of shoaling in teleosts. Pages 294-337 in: Fisher, T.J., editor. The 

behavior of teleost fishes. Johns Hopkins University Press, Baltimore, Maryland. 

Pitman R. L. & Dutton P. H. (2004) Killer whale predation on a leatherback turtle in the 

Northeast Pacific. Northwest Science 58, 497-498. 

Pivorunas A. (1979) The feeding mechanisms of baleen whales. American Scientist 67, 432-440. 

Plotkin P. (2003) Adult migrations and habitat use. In: Biology of sea turtles (eds P. L. Lutz, J. 

A. Musick and J. Wyneken) pp. 225-241. CRC Press, Boca Raton, Florida. 

Plotkin P. T., (Ed). (1995) National Marine Fisheries Service and the U.S. Fish and Wildlife 

Service Status Reviews for Sea Turtles Listed under the Endangered Species Act of 1973. 

National Marine Fisheries Service, Silver Spring, Maryland. 
479

Podreka S., Georges A., Maher B. & Limpus C. J. (1998) The environmental contaminant DDE 

fails to influence the outcome of sexual differentiation in the marine turtle Chelonia mydas. 

Environmental Health Perspectives 106, 185-188. 

Poloczanska E. S., Limpus C. J. & Hays G. C. (2009) Vulnerability of marine turtles in climate 

change. In: Advances in Marine Biology pp. 151-211. Academic Press, New York. 

Polovina J., Uchida I., Balazs G., Howell E. A., Parker D. & Dutton P. (2006) The Kuroshio 

Extension Bifurcation Region: a pelagic hotspot for juvenile loggerhead sea turtles. Deep-Sea 

Research II 53, 362-339. 

Polovina J. J., Balazs G. H., Howell E. A., Parker D. M., Seki M. P. & Dutton P. H. (2004) 

Forage and migration habitat of loggerhead (Caretta caretta) and olive ridley (Lepidochelys 

olivacea) sea turtles in the central North Pacific Ocean. Fisheries Oceanography 13, 36-51. 

Polovina J. J., Kobayashi D. R., Parker D. M., Seki M. P. & Balazs G. H. (2000) Turtles on the 

edge: Movement of loggerhead turtles (Caretta caretta) along oceanic fronts, spanning longline 

fishing grounds in the central North Pacific, 1997-1998. Fisheries Oceanography 9, 71-82. 

Pomilla C. & Rosenbaum H. C. (2005) Against the current: an inter-oceanic whale migration 

event. Biology Letters 1, 476-479. 

Pool S. A., Brodeur R. D., Goodman N. L. & Daly E. A. (2008) Abundance, distribution, and 

feeding patterns of juvenile coho salmon (Oncorhynchus kisutch) in the Juan de Fuca Eddy 

Estuarine, Coastal and Shelf Science 80, 85-94. 

Porter B. (1997) Winter ecology of Steller sea lions (Eumetopias jubatus) in Alaska. Master’s 

Thesis. University of British Columbia. 

Porter J. W. (1976) Autotrophy, heterotrophy, and resource partitioning in Caribbean reef corals. 

American Naturalist 110, 731-742. 

Posey M. H. (1988) Community changes associated with the spread of an introduced seagrass, 

Zostera japonica. Ecology 69, 974-983. 

Posluszny U. & Tomlinson P. B. (1990) Shoot organization in the seagrass Halophila 

(Hydrocharitaceae). Canadian Journal of Botany 69, 1600-1615. 

Pothaven S. A., Nalepa T. F., Schneeberger P. J. & Brandt S. B. (2001) Changes in diet and body 

condition of lake whitefish in southern Lake Michigan associated with changes in benthos. North 

American Journal of Fisheries Management 21, 876–883. 

Poulakis G. R. & Seitz J. C. (2004a) Recent occurrence of the smalltooth sawfish, Pristis 

pectinata (Elasmobranchiomorphi: Pristidae), in Florida Bay and the Florida Keys, with 

comments on sawfish ecology. Florida Scientist 67:227-35. 

Poulakis G. R. & Seitz J. C. (2004b) Recent occurrence of the smalltooth sawfish, Pristis 

pectinata (Elasmobranchiomorphi: Pristidae), in Florida Bay and the Florida Keys, with 

comments on sawfish ecology. Florida Scientist 67, 27-35. 

Power M. E. (1990) Effects of fish in river food webs. Science 250, 811-814. 

Power M. E. (1992) Top-down and bottom-up forces in food webs: do plants have primacy. 

Ecology 73, 733-746. 

PRFR. (1998) Eulachon: A significant fish for First Nations communities. Forest Sciences, 

Prince Rupert Forest Region, Smithers, British Columbia. 

Prieto A., Prietol A., Moncadal F., Nodarse G., Pugal R., Leon M. E. d., Diaz-Fernandez R., 

Espinoru G., Castillo D., Hernandez M., Peregrin E., Arazoza M. d., Salabarria D., Morales E., 

Webbs G., Manolis C. & Gomez R. (2001) Biological and ecological aspects of the hawksbill 

turtle population in Cuban waters. In: Report from the Republic of Cuba. First CITES wider 

Caribbean hawksbill turtle dialogue meeting, Mexico City. 
480

Pritchard P. C. H. (1971) The leatherback or leathery turtle, Dermochelys coriacea. IUCN 

Monograph 1, 1-39. 

Pritchard P. C. H. (1982) Nesting of the leatherback turtle, Dermochelys coriacea in Pacific 

Mexico, with a new estimate of the world population status. Copeia 1982, 741-747. 

Pritchard P. C. H. (1997) Evolution, Phylogeny, and Current Status. In: The Biology of Sea 

Turtles (eds P. L. Lutz and J. A. Musick) pp. 1-28. CRC Press, Inc., Boca Raton, FL. 

PSAT. (2004) State of the Sound 2004. State of Washington, Office of the Governor, Puget 

Sound Action Team, Olympia, Washington. 

PSAT. (2007a) State of the Sound 2007. Puget Sound Action Team, Office of the Governor, 

State of Washington. 

Psat. (2007b) State of the Sound 2007. State of Washington, Office of the Governor, Publication 

Number PSAT 07-01, Olympia, Washington. Available: 

http://www.psat.wa.gov/Publications/state_sound07/2007_stateofthesound.pdf (February 2008). 

Puget Sound Action Team. (2005) State of the sound 2004. Puget Sound Action Team, Olympia, 

Washington. 

Puget Sound Water Quality Action Team. (2002) Puget Sound’s health 2002. Puget Sound Water 

Quality Action Team, Olympia, Washington. 

Pugh R. S. & Becker P. R. (2001a) Sea turtle contaminants: A review with annotated 

bibliography. U.S. Department of Commerce, National Institute of Standards and Technology, 

Chemical Science and Technology Laboratory, Charleston, South Carolina. 

Pugh R. S. & Becker P. R. (2001b) Sea turtle contaminants: A review with annotated 

bibliography. U.S. Department of Commerce, National Institute of Standards and Technology, 

Chemical Science and Technology Laboratory, Charleston, South Carolina. 

Pughiuc D. (2010) Invasive species: Ballast water battles. Seaways. 

Pullin A. S. & Stewart G. B. (2006) Guidelines for systematic review in conservation and 

environmental management. Conservation Biology 20, 1647-1656. 

Purvis A., Gittleman J. L., Cowlishaw G. & Mace G. M. (2000) Predicting extinction risk in 

declining species. Proceedings of the Royal Society B-Biological Sciences 267, 1947-1952. 

Putrawidjaja M. (2000) Marine Turtles in Irian Jaya, Indonesia. Marine Turtle Newsletter 90, 8-

10. 

Quakenbush L. T. & Nelson. M. A. (2010) Cook Inlet beluga whale diet using stable isotope 

analysis. In: Cook Inlet Belgua Whale Science Conference pp. 20-21, Anchorage Downtown 

Marriott, Anchorage, Alaska. 

Quattro J. M. (2002) Genetic issues in aquatic species management: the shortnose sturgeon 

(Acipenser brevirostrum) in the southeastern United States. Conservation Genetics 3, 155-166. 

Quigley T. M. S. J. A. & Graham R. T. (1997) Assessment of ecosystem components in the 

interior Columbia River Basin and portions of the Klamath and Great Basins: An introduction. 

Pages 1-32 in: T.M. Quigley and S.J. Arbelbide, eds. U.S. Department of Agriculture, Forest 

Service, Pacific Northwest Research Station. General Technical Report PNW-GTR, 405. 

Quinn T. P. (2005) The behavior and ecology of Pacific salmon and trout. University of 

Washington Press, Seattle, Washington. 

Raaymakers S. (2003) The GEF/UNDP/IMO global ballast water management programme 

integrating science, shipping and society to save our seas. Proceedings of the Institute of Marine 

Engineering, Science and Technology Part B: Journal of Design and Operations, 2-10. 

Raaymakers S. & Hilliard R. (2002) Harmful aquatic organisms in ships' ballast water - Ballast 

water risk assessment. In: Alien marine organisms introduced by ships in the Mediterranean and 
481

Black seas (ed F. Briand) pp. 103-110. CIESM Workshop Monographs [CIESM Workshop 

Monogr.]. 2002., Istanbul, Turkey. 

Rabalais N. N., Turner R. E. & Scavia D. (2002) Beyond science into policy: Gulf of Mexico 

hypoxia and the Mississippi River. Bioscience 52, 129-142. 

Rabalais S. C. & Rabalais N. N. (1980) The Occurrence of Sea Turtles on the South Texas Coast. 

Contributions in Marine Science Vol. 23, 123-129. 

Råberg S., Berger-Jönsson R., Björn A., Granéli E. & Kautsky L. (2005) Effects of Pilayella 

littoralis on Fucus vesiculosus recruitment: implications for community composition. Marine 

Ecology-Progress Series 289, 131–139. 

Radtke L. D. (1966) Distribution of smelt, juvenile sturgeon, and starry flounder in the 

Sacramento-San Joaquin Delta with observations on food of sturgeon. Pages 115-119 in: J.L. 

Turner and D.W. Kelly, eds. Ecological studies of the Sacramento-San Joaquin Estuary. Part II; 

California Department of Fish and Game, Fish Bulletin, 136. 

RAE/ERF. (1999) Principles of estuarine habitat restoration: Working together to restore 

America's estuaries. Report on the RAE-ERF partnership - year one - September 1999. Restore 

America's Estuaries/Estuarine Research Foundation, Arlington, Virginia/Port Republic, 

Maryland. 

Ragen T. J. (1993) Status of the Hawaiian monk seal in 1992. 

Ragen T. J. & Lavigne D. M. (1999) The Hawaiian monk seal: Biology of an endangered 

species. Pages 224-245 in: Twiss, Jr., J.R. and R.R. Reeves, editors. Conservation and 

management of marine mammals. Smithsonian Institution Press, Washington, D.C. 

Ragueneau O., Chauvaud L., Moriceau B., Leynaert A., Thouzeau G., Donval A., Le Loc'h F. & 

Jean F. (2005) Biodeposition by an invasive suspension feeder impacts the biogeochemical cycle 

of Si in a coastal ecosystem (Bay of Brest, France). Biogeochemistry 75, 19–41. 

Rahel F. J. (2000) Homogenization of Fish Faunas Across the United States. Science 288, 854-

856. 

Ralston S. & Ianelli J. N. (1998) When lengths are better than ages: The complex case of 

bocaccio. University of Alaska, Fairbanks, Alaska. 

Ramp C., Hagen W., Palsboll P., Berube M. & Sears. R. (2010) Age-related multi-year 

associations in female humpback whales (Megaptera novaeangliae). Behavioral Ecology and 

Sociobiology 64, 1563-1576. 

Randall R. G. (1982) Emergence, population densities, and growth of salmon and trout fry in two 

New Brunswick streams. Canadian Journal of Zoology 60, 2239-2244. 

Rankin-Baransky K. (1997) Origin of loggerhead turtles (Caretta caretta) in the western North 

Atlantic Ocean as determined by mtDNA analysis. Masters Thesis submitted to Drexel 

University, June 1997. 1949p. 

Ranwell D. S. (1964) Spartina salt marshes in Southern England: II. Rate and seasonal pattern of 

sediment accretion. Journal of Ecology 52, 79-94. 

Ranwell D. S. (1967) World resources of Spartina townsendii (sensu lato) and economic use of 

Spartina marshland. Journal of Applied Ecology 4, 239-256. 

Rasmussen K., Palacios D. M., Calambokidis J., Saborío M. T., Rosa L. D., Secchi E. R., Steiger 

G. H., Allen J. M. & Stone G. S. (2007) Southern Hemisphere humpback whales wintering off 

Central America: insights from water temperature into the longest mammalian migration. 

Biology Letters 3, 302-305. 
482

Raum-Suryan K. L., Jemison L. A. & Pitcher K. W. (2009a) Entanglement of Steller sea lions 

(Eumetopias jubatus) in marine debris: Identifying causes and finding solutions. Marine 

Pollution Bulletin. 

Raum-Suryan K. L., Jemison L. A. & Pitcher. K. W. (2009b) Entanglement of Steller sea lions 

(Eumetopias jubatus) in marine debris: Identifying causes and finding solutions. Marine 

Pollution Bulletin 58, 1487-1495. 

Raum-Suryan K. L., Pitcher K. W., Calkins D. G., Sease J. L. & Loughlin T. R. (2002) 

Dispersal, rookery fidelity, and metapopulation structure of Steller sea lions (Eumetopias 

jubatus) in an increasing and a decreasing population in Alaska. Marine Mammal Science 18, 

746-764. 

Raum-Suryan K. L., Rehberg M. J., Pendleton G. W., Pitcher K. W. & Gelatt T. S. (2004) 

Development of dispersal, movement patterns, and haul-out use by pup and juvenile Steller sea 

lions (Eumetopias jubatus) in Alaska. Marine Mammal Science 20, 823-850. 

Rawson K., Sands N. J., Currens K. P., Graeber W. H., Ruckelshaus M. H., Fuerstenberg R. R. 

& Scott J., B. (2009) Viability criteria for the Lake Ozette sockeye salmon evolutionarily 

significant unit. In: NOAA Technical Memorandum p. 38. Department of Commerce, Seattle, 

Washington. 

Ray G. C., Wartzok D. & Taylor. G. (1984) Productivity and behavior of bowheads, Balaena 

mysticetus,and white whales, Delphinapterus leucas, as determined from remote sensing. Report 

of the International Whaling Commission Special Issue 6, 199-209. Reproduction in Whales, 

Dolphins and Porpoises. Proceedings of the Conference Cetacean Reproduction Estimating 

Parameters For stock Assessment and Management. 

RBCC. (2006) Watershed health factors assessment: Rogue River basin. Rogue Basin 

Coordinating Council, Talent, Oregon. 

Read A. J. (2010) Conservation biology. In: Marine Mammal Ecology and Conservation: A 

Handbook of Techniques (eds I. L. Boyd, W. D. Bowen and S. J. Iverson) pp. 340-359. Oxford 

University Press. 

Reddin D. G. (1985) Atlantic salmon (Salmo salar) on and east of the Grand Bank. Journal of 

Northwest Atlantic Fisheries Science 6, 157-164. 

Reed D. H., O’Grady J. J., Ballou J. D. & Frankham R. (2003) The frequency and severity of 

catastrophic die-offs in vertebrates. Animal Conservation 6, 109-114. 

Rees A. F. & Margaritoulis D. (2004) Beach temperatures, incubation durations, and estimated 

hatchling sex ratio for loggerhead sea turtle nests in southern Kyparissia Bay, Greece. British 

Chelonia Group Testudo 6, 23-36. 

Rees A. F., Saad A. & Jony M. (2005) Marine turtle nesting survey, Syria 2004: discovery of a 

"major" green turtle nesting area. Page 38 in Book of Abstracts of the Second Mediterranean 

Conference on Marine Turtles. Antalya, Turkey, 4-7 May 2005. 

Reeves R. R. (1992) Whale responses to anthropogenic sounds: a literature review. New Zealand 

Department of Conservation, Wellington, New Zealand. 

Reeves R. R. (2001) Overview of catch history, historic abundance, and distribution of right 

whales in the western North Atlantic, and in Cintra Bay, West Africa. Journal of Cetacean 

Research and Management 2, 187-192. 

Reeves R. R., Clapham P. J. & Wetmore S. E. (2002) Humpback whale (Megaptera 

novaeangliae) occurrence near the Cape Verde Islands, based on American 19th century whaling 

records. Journal of Cetacean Research and Management 4, 235-253. 
483

Reeves R. R., Khan J. A., Olsen R. R., Swartz S. L. & Smith T. D. (2001a) History of whaling in 

Trinidad and Tobago. Journal of Cetacean Research and Management 3, 45-54. 

Reeves R. R., Leatherwood S. & Papastavrou. V. (1991) Possible stock affinities of humpback 

whales in the northern Indian Ocean. In: Cetaceans and Cetacean Research in the Indian Ocean 

Sanctuary. S. Leatherwood and G. P. Donovan (eds.). p.259-269. United Nations Environment 

Programme, Marine Mammal Technical Report No. 3. 

Reeves R. R., Mead J. & Katona S. (1978) The right whale, Eubalaena glacialis, in the western 

North Atlantic. Report of the International Whaling Commission 28, 303-312. 

Reeves R. R., Smith T. D. & Josephson E. (2007) Near annihilation of a species: right whaling in 

the North Atlantic. In: The urban whale: North Atlantic right whales at the crossroads (eds S. D. 

Kraus and R. M. Rolland) pp. 39-74. Harvard University Press, Cambridge, Massachussetts. 

Reeves R. R., Swartz S. L., Wetmore S. E. & Clapham P. J. (2001b) Historical occurrence and 

distribution of humpback whales in the eastern and southern Caribbean Sea, based on data from 

American whaling logbooks. Journal of Cetacean Research and Management 3, 117-129. 

Reich K. J., Bjorndal K. A., Frick M. G., Witherington B., Johnson C. & Bolten A. B. (2010) 

Polymodal foraging in adult female loggerheads (Caretta caretta). Marine Biology 157, 113-121. 

Reid K. & Croxall J. (2001) Environmental response of upper trophic-level predators reveals a 

system change in an Antarctic marine ecosystem. Proceedings of the Royal Society London 

Series B 268, 377–384. 

Reid K. A., Margaritoulis D. & Speakman J. R. (2009) Incubation temperature and energy 

expenditure during development in loggerhead sea turtle embryos. Journal of Experimental 

Marine Biology and Ecology 378, 62-68. 

Reid S. D., Dockray J. J., Linton T. K., McDonald D. G. & Wood C. M. (1997) Effects of 

chronic environmental acidification and a summer global warming scenario: protein synthesis in 

juvenile rainbow trout (Oncorhynchus mykiss) Canadian Journal of Fisheries and Aquatic 

Science 54, 2014-2024. 

Reilly C. A., Echeverria T. W. & Ralston S. (1992) Interannual variation and overlap in the diets 

of pelagic juvenile rockfish (genus Sebastes) off central California. Fishery Bulletin 90, 505-515. 

Reina R. D., Spotila J. R., Paladino F. V. & Dunham A. E. (2008) Changed reproductive 

schedule of eastern Pacific leatherback turtles Dermochelys coriacea following the 1997–98 El 

Niño to La Niña transition. Endangered Species Research. 

Reiner F., Dos Santos M. E. & Wenzel F. W. (1996) Cetaceans of the Cape Verde archipelago. 

Marine Mammal Science 12, 434-443. 

Reisenbichler R. R. (1997) Genetic factors contributing to declines of anadromous salmonids in 

the Pacific Northwest. In: Pacific salmon and their ecosystems: status and future options (eds D. 

J. Stouder, P. A. Bisson and R. J. Naiman) pp. 223-244. Chapman and Hall, New York. 

Renaud M. L. (1995a) Movements and submergence patterns of Kemp's ridley turtles 

(Lepidochelys kempii). Journal of Herpetology 29, 370-374. 

