PETITION TO LIST OLYMPIC PENINSULA STEELHEAD (Oncorhynchus Mykiss) AS A THREATENED OR ENDANGERED SPECIES

PETITION TO LIST OLYMPIC PENINSULA

STEELHEAD (Oncorhynchus Mykiss) AS A

THREATENED OR ENDANGERED

SPECIES

August 1, 2022

Petitioners

NOTICE OF PETITION

August 1, 2022

The Honorable Gina Raimondo

Secretary of Commerce

U.S. Department of Commerce

1401 Constitution Ave. NW

Washington, D.C. 20230

TheSec@doc.gov

via email

The Honorable Janet Coit

Assistant Administrator for Fisheries

NOAA Fisheries

1315 East-West Highway

Silver Spring, MD 20910

janet.coit@noaa.gov

via email

Dear Secretary Raimondo:

Pursuant to Section 4(b) of the Endangered Species Act (16 U.S.C. § 1533(b)), Section 553(e) of

the Administrative Procedures Act (5 U.S.C. § 553(e)), and 50 C.F.R. § 424.14(a), The

Conservation Angler and Wild Fish Conservancy (the “Petitioners”) hereby petition the

Secretary of Commerce, through the National Marine Fisheries Service (“NMFS”), to list the

Olympic Peninsula Steelhead Distinct Population Segment (Oncorhynchus mykiss) as a

threatened or endangered species. Petitioners also request that NMFS designate critical habitat

for Olympic Peninsula steelhead concurrent with the distinct population segment being listed

as threatened or endangered.

The Conservation Angler (TCA) is a Washington-based, nonprofit, public interest organization

that uses scientific expertise and legal advocacy to protect and conserve wild steelhead,

salmon, trout, and char throughout their Pacific range. TCA also operates the Kamchatka

Steelhead Program, an international research program that has advanced scientific knowledge

of wild steelhead for over 25 years.

Wild Fish Conservancy (WFC) is a Washington-based, nonprofit, public interest organization

that uses science, education, and advocacy to promote technically and socially responsible

habitat, hatchery, and harvest management to better sustain wild fish heritage in Washington

State.

NMFS has jurisdiction over this petition. This petition sets in motion a specific process, placing

definite response requirements on NMFS. Specifically, NMFS must “make a finding as to

whether the petition presents substantial scientific or commercial information indicating that

the petitioned action may be warranted.” 16 U.S.C. § 1533(b)(3)(A). NMFS must make this

finding “[t]o the maximum extent practicable, within 90 days after receiving the petition.” Id.

Petitioners need not demonstrate that the petitioned action is warranted. Rather, Petitioners

must only present information demonstrating that such action may be warranted. While

Petitioners believe that the best available science demonstrates that listing Olympic Peninsula

steelhead as threatened or endangered is in fact warranted, there can be no reasonable

dispute that the available information indicates that listing the species as either threatened or

endangered may be warranted. As such, NMFS should promptly make a positive initial finding

on the petition and commence a status review as required by 16 U.S.C. § 1533(b)(3)(B) and 50

C.F.R. § 424.14(h)(1)-(2).

Thank you.

THE CONSERVATION ANGLER WILD FISH CONSERVANCY

___________________________________ ____________________________________

David Moskowitz Emma Helverson

Executive Director Executive Director

The Conservation Angler Wild Fish Conservancy

P.O. Box 13121 P.O. Box 402

Portland, OR 97213 Duvall, WA 98019

(971) 235-8953 (425) 788-1167

david@theconservationangler.org emma@wildfishconservancy.org

Table of Contents

EXECUTIVE SUMMARY ....................................................................................................................................... 5

LEGAL BACKGROUND ........................................................................................................................................ 10

THE STATUS OF OLYMPIC PENINSULA STEELHEAD ............................................................................................. 13

I. DESCRIPTION ................................................................................................................................................. 13

II. TAXONOMY ................................................................................................................................................... 14

III. HABITAT AND RANGE .................................................................................................................................... 14

IV. LIFE HISTORY .................................................................................................................................................. 24

V. HABITAT REQUIREMENTS .............................................................................................................................. 25

VI. DIET ............................................................................................................................................................... 26

VII. NATURAL MORTALITY ............................................................................................................................... 26

VIII. POPULATION STRUCTURE ......................................................................................................................... 27

IX. PREVIOUS STOCK ASSESSMENTS ................................................................................................................... 28

X. ABUNDANCE AND POPULATION TRENDS ...................................................................................................... 33

XI. PRODUCTIVITY ............................................................................................................................................... 59

XII. DIVERSITY ................................................................................................................................................. 66

XIII. SPATIAL STRUCTURE ................................................................................................................................. 71

ENDANGERED SPECIES ACT LISTING FACTORS .................................................................................................... 72

I. PRESENT OR THREATENED DESTRUCTION, MODIFICATION, OR CURTAILMENT ........................................... 72

OF ITS HABITAT OR RANGE ...................................................................................................................................... 72

II. OVERUTILIZATION FOR COMMERCIAL AND RECREATIONAL PURPOSES ....................................................... 92

III. DISEASE AND PREDATION ............................................................................................................................ 109

IV. INADEQUACY OF EXISTING REGULATORY MECHANISMS ............................................................................ 110

V. OTHER NATURAL OR ANTHROPOGENIC FACTORS ....................................................................................... 130

REQUEST FOR CRITICAL HABITAT DESIGNATION .............................................................................................. 138

REFERENCES .................................................................................................................................................... 139

EXECUTIVE SUMMARY

Olympic Peninsula steelhead are at risk of becoming an endangered species within the

foreseeable future. The summer-run component is nearly extinct, and the winter-run

component is declining and losing its life history diversity. The fate of the species now rests on a

depressed and contracted mid- to late-spring component of wild fish whose productivity is

limited or declining depending on the population. The remnants of these runs that historically

numbered in the tens of thousands face declining freshwater and marine habitat conditions,

increasing recreational fishing pressure, and ongoing commercial harvest. Because of these and

other demographic and ecological threats, Olympic Peninsula steelhead are likely to become

endangered within the foreseeable future. Olympic Peninsula steelhead warrant protection

under the Endangered Species Act (ESA).

Abundance

Almost every population of winter steelhead on the Olympic Peninsula is in long-term

decline over their period of record. For example, mean annual run sizes from 1980 – 2017

declined by 37-46% in the Hoh, Queets, and Quinault Rivers (McMillan et al. 2022), resulting in

the Hoh and Queets River populations increasingly failing to meet their escapement goals. The

Quillayute River system population has declined sharply since the 1990s (McMillan et al. 2022).

All populations have continued to decline since 2017 and run sizes have been so small in recent

years that fisheries were closed. Trends appear similar for the smaller populations, with some

not meeting their escapement goals for decades, though the data sets are not always as

extensive. These patterns underscore why the Olympic Peninsula DPS has the second lowest

proportion of populations with increasing trends in Washington State (Cram et al. 2018).

The plight of the winter runs is even more revealing when compared to historical

estimates from circa 1940-1960. Based on those historical estimates (McMillan et al. 2022), the

Quillayute, Queets, Hoh, and Quinault River wild winter steelhead populations have declined by

61%, 69%, 79%, and 81%, respectively, in relation to their most recent five-year mean run size.

Those estimates do not necessarily capture the full extent of decline, however, because they

came after decades of harvest and many years of habitat alterations. When compared to

cannery data in 1923, for instance, the decline in the Queets River increases to 86%. Hence, it is

likely that all the populations are now only a fraction of their former abundance.

The status of the summer-run component is dire. Summer steelhead are not monitored

or managed, but snorkel surveys by Brenkman et al. (2012) and McMillan (2022) suggest

populations in the Quillayute, Hoh, and Quinault Rivers have declined dramatically and are at

critically low levels of abundance, if not already functionally extinct. Unfortunately, almost no

information exists for summer steelhead in the Queets River. The long-term, chronic declining

trends in winter steelhead coupled with the dire plight of summer steelhead puts the Olympic

Peninsula Steelhead DPS at greater risk of extinction than when Busby et al. (1996) conducted

the last status review for the DPS (ESU at the time).

Diversity

The diversity of Olympic Peninsula steelhead has been altered in different ways that

could impact their productivity, resilience, and capacity to adapt with changing climate effects.

First, the early-returning component of winter steelhead populations in the Quillayute, Hoh,

and Queets Rivers is severely depleted and consequently, the breadth of run timing is now far

more compressed than it was historically (McMillan et al. 2022). This change in run timing was

not accounted for in the prior status review (Busby et al. 1996), but it could be critical to their

persistence. Steelhead enter and spawn earlier in the winter in warmer, more southerly regions

of their native range (Busby et al. 1996), and as climate impacts progress, winter steelhead on

the Olympic Peninsula will also need to enter and spawn earlier to keep pace with climate

change. Run timing provides one way in which salmonids can adapt to changes in stream flow

and water temperature (Manhard et al. 2017), but that could be impossible for winter

steelhead in the Olympic Peninsula DPS if their original run timing is not restored.