Renaud M. L. (1995b) Movements and submergence patterns of Kemp's ridley turtles 

(Lepidochelys kempii). Journal of Herpetology 29, 370-374. 

Renaud M. L., Carpenter J. A., Williams J. A. & A.M. Landry, Jr. (1996) Kemp's ridley sea 

turtle (Lepidochelys kempii) tracked by satellite telemetry from Louisiana to nesting beach at 

Rancho Nuevo, Tamaulipas, Mexico. Chelonian Conservation and Biology 2, 108-109. 

Renault T., Stokes N. A., Chollet B., Cochennec N., Berthe F., Gérard A. & Burreson E. M. 

(2000) Haplosporidiosis in the Pacific oyster Crassostrea gigas from the French Atlantic coast. 

Diseases of Aquatic Organisms 42, 207–214. 
484

Rennie M. D., Strecker A. L. & Palmer M. E. (2011) Bythotrephes invasion elevates trophic 

position of zooplankton and fish: implications for contaminant biomagnification. Biological 

Invasions 13, 2621-2634. 

Resendiz A., Resendiz B., Nichols W. J., Seminoff J. A. & Kamezaki N. (1998) First confirmed 

eastwest transpacific movement of a loggerhead sea turtle, Caretta caretta, released in Baja 

California, Mexico. Pacific Science 52, 151-153. 

Rester J. & Condrey R. (1996) The occurrence of the hawksbill turtle, Eretmochelys imbricata, 

along the Louisiana coast. Gulf of Mexico Science 1996, 112-114. 

Rexstad E. A. & Pikitch E. K. (1986) Stomach contents and food consumption estimates of 

Pacific hake, Merluccius productus. Fishery Bulletin 84. 

Reyff J. A. (2003) Underwater sound levels associated with constnlction of the Benicia-Martinez 

Bridge. p. 26. lllingworth & Rodkin, Inc. 

Reynolds F. L., Mills T., Benthin R. & Low A. (1993) Central Valley anadromous fisheries and 

associated riparian and wetlands areas protection and restoration action plan draft. 

Ricciardi A. (2001) Facilitative interactions among aquatic invaders: is an "invasional 

meltdown" occurring in the Great Lakes? Canadian Journal of Fisheries and Aquatic Sciences

58, 2513-2525. 

Rice D. W. (1960) Population dynamics of the Hawaiian monk seal. Journal of Mammalogy 41, 

376-385. 

Rice D. W. (1974) Whales and whale research in the eastern North Pacific. Pages 170-195 in: 

Schevill, W.E. editor. The whale problem, a status report. Harvard University Press, Cambridge, 

Massachusetts. 

Rice D. W. (1978) The humpback whale in the North Pacific: distribution, exploitation, and 

numbers. In K. S. Norris and R. R. Reeves (Editors), Report on a Workshop on Problems Related 

to Humpback Whales (Megaptera novaeangliae) in Hawaii. Contr. Rep. to U.S. Mar. Mammal 

Comm., NTIS PB-280-794. pp. 29–44. 

Rice D. W. (1998a) Marine Mammals of the World. Systematics and Distribution. Special 

Publication Number 4. The Society for Marine Mammalogy, Lawrence, Kansas. 

Rice D. W. (1998b) Marine mammals of the world.: Systematics and distribution. Special 

Publication Number 4. The Society for Marine Mammalogy, Lawrence, Kansas. 

Rich W. H. (1942) The salmon runs of the Columbia River in 1938. Fishery Bulletin 50, 103-

147. 

Richards D. V. (2000) The Status of rocky intertidal communities in Channel Islands National 

Park. California Islands Symposium, Santa Barbara. 

Richards D. V. & Davis G. E. (1993) Early warnings of modern population collapse in black 

abalone Haliotis cracherodii, Leach, 1814 at the California Channel Islands. Journal of Shellfish 

Research 12, 189-194. 

Richards L. J. (1986) Depth and habitat distributions of three species of rockfish (Sebastes) in 

British Columbia: observations from the submersible PISCES IV. Environmental Biology of 

Fishes 17, 13-21. 

Richards L. J., Paul J., Cass A. J., Fitzpatrick L., Broek R. v. d. & Lauridsen C. (1985) SCUBA 

survey of rockfish assemblages in the Strait of Georgia, July to October 1984. Department of 

Fisheries and Oceans, Fisheries Research Branch, Pacific Biological Station. 

Richardson J. I., Bell R. & Richardson T. H. (1999) Population ecology and demographic 

implications drawn from an 11-year study of nesting hawksbill turtles, Eretmochelys imbricata, 
485

at Jumby Bay, Long Island, Antigua, West Indies. Chelonian Conservation and Biology 3, 244-

250. 

Richardson J. S., Lissimore T. J., Healey M. C. & Northcote T. G. (2000) Fish communities of 

the lower Fraser River (Canada) and a 21-year contrast. Environmental Biology of Fishes 59, 

125-140. 

Richardson S. L. & Laroche W. A. (1979) Development and occurrence of larvae and juveniles 

of the rockfishes Sebastes crameri, Sebastes pinniger, and Sebastes helvomaculatus (Family 

Scorpaenidae) off Oregon. Fishery Bulletin 77, 1-46. 

Richardson T. H., Richardson J. I., Ruckdeshel C. & Dix M. W. (1978) Remigration patterns of 

loggerhead sea turtles (Caretta caretta) nesting on Little Cumberland and Cumberland Islands, 

Georgia. Florida Marine Research Publications 33, 39-44. 

Richardson W. J., C. R. Greene Jr., Malme C. I. & Thomson D. H. (1995a) Marine mammals 

and noise. Academic Press, San Diego, California. 

Richardson W. J., Greene Jr. C. R., Malme C. I. & Thomson D. H. (1995b) Marine mammals 

and noise. Academic Press; San Diego, California. 

Richmond R. H. & Hunter C. L. (1990) Reproduction and recruitment of corals: comparisons 

among the Caribbean, the tropical Pacific, and the Red Sea. Marine Ecology Progress Series 60, 

185-203. 

Richter B. D., Braun D. P., Mendelson M. A. & Master L. L. (1997) Threats to imperiled 

freshwater fauna. Conservation Biology 11, 1081-1093. 

Richter C. F., Dawson S. M. & Slooten. E. (2003) Sperm whale watching off Kaikoura, New 

Zealand: Effects of current activities on surfacing and vocalisation patterns. Department of 

Conservation, Wellington, New Zealand. Science For Conservation 219. 78p. 

Ricker W. E., Manzer D. F. & Neave E. A. (1954) The Fraser River eulachon fishery, 1941-

1953. In: Fisheries Research Board of Canada Bulletin. 

Rien T. A. & Beamesderfer R. C. (1994) Accuracy and precision of white sturgeon age estimates 

from pectoral fin rays. Transactions of the American Fisheries Society 123, 255-265. 

Rien T. A., Burner L. C., Farr R. A., Howell M. D. & North J. A. (2001a) Green sturgeon 

population characteristics in Oregon. Annual Progress Report. Sportfish Restoration Project F178-R. 

Rien T. A., Burner L. C., Farr R. A., Howell M. D. & North J. A. (2001b) Green sturgeon 

population characteristics in Oregon. 

Ritson-Williams R., Paul V. J., Arnold S. N. & Steneck R. S. (2009) Larval settlement 

preferences and post-settlement survival of the threatened Caribbean corals Acropora palmata 

and A. cervicornis. Coral Reefs 29, 71-81. 

Rivalan P., Dutton P. H., Baudry E., Roden S. E. & Girondot M. (2006) Demographic scenario 

inferred from genetic data in leatherback turtles nesting in French Guiana and Suriname. 

Biological Conservation 130, 1-9. 

Rizzo L. Y. & Schulte D. (2009) A review of humpback whales' migration patterns worldwide 

and their consequences to gene flow. Journal of the Marine Biological Association of the United 

Kingdom 89, 995-1002. 

Roark A. M., Bjorndal K. A. & Bolten A. B. (2009) Compensatory responses to food restriction 

in juvenile green turtles (Chelonia mydas). Ecology 90, 2524-2534. 

Robeck T. R., Monfort S. L., Calle P. P., Dunn J. L., Jensen E., Boehm J. R., Young S. & Clark. 

S. T. (2005) Reproduction, growth and development in captive beluga (Delphinapterus leucas). 

Zoo Biology 24, 29-49. 
486

Roberts D. A. (1979) Food habits as an ecological partitioning mechanism in the nearshore 

rockfishes (Sebastes) of Carmel Bay, California. San Francisco State University/Moss Landing 

Marine Laboratories 

Roberts H. H., Rouse J. R., Walker N. D. & Hudson J. H. (1982) Coldwater stress in Florida Bay 

and Northern Bahamas: a product of winter cold-air outbreaks. Journal of Sedimentary 

Petroleum 52, 145-155. 

Roberts L. (1990) Zebra mussel invasion threatens U.S. waters. Science 249, 1370-1372. 

Robinson R. A., Crick H. Q. P., Learmonth J. A., Maclean I. M. D., Thomas C. D., Bairlein F., 

Forchhammer M. C., Francis C. M., Gill J. A., Godley B. J., Harwood J., Hays G. C., Huntley 

B., Hutson A. M., Pierce G. J., Rehfisch M. M., Sims D. W., Santos M. B., Sparks T. H., Stroud 

D. A. & Visser M. E. (2008) Travelling through a warming world: climate change and migratory 

species. Endangered Species Research. 

Rodriguez-Lanetty M., Harii S. & Hoegh-Guldberg O. (2009) Early molecular responses of coral 

larvae to hyperthermal stress. Molecular Ecology 18, 5101-5114. 

Roemmich D. & McGowan J. (1995) Climate warming and the decline of zooplankton in the 

California Current. Science 267, 1324-1326. 

Rogers C., Gladfelter W., Hubbard D., Gladfelter E., Bythell J., Dunsmore R., Loomis C., 

Devine B., Hillis-Starr Z. & Phillips B. (2002) Acropora in the U.S. Virgin Islands: A wake or an 

awakening? In: Caribbean Acropora workshop: Potential application of the U.S. Endangered 

Species Act as a conservation strategy (ed A. W. Bruckner) pp. 99-122. 

Rogers C. S., Fitz H. C., Gilnack M., Beets J. & Hardin J. (1984) Scleractinian coral recruitment 

patterns at Salt River submarine canyon, St. Croix, U.S. Virgin Islands. Coral Reefs 3, 69-76. 

Rogers C. S., Miller J., Muller E. M., Edmunds P., Nemeth R. S., Beets J. P., Friedlander A. M., 

Smith T. B. B., R., Jeffrey C. F. G., Menza C., Caldow C., Idrisi N., Kojis B., Monaco M. E., 

Spitzack A., Gladfelter E. H. O., J.C., Hillis-Starr Z., Lundgren I., Schill W. C., Kiffner I. B., 

Richardson L. L., Devine B. E. & Voss J. D. (2008) Ecology of coral reefs in the U.S. Virgin 

Islands. In: Coral Reefs of the World (eds B. M. Riegl and R. E. Dodge) pp. 303-373. Springer 

Science + Business Media. 

Rogers C. S., Suchanek T. & Pecora F. (1982) Effects of Hurricanes David and Frederic (1979) 

on shallow Acropora palmata reef communities: St. Croix, USVI. Bulletin of Marine Science 32, 

532-548. 

Rogers I. H., Birtwell I. K. & Kruzynski G. M. (1990) The Pacific eulachon (Thaleichthys 

pacificus) as a pollution indicator organism in the Fraser River estuary, Vancouver, British 

Columbia. Science of the Total Environment 97-98, 713-727. 

Rogers S. G. & Weber W. (1995a) Status and restoration of Atlantic and shortnose sturgeons in 

Georgia. Final Report to the National Marine Fisheries Service, Southeast Regional Office, St. 

Petersburg, Florida. 

Rogers S. G. & Weber W. (1995b) Status and restoration of Atlantic and shortnose sturgeons in 

Georgia. . National Marine Fisheries Service, Southeast Regional Office, St. Petersburg, Florida. 

Roman J. (2006) Diluting the founder effect: cryptic invasions expand a marine invader's range. 

Proceedings of the Royal Society B-Biological Sciences 273, 2453-2459. 

Roman J. & Darling J. A. (2007) Paradox lost: genetic diversity and the success of aquatic 

invasions. Trends in Ecology & Evolution 22, 454-464. 

Roman J. & Palumbi S. R. (2003) Whales before whaling in the North Atlantic. Science 301, 

508-510. 
487

Roni P. & Quinn T. P. (1995) Geographic variation in size and age of North American chinook 

salmon. North American Journal of Fisheries Management 15, 325–345. 

Ronnback P. (1999) The ecological basis for economic value of seafood production supported by 

mangrove ecosystems. Ecological Economics 29, 235-252. 

Rosenbaum H. C., Brownell Jr. R. L., Brown M. W., Schaeff C. M., Portway V., White B. N., 

Malik S., Pastene L. A., Patenaude N. J., Baker C. S., Goto M., Best P. B., Clapham P. J., 

Hamilton P. K., Moore M., Payne R., Rowntree V., Tynan C. T., Bannister J. L. & DeSalle R. 

(2000) World-wide genetic differentiation of Eubalana: questioning the number of right whale 

species. Molecular Ecology 9, 1793-1802. 

Rosenthal R. J., Haldorson L., Field L. J., Moran-O'Connell V. & LaRiviere M. G. (1982) 

Inshore and shallow offshore bottomfish resources in the southeastern Gulf of Alaska. Alaska 

Coastal Research, Sitka, Alaska. 

Rosenthal R. J., Moran-O'Connell V. & Murphy M. C. (1998) Feeding ecology of ten species of 

rockfishes (Scorpaenidae) from the Gulf of Alaska. California Department of Fish and Game 74, 

16-37. 

Rosetta T. & Borys D. (1996) Identification of sources of pollutants to the lower Columbia River 

basin. Draft report. . Lower Columbia River Bi-State Program, Oregon Department of 

Environmental Quality. 

Rosman I., Boland G. S., Martin L. & Chandler C. (1987) Underwater Sightings of Sea Turtles 

in the Northern Gulf of Mexico. OCS Study; MMS 87-0107, 37. 

Ross J. P. (1979) Sea turtles in the Sultanate of Oman. 

Ross J. P. (1998) Estimations of the nesting population size of loggerhead sea turtles, Caretta 

caretta, Masirah Island, Sultanate of Oman. In: Proceedings of the Seventeenth Annual Sea 

Turtle Symposium. (eds S. P. Epperly and J. Braun) pp. 84-87. 

Ross J. P. & Barwani M. A. (1982) Review of sea turtles in the Arabian area. In: Biology and 

conservation of sea turtles (ed K. A. Bjorndal) pp. 373-383. Smithsonian Institution Press, 

Washington, D.C. 

Ross P. S., Ellis G. M., Ikonomou M. G., Barrett-Lennard L. G. & Addison R. F. (2000a) High 

PCB concentrations in free-ranging Pacific killer whales, Orcinus orca: effects of age, sex, and 

dietary preference. Marine Pollution Bulletin 40, 504-515. 

Ross P. S., Ellis G. M., Ikonomou M. G., Barrett-Lennard L. G. & Addison. R. F. (2000b) High 

PCB concentrations in free-ranging Pacific killer whales, Orcinus orca: Effects of age, sex and 

dietary preference. Marine Pollution Bulletin 40, 504-515. 

Ross P. S., Vos J. G., Birnbaum L. S. & Osterhaus A. D. M. E. (2000c) PCBs are a health risk 

for humans and wildlife. Science 289, 1878-1879. 

Rossiter A., Noakes D. L. G. & Beamish F. W. H. (1995) Validation of Age Estimation for the 

Lake Sturgeon. Transactions of the American Fisheries Society 124, 5. 

Rostal D. C. (2007) Reproductive physiology of the ridley sea turtle. Pages 151-165 in: Plotkin 

P.T., editor. Biology and conservation of ridley sea turtles. Johns Hopkins University Press, 

Baltimore, Maryland. 

Rostal D. C., Grumbles J. S., Byles R. A., Márquez M. R. & Owens D. W. (1997) Nesting 

physiology of wild Kemp’s ridley turtles, Lepidochelys kempii, at Rancho Nuevo, Tamaulipas, 

Mexico. Chelonian Conservation and Biology 2, 538-547. 

Rothenberger P., Blondeau J., Cox C., Curtis S., Fisher W. S., Garrison V., Hillis-Starr Z., 

Jeffrey C. F. G., Kladison E., Lundgren I., Miller J. W., Muller E., Nemeth R., Patterson S., 

Rogers C., Smith T., Spitzack A., Taylor M., Toller W., Wright J., Wusinich-Mendez D., 
488

Waddell J., Gass J., Noorhasan N., Olsen D. & Westphal D. (2008) The State of Coral Reef 

Ecosystems of the U.S. Virgin Islands. In: The State of Coral Reef Ecosystems of the United 

States and Pacific Freely Associated States: 2008 (eds J. E. Waddell and A. M. Clarke) pp. 29-

73. 

Rowley J. (1929) Life history of the sea-lions on the California coast. Journal of Mammalogy 10, 

1-39. 

Royce W. F., Smith L. S. & Hartt A. C. (1968) Models of oceanic migrations of Pacific salmon 

and comments on guidance mechanisms. Fisheries Bulletin 66, 441-462. 

Royer T. C. (2005) Hydrographic responses at a coastal site in the northern Gulf of Alaska to 

seasonal and interannual forcing. Deep-Sea Research II 52, 267-288. 

Royer T. C. & Weingartner T. (1999) Coastal hydrographic responses in the northern Gulf of 

Alaska to the 1997-98 ENSO event. In: Proceedings of the 1998 Science Board Symposium on 

the impacts of the 1997/98 El Niño event on the North Pacific Ocean and its marginal seas. 

Ruckelshaus M. H. & McClure M. M. (2007) Sound Science: Synthesizing ecological and 

socioeconomic information about the Puget Sound ecosystem. Prepared in Cooperation with the 

Sound Science collaborative team. U.S. Department of Commerce, National Oceanic and 

Atmospheric Administration, National Marine Fisheries Service, Northwest Fisheries Science 

Center, Seattle, Washington. Available: 

http://www.nwfsc.noaa.gov/research/shared/sound_science/index.cfm (February 2008). 

Rudnick D. A., Halat K. M. & Resh V. H. (2000) Distribution, ecology and potential impacts of 

the Chinese mitten crab (Eriocheir sinensis) in San Francisco Bay. Department of Environmental 

Science, Policy and Management University of California, Berkeley. 

Rugh D. J., Shelden K. E. W. & Hobbs R. C. (2010) Range contraction in a beluga whale 

population. Endangered Species Research 12, 69-75. 

Rugh D. J. K., Shelden E. W. & Mahoney B. A. (2000) Distribution of belugas, Delphinapterus 

leucas, in Cook Inlet, Alaska, during June/July, 1993-2000. Marine Fisheries Review 62, 6-21. 

Ruiz G., Carleton J. T., Grosholz E. D. & Hines A. H. (1997) Global Invasions of Marine and 

Estuarine Habitats by Non-Indigenous Species: Mechanisms, Extent, and Consequences. 

American Zoologist, 621-632. 

Ruiz G. M., Fofonoff P. & Hines A. H. (1999) Non-indigenous species as stressors in estuarine 

and marine communities: Assessing invasion impacts and interactions. Limnology and 

Oceanography 44, 950–972. 

Ruiz G. M., Fofonoff P. W., Carlton J. T., J.Wonham M. & Hines A. H. (2000) Invasion of 

coastal marine communities in North America: Apparent patterns, processes, and biases. Annual 

Reviews of Ecology and Systematics 31, 481-531. 

Russell D. J. & Balazs G. H. (2009) Dietary shifts by green turtles (Chelonia mydas) in the 

Kane‘ohe Bay region of the Hawaiian Islands: A 28-year study. Pacific Science 63, 181-192. 

Rybitski M. J., Hale R. C. & Musick J. A. (1995) Distribution of organochlorine pollutants in 

Atlantic sea turtles. Copeia 1995, 379-390. 

Rylaarsdam K. W. (1983) Life histories and abundance patterns of colonial corals on Jamaican 

reefs. Marine Ecology Progress Series 13, 249-260. 

Saad M. A. (1999) Title Hadramaut coast importance in conservation of endangered green turtle. 

Marine Sciences Resources Research Center, Aden. Unpublished report. 8. 

Saba V. S., Spotila J. R., Chavez F. P. & Musick J. A. (2007) Bottom-up and climatic forcing on 

the global population of leatherback turtles. Pages 162-163 in: Frick, M., A. Panagopoulou, A.F. 
489

Rees, and K. Williams, editors. Twenty-seventh Annual Symposium on Sea Turtle Biology and 

Conservation. International Sea Turtle Society, Myrtle Beach, South Carolina. 

Saeki K., Sakakibara H., Sakai H., Kunito T. & Tanabe S. (2000) Arsenic accumulation in three 

species of sea turtles. Biometals 13, 241-250. 

Sakuma K. M. & Ralston S. (1995) Distributional patterns of late larval groundfish off central 

California in relation to hydrographic features during 1992 and 1993. California Cooperative 

Oceanic Fisheries Investigations Reports 36, 179-192. 

Sale A. & Luschi P. (2009) Navigational challenges in the oceanic migrations of leatherback sea 

turtles. Proceedings of the Royal Society B-Biological Sciences 276, 3737-3745. 

Sale A., Luschi P., Mencacci R., Lambardi P., Hughes G. R., Hays G. C., Benvenuti S. & Papi F. 

(2006) Long-term monitoring of leatherback turtle diving behaviour during oceanic movements. 

Journal of Experimental Marine Biology and Ecology 328, 197-210. 

Salm R. V. (1991) Turtles in Oman: Status, threats, and management options. In: Report of the 

scientific results of the IUCN coastal zone management project CZMP4:F11. 

Salm R. V., Jensen R. A. C. & Papastavrou V. A. (1993) Marine fauna of Oman: Cetaceans, 

turtles, seabirds, and shallow water corals. IUCN, Gland, Switzerland. 

Salo E. O. (1991) Life history of chum salmon (Oncorhynchus keta). Pages, editor Pacific 

salmon life histories. University of British Columbia Press, Edmonton, Alberta. 

Sammarco P. W. (1980) Diadema and its relationship to coral spat mortality: Grazing, 

competition, and biological disturbance. Journal of Experimental Marine Biology and Ecology

45, 245-272. 

Sammarco P. W. (1985) The Great Barrier Reef vs. the Caribbean: Comparisons of grazers, coral 

recruitment patterns, and reef recovery. Proceedings of the 5th International Coral Reef 

Congress 4, 391-397. 

Sanchez I., Fernandez C. & Arrontes J. (2005) Long-Term Changes in the Structure of Intertidal 

Assemblages after Invasion by Sargassum Muticum (Phaeophyta)1. Journal of Phycology 41, 

942-949. 

Sandegren F. E. (1970) Breeding and maternal behavior of the Steller sea lion (Eumatopias 

jubatus) in Alaska. University of Alaska. 

Sandercock F. K. (1991) The History of Coho Salmon (Oncorhynchus kisutch). Pages xx-xx in, 

C. Groot and L. Margolis, eds. Pacific Salmon Life History. Vancouver: University of British 

Columbia Press. 

Sandøy S. & Langåker R. M. (2001) Atlantic salmon and acidification in southern Norway: a 

disaster in the 20th century, but a hope for the future? Water Air and Soil Pollution 130(1-4, 

1343-1348. 

Santana Garcon J., Grech A., Moloney J. & Hamann M. (2010) Relative Exposure Index: an 

important factor in sea turtle nesting distribution. Aquatic Conservation: Marine and Freshwater 

Ecosystems 20, 140-149. 

Santidrián Tomillo P., Vélez E., Reina R. D., Piedra R., Paladino F. V. & Spotila J. R. (2007) 

Reassessment of the leatherback turtle (Dermochelys coriacea) nesting population at Parque 

Nacional Marino Las Baulas, Costa Rica: Effects of conservation efforts. Chelonian 

Conservation and Biology 6, 54-62. 

Sapota M. R. & Skora K. E. (1999) Invasion of round goby (Neogobius melanostomus) in the 

Baltic. In: Conference of the Interantional Association for Great Lakes Research pp. 95-96, 

Cleveland, Ohio. 
490

Sapota M. R. & Skora K. E. (2005) Spread of alien (non-indigenous) fish species Neogobius 

melanostomus in the Gulf of Gdansk (south Baltic). Biological Invasions 7, 157-164. 

Sarti L., Eckert S., Dutton P., Barragán A. & García N. (2000) The current situation of the 

leatherback population on the Pacific coast of Mexico and Central America, abundance and 

distribution of the nestings: an update. In: 19th Annual Symposium on Sea Turtle Conservation 

and Biology (eds H. Kalb and T. Wibbels) pp. 85-87, South Padre Island, Texas. 

Sarti L. M., Eckert S. A., Garcia N. T. & Barragan A. R. (1996) Decline of the world’s largest 

nesting assemblage of leatherback turtles. Marine Turtle Newsletter 74, 2-5. 

Sato K., Matsuzawa Y., Tanaka H., Bando T., Minamikawa S., Sakamoto W. & Naito Y. (1998) 

Internesting intervals for loggerhead turtles, Caretta caretta, and green turtles, Chelonia mydas, 

are affected by temperature. Canadian Journal of Zoology 76, 1651-1662. 

Saulitis E., Matkin C., Barrett-Lennard L., Heise K. & Ellis G. (2000) Foraging strategies of 

sympatric killer whale (Orcinus orca) populations in Prince William Sound, Alaska. Marine 

Mammal Science 16, 94-109. 

Saunders R. L. (1981) Atlantic salmon (Salmo salar) stocks and management implications in the 

Atlantic Provinces and New England, USA. Canadian Journal of Fisheries and Aquatic Sciences

38, 1612-1625. 

Saunders T., Mayfield S. & Hogg A. (2009a) Using a simple morphometric marker to identify 

spatial units for abalone fishery management. Ices Journal of Marine Science 66, 305-314. 

Saunders T. M., Connell S. D. & Mayfield S. (2009b) Differences in abalone growth and 

morphology between locations with high and low food availability: morphologically fixed or 

plastic traits? Marine Biology 156, 1255-1263. 

Savino J. F. & Kolar C. S. (1996) Competition between Nonindigenous Ruffe and Native Yellow 

Perch in Laboratory Studies. Transactions of the American Fisheries Society 125, 562-571. 

Savoy T. (2007) Prey Eaten by Atlantic Sturgeon in Connecticut Waters. American Fisheries 

Society Symposium 56, 157. 

Scarff J. E. (1986) Historic and present distribution of the right whale (Eubalaena glacialis) in 

the eastern North Pacific south of 50o

N and east of 180oW. Report of the International Whaling 

Commission, 43-63. 

Scatterday J. W. (1974) Reefs and associated coral assemblages off Bonaire, Netherlands 

Antilles and their bearing on Pleistocene and Recent reef models. Proceedings of the 2nd 

International Coral Reef Symposium 2, 85-106. 

Schaffer W. M. & Elson P. F. (1975) The adaptive significance of variations in life history 

among local populations of Atlantic salmon in North America. Ecology 56, 577-590. 

Scheffer V. B. (1945) Growth and behavior of young sea lions. Journal of Mammalogy 26, 390-

392. 

Scheffer V. B. (1950) Mammals of the Olympic National Park and vicinity. Northwest Fauna 2, 

192-225. 

Scheffer V. B. & Slipp J. W. (1948) The whales and dolphins of Washington State with a key to 

the cetaceans of the west coast of North America. American Midland Naturalist 39, 257-337. 

Scheffer V. B. & Slipp. J. W. (1948a) The whales and dolphins of Washington State with a key 

to the cetaceans of the west coast of North America. American Midland Naturalist 39, 257-337. 

Scheffer V. B. & Slipp. J. W. (1948b) The whales and dolphins of Washington State with a key 

to the cetaceans of the west coast of North America. American Midland Naturalist 39, 257-337. 
491

Scheidat M., Castro C., Gonzalez J. & Williams. R. (2004) Behavioural responses of humpback 

whales (Megaptera novaeangliae) to whalewatching boats near Isla de la Plata, Machalilla 

National Park, Ecuador. Journal of Cetacean Research and Management 6, 63-68. 

Scheuerell M. D., Zabel R. W. & Sandford B. P. (2009) Relating juvenile migration timing and 

survival to adulthood in two species of threatened Pacific salmon (Oncorhynchus spp.). Journal 

of Applied Ecology 46, 983-990. 

Schloesser D. W., Kovalek W. P., Longton G. D., Ohnesorg K. L. & Smithee R. D. (1998) 

Impact of zebra and quagga mussels (Dreissena spp.) on freshwater unionids (Bivalvia: 

Unionidae) in the Detroit River of the Great Lakes. American Midland Naturalist 140, 299-313. 

Schmid J. R. (1998) Marine turtle populations on the west central coast of Florida: Results of 

tagging studies at the Cedar Keys, Florida, 1986-1995. Fishery Bulletin 96, 589-602. 

Schmid J. R., Bolten A. B., Bjorndal K. A. & Lindberg W. J. (2002) Activity patterns of Kemp's 

ridley turtles, Lepidochelys kempii, in the coastal waters of the Cedar Keys, Florida. Marine 

Biology 140, 215-228. 

Schmid J. R. & Witzell W. N. (1997a) Age and growth of wild Kemp's ridley turtles 

(Lepidochelys kempi): Cumulative results of tagging studies in Florida. Chelonian Conservation 

and Biology 2, 20 pp. 

Schmid J. R. & Witzell W. N. (1997b) Age and growth of wild Kemp's ridley turtles 

(Lepidochelys kempii): Cumulative results of tagging studies in Florida. Chelonian Conservation 

and Biology 2, 532-537. 

Schmidt J. (1916) Marking experiments with turtles in the Danish West Indies. Meddelelser Fra 

Kommissionen For Havundersogelser. Serie: Fiskeri. Bind V. Nr. 1. Kobenhavn. 

Schofield G., Bishop C. M., Katselidis K. A., Dimopoulos P., Pantis J. D. & Hays G. C. (2009) 

Microhabitat selection by sea turtles in a dynamic thermal marine environment. Journal of 

Animal Ecology 78, 14-21. 

Schoonmaker P. K., Gresh T., Lichatowich J. & Radtke H. D. (2003) Past and present salmon 

abundance: Bioregional estimates for key life history stages. Pages 33-40 in: J.G. Stockner, 

editor. Nutrients in salmonid ecosystems: Sustaining production and biodiversity. American 

Fisheries Society, Symposium 34, Bethesda, Maryland. 

Schroeder B. A. & Thompson N. B. (1987) Distribution of the loggerhead turtle, Caretta caretta, 

and the leatherback turtle, Dermochelys coriacea, in the Cape Canaveral, Florida area: Results of 

aerial surveys. In: Proceedings of the Cape Canaveral, Florida Sea Turtle Workshop (ed W. N. 

Witzell) pp. 45-53. 

Schroeder D. M. & Love M. S. (2002) Recreational fishing and marine fish populations in 

California. California Cooperative Oceanic Fisheries Investigations Reports 43, 182-190. 

Schueller P. & Peterson D. L. (2010) Abundance and Recruitment of Juvenile Atlantic Sturgeon 

in the Altamaha River, Georgia. Transactions of the American Fisheries Society 139, 1526-1535. 

Schultz J. P. (1975) Sea turtles nesting in Surinam. Zoologische Verhandelingen 143. 

Schulz J. P. (1984) Turtle conservation strategy in Indonesia. IUCN/WWF Report. 

Schulz J. P. (1987) Status of and trade in Chelonia mydas and Eretmochelys imbricata in 

Indonesia. Consultancy report prepared for IUCN Conservation Monitoring Centre. 

Schwartz F. J. (2003a) Bilateral asymmetry in the rostrum of the smalltooth sawfish, Pristis 

pectinata (Pristiformes: Family Pristidae). Journal of the North Carolina Academy of Science

119, 41-47. 
492

Schwartz F. J. (2003b) Bilateral asymmetry in the rostrum of the smalltooth sawfish, Pristis 

pectinata (Pristiformes: Family Pristidae). Journal of the North Carolina Academy of Science

119, 41-47. 

Schweigert J., McCarter B., Therriault T., Flostrand L., Hrabok C., Winchell P. & Johannessen 

D. (2007) Ecosystem overview: Pacific North Coast integrated management area (PNCIMA), 

Appendix H: pelagic fishes. 

Scott W. B. & Crossman E. J. (1973a) Atlantic salmon. Pages 192-197 in: Freshwater fishes of 

Canada. Department of Fisheries and Oceans, Scientific Information and Publications Branch, 

Ottawa. Bulletin, 184. 

Scott W. B. & Crossman E. J. (1973b) Freshwater fishes of Canada. Bulletin of the Fisheries 

Research Board of Canada 184, 1-966. 

Scott W. B. & Crossman E. J. (1973c) Freshwater fishes of Canada. Bulletin of the Fisheries 

Research Board of Canada 184, 1-966. 

Sea Turtle Association of Japan. (2010) Sea turtles of Japan. 

Seabloom E. W. (2003) Invasion, competitive dominance, and resource use by exotic and native 

California grassland species. Proceedings of the National Academy of Sciences 100, 13384-

13389. 

Seaman G. A. & Burns. J. J. (1981) Preliminary results of recent studies of Belukhas in Alaskan 

waters. (Delphinapterus leucas). Report of the International Whaling Commission 31, 567-574.-

Sc/532/Sm513). 

Sears C. J. (1994) Preliminary genetic analysis of the population structure of Georgia loggerhead 

sea turtles. NOAA Technical Memorandum NMFS-SEFSC-351. U.S. Department of Commerce, 

National Oceanic and Atmospheric Administration, National Marine Fisheries Service, 

Southeast Fisheries Science Center, Miami, Florida. 

Sears C. J., Bowen B. W., Chapman R. W., Galloway S. B., Hopkins-Murphy S. R. & Woodley 

C. M. (1995) Demographic composition of the feeding population of juvenile loggerhead sea 

turtles (Caretta caretta) off Charleston, South Carolina: evidence from mitochondrial DNA 

markers. Marine Biology 123, 869-874. 

Sease J. L. & Gudmundson C. J. (2002) Aerial and land-based surveys of Steller sea lions 

(Eumetopias jubatus) from the western stock in Alaska, June and July 2001 and 2002. NOAA. 

Sease J. L., Taylor W. P., Loughlin T. R. & Pitcher K. W. (2001) Aerial and land-based surveys 

of Steller sea lions (Eumetopias jubatus) in Alaska. June and July 1999 and 2000. U.S. 

Department of Commerce. 

Secor D., Anders P., Webster V. W. & Dixon D. (2002) Can we study sturgeon to extinction? 

What we do and don't know about the conservation of North American sturgeon. Pages 3-9 in: 

Webster, V.W., editor. Biology, management, and protection of North American sturgeon, 

Symposium 28. American Fisheries Society, Bethesda, Maryland. 

Sedell J. R. & Froggatt J. L. (1984) Importance of streamside forests to large rivers: The 

isolation of the Willamette River, Oregon, USA from its floodplain by snagging and streamside 

forest removal. Internationale Vereinigung für Theoretische und angewandte Limnologie 

Verhandlungen 22, 1828-1834. 

Seitz J. C. & Poulakis G. R. (2002) Recent occurrences of sawfishes (Elasmobranchiomorphi: 

Pristidae) along the southwest coast of Florida (USA). Florida Scientist 65, 256–266. 

Selgeby J. (1998) Predation by ruffe (Gymnocephalus cernuus) on fish eggs in Lake Superior. 

Journal of Great Lakes Research 24, 304-308. 
493

Seminoff J. A. (2004a) 2004 global status assessment: Green turtle (Chelonia mydas). p. 71. 

IUCN Marine Turtle Specialist Group Review. 

Seminoff J. A. (2004b) 2004 global status assessment: Green turtle (Chelonia mydas). IUCN 

Marine Turtle Specialist Group Review. 

Seminoff J. A. & Jones T. T. (2006) Diel movements and activity ranges of green turtles 

(Chelonia mydas) at a temperate foraging area in the Gulf of California, Mexico. Herpetological 

Conservation and Biology 1, 81-86. 

Seminoff J. A., Jones T. T., Resendiz A., Nichols W. J. & Chaloupka M. Y. (2003) Monitoring 

green turtles (Chelonia mydas) at a coastal foraging area in Baja California, Mexico: Multiple 

indices to describe population status. Journal of the Marine Biological Association of the United 

Kingdom 83, 1355-1362. 

Seminoff J. A., Resendiz A. & Nichols W. J. (2002a) Diet of East Pacific green turtles (Chelonia 

mydas) in the central Gulf of California, Mexico. Journal of Herpetology 36, 447-453. 

Seminoff J. A., Resendiz A., Nichols W. J. & Jones T. T. (2002b) Growth rates of wild green 

turtles (Chelonia mydas) at a temperate foraging area in the Gulf of California, México. Copeia

2002, 610-617. 

Senko J., Koch V., Megill W. M., Carthy R. R., Templeton R. P. & Nichols W. J. (2010a) Fine 

scale daily movements and habitat use of East Pacific green turtles at a shallow coastal lagoon in 

Baja California Sur, Mexico. Journal of Experimental Marine Biology and Ecology in press, in 

press. 

Senko J., Lopez-Castro M. C., Koch V. & Nichols W. J. (2010b) Immature East Pacific green 

turtles (Chelonia mydas) use multiple foraging areas off the Pacific Coast of Baja California Sur, 

Mexico: First evidence from mark-recapture data. Pacific Science 64, 125-130. 

Sergeant D. E. (1973) Biology of white whales (Delphinapterus leucas) in western Hudson Bay. 

Journal of the Fisheries Research Board of Canada 30, 1065-1090, +1062Pls. 

Sergeant D. E. & Brodie P. F. (1975) Identity, abundance, and present status of white whales, 

Delphinapterus leucas, in North America. Journal of the Fisheries Research Board of Canada

32, 1047-1054. 

Shaffer J. A., Penttila D., McHenry M. & Vilella D. (2007) Observations of eulachon, 

Thaleichthys pacificus, in the Elwha River, Olympic Peninsula Washington. Northwest Science

81, 76-81. 

Shane S. H. (1994) Occurrence and habitat use of marine mammals at Santa Catalina Island, 

California from 1983-91. Bulletin of the Southern California Academy of Sciences 93, 13-29. 

Shane S. H. (1995) Behavior patterns of pilot whales and Risso's dolphins off Santa Catalina 

Island, California. Aquatic Mammals 21, 195-197. 

Shapovalov L. & Taft A. C. (1954) The life histories of the steelhead rainbow trout (Salmo 

gairdneri gairdneri) and silver salmon (Oncorhynchus kisutch) with special reference to Waddell 

Creek, California, and recommendations regarding their management. California Department of 

Fish Game Bulletin 98. Available: 

http://content.cdlib.org/xtf/view?docId=kt9x0nb3v6&brand=calisphere&doc.view=entire_text

(May 2008). 

Shaver D. J. (1999) Kemp's ridley sea turtle project at Padre Island National Seashore, Texas. 