Second, the data on repeat spawners suggest that rates of iteroparity have declined in

most of the four major populations over their period of record. The Queets River population

declined from a high of over 50% down to lows of approximately 10% in recent years, while the

Quillayute River peaked at over 20% in the late-1970s and is now down to around 10%. The

Quinault River population declined from 10-15% to approximately 1-2% in the last couple years.

Repeat spawning rates do not show a declining trend in the Hoh River, but that is likely because

its period of record does not extend back to the late-1970’s to 1980 as it does in the Queets and

Quillayute Rivers.

Third, summer runs are not monitored or managed, but they are caught in fisheries.

Snorkel surveys by Brenkman et al. (2012) and McMillan (2022) indicate summer steelhead are

at critically low levels of abundance in the Quillayute, Hoh, and Quinault Rivers, and they are

exposed to high levels of stray hatchery summer steelhead. There is a strong heritable basis for

the early maturing summer run life history and such allelic variants do not arise independently

via new mutations (Prince et al. 2017; Waples et al. 2022). Given their reduced abundance, it is

likely the populations are already genetically compromised, but they are still producing some

adults. If the declines continue and summer runs go extinct, the genetic basis for the life history

could be lost entirely, leaving behind a massive void in the timing of migration of the overall

steelhead populations in the Quillayute, Hoh, Quinault, and Queets Rivers.

Productivity

Information on productivity is very limited for steelhead in the Olympic Peninsula DPS,

but what data is available suggests that productivity is declining, and recruitment has

increasingly failed to reach replacement levels (Cram et al. 2018). The most recent productivity

estimates (Cram et al. (2018) only account for data up to 2010, so they miss many of the worst

returns on record that have occurred most recently. Considering the increasingly low levels of

productivity in the Quillayute River population, and that peaks and troughs in annual run size in

the Hoh, Queets, and Quinault River populations have generally declined each decade over the

period of record (lower peaks, deeper troughs) leading to high levels of depletion in recent

years, it appears the populations are no longer compensating as effectively as they did 20-40

years ago. If a declining productivity trend continues at the rates estimated by McMillan et al.

(2022) and summer runs decline further, which seems conservative given the recent and rapid

declines in run size, winter steelhead run sizes will soon be too small to allow fisheries and

summer steelhead will be extinct.

Spatial structure

While there are no large dams affecting Olympic Peninsula steelhead, there is a

dendritic network of roads and culverts constructed for forest practices, many of which cross

salmon bearing streams and create blockages to upstream migration (Smith 2000). Juvenile

steelhead are often distributed right up to barriers in many small creeks (McMillan and Starr

2008; McMillan et al. 2013), so the blockages have likely truncated their distribution. It is likely

the spatial distribution has been reduced or altered due to the depletion of early-returning

winter steelhead and the very low abundance of summer steelhead. Those components of the

steelhead population spawn in habitats that are otherwise used less or not at all by other life

histories (e.g., Cederholm 1983; McMillan et al. 2007). Hence, changes in run timing of winter

steelhead and depletion of summer steelhead, combined with barriers, have all plausibly

reduced the populations’ spatial distributions.

Harvest impacts

Several winter steelhead populations in the Olympic Peninsula DPS support intensive

recreational and commercial fisheries, and they experience the highest harvest levels of any

steelhead populations in Washington State (Cram et al. 2018). Harvest has had two apparent

effects. First, it has increasingly reduced spawning abundance below management goals. For

instance, due to overharvest, winter steelhead in the Hoh and Queets River have failed to meet

their minimum escapement goals for abundance in 50% of the years since 2003 and 50% of the

years over the past decade, respectively. The Quillayute River system population has generally

met its overall escapement goal, but individually, the Bogachiel and Sol Duc Rivers have failed

to meet their goals in 60% and 70% of the last ten years, respectively. Unlike the larger

populations, many smaller populations no longer experience direct harvest, but nearly all of

them have also failed to meet their escapement goals, including some that have not met their

goals for 20 years or longer.

Second, the strong focus on maximizing harvest of early-timed hatchery winter

steelhead from late-November through early-January has apparently contributed to depletion

of early-returning wild winter steelhead that were formerly very abundant in the 1940’s –

1950’s (McMillan et al. 2022). As a result, the breadth of run timing is now more compressed

than it was historically. These new results challenge the claim by Busby et al. (1996) that

hatchery and wild winter steelhead were temporally segregated due to differences in run

timing. Rather, the hatchery fish have simply replaced the wild fish, underscoring the potential

for a Shifting Baseline (Pauley 1995). This should not be a surprise, however, because according

to the traditional ecological knowledge of the Quileute Tribe, the month of January meant the

“time of steelhead running” and February the “time of steelhead spawning” (Frachtenberg

1916).

Hatchery impacts

Hatcheries can exert a range of genetic and ecological impacts on wild salmon and

steelhead (Araki and Schmid 2011), and each year, large numbers of hatchery steelhead are

released into several streams in the Olympic Peninsula DPS with the largest number of fish

being released into the Quinault River (Duda et al. 2018). Unfortunately, hatchery effects are

relatively unstudied, and their impacts are therefore uncertain in the DPS. However, as

mentioned earlier, fisheries focused on maximizing harvest of hatchery steelhead and

interbreeding between hatchery and wild steelhead have both likely contributed to the

depletion of early-timed wild winter steelhead (McMillan et al. 2022). For example, Seamons et

al. (2012) found that after three generations of stocking segregated hatchery winter steelhead

the proportion of wild ancestry smolts and adults declined by 10-20% and up to 80% of

naturally produced winter steelhead were hatchery x wild hybrids. Whatever the relative

contributions of each factor are, it is clear that hatchery and wild steelhead once overlapped

substantially, and the depletion of wild fish coincides with the onset of the hatchery programs.

There are also concerns about overall genetic impacts due to the number of hatchery

steelhead that stray and spawn in nature. Cederholm (1983) noted there was a high degree of

within and between river straying of hatchery steelhead, which could change the long-term

spawning, timing, growth, and survival of wild fish. Similar concerns were raised in the status

review by Busby et al. (1996) because of the “widespread production of hatchery steelhead” in

the DPS. Indeed, a review in 2009 indicated that hatchery steelhead greatly exceeded their

pHOS goals for winter and summer steelhead in the Bogachiel River (WDFW 2022b). And

WDFW suggests that hatchery winter and summer steelhead pose substantial risks to among-

population diversity and fitness of wild steelhead due to introgression that has likely occurred

in the Pysht, Hoko, and Sol Duc Rivers (WDFW 2008).

The record shows that concerns by Cederholm (1983) and others were warranted,

particularly for summer steelhead. As evidenced by snorkel surveys by Brenkman et al. (2012)

and McMillan (2022), a substantial number of hatchery summer steelhead stray into the Hoh

and Quinault Rivers, both of which do not receive any releases of hatchery summer steelhead.

There is also a substantial amount of straying within the Quillayute River system, particularly

into the upper Calawah and Bogachiel Rivers (Brenkman et al. 2012). The high levels of pHOS,

which frequently reach and even exceed 30-50%, combined with very low abundance of wild

summer steelhead, suggest that hatchery impacts could be a significant limiting factor.

Changes to Freshwater and Marine Habitat

The depleted stocks of winter and summer steelhead in the Olympic Peninsula DPS are

threatened with a myriad of environmental challenges, including degraded freshwater habitat,

climate impacts, and a changing ocean. Although the largest watersheds in the DPS have their

headwaters in Olympic National Park, a long history of extensive and unsustainable logging has

greatly degraded freshwater habitat outside of the park. Logging and roadbuilding have

increased the frequency of mass wasting events, which cause excessive sedimentation of

stream channels, diminish supplies of large woody debris, impair water quality, increase the

frequency of peak flow events, and reduce habitat connectivity (Smith 2000; East et al. 2017).

Climate change is predicted to degrade freshwater habitat even further, contributing to an

earlier onset of warmer summer water temperatures, more extreme fluctuations in stream

flow, and eventual elimination of all the glaciers that feed the largest, most productive

watersheds (Wade et al. 2013; East et al. 2017; Fountain et al. 2022). Climate impacts will also

alter the productivity potential of the marine environment where adult steelhead spend a

substantial portion of their lives. For example, sea surface temperatures are warming and will

continue to do so, while alterations to ocean currents could alter the extent of upwelling and

there is the potential for increased acidification and changes in the food web (Klingler et al.