Pages 342-347 in: McKay, M., and J. Nides, editors. Proceedings of the Seventeenth Annual 

Gulf of Mexico Information Transfer Meeting, U.S. Department of the Interior, Minerals 

Management Service, Gulf of Mexico OCS Region, MMS 99-0042. 
494

Shaver D. J. (2002) Kemp's ridley sea turtle project at Padre Island National Seashore and Texas 

sea turtle nesting, and stranding 2001 report. U.S. Department of the Interior, U.S. Geological 

Survey, Corpus Christi, Texas. 

Shaver D. J., Amos A. F., Higgins B. & Mays J. (2005) Record 42 Kemp's ridley nests found in 

Texas in 2004. Marine Turtle Newsletter 108, 1-3. 

Shaver D. J. & Wibbels T. (2007a) Head-starting the Kemp's ridley sea turtle. Pages 297-323 in: 

Plotkin P.T., editor. Biology and conservation of ridley sea turtles. Johns Hopkins University 

Press, Baltimore, Maryland. 

Shaver D. J. & Wibbels T. (2007b) Head-starting the Kemp's ridley sea turtle. In: Biology and 

Conservation of Ridley Sea Turtles (ed P. T. Plotkin) pp. 297-323. The Johns Hopkins University 

Press, Baltimore, MD. 

Shea K. & Mangel M. (2001) Detection of population trends in threatened coho salmon 

(Oncorhynchus kisutch). Canadian Journal of Fisheries and Aquatic Sciences 58, 375-385. 

Shelden K. E. W., Rugh D. J., Mahoney B. A. & Dahlheim M. E. (2003) Killer whale predation 

on belugas in Cook Inlet, Alaska: Implications for a depleted population. Marine Mammal 

Science 19, 529-544. 

Sherr E. B., Sherr B. F. & Wheeler P. A. (2005) Distribution of coccoid cyanobacteria and small 

eukaryotic phytoplankton in the upwelling ecosystem off the Oregon coast during 2001 and 

2002. Deep-Sea Research II 52, 317-330. 

Sherwood C. R., Jay D. A., Harvey R. B., Hamilton P. & Simenstad C. A. (1990) Historical 

changes in the Columbia River estuary. Progress in Oceanography 25, 299-357. 

Shillinger G. L., Palacios D. M., Bailey H., Bograd S. J., Swithenbank A. M., Spotila J. R., 

Wallace B. P., Paladino F. V., Eckert S. A., Piedra R. & Block B. A. (2010) Four years and 

fourty-six turtles: tracking the movements and behaviors of leatherback sea turtles in the eastern 

Pacific. In: 28th Annual Symposium on Sea Turtle Biology and Conservation (eds K. Dean and 

M. C. L. Castro) p. 53. National Marine Fisheries Service. 

Shinn E. A. (1963) Spur and groove formation on the Florida Reef Tract. Journal of Sedimentary 

Petroleum 33, 291-303. 

Shinn E. A. (1976) Coral reef recovery in Florida and the Persian Gulf. Environmental Geology

1, 241-254. 

Shoop C. R. & Kenney R. D. (1992a) Seasonal distributions and abundances of loggerhead and 

leatherback sea turtles in waters of the northeastern United States. Herpetological Monographs

6, 43-67. 

Shoop C. R. & Kenney R. D. (1992b) Seasonal distributions and abundances of loggerhead and 

leatherback seaturtles in waters of the northeastern United States. Herpetological Monographs 6, 

43-67. 

Sigler M. F., Tollit D. J., Vollenweider J. J., Thedinga J. F., Csepp D. J., Womble J. N., M, 

Wong y. A., Rehberg M. J. & Trites. r. W. (2009) Steller sea lion foraging response to seasonal 

changes in prey availability. Marine Ecology Progress Series 388, 243-261. 

Simenstad C. A. & Salo E. O. (1982) Foraging success as a determinant of estuarine and 

nearshore carrying capacity of juvenile chum salmon (Oncorhychus keta) in Hood Canal, 

Washington. Pages 21-37 in B.R. Melteff and R.A. Nevé, eds. Proceedings of the North Pacific 

Aquaculture Symposium, Alaska Sea Grant Report 82-2. 

Simkanin C., Davidson I., Falkner M., Sytsma M. & Ruiz G. (2009) Intra-coastal ballast water 

flux and the potential for secondary spread of non-native species on the US West Coast. Marine 

Pollution Bulletin 58, 366-374. 
495

Simmonds M. P. & Eliott. W. J. (2009) Climate change and cetaceans: Concerns and recent 

developments. Journal of the Marine Biological Association of the United Kingdom 89, 203-210. 

Simpfendorfer C. (2001) Essential habitat of smalltooth sawfish (Pristis pectinata). Mote Marine 

Library Technical Report 786. Mote Marine Laboratory, Sarasota, Florida. 

Simpfendorfer C. A. (2000a) Predicting population recovery rates for endangered western 

Atlantic sawfishes using demographic analysis. Environmental Biology of Fishes 58, 371-377. 

Simpfendorfer C. A. (2000b) Predicting recovery rates for endangered western Atlantic 

sawfishes using demographic analysis. Environmental Biology of Fishes 58, 371-377. 

Simpfendorfer C. A. (2002) Smalltooth sawfish: the USA’s first endangered elasmobranch? 

Endangered Species Update 19, 45-49. 

Simpfendorfer C. A. (2003a) Abundance, movement and habitat use of the smalltooth sawfish. 

Final Report to the National Marine Fisheries Service, Grant number WC133F-02-SE-0247. 

Mote Marine Laboratory, Sarasota, Florida. Mote Marine Laboratory Technical Report, 929. 

Simpfendorfer C. A. (2003b) Abundance, movement and habitat use of the smalltooth sawfish. 

p. 20 pp. Mote Marine Lab, Sarasota, FL. 

Simpfendorfer C. A., Poulakis, G. R., O’Donnell, P. M. and Wiley, T. R. (2008) Growth rates of 

juvenile smalltooth sawfish Pristis pectinata Latham in the western Atlantic. Journal of Fish 

Biology 72: 711-723. 

Simpfendorfer C. A. & Wiley T. R. (2004) Determination of the distribution of Florida’s 

remnant sawfish population, and identification of areas critical to their conservation. Mote 

Marine Laboratory Technical Report. Mote Marine Laboratory, Sarasota, Florida. 

Simpfendorfer C. A., Wiley T. R. & Yeiser B. G. (2010) Improving conservation planning for an 

endangered sawfish using data from acoustic telemetry. Biological Conservation. 

Sinclair E. & Zeppelin T. (2002) Seasonal and spatial differences in diet in the western stock of 

Steller sea lions (Eumetopias jubatus). Journal of Mammalogy 83, 973-990. 

Sirovic A., Cutter G. R., Butler J. L. & Demer D. A. (2009) Rockfish sounds and their potential 

use for population monitoring in the Southern California Bight. pp. 981-990. 

Skjeveland J. E., Welsh S. A., Mangold M. F., Eyler S. M. & Nachbar S. (2000) A report of 

investigations and research on Atlantic and shortnose sturgeon in Maryland waters of 

Chesapeake Bay (1996-2000). U.S. Fish and Wildlife Service, Annapolis, Maryland. 

Slaney T. L., Hyatt K. D., Northcote T. G. & Fielden R. J. (1996) Status of anadromous salmon 

and trout in British Columbia and Yukon. Fisheries 21, 20-35. 

Slijper E. J., Utrecht W. L. V. & Naaktgeboren C. (1964) Remarks on the distribution and 

migration of whales, based on observations from Netherlands ships. Bijdragen Tot De Dierkunde

34, 3-93. 

Smith G., Stamm C. & Petrovic F. (2003) Haliotis cracherodii (IUCN Red List of Threatened 

Species). 

Smith L. D., Wonham M. J., McCann L. D., Ruiz G. M., Hines A. H. & Carlton J. T. (1999a) 

Invasion pressure to a ballast-flooded estuary and an assessment of inoculant survival. Biological 

Invasions 1, 67-87. 

Smith T. D., Allen J., Clapham P. J., Hammond P. S., Katona S., Larsen F., Lien J., Mattila D. & 

Palsbøll P. J. (1999b) An ocean-basin-wide mark-recapture study of the North Atlantic 

humpback whale (Megaptera novaeangliae). Marine Mammal Science 15, 1-32. 

Smith T. D., Barthelmess K. & Reeves R. R. (2006) Using historical records to relocate a longforgotten summer feeding ground of North Atlantic right whales. Marine Mammal Science 22, 

723-734. 
496

Smith T. D. & G.Pike D. (2009) The enigmatic whale: the North Atlantic humpback. Nammco 

Scientific Publications 7, 161-178. 

Smith T. D. & Reeves R. R. (2003) Estimating American 19th century catches of humpback 

whales in the West Indies and Cape Verde Islands. Caribbean Journal of Science 39, 286-297. 

Smith T. I. J. (1985) The fishery, biology, and management of Atlantic sturgeon, Acipenser 

oxyrhynchus, in North America. Environmental Biology of Fishes 14, 61-72. 

Smith T. I. J. & Clugston J. P. (1997) Status and management of Atlantic sturgeon, Acipenser 

oxyrinchus, in North America. Environmental Biology of Fishes 48, 335-346. 

Smith W. E. & Saalfeld R. W. (1955) Studies on Columbia River smelt Thaleichthys pacificus 

(Richardson). Washington Department of Fisheries, Fisheries Research Paper 1, 3-26. 

Smock L. A. A. B. W. & Benke A. C. (2005) Atlantic coast rivers of the southeastern United 

States. Pages 72-122 in A.C. Benke and C.E. Cushing, eds. Rivers of North America. Elsevier 

Academic Press, Burlington, Massachusetts. Available: 

http://books.google.com/books?id=faOU1wkiYFIC&pg=RA3-PA541&lpg=RA3-

PA541&dq=pacific+coast+rivers+of+the+coterminous+united+states&source=web&ots=-

pMpyECFaA&sig=FkGrIiwgkfDyHxXCWXRalK_XSvU#PPR1,M1 (February 2008). 

Snover M. L., Hohn A. A., Crowder L. B. & Heppell S. S. (2007a) Age and growth in Kemp's 

ridley sea turtles: evidence from mark-recapture and skeletochronology. In: Biology and 

Conservation of Ridley Sea Turtles (ed P. T. Plotkin). The Johns Hopkins University Press, 

Baltimore, MD. 

Snover M. L., Hohn A. A., Crowder L. B. & Heppell S. S. (2007b) Age and growth in Kemp's 

ridley sea turtles: Evidence from mark-recapture and skeletochronology. Pages 89-106 in: 

Plotkin P.T., editor. Biology and conservation of ridley sea turtles. Johns Hopkins University 

Press, Baltimore, Maryland. 

Sogard S. M., Berkeley S. A. & Fisher R. (2008) Maternal effects in rockfishes Sebastes spp.: a 

comparison among species. Marine Ecology-Progress Series 360, 227-236. 

Sogard S. M., Williams T. H. & Fish H. (2009) Seasonal Patterns of Abundance, Growth, and 

Site Fidelity of Juvenile Steelhead in a Small Coastal California Stream. Transactions of the 

American Fisheries Society 138, 549-563. 

Solbakken V., Hansen A. T. & Stefansson S. O. (1994) Effects of photoperiod and temperature 

on growth and parr– smolt transformation in Atlantic salmon (Salmo salar L.) and subsequent 

performance in seawater. Aquaculture 121:13– 27. 

Solow A. R., Bjorndal K. A. & Bolten A. B. (2002) Annual variation in nesting numbers of 

marine turtles: The effect of sea surface temperature on re-migration intervals. Ecology Letters 5, 

742-746. 

Somero G. N. & Hofmann G. E. (1997) Global warming: Implications for freshwater and marine 

fish. Society of Experimental Biology Seminar Series 61, 1-24. 

Soong K. & Lang J. C. (1992) Reproductive integration in coral reefs. Biological Bulletin 183, 

418-431. 

Spangler E. A. K. (2002) The ecology of eulachon (Thaleichthys pacificus) in Twentymile River, 

Alaska. University of Alaska, Fairbanks, Alaska. 

Spaulding G. C. (1964) Comparative feeding habits of the fur seal, sea lion, and harbour seal on 

the British Columbia coast. 

Speckman S. G. & Piatt. J. F. (2000) Historic and current use of Lower Cook Inlet, Alaska, by 

belugas, Delphinapterus leucas. Marine Fisheries Review 62, 22-26. 
497

Spence B. C., Lomnicky G. A., Hughes R. M. & Novitzki R. P. (1996) An ecosystem approach 

to salmonid conservation. NOAA, National Marine Fisheries Service. 

Spotila J. R. (2004a) Sea turtles: A complete guide to their biology, behavior, and conservation. 

The Johns Hopkins University Press and Oakwood Arts, Baltimore, Maryland. 

Spotila J. R. (2004b) Sea turtles: A complete guide to their biology, behavior, and conservation. 

John Hopkins University Press, Baltimore. 227p. 

Spotila J. R., Dunham A. E., Leslie A. J., Steyermark A. C., Plotkin P. T. & Paladino F. V. 

(1996) Worldwide population decline of Dermochelys coriacea: Are leatherback turtles going 

extinct? Chelonian Conservation and Biology 2, 209-222. 

Spotila J. R., Reina R. D., Steyermark A. C., Plotkin P. T. & Paladino F. V. (2000) Pacific 

leatherback turtles face extinction. Nature 405, 529-530. 

Spruill T. & Survey G. (1998) Water quality in the Albemarle-Pamlico drainage basin, North 

Carolina and Virginia, 1992-95. In: U.S. Geological Survey Circular 1157. US Dept. of the 

Interior, US Geological Survey. 

Squiers T. S. (2003) Completion report Kennebec River shortnose sturgeon population study 

(1997-2001). National Marine Fisheries Service. 

Squiers T. S., Smith M. & Flagg L. (1982) American shad enhancement and status of sturgeon 

stocks in selected Maine waters. National Marine Fisheries Service, Gloucester. 

SSPS. (2007) Puget Sound salmon recovery plan. Volume 1, recovery plan. Shared Strategy for 

Puget Sound, Seaatle, Washington. 

Stabeno P. J., Bond N. A., Hermann A. J., Kachel N. B., Mordy C. W. & Overland. J. E. (2004) 

Meteorology and oceanography of the northern Gulf of Alaska. Continental Shelf Research 24, 

859-897. 

Stachowicz J. J. (2002) Linking climate change and biological invasions: Ocean warming 

facilitates nonindigenous species invasions. Proceedings of the National Academy of Sciences

99, 15497-15500. 

Stachowicz J. J., Whitlatch R. B. & Osman R. W. (1999) Species Diversity and Invasion 

Resistance in a Marine Ecosystem. Science 286, 1577-1579. 

Stafford K. M., Mellinger D. K., Moore S. E. & Fox C. G. (2007) Seasonal variability and 

detection range modeling of baleen whale calls in the Gulf of Alaska, 1999–2002. The Journal of 

the Acoustical Society of America 122, 3378-3390. 

Stahl-Johnson K. L. (1985) Descriptive characteristics of reared Sebastes caurinus and S. 

auriculatus larvae. In: 3rd International Symposium on the Early Life History of Fishes & 8th 

Annual Larval Fish Conference (eds A. W. Kendall Jr. and J. B. Marliave) pp. 65-76. 

STAJ. (2002) Population trends and mortality of Japanese loggerhead turtles, Caretta caretta in 

Japan. In: Western Pacific Sea Turtle Cooperative Research and Management Workshop, 

Honolulu, Hawaii. 

Stamper M. A., Spicer C. W., Neiffer D. L., Mathews K. S. & Fleming G. J. (2009) 

MORBIDITY IN A JUVENILE GREEN SEA TURTLE (CHELONIA MYDAS) DUE TO 

OCEAN-BORNE PLASTIC. Journal of Zoo and Wildlife Medicine 40, 196-198. 

Stanford J. A. F. R. H. S. V. G. & Synder E. B. (2005) Columbia River basin. Pages 591-653 in 

A.C. Benke and C.E. Cushing, eds. Rivers of North America. Elsevier Academic Press, 

Burlington, Massachusetts. Available: 

http://books.google.com/books?id=faOU1wkiYFIC&pg=RA3-PA541&lpg=RA3-

PA541&dq=pacific+coast+rivers+of+the+coterminous+united+states&source=web&ots=-

pMpyECFaA&sig=FkGrIiwgkfDyHxXCWXRalK_XSvU#PPR1,M1 (February 2008). 
498

Stanley D. R. & Wilson C. A. (2003) Utilization of offshore platforms by recreational fishermen 

and scuba divers off the Louisiana coast. Bulletin of Marine Science 44, 767-775. 

Starbird C. H., Baldridge A. & Harvey J. T. (1993) Seasonal occurrence of leatherback sea 

turtles (Dermochelys coriacea) in the Monterey Bay region, with notes on other sea turtles, 1986-

1991. California Fish and Game 79, 54-62. 

Starr R. M. (1998) Design principles for rockfish reserves on the U. S. West Coast. (ed M. M. 

Yoklavich) pp. 50-63. Marine harvest refugia for west coast rockfish: a workshop. 

Starr R. M., Heine J. N., Felton J. M. & Cailliet G. M. (2002) Movements of bocaccio (Sebastes 

paucispinis) and greenspotted (S. chlorostictus) rockfishes in a Monterey submarine canyon: 

implications for the design of marine reserves. Fishery Bulletin 100, 324-337. 

STAT. (1999) Status of the canary rockfish resource off Oregon and Washington in 1999. 

National Marine Fisheries Service, Newport, Oregon. 

Stat M. & Gates R. D. (2008) Vectored introductions of marine endosymbiotic dinoflagellates 

into Hawaii. Biological Invasions 10, 579-583. 

State of Oregon. (2005) Coho assessment part 1: Synthesis final report. Eugene, Oregon. 

State O. S. o. (2004) Oregon Blue Book. Oregon Secretary of State, Salem, Oregon. 

Staurnes M., Blix P. & Reite O. B. (1993) Effects of acid water and aluminum on parr smolt 

transformation and seawater tolerance in Atlantic salmon, Salmo salar. Canadian Journal of 

Fisheries and Aquatic Sciences 50, 1816-1827. 

Stearns S. C. (1992) The evolution of life histories. Oxford University Press, 249p. 

Steiger G. H., Calambokidis J., Straley J. M., Herman L. M., Cerchio S., Salden D. R., R. J. U.-., 

Jacobsen J. K., Ziegesar O. v., Balcomb K. C., Gabriele C. M., Dahlheim M. E., Uchida S., Ford 

J. K. B., P. P. L. d. G.-., Yamaguchi M. & Barlow. J. (2008) Geographic variation in killer 

whale attacks on humpback whales in the North Pacific: Implications for predation pressure. 

Endangered Species Research 4, 247-256. 

Stein A. B., Friedland K. D. & Sutherland M. (2004) Atlantic Sturgeon Marine Bycatch and 

Mortality on the Continental Shelf of the Northeast United States. North American Journal of 

Fisheries Management 24, 171-183. 

Stein J. E., Collier T. K., Reichert W. L., Casillas E., Hom T. & Varanasi U. (1992) 

Bioindicators of contaminant exposure and sublethal effects: Studies with benthic fish in Puget 

Sound, WA. Environmental Toxicology and Chemistry 11, 701-714. 

Stenseth N. C., Mysterud A., Ottersen G., Hurrell J. W., Chan K.-S. & Lima. M. (2002) 

Ecological effects of climate fluctuations. Science 297, 1292-1296. 

Stevens D. E. & Miller L. W. (1983) Effects of river flow on abundance of young Chinook 

salmon, American shad, longfin smelt, and delta smelt in the Sacramento-San Joaquin river 

system. Journal of Fisheries Management 3, 425-437. 

Stevenson J. T. & Secor D. H. (1999) Age determination and growth of Hudson River Atlantic 

sturgeon Acipenser oxyrinchus. Fishery Bulletin 98, 153-166. 

Stevick P., Allen J., Clapham P. J., Friday N., Katona S. K., Larsen F., Lien J., Mattila D. K., 

Palsbøll P. J., Sigujónsson J., Smith T. D., Øien N. & Hammond P. S. (2003a) North Atlantic 

humpback whale abundance and rate of increase four decades after protection from whaling. 

Marine Ecology Progress Series 258, 263-273. 

Stevick P. T., Allen J., Clapham P. J., Friday N., Katona S. K., Larsen F., Lien J., Mattila D. K., 

Palsbøll P. J., Sears R., Sigurjónsson J., Smith T. D., Vikingsson G., Øien J. & Hammond P. S. 

(2001) Trends in abundance of North Atlantic humpback whales, 1979-1993. Paper 
499

SC/53/NAH52 presented to the International Whaling Commission Scientific Committee. 

Available from IWC, 135 Station Road, Impington, Cambridge, UK. 

Stevick P. T., J. Allen, M. Berube, P. J. Clapham, S. K. Katona, F. Larsen, J. Lien, D. K. Mattila, 

Palsbøll, J. Robbins, J. Sigurjónsson, T. D. Smith, N. Øien & Hammond P. S. (2003b) 

Segregation of migration by feeding ground origin in North Atlantic humpback whales 

(Megaptera novaeangliae). Journal of Zoology 259, 231-237. 

Stevick P. T., Neves M. C., Johansen F., Engel M. H., Allen J., Marcondes M. C. C. & Carlson 

C. (2011) A quarter of a world away: Female humpback whale moves 10,000 km between 

breeding areas. Biology Letters 7, 299-302. 

Stewart K., Sims M., Meylan A., Witherington B., Brost B. & Crowder L. B. (2011) Leatherback 

nests increasing significantly in Florida, USA; trends assessed over 30 years using multilevel 

modeling. Ecological Applications 21, 263-273. 

Stinson M. L. (1984a) Biology of sea turtles in San Diego Bay, California, and in northeastern 

Pacific Ocean. Master’s thesis. San Diego State University, San Diego, California. 

Stinson M. L. (1984b) Biology of sea turtles in San Diego Bay, California, and in northeastern 

Pacific Ocean. San Diego State University, San Diego, California. 

Stolte L. (1981) The forgotten salmon of the Merrimack. Department of the Interior, Northeast 

Region, Washington, D.C. 

Stolz J. & Schnell J. (1991) Trout. Stackpole Books, Harrisburg, Pennsylvania. 

Stone I. (1874) Report of operations during 1872 at the United States salmon hatching 

establishment on the McCloud River. U.S. Commission on Fish and Fisheries, Report for 1872 

and 1873. Part II: 168-215. Washington, D.C. Available: 

http://penbay.org/cof/cof_1872_1873.html (April 2008). 

Storelli M., Barone M. G. & Marcotrigiano G. O. (2007a) Polychlorinated biphenyls and other 

chlorinated organic contaminants in the tissues of Mediterranean loggerhead turtle Caretta 

caretta. Science of the Total Environment 273, 456-463. 

Storelli M., Barone M. G. & Marcotrigiano G. O. (2007b) Polychlorinated biphenyls and other 

chlorinated organic contaminants in the tissues of Mediterranean loggerhead turtle Caretta 

caretta. Science of the Total Environment 273 (2-3, 456-463. 

Storelli M., Barone M. G., Storelli A. & Marcotrigiano G. O. (2008) Total and subcellular 

distribution of trace elements (Cd, Cu and Zn) in the liver and kidney of green turtles (Chelonia 

mydas) from the Mediterranean Sea. Chemosphere 70, 908-913. 

Storr J. F. (1964) Ecology and oceanography of the coral-reef tract, Abaco Island, Bahamas. 

Geological Society of America Special Paper 79. 

Stout H. A., Lawson P. W., Bottom D., Cooney T., Ford M., Jordan C., Kope R., Kruzic L., Pess 

G., Reeves G., Scheuerell M., Wainwright T., Waples R., Weitkamp L., Williams J. & Williams 

T. (2011) Scientific conclusions of the status review for Oregon Coast coho salmon 

(Oncorhynchus kisutch). Oregon Coast Coho Salmon Biological Review Team, NMFS, 

NWFSC. 

Strain E. M. A. & Johnson C. R. (2009) Competition between an invasive urchin and 

commercially fished abalone: effect on body condition, reproduction and survivorship. Marine 

Ecology-Progress Series 377, 169-182. 

Straley J. M. (1990) Fall and winter occurrence of humpback whales (Megaptera novaeangliae) 

in southeastern Alaska. Report of the International Whaling Commission Special Issue 12, 319-

323.-Individual Recognition of Cetaceans Use of Photo-Identification and Other Techniques To 

Estimate Population Parameters). 
500

Strange J. S. (2010) Upper thermal limits to migration in adult Chinook salmon: Evidence from 

the Klamath River Basin. Transactions of American Fisheries Society 139, 1091-1108. 

Strayer D. L. (1999) Effects of alien species on freshwater mollusks in North America. Journal 

of the North American Benthological Society 18, 74-98. 

Strayer D. L. (2010) Alien species in fresh waters: ecological effects, interactions with other 

stressors, and prospects for the future. Freshwater Biology 55, 152-174. 

Sturdevant M. V., Willette T. M., Jewett S. & Deberec E. (1999) Diet composition, diet overlap, 

and size of 14 species of forage fish collected monthly in PWS, Alaska, 1994-1995. (ed M. V. 

Sturdevant) pp. 12-36. 

Suárez-Morales E., Paredes-Trujillo A. & González-Solís D. (2010) The introduced asian 

parasitic copepod Neoergasilusjaponicus (Harada) (Cyclopoida: Ergasilidae) from endangered 

cichlid teleosts in Mexico. Zoological Science 27, 851-855. 

Suárez A., Dutton P. H. & Bakarbessy J. (2000) Leatherback (Dermochelys coriacea) nesting on 

the North Vogelkop Coast of Irian Jaya, Indonesia. P.260 In: Kalb, H. and T. Wibbels (eds), 19th 

Annual Symposium on Sea Turtle Conservation and Biology. 2-6 March 1999, South Padre 

Island, Texas. 

Sunderraj S. F., Joshua W. J. & Kumar V. V. (2006) Sea turtles and their nesting habitats in 

Gujarat. Pages 156-169 in Shanker, K. and B.C. Choudhury (editors). Marine Turtles of the 

Indian Subcontinent. Universities Press, India. 

Sutherland K. P., Shaban S., Joyner J. L., Porter J. W. & Lipp E. K. (2011) Human pathogen 

shown to cause disease in the threatened elkhorn coral Acropora palmata. PlosOne 6. 

Suydam R., Burns J. J. & Carroll G. (1999) Age, growth, and reproduction of beluga whales 

from the eastern Chukchi Sea, Alaska. Paper presented to the Alaska Beluga Whale Committee 

workshop, March 30-April 1, 1999. 

Swain D. P., Sinclair A. F. & Hanson J. M. (2007) Evolutionary response to size-selective 

mortality in an exploited fish population. Proceedings of the Royal Society of London Series B: 

Biological Sciences 274, 1015-1022. 

Swartz S. L., Cole T., McDonald M. A., Hildebrand J. A., Oleson E. M., Martinez A., Clapham 

P. J., Barlow J. & Jones M. (2003) Acoustic and visual survey of humpback whales (Megaptera 

novaeangliae) distribution in the Eastern and Southeastern Caribbean Sea. Caribbean Journal of 

Science 39, 195-208. 

Symons P. E. K. (1971) Behavioural Adjustment of Population Density to Available Food by 

Juvenile Atlantic Salmon. The Journal of Animal Ecology 40, 569-587. 

Szmant A. M. (1986) Reproductive ecology of Caribbean reef corals. Coral Reefs 5, 43-53. 

Takano H. (1956) Harmful blooming of minute cells of Thalassiosira decipiens in coastal water 

in Tokyo Bay. Journal of the Oceanography Society Japan 12, 63-67. 

Talavera-Saenz A., Gardner S. C., Rodriquez R. R. & Vargas B. A. (2007) Metal profiles used as 

environmental markers of green turtle (Chelonia mydas) foraging resources. Science of the Total 

Environment 373, 94-102. 

Tamburri M. N., Wasson K. & Matsuda M. (2002) Ballast water deoxygenation can prevent 

aquatic introductions while reducing ship corrosion. Biological Conservation 103, 331-341. 

Tanaka E. (2009) Estimation of temporal changes in the growth of green turtles Chelonia mydas 

in waters around the Ogasawara Islands. Fisheries Science 75, 629-639. 

Taquet C., Taquet M., Dempster T., Soria M., Ciccione S., Roos D. & Dagorn L. (2006) 

Foraging of the green sea turtle Chelonia mydas on seagrass beds at Mayotte Island (Indian 

Ocean), determined by acoustic transmitters. Marine Ecology Progress Series 306, 295-302. 
501

Taylor A. H., Jordon M. B. & Stephens J. A. (1998) Gulf Stream shifts following ENSO events. 

Nature 393, 68. 

Taylor B. L., Martien K. & Morin P. (2010) Identifying units to conserve using genetic data. In: 

Marine Mammal Ecology and Conservation: A Handbook of Techniques (eds I. L. Boyd, W. D. 

Bowen and S. J. Iverson) pp. 306-324. Oxford University Press. 

Taylor E. B. & McPhail J. D. (1985) Burst Swimming and Size-Related Predation of Newly 

Emerged Coho Salmon Oncorhynchus kisutch. Transactions of the American Fisheries Society

114, 546-551. 

Tegner M. J. (1989) The California abalone fishery: Production, ecological interactions, and 

prospects for the future. In J.F. Caddy (ed.), Marine Invertebrate Fisheries: Their Assessment 

and Management. John Wiley and Sons, Inc. 

Tegner M. J. & Levin L. A. (1982) Do sea urchins and abalones compete in California kelp 

forest communities? In J.M. Lawrence (ed.), International Echinoderm Conference. A.A. 

Balkema. Rotterdam. 

Telesnicki G. J. & Goldberg W. M. (1995) Effects of turbidity on the photosynthesis of two 

south Florida reef coral species. Bulletin of Marine Science 57, 527-539. 

Terdalkar S., Kulkarni A. S., Kumbhar S. N. & Matheickal J. (2005) Bio-economic risks of 

ballast water carried in ships, with special reference to harmful algal blooms. Nature, 

Environment and Pollution Technology 4, 43-47. 

TEWG. (1998a) An assessment of the Kemp's ridley (Lepidochelys kempii) and loggerhead 

(Caretta caretta) sea turtle populations in the Western North Atlantic. In: Proceedings of the 

Sixteenth Annual Symposium on Sea Turtle Biology and Conservation. Turtle Expert Working 

Group. 

TEWG. (1998b) An assessment of the Kemp's ridley (Lepidochelys kempii) and loggerhead 

(Caretta caretta) sea turtle populations in the western North Atlantic. A report of the Turtle 

Expert Working Group (TEWG); NOAA Technical Memorandum NMFS-SEFSC-409. 96p. 

TEWG. (2000a) Assessment update for the Kemp's ridley and loggerhead sea turtle populations 

in the western North Atlantic. NOAA Technical Memorandum NMFS-SEFSC-444. 

TEWG. (2000b) Assessment update for the Kemp's ridley and loggerhead sea turtle populations 

in the western North Atlantic. In: NOAA Technical Memorandum p. 115. Turtle Expert Working 

Group (TEWG). 

TEWG. (2007a) An assessment of the leatherback turtle population in the Atlantic Ocean. NOAA 

Technical Memorandum NMFS-SEFSC-555. 116p. 

TEWG. (2007b) An assessment of the leatherback turtle population in the Atlantic Ocean. In: 

NOAA Technical Memorandum p. 116. Turtle Expert Working Group, Department of 

Commerce. 

TGLO. (2010) Adopt a beach newletter. Texas General Land Office. 

Thomas A. & Strub P. T. (2001) Cross-shelf phytoplankton pigment variability in the California 

Current. Continental Shelf Research 21, 1157-1190. 

Thomas M. V. (1997) Diet of the Round Goby in the St. Clair River and Lake St. Clair, 1993. 

State of Michigan Department of Natural Resources. 

Thompson J. D. (1991) The biology of an invasive plant. Bioscience 41, 393-401. 

Thorbjarnarson J. B., Platt S. G. & Khaing S. T. (2000) Sea turtles in Myanmar: Past, and 

present. Marine Turtle Newsletter 88, 10-11. 

Thorpe J. E. (1977) Bimodal distribution of lengths of juvenile Atlantic salmon (Salmo salar L.) 

under artificial rearing conditions. Journal of Fish Biology 11, 175–184. 
502

Thorpe J. E. (1994) Reproductive strategies in Atlantic salmon, Salmo salar L. Aquaculture and 

Fisheries Management 25, 77-87. 

Thorpe J. E., Morgan R. I. G., Ottaway E. M. & Miles M. S. (1980) Time of divergence of 

growth groups between potential 1+ and 2+ smolts among sibling Atlantic salmon. Journal of 

Fish Biology 17:13–21. 

Thorson T. B. (1976) Observations on the reproduction of sawfish, Pristis perotteti, in Lake 

Nicaragua, with recommendations for its conservation. Pages 641-650 in: Thorson, T.B., editor. 

Investigations of the ichthyofauna of Nicaraguan lakes, University of Nebraska, Lincoln. 

Thorson T. B. (1982) Life history implications of a tagging study of the largetooth sawfish, 

Pristis perotteti, in the Lake Nicaragua-Río San Juan system. Environmental Biology of Fishes 7, 

207-228. 

Threlfall W. (1978) First record of the Atlantic leatherback turtle (Dermochelys coriacea) from 

Labrador. Canadian Field Naturalist 92(3):287. 

Tiffan K. F., Kock T. J., Haskell C. A., Connor W. P. & Steinhorst R. K. (2009) Water velocity, 

turbulence, and migration rate of subyearling fall Chinook salmon in the free-flowing and 

impounded Snake River. Transactions of the American Fisheries Society 138, 373-384. 

Tissot B. N., Hixon M. A. & Stein D. L. (2007) Habitat-based submersible assessment of macroinvertebrate and groundfish assemblages at Heceta Bank, Oregon, from 1988 to 1990. Journal of 

Experimental Marine Biology and Ecology 352, 50-64. 

Tiwari M., Bjorndal K. A., Bolten A. B. & Bolker B. M. (2005) Intraspecific application of the 

mid-domain effect model: Spatial, and temporal nest distributions of green turtles, Chelonia 

mydas, at Tortuguero, Costa Rica. Ecology Letters 8, 918-924. 

Tiwari M., Bjorndal K. A., Bolten A. B. & Bolker B. M. (2006) Evaluation of density-dependent 

processes, and green turtle Chelonia mydas hatchling production at Tortuguero, Costa Rica. 

Marine Ecology Progress Series 326, 283-293. 

Tolimieri N. & Levin P. S. (2005) The roles of fishing and climate in the population dynamics of 

bocaccio rockfish. Ecological Applications 15, 458-468. 

Tomas J., Castroviejo J. & Raga J. A. (1999) Sea turtles in the south of Bioko Island (Equatorial 

Guinea). Marine Turtle Newsletter 84, 4-6. 

Tomás J., Gozalbes P., Raga J. A. & Godley B. J. (2008) Bycatch of loggerhead sea turtles: 

insights from 14 years of stranding data. Endangered Species Research 5, 161-169. 

Tomas J. & Raga J. A. (2008) Occurrence of Kemp's ridley sea turtle ( Lepidochelys kempii) in 

the Mediterranean. Marine Biodiversity Records 1. 

Tomascik T. & Sander F. (1987) Effects of eutrophication on reef-building corals. I. Structure of 

scleractinian coral communities on fringing reefs, Barbados, West Indies. Marine Biology 94, 

53-75. 

Tomilin A. G. (1967) Mammals of the USSR, and adjacent countries. Volume 9, Cetacea. Israel 

Program Sci. Transl. No. 124, National Technical Information Service TT 65-50086. Springfield, 

Virginia (Translation of Russian text published in 1957). 

Tomy G. T., Pleskach K., Ferguson S. H., Hare J., Stern G., Macinnis G., Marvin C. H. & 

Loseto. L. (2009) Trophodynamics of some PFCs and BFRs in a western Canadian Arctic marine 

food web. Environmental Science AND Technology 43, 4076-4081. 

Torquemada Y. F., Durako M. J. & Lizaso J. L. S. (2005) Effects of salinity and possible 

interactions with temperature, and pH on growth, and photosynthesis of Halophila johnsonii 

Eiseman. Marine Biology 148, 251-260. 
503

Tourinho P. S., Sul J. A. I. d. & Fillmann G. (2009) Is marine debris ingestion still a problem for 

the coastal marine biota of southern Brazil? Marine Pollution Bulletin in press, in press. 

Trites A. W. & Donnelly C. P. (2003) The decline of Steller sea lions in Alaska: A review of the 

nutritional stress hypothesis. Mammal Review 33, 3-28. 

Trites A. W. & Larkin P. A. (1992) The status of Steller sea lion populations, and the 

development of fisheries in the Gulf of Alaska, and Aleutian Islands. Report contract 

NA17FD0177 to Pacific States Marine Fisheries Commission, Gladstone, Oregon. 

Trites A. W. & Larkin P. A. (1996) Changes in the abundance of Steller sea lions (Eumetopias 

jubatus) in Alaska from 1956 to 1992: how many were there? Aquatic Mammals 22. 

Trites A. W., Porter B. P., Deecke V. B., Coombs A. P., Marcotte M. L. & Rosen D. A. S. (2006) 

Insights into the timing of weaning and the attendance patterns of lactating Steller sea lions 

(Eumetopias jubatus) in Alaska during winter, spring, and summer. Aquatic Mammals 32, 85-97. 

Troeng S., Chacon D. & Dick B. (2004) Leatherback turtle Dermochelys coriacea nesting along 

the Caribbean coast of Costa Rica. In: Twenty-First Annual Symposium on Sea Turtle Biology 

and Conservation (eds M. S. Coyne and R. D. Clark) p. 13. 

Troëng S. & Chaloupka M. (2007) Variation in adult annual survival probability and remigration 

intervals of sea turtles. Marine Biology 151, 1721-1730. 

Troëng S., Harrison E., Evans D., Haro A. d. & Vargas E. (2007) Leatherback turtle nesting 

trends and threats at Tortuguero, Costa Rica. Chelonian Conservation and Biology 6, 117-122. 

Troëng S. & Rankin E. (2005) Long term conservation efforts contribute to positive green turtle 

Chelonia mydas nesting trend at Tortuguero, Costa Rica. Biological Conservation 121, 111–116. 

Tsikhon-Lukanina E. I. & Reznichenko O. G. (1991) Diet of the ctenophore Mnemiopsis in the 

Black Sea as a function of size. Oceanology 31, 320-323. 

Tuato'o-Bartley N., Morrell T. E. & Craig P. (1993) Status of sea turtles in American Samoa in 

1991. Pacific Science 47, 215-221. 

Tucker A. D. (2009) Nest site fidelity and clutch frequency of loggerhead turtles are better 

elucidated by satellite telemetry than by nocturnal tagging efforts: Implications for stock 

estimation. Journal of Experimental Marine Biology and Ecology in press, in press. 

Tunnicliffe V. (1981) Breakage, and propagation of the stony coral Acropora cervicornis. 

Proceedings of the National Academy of Science 78, 2427-2431. 

Turekian K. K., Harriss R. C. & Johnson D. G. (1967) The variations of Si, Cl, Na, Ca, Sr, Ba, 

Co, and Ag in the Neuse River, North Carolina. Limnology and Oceanography 12, 702-706. 

Turgeon D. D. & O'Connor T. P. (1991) Long Island Sound: Distributions, Trends, and Effects 

of Chemical Contamination. Estuaries 14, 279-289. 