2008; Miller et al. 2013; Dalton et al. 2016). Each of these changes, past and future, singular

and cumulative, are expected to negatively affect the survival and productivity of steelhead if

they do not have sufficient abundance, diversity, and spatial structure to adapt and keep pace

with climate change.

Summary

This petition demonstrates that the Olympic Peninsula Steelhead DPS warrants

protection under the ESA. It is divided in two parts. Part One covers the Olympic Peninsula

Steelhead DPS’s description, taxonomy, life history, distribution, and population status. Part

Two describes the current and future threats to the DPS in the context of the ESA’s five listing

factors. Based on the substantial information provided herein, NMFS should list Olympic

Peninsula steelhead as a threatened or endangered species and designate its critical habitat.

LEGAL BACKGROUND

The Endangered Species Act

The ESA defines “species” to mean “any subspecies of fish or wildlife or plants, and any

distinct population segment of any species of vertebrate fish or wildlife which interbreeds when

mature.” 16 U.S.C. 1532(16) (Emphasis added).

Olympic Peninsula steelhead are a distinct population segment (DPS) of steelhead

(Busby et al. 1996; 61 Fed Reg 41544 (Aug. 9, 1996)). Genetic data from the Washington

Department of Fish and Wildlife (WDFW) indicate that the Olympic Peninsula steelhead DPS is

“substantially isolated from other regions in western Washington” (Busby et al. 1996). In

addition to genetic differences, Olympic Peninsula steelhead are further characterized by

habitat, climatic, and zoogeographical differences between it and adjacent DPSs, including the

Southwest Washington and Puget Sound DPSs (Busby et al. 1996). NMFS delineates the Olympic

Peninsula steelhead DPS to include populations that occur in river basins to the west of the

Elwha River and south to, but not including, the rivers that flow into Grays Harbor (Busby et al.

1996; 61 Fed Reg 41544 (Aug. 9, 1996)).

Figure 1. Map of the Olympic Peninsula Steelhead Distinct Population Segment (Source: NMFS).

When making a listing determination, NMFS must analyze the status of a species in

conjunction with five statutory listing factors, relying “solely on the best scientific and

commercial data available.” 16 U.S.C. § 1533(b)(1)(A). The five listing factors include:

1. The present or threatened destruction, modification, or curtailment of its

habitat or range;

2. Overutilization for commercial, recreational, scientific, or educational

purposes;

3. Disease or predation;

4. The inadequacy of existing regulatory mechanisms; and

5. Other natural or manmade factors that affect its continued existence.

Id. at § 1533(a)(1).

Under the ESA, a species is “endangered” if it “is in danger of extinction throughout all

or a significant portion of its range.” Id. at § 1532(6). A species is “threatened” if it is “likely to

become an endangered species within the foreseeable future throughout all or a significant

portion of its range.” Id. at § 1532(20) (Emphasis added).

The Foreseeable Future and Climate Change

The foreseeable future extends as far into the future as NMFS can reasonably determine

that both future threats and the species’ responses to those threats are likely. 50 C.F.R. §

424.11(d). When analyzing the threats of climate change, it is NMFS policy to “project effects

over the longest possible period for which credible projections are available in order to ensure

the best available science is fully considered” (Tortorici 2016).

NMFS has looked out as far as the end of the 21

Century when making listing

determinations (Tortorici 2016). For example, the Ninth Circuit held that it was not arbitrary or

capricious for NMFS to list the ringed seal based on climate change models that projected as far

out as year 2100. Alaska Oil & Gas Ass’n v. Nat’l Marine Fisheries Serv., 722 F. App’x 666, 669

(9th Cir. 2018) quoting Alaska Oil & Gas Ass’n v. Pritzker, 840 F.3d 671, 680 (2016); see also

Alaska Oil & Gas Ass’n v. Pritzker, 840 F.3d 671 (9th Cir. 2016), cert denied 138 S. Ct. 924 (2018)

(upholding NMFS’s decision to list bearded seals as threatened based on climate change models

that predicted that the sea ice the seals depend on for birthing and mating would mostly

disappear by 2095) and Safari Club International v. Salazar (In re Polar Bear Endangered Species

Act Listing and § 4(d) Rule Litigation) 709 F.3d 1 (D.C. Cir. 2013), cert denied 571 U.S. 887 (2013)

(upholding the U.S. Fish & Wildlife Service’s decision to list the polar bear as threatened based,

in part, on projected climate change effects to the species and its habitat 45 years in the

future).

As an example of the feasibility of a 100-year time frame, the Intergovernmental Panel

on Climate Change (“IPCC”) provides climate change projections through 2100 under a range of

plausible emissions scenarios (IPCC 2021). NMFS recognizes the IPCC as a credible information

source. The Service’s guidance on climate change and ESA determinations requires it to “use

climate indicator values projected under the [IPCC’s] Representative Concentration Pathway 8.5

when data are available” (Tortorici 2016). For these reasons, the use of 100 years as the

foreseeable future is consistent with NMFS’s climate change policy.

As indicated by NMFS’s policy on climate change and listing decisions (Tortorici 2016),

the best available science standard does not require NMFS to be certain about climate change

and its effects on Olympic Peninsula steelhead.

“While it requires that decisions not be based on mere generalizations or

speculation, the best available science standard does not require that

information be free from uncertainty. For example, to support listing a species

on the basis of climate change related impacts, we must have information

particular to that species to demonstrate that it will be impacted by climate

change, such as through a reduction of suitable habitat within its known range. It

is not necessary, however, to have projections at a particular geographic scale or

to have a complete understanding of the biological reasons for and extent of the

species’ sensitivity to climate change.”

(Tortorici 2016). This petition supports listing Olympic Peninsula steelhead for multiple reasons,

including climate change-related impacts. The best available science, including NMFS’s own

reports, demonstrates that the species is impacted by climate change.

PART ONE

THE STATUS OF OLYMPIC PENINSULA STEELHEAD

Threats to a species’ long-term persistence are manifested demographically as risks to its

abundance, population growth rate, spatial structure and connectivity, and genetic and ecological

diversity. These demographic risks thus provide the most direct indices or proxies of extinction risk.

A species at very low levels of abundance and with few populations will be less tolerant to

environmental variation, catastrophic events, genetic processes, demographic stochasticity,

ecological interactions, and other processes (e.g., Meffe and Carroll 1994, Caughley and Gunn

1996). A population growth rate that is unstable or declining over a long period of time indicates

poor resiliency to future environmental change (e.g., Lande 1993, Middleton and Nisbet 1997, Foley

1997). A species that is not widely distributed across a variety of well-connected habitats is at

increased risk of extinction due to environmental perturbations, including catastrophic events

(Schlosser and Angermeier 1995, Hanski and Gilpin 1997, Tilman et al. 1997, Cooper and Mangel

1999). A species that has lost locally adapted genetic and ecological diversity may lack the raw

resources necessary to exploit a wide array of environments and endure short- and long-term

environmental changes (e.g., Groot and Margolis 1991, Wood 1995). Assessing extinction risk of a

species involves evaluating whether risks to its abundance, population growth rate, spatial structure

and or diversity are such that it is at or near an extinction threshold, or likely to become so in the

foreseeable future. As demonstrated by this petition, Olympic Peninsula steelhead are likely to

become an endangered species within the foreseeable future and, therefore, should be listed as a

“threatened” species under the ESA.

I. DESCRIPTION

Olympic Peninsula steelhead are a distinct population segment of steelhead (Busby et

al. 1996; 61 Fed Reg 41544 (Aug. 9, 1996)). Steelhead are the anadromous form of

Oncorhynchus mykiss and may display the most diverse life histories of any salmonid (Kendall et

al. 2015). As with other salmonids, they begin their life cycle in freshwater where they spend 1-

4 years growing as parr, which are relatively drab and natural in coloration to match their

surrounding stream environments (Busby et al. 1996). After reaching a growth threshold in size,

typically between 150 – 200 mm in length, they undergo a complex series of physiological

changes, such as increased levels of NA+/K+-ATPase in the gills, that assist with osmoregulation

and life in saltwater (Busby et al. 1996). As their internal physiology changes, so does their

outward appearance. Their body becomes more fusiform, their scales become more deciduous,

and they take on a silvery appearance (Busby et al. 1996).

Steelhead are known to display three general categories of ocean migrations. There are

typically anadromous fish that spend 1-4 years in the ocean before returning to spawn in

freshwater, attaining sizes of 45 – 1125 mm in length and 0.9 – 20.5 kg in weight (Kendall et al.