Turkozan O. & Yilmaz C. (2008) Loggerhead turtles, Caretta caretta, at Dalyan Beach, Turkey: 

Nesting activity (2004–2005) and 19-year abundance trend (1987–2005). Chelonian 

Conservation and Biology 7, 178-187. 

Tutschulte T. C. (1976) The comparative ecology of three sympatric abalones. Doctoral 

dissertation. University of California, San Diego. 

Tutschulte T. C. & Connell J. H. (1988) Growth of three species of abalones (Haliotis) in 

southern California. The Veliger 31(3/4, 204-213. 

Twiss S. D., Thomas C., Poland V., Graves J. A. & Pomeroy. P. (2007) The impact of climatic 

variation on the opportunity for sexual selection. Biology Letters 3, 12-15. 

Tyack P. L. (1981) Interactions between singing Hawaiian humpback whales and conspecifics 

nearby. Behavioral Ecology and Sociobiology 8, 105-116. 
504

Uchida I. & Nishiwaki M. (1995) Sea turtles in the waters adjacent to Japan. In: Biology and 

Conservation of Sea Turtles (ed K. A. Bjorndal) pp. 317-319. Smithsonian Institution Press, 

Washington, D.C. 

Upper Columbia Salmon Recovery Board. (2007) Final Upper Columbia spring Chinook salmon 

and steelhead recovery plan. In: Recovery Plan. National Marine Fisheries Service and U.S. Fish 

and Wildlife Service. 

Urban J., Jaramillo A., Aguayo A., Guevara P. L. D., Salinas M., Alvarez C., Medrano L., 

Jacobsen J., Balcomb K., Claridge D., Calambokidis J., Steiger G., Straley J., Ziegesar O. V., 

Waite J. M., Dahlheim M., Darling J. & Baker. C. S. (2000) Migratory destinations of humpback 

whales wintering in the Mexican Pacific. (Megaptera novaeangliae). Journal of Cetacean 

Research and Management 2, 101-110. 

USACE. (2003) Merrimack River Watershed Assessment Study: Description of Existing 

Conditions. p. 108. U.S. Army Corps of Engineers, New England District. 

USACOE. (2010) Sea turtle data warehouse. U.S. Army Corps of Engineers. 

USAF. (1996) Sea turtles in the Gulf. Air Force Material Command, Eglin Air Force Base. 

USASAC. (2004) Annual Report of the U.S. Atlantic Salmon Assessment Committee United 

States Atlantic Salmon Assessment Committee. 

USCG. (2003) 2002 national recreational boating survey state data report. United States Coast 

Guard, Columbus, Ohio. 

USCG. (2005) Boating statistics—2005. United States Coast Guard, Washignton D.C. 

USCOP. (2004) An ocean blueprint for the 21st century. Final report. U.S. Commission on 

Ocean Policy, Washington, D. C. 

USDC. (2009a) Endangered and threatened species; recovery plans for Lake Ozette sockeye 

salmon. Federal Register 74, 25706-25710. 

USDC. (2009b) Endangered and threatened wildlife and plants: Final rulemaking to designate 

critical habitat for the threatened southern distinct population segment of North American green 

sturgeon. Federal Register 74, 52300-52351. 

USFWS. (1999) South Florida multi-species recovery plan. United States Fish and Wildlife 

Service, Atlanta, Georgia. 

USFWS. (2000) Report on the Mexico/United States of America population restoration project 

for the Kemp's ridley sea turtle, Lepidochelys kempii, on the coasts of Tamaulipas, and Veracruz, 

Mexico. United States Fish and Wildlife Service. 

USFWS. (2001) Report on the Mexico/United States of America population restoration project 

for the Kemp's ridley sea turtle, Lepidochelys kempii, on the coasts of Tamaulipas, and Veracruz, 

Mexico. United States Fish and Wildlife Service. 

USFWS. (2002) Report on the Mexico/United States of America population restoration project 

for the Kemp's ridley sea turtle, Lepidochelys kempii, on the coasts of Tamaulipas, and Veracruz, 

Mexico. United States Fish and Wildlife Service. 

USFWS. (2003) Report on the Mexico/United States of America population restoration project 

for the Kemp's ridley sea turtle, Lepidochelys kempii, on the coasts of Tamaulipas, and Veracruz, 

Mexico. United States Fish and Wildlife Service. 

USFWS. (2004) Report on the Mexico/United States of America population restoration project 

for the Kemp's ridley sea turtle, Lepidochelys kempii, on the coasts of Tamaulipas, and Veracruz, 

Mexico. United States Fish and Wildlife Service. 
505

USFWS. (2005a) Report on the Mexico/United States of America population restoration project 

for the Kemp's ridley sea turtle, Lepidochelys kempii, on the coasts of Tamaulipas, and Veracruz, 

Mexico. United States Fish and Wildlife Service. 

USFWS. (2006) Report on the Mexico/United States of America population restoration project 

for the Kemp's ridley sea turtle, Lepidochelys kempii, on the coasts of Tamaulipas, and Veracruz, 

Mexico. United States Fish and Wildlife Service. 

Usfws. (2007) Restoring migratory fish to the Connecticut River basin. Available: 

http://www.fws.gov/r5crc/Habitat/Contaminants.html (February 2008). 

USFWS & NMFS. (1992) Recovery plan for the Kemp's ridley sea turtle (Lepidochelys kempii). 

National Marine Fisheries Service, St. Petersburg, Florida. 

USFWS N. a. (1995) Status reviews for sea turtles listed under the Endangered Species Act of 

1973. National Marine Fisheries Service, Silver Spring, Maryland. 

USFWS N. a. (1998) Recovery Plan for U.S. Pacific Populations of the Loggerhead Turtle 

(Caretta caretta). National Marine Fisheries Service, Silver Spring, Maryland. 

USFWS N. a. (2005b) Final recovery plan for the Gulf of Maine distinct population segment of 

Atlantic Salmon (Salmo salar). NMFS and USFWS, SIlver Spring, Maryland and Hadley, 

Massachussetts. 

Usgs. (2004) Evaluation of passage performance of adult American shad at lower Connecticut 

River mainstream fish passage facilities. U.S. Department of the Interior, U.S. Geological 

Survey, project 09042 Kearneysville, West Virginia. Available: 

http://www.lsc.usgs.gov/CAFL/Fish%20Passage/Projects/09042%20Turners%20Fishways/Turn

ers%20Fishways.htm (February 2008). 

USGS. (2005) Continuous resistivity profiling data from the Upper Neuse River Estuary, North 

Carolina, 2004-2005. United States Geological Survey. 

USGS. (2010) Hypoxia in the Gulf of Mexico. 

USN. (2006) Marine resource assessment for the Pacific Northwest operating area. U.S. Navy, 

Pearl Harbor, Hawaii. 

USN. (2008) Biological evaluation for the Gulf of Mexico rangle complex. U.S. Navy. 

USN. (2009) Gulf of Mexico range complex final environmental impact statement/overseas 

environmental impact statement (EIS/OEIS) volume 1 (version 3). United States Navy, Norfolk, 

Virginia. 

Vaclavik T. & Meentemeyer R. K. (2009) Invasive species distribution modeling (iSDM): are 

absence data and dispersal constraints needed to predict actual distributions. Ecological 

Modelling 220, 3248-3258. 

Van Banning P. (1987) Further results of the Bonamia ostreae challenge tests in Dutch oyster 

culture. Aquaculture 67, 191-194. 

van de Merwe J. P. V., Hodge M., Olszowy H. A., Whittier J. M., Ibrahim K. & Lee S. Y. (2009) 

Chemical Contamination of Green Turtle (Chelonia mydas) Eggs in Peninsular Malaysia: 

Implications for Conservation and Public Health. Environmental Health Perspectives 117, 1397-

1401. 

Van der Elst R. (1981) A guide to the common sea fisheries of southern Africa. Editor C. Struik, 

Cape Town, South Africa. 

Van der Kraak G. & Pankhurst N. W. (1997) Temperature effects on the reproductive 

performance of fish. Pages 159-176 in: McDonald, D.G. and C.M. Wood. Global warming: 

Implications for freshwater and marine fish. Society for Experimental Biology Seminar Series 

61. 
506

Van Eenennaam J., Doroshov S., Moberg G., Watson J., Moore D. & Linares J. (1996) 

Reproductive conditions of the Atlantic sturgeon (Acipenser oxyrinchus) in the Hudson River. 

Estuaries and Coasts 19, 769-777. 

Van Eenennaam J. P. & Doroshov S. I. (1998) Effects of age and body size on gonadal 

development of Atlantic sturgeon. Journal of Fish Biology 53, 624-637. 

Van Eenennaam J. P., Linares J., Doroshov S. I., Hillemeier D. C., Willson T. E. & Nova A. A. 

(2006) Reproductive conditions of the Klamath River green sturgeon. Transactions of the 

American Fisheries Society 135, 151-163. 

Van Engel W. A., Dillon W. A., Zwerner D. & Eldridge D. (1966) Loxothylacus panopaei 

(Cirripedia, Sacculinidae), an instroduced parasite on xanthid crab in Chesapeake Bay, U.S.A. . 

Crustaceana 10, 111-112. 

Van Scheppingen W. B., Verhoeven A. J. I. M., Mulder P., Addink M. J. & Smeenk C. (1996) 

Polychlorinated-biphenyls, dibenzo-p-dioxins, and dibenzofurans in harbor porpoises Phocoena 

phocoena stranded on the Dutch coast between 1990 and 1993. Archives of Environmental 

Contamination and Toxicology 30, 492-502. 

Vander Zanden H. B., Bjorndal K. A., Reich K. J. & Bolten A. B. (2010) Individual specialists in 

a generalist population: results from a long-term stable isotope series. Biology Letters in press, 

in press. 

Vanderlaan A. S. M., Taggart C. T., Serdynska A. R., Kenney R. D. & Brown M. W. (2008a) 

Reducing the risk of lethal encounters: vessels and right whales in the Bay of Fundy and on the 

Scotian Shelf. Endangered Species Research 4, 283-283. 

Vanderlaan A. S. M., Taggart C. T., Serdynska A. R., Kenney R. D. & Brown M. W. (2008b) 

Reducing the risk of lethal encounters: Vessels and right whales in the Bay of Fundy and on the 

Scotian Shelf. Endangered Species Research 4, 283-297. 

Vanderploeg H. A., Liebig J. R., Carmichael W. W., Agy M. A., Johengen T. H., Fahnenstiel G. 

L. & Nalepa T. F. (2001) Zebra mussel (Dreissena polymorpha) selective filtration promoted 

toxic Microcystis blooms in Saginaw Bay (Lake Huron) and Lake Erie. Canadian Journal of 

Fisheries and Aquatic Sciences 58, 1208-1221. 

Varanasi U., Casillas E., Arkoosh M. R., Hom T., Misitano D. A., Brown D. W., Chan S. L., C. 

T. K., McCain B. B. & Stein J. E. (1993a) Contaminant exposure and associated biological 

effects in juvenile Chinook salmon (Oncorhynchus tshawytscha) from urban and nonurban 

estuaries of Puget Sound. NOAA Technical Memorandum NMFS-NWFSC-8. Seattle, 

Washington. 

Varanasi U., Stein J. E. & Nishimoto M. (1989) Biotransformation and disposition of polycyclic 

aromatic hydrocarbons (PAH) in fish. In: Metabolism of polycyclic aromatic hydrocarbons in the 

aquatic environment (ed U. Varanasi) pp. 94-149. CRC Press, Boca Raton, Florida. 

Varanasi U. E., Casillas M. R., Arkoosh T., Hom D. A., Misitano D. W., Brown S. L., Chan T. 

K. C., McCain B. B. & Stein J. E. (1993b) Contaminant exposure and associated biological 

effects in juvenile Chinook salmon (Oncorhynchus tshawytscha) from urban and nonurban 

estuaries of Puget Sound. NOAA Technical Memorandum NMFS-NWFSC-8. Seattle, 

Washington. 

Vargo S., Lutz P., Odell D., Vleet E. V. & Bossart G. (1986a) Study of the effects of oil on 

marine turtles. Minerals Management Service, Vienna, Virginia. 

Vargo S., Lutz P., Odell D., Vleet E. V. & Bossart G. (1986b) Study of the effects of oil on 

marine turtles. Minerals Management Service, Vienna, Virginia. 
507

Vargo S., Lutz P., Odell D., Vleet E. V. & Bossart G. (1986c) Study of the effects of oil on 

marine turtles. Minerals Management Service, Vienna, Virginia. 

Vaughan T. W. (1915) The geological significance of the growth rate of the Floridian and 

Bahamian shoal-water corals. Journal of the Washington Academy of Science 5, 591-600. 

VDCR. (2006) James River watershed. Virginia Department of Conservation and Recreation. 

Velez-Zuazo X., Ramos W. D., Van Dam R., Diez C. E., Arreu-Grobois A. & McMillan O. 

(2008) Dispersal, recruitment and migratory behavior in a hawksbill sea turtle aggregation. 

Molecular Ecology 17, 839-853. 

Vera V. (2007) Nesting of green turtles in Aves Island Wildlife Refuge. 2006 season. In: Twentyseventh Annual Symposium on Sea Turtle Biology and Conservation (eds M. Frick, A. 

Panagopoulou, A. F. Rees and K. Williams) p. 275. International Sea Turtle Society, Myrtle 

Beach, South Carolina. 

Veron J. E. N. (2000) Corals of the World. Australian Institute of Marine Science 1:463. 

Vetter R. D. & Lynn E. A. (1997) Bathymetric demography, enzyme activity patterns, and 

bioenergetics of deep-living scorpaenid fishes (genera Sebastes and Sebatolobus): paradigms 

revisited. Marine Ecology Progress Series 155, 173-188. 

Viant M. R., Walton J. H. & Tjeerdema R. S. (2001) Comparative sublethal actions of 3-

trifluoromethyl-4-nitrophenol in marine molluscs, as measured by in vivo P-31 NMR. Pesticide 

Biochemistry and Physiology 71, 40-47. 

Villegas-Amtmann S. & Costa. D. P. (2010) Oxygen stores plasticity linked to foraging 

behaviour and pregnancy in a diving predator, the Galapagos sea lion. Functional Ecology 24, 

785-795. 

Vincent-Lang D. S. & Queral I. (1984) Eulachon spawning in the lower Susitna River. (ed C. C. 

E. a. D. S. Vincent-Lang). Alaska Department of Fish and Game, Anchorage, Alaska. 

Vinogradov M. Y., Shuskina E. A., Musayeva E. I. & Sorokin P. Y. (1989) A newly acclimated 

species in the Black Sea: the ctenophore Mnemiopsis leidyi (Ctenophora: Lobata). Oceanology

29, 220-224. 

Virnstein R. W. & Hall L. M. (2009) Northern range extension of the seagrasses Halophila 

johnsonii and Halophila decipienis along the east coast of Florida, USA. Aquatic Botany 90, 89-

92. 

Virnstein R. W., Hayek L. A. C. & Morris L. J. (2009) Pulsating patches: a model for the spatial 

and temporal dynamics of the threatened seagrass Halophila johnsonii. Marine Ecology-Progress 

Series 385, 97-109. 

Virnstein R. W. & Morris L. J. (2007) Distribution and abundance of Halophila johnsonii in the 

Indian River Lagoon: An update. Technical Memorandum # 51. St. Johns River Water 

Management District, Palatka, Florida. 

Virnstein R. W., Morris L. J., Miller J. D. & Miller-Myers R. (1997) Distribution and abundance 

of Halophila johnsonii in the Indian River Lagoon. Technical Memorandum # 24. St. Johns River 

Water Management District, Palatka, Florida. 

Visbeck M. (2002) The ocean’s role in Atlantic climate variability. Science 297, 2223-2225. 

Visser I. N. (1999) Propeller scars on and known home range of two orca (Orcinus orca) in New 

Zealand waters. New Zealand Journal of Marine and Freshwater Research 33, 635-642. 

Visser I. N. & Fertl. D. (2000) Stranding of a New Zealand killer whale (Orcinus orca) and 

information on post-stranding sightings, including a probable boat strike of the individual. 

European Research on Cetaceans 14, 208-209. Proceedings of the Fourteenth Annual 
508

Conference of the European Cetacean Society. P. G. H. Evans, R. Pitt-Aiken & E. Rogan-Eds.). 

Cork, Ireland, 202-205 April. 

Vollmer S. (2002) Genetics of Acropora cervicornis. In: Caribbean Acropora workshop: 

Potential application of the U.S. Endangered Species Act as a conservation strategy (ed A. W. 

Bruckner) pp. 163-164. 

Vollmer S. V. & Palumbi S. R. (2007) Restricted gene flow in the Caribbean staghorn coral 

Acropora cervicornis: Implications for the recovery of endangered reefs. Journal of Heredity 98, 

40-50. 

Von Halle B. & Simberloff D. (2005) Ecological Resistance to Biological Invasion 

Overwhelmed by Propagule Pressure. Ecology 86, 3212-3218. 

Waite J. (2003) Cetacean Survey. National Marine Mammal Laboratory (NMML), Cetacean 

Assessment and Ecology Program, Quarterly Report. Available online at: 

http://www.afsc.noaa.gov/Quarterly/jas2003/divrptsNMML2.htm Accessed 5/30/08. 

Waite J. M., Dahlheim M. E., Hobbs R. C., Mizroch S. A., von Ziegesar-Matkin O., Straley J. 

M., Herman L. M. & Jacobsen J. (1999) Evidence of a feeding aggregation of humpback whales 

(Megaptera novaeangliae) around Kodiak Island, Alaska. Marine Mammal Science 15, 210-220. 

Waknitz & F. (2003) Interactions of Atlantic salmon in the Pacific Northwest IV. Impacts on the 

local ecosystems. Fisheries Research 62, 307-328. 

Waldman J. R. & Wirgin I. I. (1998) Status and restoration options for Atlantic sturgeon in North 

America. Conservation Biology 12, 631-638. 

Waldron K. D. (1968) Early larvae of the canary rockfish, Sebastodes pinniger. Journal of the 

Fisheries Research Board of Canada 25, 801-803. 

Wall G. R., Riva-Murray K. & Phillips P. J. (1998) Water Quality in the Hudson River Basin, 

New York and Adjacent States, 1992-95. p. 37. U. S. Geological Survey Water Resources 

Division. 

Wallace B. P., Avens L., Braun-McNeill J. & McClellan C. M. (2009) The diet composition of 

immature loggerheads: Insights on trophic niche, growth rates, and fisheries interactions. Journal 

of Experimental Marine Biology and Ecology 373, 50-57. 

Wallace B. P., Lewison R. L., McDonald S. L., McDonald R. K., Kot C. Y., Kelez S., Bjorkland 

R. K., Finkbeiner E. M., Helmbrecht S. r. & Crowder L. B. (2010) Global patterns of marine 

turtle bycatch. Convervation Letters in press, in press. 

Wallace J. R. (2001) Status of the yelloweye rockfish resource in 2001 for northern California 

and Oregon waters. Washignton Department of Fish and Wildlife, Montesano, Washignton. 

Wallace J. R. (2007) Update to the status of yelloweye rockfish (Sebastes ruberrimus) off the 

U.S. West Coast in 2007. Pacific Fishery Management Council, Portland, Oregon. 

Wallentinus I. (2002) Introduced marine algae and vascular plants in European aquatic 

environments. In: Invasive aquatic species of Europe: Distribution, impacts and management

(eds E. Leppakoski, S. Gollasch and S. Olenin) pp. 27-52. 

Wallentinus I. (2006a) Alien Species Alert: Undaria pinnatifda (wakame or Japanese kelp). ICES 

WGITMO Report 2006, Annex 9, ICES Advisory Committee on the Marine Environment, 292-

328. 

Wallentinus I. (2006b) Introductions and transfers of plants. In: Status of Introductions of Nonindigenous Marine Species to North Atlantic and Adjacent Waters According to National 

Reports Considered at Meetings of the Working Group on Introductions and Transfers of Marine 

Organisms 1992–2002 (eds S. Gollasch, D. Kieser, D. Minchin and I. Wallentinus). 
509

Wallentinus I. & Nyberg C. D. (2007) Introduced marine organisms as habitat modifiers. Marine 

Pollution Bulletin 55, 323-332. 

Wambiji N., Gwada P., Fondo E., Mwangi S. & Osore M. K. (2007) Preliminary results from a 

baseline survey of the port of Mombasa: with focus on molluscs. In: 5th Western Indian Ocean 

Marine Science Association Scientific Symposium; Science, Policy and Management pressures 

and responses in the Western Indian Ocean region, Durban, South Africa. 

Wang J., Hulck K., Hong S.-M., Atkinson S. & Li Q. X. (2011) Accumulation and maternal 

transfer of polychlorinated biphenyls in Steller sea lions (Eumetopias jubatus) from Prince 

William Sound and the Bering Sea, Alaska. Environmental Pollution 159, 71-77. 

Wang Y., Kang W. L., Xin S. G. & Xing W. Q. (2009) Analysis Study of Trace Elements in 

Abalone and Sea Cucumber. Spectroscopy and Spectral Analysis 29, 511-514. 

Waples R. S., Aebersold P. B. & Winans G. A. (1997) Population genetic structure and life 

history variability in Oncorhynchus nerka from the Snake River basin. Report to Bonneville 

Power Administration, Contract No. 1993BP05326, Project No. 199306800. BPA Report 

DOE/BP-05326-1. 

Waples R. S., Johnson O. W. & R.P. Jones, Jr. (1991) Status review for Snake River sockeye 

salmon. U.S. Department of Commerce, NOAA Technical Memorandum NMFS F/NWC-195. 

Wapnick C. M., Precht W. F. & Aronson R. B. (2004) Millennial-scale dynamics of staghorn 

coral in Discovery Bay, Jamaica. Ecology Letters 7, 354-361. 

Ward E. J., Holmes E. E. & Balcomb. K. C. (2009a) Quantifying the effects of prey abundance 

on killer whale reproduction. Journal of Applied Ecology 46, 632-640. 

Ward E. J., Parsons K., Holmes E. E., III K. C. B. & Ford. J. K. B. (2009b) The role of 

menopause and reproductive senescence in a long-lived social mammal. Frontiers in Zoology 6. 

Ward G. M. P. M. H. & Ward A. K. (2005) Gulf Coast rivers of the southeastern United States. 

Pages 125-179 in A.C. Benke, and C.E. Cushing, editor. Rivers of North America. Elsevier 

Academic Press, Burlington, Massachusetts. Available: 

http://books.google.com/books?id=faOU1wkiYFIC&pg=RA3-PA541&lpg=RA3-

PA541&dq=pacific+coast+rivers+of+the+coterminous+united+states&source=web&ots=-

pMpyECFaA&sig=FkGrIiwgkfDyHxXCWXRalK_XSvU#PPR1,M1 (February 2008). 

Waring C. P. & Moore A. (1998) Mechanistic effects of a triazine pesticide on reproductive 

endocrine function in mature male Atlantic salmon (Salmo salar) parr. Pesticide Biochemistry 

and Physiology 62, 41-50. 

Waring G. T., Josephson E., Fairfield C. P. & (Eds). K. M.-F. (2009) U.S. Atlantic and Gulf of 

Mexico Marine Mammal Stock Assessments -- 2008. NOAA Technical Memorandum NMFSNE-210. 440pp. 

Waring G. T., Pace R. M., Quintal J. M., Fairfield C. P. & (Eds.). K. M.-F. (2004a) U.S. Atlantic 

and Gulf of Mexico Marine Mammal Stock Assessments -- 2003. NOAA Technical 

Memorandum NMFS-NE-182. 287pp. 

Waring G. T., Pace R. M., Quintal J. M., Fairfield C. P. & Maze-Foley K. (2004b) U.S. Atlantic 

and Gulf of Mexico Marine Mammal Stock Assessments - 2003. NOAA Technical Memorandum 

NMFS-NE-182, Woods Hole, Massachusetts, 300p. 

Waring G. T., Quintal J. M., Swartz S. L., Clapham P. J., Cole T. V. N., Fairfield C. P., Hohn A., 

Palka D. L., Rossman M. C. & Yeung C. (2001) U.S. Atlantic and Gulf of Mexico Marine 

Mammal Stock Assessments - 2001. NOAA Technical Memorandum NMFS-NE-168, Woods 

Hole, Massachusetts. 318p. 
510

Warkentine B. E. & Rachlin J. W. (2010) The first record of Palaemon macrodactylus (oriental 

shrimp) from the eastern coast of North America. Northeastern Naturalist 17, 91-102. 

Washington Department of Fisheries W. D. o. W. & Western Washington Treaty Indian T. 

(1993) 1992 Washington State salmon and steelhead stock inventory (SASSI). Washington 

Department of Fish and Wildlife, Olympia, Washington. 

Washington P. M., Gowan R. & Ito D. H. (1978) A biological report on eight species of rockfish 

(Sebastes spp.) from Puget Sound, Washington. In: Northwest and Alaska Fisheries Center 

Processed Report. National Marine Fisheries Service, Seattle, Washington. 

Watkins W. A. (1986) Whale Reactions to Human Activities in Cape-Cod Waters. Marine 

Mammal Science 2, 251-262. 

Watkins W. A. & Schevill W. E. (1976) Right whale feeding and baleen rattle. Journal of 

Mammalogy 57, 58-66. 

Watkins W. A. & Schevill W. E. (1979) Aerial observation of feeding behavior in four baleen 

whales: Eubalaena glacialis, Balaenoptera borealis, Megaptera novaeangliae, and Balaenoptera 

physalus. Journal of Mammalogy 60, 155-163. 

Watt W. D. (1981) Present and potential effects of acid precipitation on the Atlantic salmon of 

eastern Canada. International. Atlantic Salmon Foundation Special Publication 10, 39-46. 

Watt W. D., Scott C. D. & White W. J. (1983) Evidence of acidification of some Nova Scotian 

rivers and its impact on Atlantic salmon, Salmo salar. Canadian Journal of Fisheries and 

Aquatic Sciences 40, 462-473. 

Watt W. D., Scott C. D., Zamora P. J. & White W. J. (2000) Acid toxicity levels in Nova Scotian 

Rivers have not declined in synchrony with the decline in sulfate levels. Water Air and Soil 

Pollution 118(3-4, 203-229. 

Waycott M., Duarte C. M., Carruthers T. J. B., Orth R. J., Dennison W. C., Olyarnik S., 

Calladine A., Fourqurean J. W., Kenneth L. Heck J., Hughes A. R., Kendrick G. A., Kenworthy 

W. J., Short F. T. & Williams S. L. (2009) Accelerating loss of seagrasses across the globe 

threatens coastal ecosystems. Proceedings of the National Academy of Sciences 106, 12377-

12381. 

Waycott M. B., Longstaff J. & Mellors J. (2005) Seagrass population dynamics and water quality 

in the Great Barrier Reef region: A review and future research directions. Marine Pollution 

Bulletin 51, 343-350. 

WDF. (1975-1986) Washington state sport catch reports. Washington Department of Fisheries. 

WDF (Washington Department of Fisheries) W. W. D. o. W., and WWTT (Western Washington 

Treaty Indian Tribes), . (1993) 1992 Washington state salmon and steelhead stock inventory 

(SASSI) available from Washington Deparmtnet of Fish and Wildlfie, 600 Captiol Way N., 

Olympia, Washington 98501. 

WDFW. (2007) 2007 Commercial Landed Catch. 

WDFW. (2009) Fish passage and surface water diversion screening assessment and prioritization 

manual. Washington Department of Fish and Wildlife, Olympia, Washington. 

WDFW & ODFW. (2001) Washington and Oregon eulachon management plan. Washington 

Department of Fish and Wildlife and Oregon Department of Fish and Wildlife. 

Webb M. A., Feist H. G. W., Fitzpatrick M. S., Foster E. P., Schreck C. B., Plumlee M., Wong 

C. & Gundersen D. T. (2006) Mercury concentrations in gonad, liver, and muscle of white 

sturgeon Acipenser transmontanus in the Lower Columbia River. Archives of Environmental 

Contamination and Toxicology 50, 443–451. 
511

Webber J. D., Chun S. N., MacColl T. R., Mirise L. T., Kawabata A., Anderson E. K., Cheong T. 

S., Kavvas L., McRotondo M. G., Hochgraf K. L., Churchwell R. & Cech J. J. (2007) Upstream 

Swimming Performance of Adult White Sturgeon: Effects of Partial Baffles and a Ramp. 

Transactions of the American Fisheries Society 136, 402-408. 

Weber W. (1996a) Population size and habitat use of shortnose sturgeon, Acipenser 

brevirostrum, in the Ogeechee River system. University of Georgia, Athens. 

Weber W. (1996b) Population size and habitat use of shortnose sturgeon, Acipenser 

brevirostrum, in the Ogeechee River system, Georgia. Master’s thesis. University of Georgia, 

Athens, Georgia. 

Wei G., McCulloch M. T., Mortimer G., Deng W. & Xie L. (2009) Evidence for ocean 

acidification in the Great Barrier Reef of Australia. Geochimica Et Cosmochimica Acta 73, 

2332-2346. 

Weijerman M. L., Tienen H. G. v., Schouten A. D. & Hoekert W. E. J. (1998) Sea turtles of 

Galibi, Suriname. In: Proceedings of the Sixteenth Annual Symposium on Sea Turtle Biology and 

Conservation. (eds R. Byles and Y. Fernandez) pp. 142-144. 

Weinberg K. L. (1994) Rockfish assemblages of the middle shelf and upper slope off Oregon 

and Washington. Fishery Bulletin 92, 620-632. 

Weinrich M. T., Bove J. & Miller N. (1993) Return and survival of humpback whale (Megaptera 

novaeangliae) calves born to a single female in three consecutive years. Marine Mammal 

Science 9, 325-328. 

Weinrich M. T., Lambertsen R. H., Belt C. R., Schilling M., Iken H. J. & Syrjala S. E. (1992) 

Behavioral reactions of humpback whales Megaptera novaeangliae to biopsy procedures. 

Fishery Bulletin 90, 588-598. 

Weir C. R. (2007) Observations of marine turtles in relation to seismic airgun sound off Angola. 

Marine Turtle Newsletter 116, 17-20. 

Weisbrod A. V., D. S., J. M. M. & J. S. J. (2000) Organochlorine exposure and bioaccumulation 

in the endangered northwest Atlantic right whale (Eubalaena glacialis) population. 

Environmental Toxicology and Chemistry 19, 654–666. 

Weitkamp L. A. (2009) Marine distributions of Chinook salmon from the West Coast of North 

America determined by coded wire tag recoveries. Transactions of American Fisheries Society

139, 147-170. 

Weitkamp L. A., Wainwright T. C., Bryant G. J., Milner G. B., Teel D. J., Kope R. G. & Waples 

R. S. (1995a) Status review of coho salmon from Washington, Oregon, and California. NMFS, 

NWFSC, Seattle, Washignton. 

Weitkamp L. A., Wainwright T. C., Bryant G. J., Milner G. B., Teel D. J., Kope R. G. & Waples 

R. S. (1995b) Status review of coho salmon from Washington, Oregon, and California. 

Northwest Fisheries Science Center, Seattle, Washington. 

Welch D. W., Ishida Y. & Nagasawa K. (1998) Thermal limits and ocean migrations of sockeye 

salmon (Oncorhynchus nerka): long-term consequences of global warming. Canadian Journal of 

Fisheries and Aquatic Sciences 55, 937-948. 

Welsh S. A., Mangold M. F., Skjeveland J. E. & Spells A. J. (2002) Distribution and movement 

of shortnose sturgeon (Acipenser brevirostrum) in Chesapeake Bay. Estuaries 25, 101-104. 

Wentz D. A., Bonn B. A., Carpenter K. D., Hinkle S. R., Janet M. L., Rinella F. A., Uhrich M. 

A., Waite I. R., Laenen A. & Bencala K. E. (1998) Water quality in the Willamette Basin, 

Oregon, 1991-95. USFWS. 
512

Westlake R. L. & Gilmartin W. G. (1990) Hawaiian monk seal pupping locations in the 

Northwestern Hawaiian Islands. Pacific Science 44, 366-383. 

Westrheim S. J. (1975) Reproduction, maturation, and identification of larvae of some Sebastes 

(Scorpaenidae) species in the northeast Pacific Ocean. Journal of the Fisheries Research Board 

of Canada 32, 2399-2411. 

Westrheim S. J. & Harling W. R. (1975) Age-length relationships for 26 scorpaenids in the 

northeast Pacific Ocean. Fisheries and Marine Service Research Division, Nanaimo, British 

Columbia. 

Whalen K. G., Parrish D. L., Mather M. E. & McMenemy J. R. (2000) Cross-tributary analysis 

of parr to smolt recruitment of Atlantic salmon (Salmo salar). Canadian Journal of Fisheries and 

Aquatic Sciences 57, 1607-1616. 

Wheaton J. W. & Jaap W. C. (1988) Corals and other prominent benthic cnidaria of Looe Key 

National Marine Sanctuary, FL. Florida Marine Research Publication 43. 

Wheeler P. A. & Hill J. (1999) Biological effects of the 1997-1998 El Niño event off Oregon: 

Nutrient and chlorophyll distributions. In: Proceedings of the 1998 Science Board Symposium on 

the impacts of the 1997/98 El Niño event on the North Pacific Ocean and its marginal seas (eds 

H. J. Freeland, W. T. Peterson and A. Tyler). North Pacific Marine Science Organization 

(PICES): PICES Scientific Report No. 10. 

White H. C. (1942) Atlantic salmon redds and artificial spawning beds. Journal of the Fisheries 

Research Board of Canada 6, 37-44. 

Whitehead H. (1997) Sea surface temperature and the abundance of sperm whale calves off the 

Galapagos Islands: Implications for the effects of global warming. Report of the International 

Whaling Commission 47, 941-944.-Sc/948/O930). 

Whitehead H. & Moore M. J. (1982) Distribution, and movements of West Indian humpback 

whales in winter. Canadian Journal of Zoology 60, 2203-2211. 

Whiteman K. W., Travnichek V. H., Wildhaber M. L., DeLonay A., Papoulias D. & Tillett D. 

(2004) Age Estimation for Shovelnose Sturgeon: A Cautionary Note Based on Annulus 

Formation in Pectoral Fin Rays. North American Journal of Fisheries Management 24, 731-734. 

Whitman Miller A., Chang A. L., Cosentino-Manning N. & Ruiz G. M. (2004) A New Record 

and Eradication of the Northern Atlantic Algaascophyllum Nodosum(Phaeophyceae) from San 

Francisco Bay, California, USA. Journal of Phycology 40, 1028-1031. 

Whitney F. A., Mackas D. L., Welch D. W. & Robert M. (1999) Impact of the 1990s El Niños on 

nutrient supply and productivity of Gulf of Alaska waters. In: Proceedings of the 1998 Science 

Board Symposium on the impacts of the 1997/98 El Niño event on the North Pacific Ocean and 

its marginal seas. PICES Scientific Report No. 10. 

Wibbels T. (2003) Critical approaches to sex determination in sea turtle biology and 

conservation. In: Biology of sea turtles (eds P. Lutz, J. Musik and J. Wynekan) pp. 103-134. 

CRC Press. 

Wibbels T. (2007) Sex determination and sex ratio in ridley turtles. Pages 167-189 in: Plotkin 

P.T., editor. Biology and conservation of ridley sea turtles. Johns Hopkins University Press, 

Baltimore, Maryland. 

Wibbels T., Marion K., Nelson D., Dindo J. & Geis A. (2005) Evaluation of the bay systems of 

Alabama (US) as potential foraging habitat for juvenile sea turtles. Pages 275-276 in: Mosier, A., 

A. Foley, and B. Brost, editors. Proceedings of the Twentieth Annual Symposium on Sea Turtle 

Biology and Conservation. NOAA Technical Memorandum NMFS-SEFSC-477. 
513

Wiese F. I. & Robertson G. J. (2004) Assessing seabird mortality froni chronic oil discharges at 

sea. Journal of Wildlife Management 68, 627-638. 

Wigand C., Stevenson J. C. & Cornwell J. C. (1997) Effects of different submersed macrophytes 

on sediment biogeochemistry. Aquatic Botany 56, 233-244. 

Wilcove D. S., Rothstein D., Dubow J., Phillips A. & Losos E. (1998) Quantifying threats to 

imperiled species in the United States. Bioscience 48, 607-615. 

Wiley D. N., Moller J. C., R. M. Pace I. & Carlson C. (2008) Effectiveness of voluntary 

conservation agreements: Case study of endangered whales and commercial whale watching. 

Conservation Biology 22, 450-457. 

Wiley T. R. & Simpfendorfer C. A. (2007) The ecology of elasmobranchs occurring in the 

Everglades National Park, Florida: implications for conservation and management. Bulletin of 

Marine Science 1980, 171-189. 

Wilkinson C. (2000) Status of coral reefs of the world: 2000. Global Coral Reef Monitoring 

Network, Australian Institute of Marine Science. 

Wilkinson C. & Souter D. (2008) Status of Caribbean coral reefs after bleaching and hurricanes 

in 2005. Global Coral Reef Monitoring Network, and Reef and Rainforest Research Centre, 

Townsville. 

Williams E. H. & Bunkley-Williams L. (1990) The world-wide coral reef bleaching cycle and 

related sources of coral mortality. Atoll Research Bulletin 335, 1-71. 

Williams E. H., Ralston S., MacCall A. D., Woodbury D. & Pearson D. E. (1999) Stock 

assessment of the canary rockfish resource in the waters off southern Oregon and California in 

1999. In: Status of the Pacific coast groundfish fishery through 1999 and recommended 

acceptable biological catches for 2000 appendix: stock assessments p. 75. Pacific Fishery 

Management Council, Portland, Oregon. 

Williams K. L., Steyermark A. C., Hurd C. C., Schwandt A. J., Paladino F. V. & Spotila J. R. 

(1996) Population ecology, nesting, and sucess of leatherback turtles, Dermochelys coriacea, at 

Las Baulas de Guanacaste National Park, Costa Rica. P.340 In: J.A. Keinath, D.E. Barnard, J.A. 

Musick, and B.A. Bell (Compilers), Proceedings of the 15th Annual Symposium on Sea Turtle 

Biology and Conservation. Department of Commerce, NOAA Technical Memorandum NMFSSEFSC-351. 

Williams R., Bain D. E., Ford J. K. B. & Trites A. W. (2002a) Behavioural responses of male 

killer whales to a ‘leapfrogging’ vessel. Journal of Cetacean Research, and Management 4, 305-

310. 

Williams R., Bain D. E., Ford J. K. B. & Trites. A. W. (2002b) Behavioural responses of male 

killer whales to a “leapfrogging” vessel. (Orcinus orca). Journal of Cetacean Research and 

Management 4, 305-310. 

Williams R., Lusseau D. & Hammond. P. S. (2006) Estimating relative energetic costs of human 

disturbance to killer whales (Orcinus orca). Biological Conservation 133, 301-311. 

Williams R., Trites A. W. & Bain D. E. (2002c) Behavioural responses of killer whales (Orcinus 

orca) to whale-watching boats: opportunistic observations and experimental approaches. Journal 

of Zoology 256, 255-270. 

Williams R., Trites r. W. & Bain. D. E. (2002d) Behavioural responses of killer whales (Orcinus 

orca) to whale-watching boats: Opportunistic observations and experimental approaches. Journal 

of Zoology 256, 255-270. 