2015). There are also half-pounders that only undertake a short, near-shore ocean migration

for a few months attaining a size of 25 – 40 mm in length and a weight of 0.1 – 0.7 kg, but

generally return in an immature state and will undertake a full ocean migration the next year

(Hodge et al. 2014; Kendall et al. 2015). Last, there are individuals that undertake short

migrations and attain sizes intermediate to the fully anadromous and half-pounder life histories

but are mature upon return to freshwater and are typically male (Kendall et al. 2015). The latter

are referred to as “estuarine” or “jack” life histories, potentially depending on what part of the

ocean they migrate to (Kendall et al. 2015). While the fully anadromous form is common

throughout the native range of steelhead, far less information exists on the distribution of the

half-pounder life history except for a few rivers where they are abundant in California and

southern Oregon, and almost no information exists on estuarine or jack life histories outside of

the Kamchatka Peninsula, Russia (Kendall et al. 2015).

Adults that return to freshwater to spawn are further delineated into two run types,

including ocean-maturing “winter run” and river-maturing “summer run” life histories (Busby et

al. 1996). Olympic Peninsula steelhead include winter and summer run life histories. Winter

runs enter freshwater sexually mature or close to sexual maturity from late-October through

early-June, while summer runs enter sexually immature from May through October and

sexually mature as they stage in freshwater. The physiological processes responsible for

osmoregulation will reverse to allow the fish to persist in freshwater and they begin to display

signs of secondary sexual characteristics as they approach the spawning season from late-

winter through spring to early-summer. Males become much darker shades of green and red,

while females become less silvery and display various shades of red and light green. Both sexes

may develop fungal infections due to stress, fights, and compromised immune systems.

Individuals that survive the rigors of mating and spawning will once again – if in

sufficient condition and health – restart the smoltification process in preparation for saltwater

entry (Buelow and Moffit 2015). During emigration to the ocean their outward appearance will

also change to become more silvery as they did when they were smolts for the first time.

II. TAXONOMY

Kingdom: Animalia

Phylum: Chordata

Class: Osteichthyes

Order: Salmoniformes

Family: Salmonidae

Genus: Oncorhynchus

Species: Oncorhynchus mykiss

III. HABITAT AND RANGE

The Olympic Peninsula is a large arm of land located in western Washington. It is

bordered by Hood Canal to the east, the Strait of Juan de Fuca to the north, and the Pacific

Ocean to the west. The Olympic Mountains are its natural centerpiece. These coastal

mountains reach 1,200 to 2,400 meters above sea level and currently hold 184 glaciers (Busby

et al. 1996; Reidel et al. 2017). The area receives copious rainfall (McHenry et al. 1996). The

west side of the peninsula receives the most precipitation, ranging from 70 to 100 inches in the

lower coastal plains (Dalton et al. 2016) and up to 240 inches in the mountains (McHenry et al.

1996). Olympic National Park protects the only temperate rainforest in the contiguous United

States (Dalton et al. 2016). The park’s foothills are dominated by undisturbed old western

hemlock (Tsuga heterophylla), douglas fir (Pseudotsuga menziesii), and sitka spruce (Picea

sitchensis) (Dalton et al. 2016). Outside of the park, the forest has been extensively logged

(McHenry et al. 1996). Running through this dichotomous landscape of pristine and degraded

habitat are some of the last large, undammed rivers in the Pacific Northwest.

Olympic Peninsula steelhead occur in three distinct water resource inventory areas

(WRIA): WRIAs 19, 20, and 21 (Table 1).

A. Water Resource Inventory Area 19

WRIA 19 begins immediately west of the Elwha River and extends eastward to Cape

Flattery, the northwesternmost point in the lower forty-eight. The area experiences a cool

maritime climate with annual precipitation ranging from 80” to 130” (HSRG 2004). Precipitation

is higher on the west side of WRIA 19 (NOPLE 2015). The Sekiu River basin on the western edge

of WRIA 19 receives 95-120 inches of precipitation per year (NOPLE 2015). The Salt Creek basin

on the eastern edge receives 35-55 inches of precipitation annually (NOPLE 2015).

The forestland has a mix of tree species that vary east to west. In the eastern portion of

WRIA 19, Douglas fir (Psuedotsuga mensiezii) is the dominant species, with red alder (Alnus

rubra), vine maple (Acer cirnicatum), and bigleaf maple (Acer macrophylla) also occurring in the

area (NOPLE 2015). In the western portion of WRIA 19, the forests are dominated by western

hemlock (Tsuga heterophylla) and Sitka spruce (Picea sitchensis) (NOPLE 2015).

Commercial forestry accounts for 76% of land use in WRIA 19 (NOPLE 2015). Private

interests own 56% of the commercial forestland, the Washington Department of Natural

Resources (WDNR) owns 28%, the U.S. Forest Service (USFS) owns 12%, and the remaining 4% is

owned by the county and small landowners (NOPLE 2015). The non-commercial forestland

occurs inside Olympic National Park (11.6%) or is classified as rural, urban, industrial, tribal

reservation, or miscellaneous (12.4%) (NOPLE 2015).

Steelhead are one of the more widely distributed salmonids in WRIA 19 (NOPLE 2015;

Table 1). Winter steelhead occur in several small to medium-sized watersheds, including the

following subbasins listed in order of size: Hoko River (71 sq mi), Lyre River (67.9 sq mi), Pysht

River (46.3 sq mi), Sekiu River (33.2 sq mi), Clallam River (31 sq mi), Salt Creek (19.1 sq mi), East

Twin River (13.6 sq mi), and West Twin River (12.6 sq miles) (NOPLE 2015). The Western Strait

Independents collectively drain 73.3 square miles (NOPLE 2015).

Summer steelhead may occur in WRIA 19 (Table 1). Summer steelhead are thought to

occur in the Lyre River (McHenry et al. 1996; Lyre-Hoko Watershed (WRIA 19) Planning Unit

2008) and may occur in the East and West Twin Rivers, Murdock Creek, Fielding Creek, and

Colville Creek (WRIA 19 Watershed Plan 2008).

The Lyre River is the only river in WRIA 19 that originates in an alpine area (approx.

5,500 feet) (NOPLE 2015). All other rivers and streams in WRIA 19 drain from low elevation

foothills ranging 2,000 to 3,500 feet in elevation (NOPLE 2015).

B. Water Resource Inventory Area 20

WRIA 20 begins at Cape Flattery and extends south to, but not including, Kalaloch Creek

(HSRG 2004). The area receives 80” to 240” of rain per year (McHenry et al. 1996). Wind and

heavy rainstorms are common (HSRG 2004). Inside Olympic National Park rests an undisturbed

temperate rainforest, which includes old-growth Sitka spruce, western hemlock, and silver fir

(Abies alba) (HSRG 2004). Outside of the park there is significant habitat disturbance, due in

large part to extensive clearcutting and roadbuilding that occurred during the 1960s-1980s

(HSRG 2004). Commercial forestry is the predominant land use in the lower reaches of WRIA 20

rivers (NPCLE 2020).

Olympic Peninsula steelhead occur in several large and small systems within WRIA 20

(Table 1). The winter and summer-run components of the Olympic Peninsula DPS are found in

the Hoh, Quillayute/Bogachiel, Calawah, and Sol Duc Rivers (Cram et al. 2018). Several other

rivers and streams support the winter run life history only, including the Dickey River, Goodman

Creek, Mosquito Creek, Ozette River, Tsoo-Yess (Sooes) River, and Wa’atch River (Cram et al.

2018).

1. Quillayute River System (Quillayute, Sol Duc, Bogachiel, Calawah, and Dickey Rivers)

The Quillayute River system includes the Quillayute, Sol Duc, Calawah, Bogachiel, and

Dickey Rivers, which collectively drain 628 square miles of land (Klinger et al. 2008). Nearly one-

third of the Quillayute River basin lies within Olympic National Park (32% of the Sol Duc River

basin, 29% of the Bogachiel River basin, and 20% of the Calawah River basin occur inside the

park (Houston and Contor 1984)). The Quillayute River proper, formed by the confluence of the

Bogachiel and Sol Duc Rivers, is only 5.6 miles long (Smith 2000).

a. Sol Duc River

The Sol Duc River originates in the northern Olympic Mountains and generally flows

west and northwest before leaving the boundary of Olympic National Park. After leaving the

park, the river flows south and west through 11.1 miles of the Olympic National Forest, where

more than half of the land is in late successional reserve (Smith 2000). Outside the Olympic

National Forest, the Sol Duc River winds through private and state-owned land. The watershed

receives 90-120 inches of precipitation per year and its upper reaches are influenced by the

rain-on-snow zone.

Table 1. Summary of watersheds and steelhead populations, including Watershed Resource Inventory Area (WRIA), whether the

watershed supports(ed) winter and/or summer run steelhead, period of record for monitoring data (usually redd counts) for winter

runs, mean annual run size and range, mean annual escapement and range, trend in annual run size or abundance, mean harvest

rate, percent of last 10-years the population achieved its escapement goal, average number of hatchery smolts released (winter and

summer run), and the population risk score generated by WDFW’s Steelhead at Risk report (Cram et al. 2018). N/A = not available or

applicable. N/A is used to describe “run size” in larger tributaries to Quillayute River and the Clearwater River (tributary of Queets

River) because the tribal fishery occurs almost solely below the tributary mouths and hence, the ultimate destination of the fish is

unknown. Consequently, only escapement can be estimated for those tributaries.

Watershed (Major

tributary)

WRIA

Winter

(W)

and/or

summer

run (S)

Monitoring

period of

record for

winter runs

Mean

(Range)

annual run

size winter

runs

Mean (Range)

annual

escapement

winter runs

Trend in

abundance

2,3,4

Mean

harvest rate

(range)

Percent last

10-years

met

escapement

goal

Average

hatchery

smolts

release 2009-

2013

(Winter,

Summer)

Population

risk score as

of 2013

(Cram et al.

2018)

Hoko River 19 W 1985-2020

N/A

566

(193 – 990)

40% decline

N/A

80% 24,000 (W) Low

Pysht River 19 W 1995-2020

N/A

499

(195 – 936)

21% decline

14%

70% 0 Mod.

Clallam River 19 W 1999-2020

N/A

158

(45 – 284)

27% decline

0.7%

No goal 2,000 (W) Mod.

Deep Creek 19 W 2010-2020

N/A

(47 – 129)

Declining

N/A

No goal 0 Insf.

West Twin River 19 W 2010-2020

N/A

(21 – 90)

Declining

N/A

0% 0 Insf.

East Twin River 19 W 2010-2020

N/A

(31 – 116)

Declining

N/A

10% 0 Insf.

Salt Creek 19 W 1995-2020

N/A

116

(32 – 237)

43% decline

3.9%

0% 0 Mod.

Lyre River 19 W & S

N/A N/A N/A N/A N/A

No goal

5,000 (W),

2,000 (S)

Insf.

Sekiu River 19 W

N/A N/A N/A N/A N/A

No goal 12,000 (W) Insf.

Sail River 19 W

N/A N/A N/A N/A N/A

No goal 10,000 (W) Insf.

Tsoo-Yess River 20 W

N/A N/A N/A N/A N/A

No goal 96,000 (W) Insf.

Ozette River 20 W

N/A N/A N/A N/A N/A

No goal 0 Insf.

Quillayute River 20 W & S 1978-2020

13,064

(6,456 –

21,615)

9,340

(5,500 – 16,919)

No trend

2,3

28%

(10% - 55%)

90%

243,000 (W),

47,000 (S)

N/A

Dickey R. (trib) 20 W 1978-2020

N/A

460

(143 – 1,607)

N/A See Quillayute 100%

0 Low

Sol Duc R. (trib) 20 W & S 1978-2020

N/A

3,864

(1,791 – 7,634)

N/A See Quillayute 70%

23,000 (W),

12,000 (S)

Low

Calawah R. (trib) 20 W & S 1978-2020

N/A

2,980

(989 – 5,985)

N/A See Quillayute 100%

40,000 (W),

35,000 (S)

Moderate

Bogachiel R. (trib) 20 W & S 1978-2020

N/A

1,975

(730 – 4,553)

N/A See Quillayute 60%

180,000 (W) Low

Goodman Creek 20 W 1995-2020

N/A

184

(45 – 374)

54% decline

6.8%

0% 4,000

High

Mosquito Creek 20 W

N/A N/A N/A N/A N/A N/A

0 Insf.

Hoh River 20 W & S 1980-2020

4,117

(2,541 – 5,783)

2,726

(1,616 – 4,593)

37% decline

33%

(7% - 55%)

60% 68,000 (W) Mod.

Kalaloch Creek 21 W

N/A N/A N/A N/A N/A No goal

0 Insf.

Queets River 21 W & S 1980-2018

7,648

(4,200 –

13,309)

4,845

(2,271 – 7,841)

45% decline

35%

(10% - 55%)

30% 157,000 (W) Mod.

Clearwater R.

(trib)

21 W & S

1980-2018

N/A

1,744

(847 – 2,966)

N/A See Queets

50% 0 Low

Quinault River 21 W & S 1978-2020

5,883

(2,179 – 9,726)

3,107

(1,366 – 5,774)

44% decline

46%

(15% - 65%)

No goal 486,000 (W) Low – Mod

Upper Quinault 21 W & S 1978-2020

N/A

1,511

(772 – 2,877)

N/A See Quinalt

1,200 0 Low

Moclips River 21 W 1988-2000

N/A

299

(130 – 560)

27% increase

N/A No goal

0 Low

Raft River 21 W

N/A N/A N/A N/A N/A No goal

0 Insf.

Copalis River 21 W

N/A N/A N/A N/A N/A No goal

0 Insf.

? = There are reports of small numbers of unclipped summer steelhead, 1) in the Lyre River, but it has never been determined whether those

were wild fish from the Lyre, strays from another river, or naturalized offspring of hatchery summer steelhead scatter plants in the Lyre River;

and 2) in the Clearwater River, which could be strays from the Queets River or naturalized offspring of hatchery summer steelhead scatter plants

in the Clearwater River.

NR = Hatchery releases have ended on these rivers.

1 = Data range is presumed to be only for winter runs. Data on run size and escapement of summer runs is almost entirely lacking, but redd

counts for winter steelhead may also include some summer steelhead since there is overlap in their spawning distribution.

2 = Source is Cram et al. (2018) Steelhead at Risk Report, with period of record ending in 2013.

3 = Source is McMillan et al. (2021).

4 = Source is North Olympic Peninsula Lead Entity (NOPLE) report (2015)

5 = Quillayute River population is not in statistically significant decline over its full period of record, but

it has declined sharply from 1996-2017 at a rate of 5,533 adults/decade (McMillan et al. 2022).

6 = Harvest rate is included in whole population estimate.

7 = Period of record 1995-2009, per Cram et al. 2018

8 = Period of record 1999-2013, per Cram et al. 2018

9 = Period of record 1995-2013, per Cram et al. 2018

b. Bogachiel River

The Bogachiel River is formed by the North and South Forks of the Bogachiel River,

which originate in the Olympic Mountains (Smith 2000). Its upper reaches are in Olympic

National Park, while its middle and lower reaches flow through timber and other agricultural

land (Smith 2000). The river’s most significant tributary is the Calawah River, an important

salmonid tributary itself (Smith 2000). The Bogachiel River’s other important salmonid

tributaries include Murphy, Maxfield, Weeden, Mill, Grader, and Dry Creeks (Smith 2000).

c. Calawah River

The Calawah River is formed by the North and South Forks of the Calawah River, which

originate in the Olympic Mountains (Smith 2000). The South Fork and its largest tributary, the

Sitkum River, provides spawning and rearing habitat for winter and summer steelhead and they

flow through late-successional reserve land inside the Olympic National Forest (Smith 2000).

The Nork Fork Calawah also supports winter and summer steelhead (Smith 2000).

d. Dickey River

The Dickey River is formed by the West, Middle, and East Forks of the Dickey River

(Smith 2000). The Dickey River is a low gradient system, and it has numerous wetlands and

sloughs (Smith 2000). Unlike the Sol Duc, Bogachiel, and Calawah Rivers, the Dickey River does

not originate inside Olympic National Park (Smith 2000). Rather, the river and its tributaries

occupy a heavily logged drainage (Smith 2000).

2. Hoh River

The Hoh River originates deep inside Olympic National Park, on the glacial slopes of Mt.

Olympus, and flows westward for 56 miles before emptying in the Pacific Ocean (Duda et al.

2018). The Hoh River basin drains nearly 299 square miles of land and has an extensive

floodplain (WCSSP 2013). The watershed receives 93-240 inches of rain per year (Klinger et al.

2008). The Hoh River has a strong glacial influence (McHenry et al. 1996).

Fifty-seven percent of the watershed is located within Olympic National Park, with the

remaining 43% flowing through state (24.4%), private, and tribal lands (Klinger et al. 2008). The

South Fork of the Hoh River, a major steelhead tributary, joins the mainstem of the Hoh River at

RM 30 (Smith 2000). Other significant salmonid tributaries to the Hoh River include Slide, Falls,

Mt Tom, Jackson, Taft, Snider, East Twin, Canyon, Spruce, Dismal, Pole, Tower, Lindner, Clear,

Willoughby, Elk, Alder, Winfield, Hell Roaring, Lost, Pins, Anderson, Nolan, Braden, and Fossil

Creeks (Smith 2000).