Williams S. L. & Grosholz E. D. (2008) The invasive species challenge in estuarine and coastal 

environments: Marrying management and science. Estuaries and Coasts 31, 3-20. 
514

Willson M. F., Armstrong R. H., Hermans M. C. & Koski K. (2006) Eulachon: a review of 

biology and an annotated bibliography. Auke Bay Laboratory, Alaska Fisheries Science Center, 

Juneau, Alaska. 

Wilson J. A. & McKinley R. S. (2004) Distribution, habitat and movements. Pages 40-69 in: 

G.T.O. LeBreton, F.W.H. Beamish, and R.S. McKinley, eds. Sturgeon and paddlefish of North 

America. Kluwer Academic Publishers. 

Wimberly M. C., Spies T. A., Long C. J. & Whitlock C. (2000) Simulating historical variability 

in the amount of old forests in the Oregon Coast Range. Conservation Biology 14, 167–180. 

Wing B. L. (1977) Salmon food observations. Pages 20-27 in: Southeast Alaska troll log book 

program: 1976 scientific report. Alaska Sea Grant Report 77-11. 

Wing B. L. & Hodge R. P. (2002) Occurrence terminology for marine turtles. Marine Turtle 

Newsletter 95, 15-16. 

Winn H. E., Edel R. K. & Taruski A. G. (1975) Population estimate of the humpback whale in 

the West Indies by visual and acoustic techniques. Journal of the Fisheries Research Board of 

Canada 32, 499–506. 

Winn H. E., Goodyear J. D., Kenney R. D. & Petricig R. O. (1995) Dive patterns of tagged right 

whales in the Great South Channel. Continental Shelf Research 15, 593-611. 

Winn H. E., Price C. A. & Sorensen P. W. (1986) The distributional biology of the right whale 

(Eubalaena glacialis) in the western North Atlantic. Report of the International Whaling 

Commission, 129-138. 

Winn H. E. & Reichley N. E. (1985) Humpback whale - Megaptera novaeangliae. Handbook of 

Marine Mammals: Vol. 3 The Sirenians and Baleen Whales, 241-274. 

Winter A., Foy R. J. & Wynne. K. (2009) Seasonal differences in prey availability around a 

Steller sea lion haulout and rookery in the Gulf of Alaska. Aquatic Mammals 35, 145-162. 

Wirgin I., Grunwald C., Carlson E., Stabile J., Peterson D. L. & Waldman J. (2005a) Range-wide 

population structure of shortnose sturgeon Acipenser brevirostrum based on sequence analysis of 

the mitochondrial DNA control region. Estuaries 28, 406-421. 

Wirgin I., Grunwald C., Carlson E., Stabile J., Peterson D. L. & Waldman J. (2005b) Rangewide population structure of shortnose sturgeon Acipenser brevirostrum based on sequence 

analysis of the mitochondrial DNA control region. Estuaries 28, 16. 

Wirgin I., Waldman J., Stabile J., Lubinski B. & King T. (2002) Comparison of mitochondrial 

DNA control region sequence and microsatellite DNA analyses in estimating population 

structure and gene flow rates in Atlantic sturgeon Acipenser oxyrinchus. Journal of Applied 

Ichthyology 18, 313-319. 

Wirgin I., Waldman J. R., Rosko J., Gross R., Collins M. R., Rogers S. G. & Stabile J. (2000) 

Genetic Structure of Atlantic Sturgeon Populations Based on Mitochondrial DNA Control 

Region Sequences. Transactions of the American Fisheries Society 129, 476-486. 

Wise J. P., Wise S. S., Kraus S., Shaffiey F., Grau M., Chen T. L., Perkins C., Thompson W. D., 

Zheng T., Zhang Y., Romano T. & O’Hara T. (2008) Hexavalent chromium is cytotoxic and 

genotoxic to the North Atlantic right whale (Eubalaena glacialis) lung and testes fibroblasts. 

Mutation Research 650:30–38. 

Wishard L. N., Utter F. M. & Gunderson D. R. (1980) Stock separation of five rockfish species 

using naturally occurring biochemical genetic markers. Marine Fisheries Review 42, 64-73. 

Wishner K., Durbin E., Durbin A., Macaulay M., Winn H. & Kenney R. (1988) Copepod patches 

and right whales in the Great South Channel off New England. Bulletin of Marine Science 43, 

825-844. 
515

Witherington B., Hirama S. & Mosier A. (2003) Effects of beach armoring structures on marine 

turtle nesting. Florida Fish and Wildlife Conservation Commission. 

Witherington B., Hirama S. & Mosier A. (2007) Change to armoring and other barriers to sea 

turtle nesting following severe hurricanes striking Florida beaches. Florida Fish and Wildlife 

Conservation Commission. 

Witherington B., Kubilis P., Brost B. & Meylan A. (2009) Decreasing annual nest counts in a 

globally important loggerhead sea turtle population. Ecological Applications 19, 30-54. 

Witherington B. E. (1992) Behavioral responses of nesting sea turtles to artificial lighting. 

Herpetologica 48, 31-39. 

Witherington B. E. & Bjorndal K. A. (1991) Influences of artificial lighting on the seaward 

orientation of hatchling loggerhead turtles Caretta caretta. Biological Conservation 55, 139-149. 

Witherington B. E., Herren R. & Bresette M. (2006) Caretta caretta – Loggerhead Sea Turtle. 

Chelonian Research Monographs 3, 74-89. 

Witherington B. E., Herren R. & Bresette M. (2006b) Caretta caretta – Loggerhead Sea Turtle. 

Chelonian Research Monographs 3, 74-89. 

Witt M. J., Baert B., Broderick A. C., Formia A., Fretey J., Gibudi A., Mounguengui G. A. M., 

Moussounda C., Ngouessono S., Parnell R. J., Roumet D., Sounguet G. P., Verhage B., Zogo A. 

& Godley B. J. (2009) Aerial surveying of the world's largest leatherback turtle rookery: A more 

effective methodology for large-scale monitoring. Biological Conservation 142, 1719-1727. 

Witteveen B., Straley J. M., Ziegesar O. V., Steel D. & Baker. C. S. (2004) Abundance and 

mtDNA differentiation of humpback whales (Megaptera novaeangliae) in the Shumagin Islands, 

Alaska. Canadian Journal of Zoology 82, 1352-1359. 

Witteveen B. H., Worthy G. A. J., Wynne K. M., Hirons A. C., Iii A. G. A. & Markel. R. W. 

(2011) Trophic levels of North Pacific humpback whales (Megaptera novaeangliae) through 

analysis of stable isotopes: Implications on prey and resource quality. Aquatic Mammals 37, 101-

110. 

Witzell W. N. (1981) Predation on Juvenile Green Sea Turtles, Chelonia mydas , By a Grouper, 

Promicrops lanceolatus (Pisces: Serranidae) in the Kingdom of Tonga, South Pacific. Bulletin of 

Marine Science. Vol. 31, no. 4. 

Witzell W. N. (1983) Synopsis of biological data on the hawksbill turtle Eretmochelys imbricata

(Linnaeus, 1766). In: Food and Agriculture Organization (FAO) Fisheries Synopsis No. 137 p. 

78. FAO. 

Witzell W. N., Geis A. A., Schmid J. R. & Wibbels T. (2005a) Sex ratio of immature Kemp's 

ridley turtles (Lepidochelys kempii) from Gullivan Bay, Ten Thousand Islands, south-west 

Florida. Journal of the Marine Biological Association of the United Kingdom 85, 205-208. 

Witzell W. N., Geis A. A., Schmid J. R. & Wibbels T. (2005b) Sex ratio of immature Kemp's 

ridley turtles (Lepidochelys kempii) from Gullivan Bay, Ten Thousand Islands, south-west 

Florida. Journal of the Marine Biological Association of the U.K. 85, 205-208. 

Witzell W. N. & Schmid J. R. (2005) Diet of immature Kemp's ridley turtles (Lepidochelys 

kempii) from Gullivan Bay, Ten Thousand Islands, southwest Florida. Bulletin of Marine Science

77, 191-199. 

Wolfe S. H., Reidenauer J. A. & Means D. B. (1988) An ecological characterization of the 

Florida Panhandle. U.S. Fish and Wildlife Service and MMS, New Orleans, Louisiana. 

Womble J. N. & Conlon. S. (2010) Observation of Steller sea lion (Eumetopias jubatus) 

predation on a harbor seal (Phoca vitulina richardii) in the Glacier Bay region of southeastern 

Alaska. Aquatic Mammals 36, 129-137. 
516

Wonham M. J., Walton W. C., Ruiz G. M., Frese A. M. & Galil B. S. (2001) Going to the 

source: role of the invasion pathway in determining potential invaders. Marine Ecology-Progress 

Series 215, 1-12. 

Woodbury D. & Ralston S. (1991) Interannual variation in growth rates and back-calculated 

birth-date distributions of pelagic juvenile rockfishes (Sebastes spp.) off the central California 

coast. Fishery Bulletin 89, 523-533. 

Woodley T. H., Brown M. W., Kraus S. D. & Gaskin D. E. (1991) Organochlorine levels in 

North Atlantic right whales (Eubalaena glacialis) blubber. Archives of Environmental 

Contamination and Toxicology 21, 141-145. 

Woodward D. F., Brumbaugh W. G., DeLonay A. J., Little E. E. & Smith C. E. (1994) Effects 

on rainbow trout fry of a metals-contaminated diet of benthic invertebrates from the Clark Fork 

River, Montana. Transactions of the American Fisheries Society 133, 51-62. 

Woody K., Atkinson A., Clark R., Jeffrey C., Lundgren I., Miller J., Monaco M., Muller E., 

Patterson M., Rogers C., Smith T., Spitzak T., Waara R., Whelan K., Witcher B. & Wright A. 

(2008) Coral Bleaching in the U.S. Virgin Islands in 2005 and 2006. In: Status of Caribbean 

Reefs after Bleaching and Hurricanes in 2005 (eds C. Wilkinson and D. Souter) pp. 68-72. 

Global Coral Reef Monitoring Network and Reef and Rainforest Research Center. 

Work P. A., Sapp A. L., Scott D. W. & Dodd M. G. (2010) Influence of small vessel operation 

and propulsion system on loggerhead sea turtle injuries. Journal of Experimental Marine Biology 

and Ecology. 

Work T. M., Dagenais J., Balazs G. H., Schumacher J., Lewis T. D., Leong J. A. C., Casey R. N. 

& Casey J. W. (2009) In vitro biology of fibropapilloma-associated turtle herpesvirus and host 

cells in Hawaiian green turtles (Chelonia mydas). Journal of General Virology 90, 1943-1950. 

Wray P. & Martin. K. R. (1983) Historical whaling records from the western Indian Ocean. 

Report of the International Whaling Commission Special Issue 5, 213-241.-Sc/232/O218). 

Special Issue on Historical Whaling Records. 

Wright S. (1999) Petition to list eulachon Thaleichthys pacificus as threatened or endangered 

under the Endangered Species Act. 

WSDE. (2006) Vessel entries and transits for Washington waters,VEAT 2005. Washington State 

Department of Ecology, Olympia, Washington. 

Wunderlich R. C., Winter B. D. & Meyer J. H. (1994) Restoration of the Elwha River ecosystem. 

Fisheries 19, 11–19. 

WWF. (2001) The status of wild Atlantic salmon: A river by river assessment. Washington, D.C. 

WWOANW. (2007) 2007 guidelines and best practices for commercial whale watching 

operators. Whale Watch Operators Association Northwest. 

Wydoski R. S. & Whitney R. R. (1979a) Inland fishes of Washington. University of Washington 

Press, Seattle, Washington. 

Wydoski R. S. & Whitney R. R. (1979b) Inland fishes of Washington. University of Washington 

Press, Seattle, Washington. 

Wyllie Echeverria T. (1987) Thirty-four species of California rockfishes: maturity and 

seasonality of reproduction. Fisheries Bulletin 85, 229-250. 

Wynne K., Foy R. J. & Buck L. (2005) Gulf apex predator-prey study. University of Alaska, 

Kodiak, Alaska. 

Wynne K. M. & Witteveen B. (2005) Opportunistic aerial sightings of large whales within 

Steller sea lion critical habitat in the Kodiak Archipelago. p. 241. NOAA. 
517

Yamanaka K. L. & Kronlund A. R. (1997) Inshore rockfish stock assessment for the west coast 

of Canada in 1996 and recommended yields for 1997. 

Yamanaka K. L., Lacko L. C., Withler R., Grandin C., Lochead J. K., Martin J. C., Olsen N. & 

Wallace S. S. (2006) A review of yelloweye rockfish Sebastes ruberrimus along the Pacific coast 

of Canada: biology, distribution, and abundance trends 

Yan N. D., Girard R. & Boudreau S. (2002) An introduced predator (Bythotrephes) reduces 

zooplankton species richness. Ecological Letters 5, 481-485. 

Yan N. D., Somers K. M., Girard R. E., Paterson A. M., Keller W., Ramcharan C. W., Rusak J. 

A., Ingram R., Morgan G. E. & Gunn J. M. (2008) Long-term trends in zooplankton of Dorset, 

Ontario, lakes: the probable interactive effects of changes in pH, total phosphorus, dissolved 

organic carbon, and predators. Canadian Journal of Fisheries and Aquatic Sciences 65, 862-877. 

Yañez I. L., Gaos A. R. & Aruaz R. M. (2010) Eastern Pacific leatherback, green, and olive 

ridley sea turtle nesting at Playa Naranjo; first census in eight years In: 28th Annual 

Symposiumon Sea Turtle Biology and Conservation (eds K. Dean and M. C. L. Castro). National 

Marine Fisheries Service. 

Yang M. S. & Nelson M. W. (1999) Food habits of the commercially important groundfishes in 

the Gulf of Alaska in 1990, 1993, and 1996. NOAA. 

Yender R., Michel J. & Lord C. (2002) Managing seafood safety after an oil spill. National 

Oceanic and Atmospheric Administration, National Ocean Service, Office of Response and 

Restoration, Seattle, Washington. 

Yentsch C. S., Yentsch C. M., Cullen J. J., Lapointe B., Phinney D. A. & Yentsch S. W. (2002) 

Sunlight and water transparency: Cornerstones in coral research. Journal of Experimental 

Marine Biology and Ecology 268, 171-183. 

Ylitalo G. M., Matkin C. O., Buzitis J., Krahn M. M., Jones L. L., Rowles T. & Stein. J. E. 

(2001) Influence of life-history parameters on organochlorine concentrations in free-ranging 

killer whales (Orcinus orca) from Prince William Sound, AK. Science of the Total Environment

7-Oct, 183-203. 

Yoklavich M. M., Greene H. G., Cailliet G. M., Sullivan D. E., Lea R. N. & Love M. S. (2000) 

Habitat associations of deep-water rockfishes in a submarine canyon: an example of a natural 

refuge. Fishery Bulletin 98, 625-641. 

York A. (1994) The population dynamics of the northern sea lions, 1975-85. Marine Mammal 

Science 10, 38-51. 

York R. S. (2005) Megagametogenesis and nuclear DNA content estimation in Halophila 

(Hydrocaritaceae). Master’s thesis, University of North Carolina at Wilmington. 

Yoshiyama R. M., Gerstung E. R., Fisher F. W. & Moyle P. B. (1996) Historical and present 

distribution of chinook salmon in the Central Valley drainage of California. In: Department of 

Wildlife, Fish, and Conservation Biology, University of California, Davis California. 

Young N. M., Iudicello S., Evans K. & Baur D. (1993) The incidental capture of marine 

mammals in U.S. fisheries: problems and solutions. Center for Marine Conservation, 

Washington, D. C. 

Zabel R. W. & Achord S. (2004) Relating size of individuals to juvenile survival within and 

among closely-related populations of chinook salmon. Ecology 85, 795-806. 

Zabel R. W. & Williams J. G. (2002) Selective mortality in chinook salmon: what is the role of 

human disturbance? Ecological Applications 12, 173-183. 
518

Zamon J. E. & Welch D. W. (2005a) Rapid shift in zooplankton community composition on the 

northeast Pacific shelf during the 1998-1999 El Niño-La Niña event. Canadian Journal of 

Fisheries and Aquatic Science 62, 133-144. 

Zamon J. E. & Welch D. W. (2005b) Rapid shift in zooplankton community composition on the 

northeast Pacific shelf during the 1998-1999 El Niño - La Niña event. Canadian Journal of 

Fisheries and Aquatic Sciences 62, 133-144. 

Zárate P. M., Cahoon S. S., Contato M. C. D., Dutton P. H. & Seminoff J. A. (2006) Nesting 

beach monitoring of green turtles in the Galapagos Islands: a 4-year evaluation. In: Twenty-sixth 

Annual Symposium on Sea Turtle Biology and Conservation (eds M. Frick, A. Panagopoulou, A. 

F. Rees and K. Williams) p. 337. International Sea Turtle Society, Athens, Greece. 

Zárate P. M., Parra M. A., Robles M., Dutton P. H. & Seminoff J. A. (2010) Sea turtle strandings 

and mortality in the Galapagos Archipelago: causes and threats In: 28th Annual Symposiumon 

Sea Turtle Biology and Conservation (eds K. Dean and M. C. L. Castro) p. 126. National Marine 

Fisheries Service. 

Zeitsev Y. & Ozturk B. (2001) Exotic species in the Aegean, Marmara, Black, Azov, and 

Caspian Seas. 1-267. 

Zerbini A. N., Waite J. M., Laake J. L. & Wade P. R. (2006) Abundance, trends and distribution 

of baleen whales off Western Alaska and the central Aleutian Islands. Deep-Sea Research I 53, 

1772-1790. 

Zettler M. L. (1997) The newcomer Marenzelleria viridis (Verrill, 1873), its development and 

influence on the indigenous macrozoobenthos in a coastal water of the southern Baltic. In: 

Proceedings of the 14th Baltic Marine Biologists Symposium (ed E. Ojaveer) pp. 280-296. 

Estonian Academy Publishers, Tallinn. 

Zimmerman C. E., Edwards G. W. & Perry K. (2008) Maternal origin and migratory history of 

Oncorhynchus mykiss captured in rivers of the Central Valley, California. California Department 

of Fish and Game. 

Zmudzinski L. (1996) The effect of the introduction of the American species Marenzelleria 

viridis (Polychaeta: spionidae) on the benthic ecosystem of Vistula Lagoon. Marine Ecology 17, 

221-226. 

Zug G. R., Balazs G. H., Wetherall J. A., Parker D. M. & Murakawa S. K. K. (2002) Age and 

growth of Hawaiian green sea turtles (Chelonia mydas): An analysis based on 

skeletochronology. Fishery Bulletin 100, 117-127. 

Zug G. R. & Glor R. E. (1998) Estimates of age and growth in a population of green sea turtles 

(Chelonia mydas) from the Indian River Lagoon system, Florida: A skeletochronological 

analysis. Canadian Journal of Zoology 76, 1497-1506. 

Zug G. R., Kalb H. J. & Luzar S. J. (1997) Age and growth on wild Kemp's ridley sea turtles 

Lepidochelys kempii from skeletochronological data. Biological Conservation 80, 261-268. 

Zug G. R. & Parham J. F. (1996) Age and growth in leatherback turtles, Dermochelys coriacea: 

A skeletochronological analysis. Chelonian Conservation and Biology 2, 244-249. 

Zurita J. C., Herrera R., Arenas A., Torres M. E., Calderón C., Gómez L., Alvarado J. C. & 

Villavicencia R. (2003) Nesting loggerhead and green sea turtles in Quintana Roo, Mexico. 

Pp.25-127 In: Seminoff, J.A. (Ed), 22nd Annual Symposium on Sea Turtle Biology and 

Conservation, 4-7 April, 2002, Miami, FL. U.S. Department of Commerce, NOAA Technical 

Memorandum NMFS-SEFSC-503.