3. Ozette River

The Ozette River originates at the northern end of Ozette Lake and flows 5.3 miles west

to the Pacific Ocean (NPCLE 2020). Coal Creek is the largest tributary to the Ozette River (NPCLE

2020). Multiple tributaries drain into Lake Ozette, including the Big River and Umbrella,

Crooked, Siwash, South, Palmquist, Quinn, Elk, and Lost Net Creeks (NPCLE 2020).

4. Tsoo-Yess (Soose) and Wa’atch Rivers

The Tsoo-Yess and Wa’atch Rivers are short, rain-fed systems that originate in coastal

foothills located in the northeastern corner of WRIA 20 (NPCLE 2020). Both rivers flow through

the Makah Reservation and a small coastal strip of Olympic National Park (NPCLE 2020).

5. Goodman and Mosquito Creeks

Goodman and Mosquito Creeks are located to the north of the Hoh River and south of

the Quillayute River system. Both creeks originate in coastal foothills and flow through state

and private timberland lands and a coastal section of Olympic National Park before emptying

into the Pacific Ocean (NPCLE 2020).

C. Water Resource Inventory Area 21

WRIA 21 begins at Kalaloch Creek in the north and ends at Connor Creek in the south

(QINLE 2011). The area receives heavy rainfall, measuring 120” to 200” per year (HSRG 2004).

Prior to European American settlement, the area was covered by old growth western red cedar

(Thuja plicata), Sitka spruce, Douglas fir, and western hemlock (QINLE 2011). Today, the only

undisturbed forestland in WRIA 21 is in Olympic National Park. WRIA 21 has an extensive

coastal plain and 65 miles of marine shoreline (QINLE 2011).

Within WRIA 21, Olympic Peninsula steelhead occur in the Queets, Clearwater, Quinault,

Raft, Copalis, and Moclips Rivers and Kalaloch Creek (HSRG 2004). The Queets and Quinault

Rivers are large glacially-influenced systems that originate inside Olympic National Park

(McHenry et al. 1996, QINLE 2011, WCSSP 2013). The Clearwater, Raft, Copalis, and Moclips

Rivers and Kalaloch Creek are rain-dominant systems (WCSSP 2013).

1. Queets River

The Queets River originates on Humes Glacier on the southeast side of Mt. Olympus.

The river measures 51 miles long, drains 450 square miles of land, and flows at an average rate

of 8,000 cubic feet per second (cfs) in the winter and 1,015 cfs in the summer (McMillan 2006).

Nearly the entire course of the Queets River flows through Olympic National Park, with only the

lower eight miles running through the Quinault Indian Reservation before reaching the Pacific

Ocean (Smith and Caldwell 2001). Its major tributaries include the Clearwater, Sams, and

Salmon Rivers and Matheny and Tshletshy Creeks (Smith and Caldwell 2001). Steelhead spawn

in these tributaries, as well as several smaller ones (e.g., Miller Creek), and in the mainstem of

the Queets River (Smith and Caldwell 2001).

Olympic National Park owns 34 miles of tributary streams, including all of Tsheltchy

Creek and the lower five miles of the Sams River (Smith and Caldwell 2001). The USFS owns

84% of Matheny Creek watershed, 73% of the Sams River watershed, and 30% of the Salmon

River watershed (Smith and Caldwell 2001). The USFS manages these lands as riparian reserves,

late successional reserves, or adaptive management areas (Smith and Caldwell 2001). The

Quinault Tribe owns the lower eight miles of the Queets River and 54% of the Salmon River

drainage (McMillan 2006).

2. Clearwater River

The Clearwater River originates in the foothills of the Olympic Mountains (WCSSP 2013)

and flows 39 miles before emptying into the Queets River several miles upstream of the Pacific

Ocean. WDNR owns 79% of the Clearwater River watershed and roughly 20% is privately owned

(McMillan 2006). Its major tributaries include the Sollecks and Snahapish Rivers and Christmas

and Stequaleho Creeks (Smith and Caldwell 2001). The upper watershed receives 120” to 160”

of rain per year (McHenry et. al 1996).

3. Quinault River

The headwaters of the Quinault River begin in the Mt. Lawson and the Enchanted Valley

watersheds (Smith and Caldwell 2001). The river measures 69 miles long, drains 434 square

miles of land, and flows at an average rate of 6,300 cfs in the winter and 1,080 cfs in the

summer (McMillan 2006). The river feeds into Lake Quinault, which spans 3,729 acres. Below

the lake, the river runs southwesterly through the Quinault Indian Reservation for 33 miles until

it meets the Pacific Ocean (Smith and Caldwell 2001).

Nearly half (47%) of the Quinault River basin occurs within Olympic National Park

(Houston and Contor 1984). The entire North Fork and most of the East Fork occur inside in the

park. Roughly one-third (32%) of the basin occurs on the Quinault Indian Reservation, with the

remaining 13% located on USFS land and 4% on private land (Smith and Caldwell 2001).

Winter and summer steelhead occur in the Quinault River (Smith and Caldwell 2001).

Winter steelhead spawn in the mainstem below Lake Quinault, the North Fork of the Quinault

River, and the following creeks: Cook, Elk, Willaby, Falls, Gatton, Zeigler, Kestner, Inner, Slough,

Alder, Big, Fox, Fletcher, Boulder, Ten O’clock, Canoe, Irely, and Bunch Creeks (Smith and

Caldwell 2001). Summer steelhead spawn in the East Fork and North Fork of the Quinault River

(Brenkman et al. 2012).

4. Raft River

The Raft River is located between the Queets and Quinault Rivers (WCSSP 2013). The

basin is 71,824 acres in size and includes the Raft River, North Fork Raft River, Red Creek and

several independent tributaries (Smith and Caldwell 2001). The river originates in the foothills

of the Olympic Mountains and flows through a coastal plain before emptying into the Pacific

Ocean. The Raft River is 11.5 miles long and occurs almost entirely within the Quinault Indian

Reservation (81%). The remaining 19% of the basin is private land (Smith and Caldwell 2001).

5. Moclips River

The Moclips River and its main tributary, the North Fork Moclips River, originate in the

foothills to the south of the Quinault River. Together, the Moclips River and its north fork

measure 17 miles long and flow west to the Pacific Ocean. The Moclips River basin is 53,528

acres in size, most of which is private land (54%) or owned by the Quinault Indian Nation (39%)

(Smith and Caldwell 2001). The state owns the remaining 7% of the basin (Smith and Caldwell

2001). Winter steelhead spawn in the Moclips River mainstem, the North Fork of the Moclips

River, and Wreck Creek (Smith and Caldwell 2001).

6. Copalis River

The Copalis River is a low gradient system located to the south of the Moclips River. The

river originates in the foothills and flows through a coastal plain for a total of 24 miles before

reaching the Pacific Ocean (WSCCP 2013). The Copalis River basin is 36,818 acres in size, nearly

all of which is privately owned (95%), with the remaining 5% owned by the state (Smith and

Caldwell 2001). Winter steelhead spawn in the river’s mainstem (Smith and Caldwell 2001).

7. Kalaloch Creek

The Kalaloch Creek Basin spans 13,649 acres and includes Kalaloch Creek and four

unnamed tributaries. Most of the land is owned by WDNR (41%) or private interests (40%).

Olympic National Park owns 18% of the basin and the Quinault Indian Nation owns 1% (Smith

and Caldwell 2001). Kalaloch Creek supports winter steelhead, which spawn in the lower

mainstem and the West Fork of Kalaloch Creek (Smith and Caldwell 2001).

IV. LIFE HISTORY

Steelhead emerge from the gravel from spring through early summer and spend one to

four years in freshwater before migrating to the ocean, where they spend one to four years

before returning to freshwater to spawn (Busby et al. 1996). Steelhead smolts on the Olympic

Peninsula typically migrate to the ocean from April through June, with peak emigration

occurring in mid- to late-May (Busby et al. 1996).

Juvenile steelhead are present off the Washington coast in May and June, with only a

few lingering into July and August (Pearcy et al. 1990). They migrate north and spend their first

summer in the Gulf of Alaska and North Pacific Ocean (Atcheson et al. 2012). Afterwards they

distribute throughout their north Pacific Ocean range (Atcheson et al. 2012). Some will go as far

west as the Kuril Islands (Atcheson et al. 2012).

Steelhead distribution in the Pacific Ocean appears to be driven by temperature and

salinity (Okazaki 1983, Sutherland 1973, Light et al. 1989). The south-north distribution of

steelhead corresponds with a 3° C to 16° C temperature range, with nearly 96% of steelhead

occurring in waters measuring at or below 12° C (Sutherland 1973, as cited in Light et al. 1989).

The northern boundary, which is slightly above the Aleutian Islands, may correspond with

factors other than sea surface temperatures, such as salinity (Light et al. 1989). Steelhead are

generally found within 10 meters of the ocean’s surface (Godfrey et al. 1975, as cited in Light et

al. 1989).

Except for half-pound steelhead stocks, “[i]nformation from tagging studies shows little

or no differences in ocean distribution among stocks, groups, or races.” (Light et al. 1989).

Therefore, Olympic Peninsula steelhead likely follow the same ocean distribution pattern as all

other steelhead, not including half-pounders.

The Olympic Peninsula steelhead DPS includes seven summer-run and twenty-four

winter-run populations (Cram et al. 2018). The summer-run populations return from May

through October (McHenry et al. 1996) and spawn from January through April; however, the

populations remain largely unstudied and, consequently, their status and trends are unknown

(WDFW 1992; Busby et al. 1996; McMillan 2006; Brenkman et al. 2012; Cram et al. 2018).

Historically, a substantial proportion of the winter-run populations returned between

December and January (McMillan et al. 2022). Currently, the winter-run component returns

and spawns between January and May, with a peak in March through April depending on the

population (Busby et al. 1996; McMillan et al. 2007; McMillan et al. 2022).

Steelhead are iteroparous (Busby et al. 1996). Steelhead on their first spawning run are

called “maiden” fish (Light et al. 1989). Maiden fish that survive spawning and return to the

ocean are known as “kelts” (Behnke 2002). Steelhead may return to spawn up to five times

within their lifetimes; however, most steelhead will only spawn once (Busby et al. 1996; Kendall

et al. 2015).

V. HABITAT REQUIREMENTS

Steelhead are poikilotherms and evolved to survive within a range of stream and ocean

temperatures. In general, steelhead prefer water temperatures between 10-15°C, but this

varies by life stage and population (McCullough et al. 1999; Hicks 2000). For instance, Fuss

(1998) suggests 5-11°C is optimal for steelhead egg survival and eggs and Rombough (1988)

found alevins that experience water temperatures over 15°C are smaller than individuals that

experience cooler water temperatures. This suggests optimal incubation temperatures are

below 11-12°C (Hicks 2000). The most favorable temperature range for juvenile growth is

between 5-17°C, with the optimal temperatures varying by season (Hicks 2000). Temperatures

over 19°C can limit the occurrence and growth of juvenile steelhead if food is not sufficient,

with lethal temperatures beginning at around 27°C (Hicks 2000). Temperatures of 11.3-12.3°C

are consistently cited as the uppermost constant temperature exposures that will not interfere

with smoltification (Zaugg et al. 1972; Zaugg and Wagner 1973; Zaugg 1981), with detrimental

effects occurring once water temperatures begin to exceed 12.7°C. Adult migration in winter

typically occurs when temperatures are relatively cool, but there is evidence that upstream

migration by adults stops when temperatures reach 20-21°C, which is a concern for summer

steelhead (Hicks 2000). Adult steelhead can begin to die when water temperatures reach 21.6-

23.8°C and generally, steelhead avoid water temperatures of 24°C and above (unless they are

genetically adapted to warmer water temperatures) (Hicks 2000).

Steelhead eggs and juveniles require clean gravel. If fine-grained sediment exceeds 12-

20% of the surface, young steelhead are buried and suffocate (Phillips et al 1975; Reiser and

White 1988; Jensen et al. 2009). Sedimentation in streams can also negatively impact growth

and survival of juvenile salmonids in different ways (Suttle et al. 2004). For example, high levels

of fine sediment can decrease prey, leading to starvation (Murphy et al. 1981).

Large woody debris (LWD) is also an important component of steelhead habitat

(Cederholm et al. 1997; Thompson et al. 2012). LWD stabilizes stream channels, creates pools,

reduces erosion, and provides cover, resting, and feeding areas for juveniles (Bilby and Bisson

1998). Complex floodplain habitat is also important to steelhead survival, as side channels and

other natural features provide feeding habitat, refuge from high water events, and other

benefits (Beechie et al. 1994; Montgomery et al. 1999; Bellmore et al. 2013).

Steelhead adapted to natural hydrographs, and juvenile and adult migrations are timed

to coincide with historic flow regimes that maximize survival (Bjornn and Reiser 1991). Changes

in flow regimes due to climate change and other anthropogenic causes (e.g., water

withdrawals) interfere with these major life history stages, increasing stress and susceptibility

to disease, and, occasionally, leading to direct mortality (e.g., stranding due to low streamflows)

(Bjornn and Reiser 1991; Wade et al. 2013).

VI. DIET

A. Freshwater

Steelhead fry consume aquatic invertebrates, including chironomids, mayflies, and

terrestrial macroinvertebrates (Davis 2015). One-year-old juveniles mostly consume insect

larvae and pupae, adult insects, and crustaceans (Davis 2015). By age two, juvenile steelhead

are larger in size and can consume larger and more diverse prey items, such as fish larvae and

small fish (Davis 2015).

B. Marine

Steelhead smolts appear to spend a relatively short period of time feeding in nearshore

waters before dispersing across the vast North Pacific Ocean. In the 1980s, researchers purse

seined the coasts of Washington and Oregon to collect juvenile steelhead and study their

stomach contents (Pearcy et al. 1990). The researchers intercepted juvenile steelhead in May

through August; however, the peak catch occurred in May and June. Their stomach contents

contained a variety of fish (juvenile rockfish, sandlance, brown Irish lord, and greenlings),

euphasids, barnacle larvae, copepods, and other crustraceans (Pearcy et al. 1990). The

researchers estimated that steelhead grow by about 1.1 mm per day between May and early

July.

Steelhead in the open ocean consume a wide variety of prey. Sampling of steelhead

diets in the Gulf of Alaska and Central North Pacific found that the dominant prey were fish and

squid, with the primary species within those categories being Atka mackerel (Pleurogrammus

monopterygius), three-spined stickleback (Gasterosteus aculeatus), lantern fish (Myctophidea spp.),

and the minimal armhook squid (Berryteuthis anonychus) (Atcheson 2010). The study also found

that steelhead consume euphausiids, copepods, amphipods, polychaetes, crustacean larvae, and,

unfortunately, some styrofoam and plastic (Atcheson 2010). Plastic and styrofoam are of particular

concern because it could result in mortality, toxicity, or delayed effects, such as heritable

alterations in gene expression (Myers et al. 2013).

VII. NATURAL MORTALITY

A. Freshwater

Although a tremendous amount of research has been conducted on the survival of adult

steelhead, far less information is available for specific mortality rates for juveniles in freshwater

across a gradient of environments. Mortality at the egg stage can be very high, with survival

from egg to emergence ranging from 18-95% depending on factors such as temperature, intra-

gravel flow, and fine sediment (Shapovalov 1937; Sheppard 1972; Phillips et al. 1975; Bjornn

1978; Biley and Moring 1988). However, Shapovalov and Taft (1954) experimentally concluded

that under ideal conditions, survival to emergence averages about 80-90%. Bjornn (1978) and

Phillips et al. (1975) also conducted experiments and found a similar upper limit as Shapovalov

and Taft (1954) when fine sediment was absent from the redd, but survival was only 18% when

70% of the redd consisted of fine sediment.

After emergence, evidence suggests - as with most salmonids - that the mortality rate is

high during the initial weeks to months of life when juvenile fish are small, weak, and fairly

immobile. For example, Burns (1971) estimated June-to-October survival of young-of-the-year

salmonids in a California stream was only 27% (range = 20-29%), while age-1 and older juveniles

survived at a rate of 56% (range = 34-94%). Bjornn (1978) estimated mortality rate of 80-90%

during the first summer of life, with an overall mortality of 94-99% upon completing the first

year of life. A review of these papers and others suggests that mortality rates almost always

exceed 50-60% early in life and thereafter, conditions being equal, survival tends to increase as

fish size and condition increases (Biley and Moring 1988).

Mortality also occurs as smolts migrate to and enter the ocean. For example, Melnychuk

et al. (2007) found that 65-73% mortality of juvenile steelhead occurred in the first month of

smolt migration. Romer et al. (2012) reported that 63-89% of the steelhead smolts they tagged

survived a short freshwater migration to the estuary, but that 50-60% of the fish died in the

estuary despite most fish spending less than one day in the estuary.

B. Marine

Quinn (2005) estimated that approximately half of steelhead life cycle mortality occurs

in the ocean. Typically, smolt to adult survival rates in the ocean is between 2% and 10% (Quinn

2005). Steelhead ocean survival is largely dependent on sea surface temperature during their

first summer in the ocean (Atcheson 2010).

VIII. POPULATION STRUCTURE

Olympic Peninsula steelhead are distinct from other distinct population segments,

including the Southwest Washington and Puget Sound DPSs (Busby et al. 1996). That distinction

still holds true. WDFW acknowledges it is “not yet able to fully evaluate genetic population

structure to aid the process of verifying the 1992 population definitions for Olympia [sic]

Peninsula and SW Washington DPSs” (Cram et al. 2018).

The Olympic Peninsula Steelhead DPS includes summer-run and winter-run populations

(Table 1; Busby et al. 1996). WDFW considers summer steelhead populations in the Sol Duc,

Bogachiel, Calawah, Hoh, Queets, Clearwater, and Quinault Rivers to be distinct from each

other based on geographic isolation (WDF et al. 1993). WDFW considers the following winter

steelhead population to be distinct stocks: Salt Creek/Independents (Salt, Whiskey, Colville, and

Field Creeks), Lyre River, Pysht/Independents (Pysht River, Deep Creek, East Twin River, and

West Twin River), Clallam River, Hoko River, Sekiu River, Sail River, Sooes/Waatch (Sooes and

Waatch Rivers), Ozette River, Sol Duc River, Quillayute/Bogachiel (Quillayute and Bogachiel

Rivers), Calawah River, Dickey River, Goodman Creek, Mosquito Creek, Hoh River, Kalaloch

Creek, Queets River, Clearwater River, Quinault/Lake Quinault (Lower Quinault and broodstock

net pet steelhead in Lake Quinault), Quinault River (above Lake Quinault), Raft River, Moclips

River, and the Copalis River (WDF et al. 1993).

IX. PREVIOUS STOCK ASSESSMENTS

There have been several stock assessments for the largest wild winter steelhead

populations in the Olympic Peninsula DPS (Table 2 and Table 3) but assessments for smaller

populations with less consistent data are more limited. Below we list and summarize the

assessments that have been conducted to date. No assessments are available for wild summer

run steelhead because data on those populations is almost entirely lacking (except for data

collected by the National Park Service (Brenkman et al. 2012) and John McMillan (Science

Director, The Conservation Angler: McMillan 2022)).

Table 2. Reports that reviewed the status and trends of wild winter steelhead populations in

the Quillayute, Hoh, Queets, and Quinault Rivers. Only the HSRG 2004 included an evaluation of

summer steelhead.

Reference Status

Nehlsen et al. 1991 Healthy, stable to increasing trend

WDF et al. 1993 Healthy

Busby et al. 1996

Stable to increasing; status review

determined ESA-listing was not warranted

McHenry et al. 1996 Healthy, stable to increasing trend

WDFW 2002 Healthy, stable to increasing trend

HSRG 2004

At risk (summer runs listed as critical, except

for Quinault population which is “at risk”)

Kendall et al. 2017

Declining trend in Hoh and Queets River

populations

Cram et al. 2018

Declining trend in Hoh, Queets, and

Quinault River populations

McMillan et al. 2022

Declining trend in Hoh, Queets, and

Quinault River populations

A. American Fisheries Society (1991)

In 1991, the American Fisheries Society published a list of 214 native, naturally spawning

salmonid stocks that were either at high or medium risk of extinction or were a species of

concern (Nehlsen et al. 1991). This list did not include Olympic Peninsula steelhead (Nehlsen et

al. 1991).

B. Washington Department of Fish and Wildlife (1992)

In 1992, the Washington Department of Fisheries (“WDF”), Washington Department of

Wildlife, and the Western Washington Treaty Indian Tribes assessed the status of Olympic

Peninsula winter and summer steelhead stocks (WDF et al. 1993). For WRIA 19 populations,

WDF et al. (1993) identified the Hoko River and Pysht/Independent winter stocks as “healthy.”

WDF et al. (1993) defined “healthy” to mean the stock was “experiencing stable escapement,

survival, and production trends and not displaying a pattern of chronically low abundance.” The

Salt Creek/Independents and Sekiu, Sail, Lyre, and Clallam River populations were listed as

“unknown” (WDF et al. 1993).

For WRIA 20 populations, WDF et al. (1993) identified five winter steelhead stocks as

“healthy” and five as “unknown.” The five healthy stocks included winter steelhead in the

Quillayute/Bogachiel, Dickey, Sol Duc, Calawah, and Hoh Rivers (WDF et al. 1993). WDF et al

(2013) could not determine the status of winter steelhead populations in the Sooes, Wa’atch,

or Ozette Rivers or Goodman, Mosquito, or Kalaloch Creeks (WDF et al. 1993).

For WRIA 21 populations, WDF et al. (1993) determined that four out of five winter

steelhead stocks were healthy, including the Quinault, Moclips, Queets, and Clearwater River

populations. The Departments and the tribes could not determine the status of the Copalis

River population (WDF et al. 1993).

Except for the Queets River summer run population, which WDF et al. (1993) designated

as “healthy,” the departments and tribes could not determine the status of summer steelhead

in the Sol Duc, Bogachiel, Calawah, Hoh, or Clearwater Rivers (WDF et al. 1993).

C. National Marine Fisheries Service (1996)

In 1996, NMFS reviewed the status of Olympic Peninsula steelhead (Busby et al. 1996).

At the time, seven of the twelve winter steelhead stocks that had population trend data were

declining or had relatively flat trends, while the other five were increasing (Busby et al. 1996).

The declining and stable stocks included the following winter steelhead populations: Pysht River

(-5.8%), Hoko River (-7.6%) Quillayute/Bogachiel River (-0.2%), Dickey River (-4.4%), Sol Duc

River (-0.1%), Clearwater River (-0.5%); and Quinault River/Lake Quinault (-2.6%) (Busby et al.

1996). The increasing stocks included the following winter steelhead populations: Calawah

River (1.1%), Hoh River (0.2%), Queets River (0.9%), Upper Quinault River (1.8%), and the

Moclips River (13.6%) (Busby et al. 1996). No population trend data were available for summer

steelhead (Busby et al. 1996).

At the time, most Olympic Peninsula steelhead populations appeared to be self-

sustaining because there were no strong abundance trends (Busby et al. 1996). Additionally,

WDFW provided information indicating that there was “substantial temporal separation

between hatchery and native winter steelhead” (Busby et al. 1996). However, NMFS noted

there were “isolated problems” with the population sustainability, including declining trends in

the Pysht/Independents and Quinault River populations (Busby et al. 1996). NMFS also noted

there was a “substantial contribution” of hatchery spawners in the Quinault River population

(Busby et al. 1996). Based on those facts, NMFS determined that Olympic Peninsula steelhead

did not warrant listing at the time (61 Fed Reg. 41541, 41550).

However, the biological review team expressed several concerns regarding Olympic

Peninsula steelhead (61 Fed Reg., at 41550).

“*** [T]he BRT has several concerns about the overall health of this ESU and

about the status of certain stocks within it. The majority of recent abundance

trends are upward (including three of the four largest stocks), although trends in

several stocks are downward. These downward trends may be largely due to

recent climate conditions. There is widespread production of hatchery steelhead

within this ESU, largely derived from a few parent stocks, and this could increase

genetic homogenization of the resource despite management efforts to

minimize introgression of the hatchery gene pool into natural populations.

Estimates of the proportion of hatchery fish on natural spawning grounds range

from 16% to 44%, with the two stocks with the largest abundance of natural

spawners (Queets and Quillayute) having the lowest hatchery proportions.”

(Busby et al. 1996).

D. Washington Department of Fish and Wildlife (2002)

In 2002, WDFW updated its status assessment of Olympic Peninsula steelhead

populations (WDFW 2002). The 2002 assessment downgraded the Quinault River “mixed”

origin winter steelhead stock from “healthy” to “depressed” (WDFW 2002). It changed the

statuses of Queets River summer steelhead and Moclips River winter steelhead from “healthy”

to “unknown” (WDFW 2002). Finally, it changed the Goodman Creek, Pysht/Independents, and

Salt Creek/Independents winter steelhead population statuses from “unknown” to “healthy”

(WDFW 2002).

E. North Olympic Peninsula Lead Entity (2004)

In 2004, a technical review team for the North Olympic Peninsula Lead Entity for Salmon

(“NOPLE”) assessed winter steelhead populations in WRIA 19. The team determined that the

following winter steelhead populations were “depressed”: Sekiu River, Clallam River, West Twin

River, East Twin River, and the Western Strait Independents populations (NOPLE 2015). The

team identified the Hoko River, Pysht River, and the aggregated Salt Creek populations as

“healthy” and the Lyre River population as “unknown” (NOPLE 2015).

F. Hatchery Scientific Review Group (2004)

In 2004, the Hatchery Scientific Review Group (HSRG) issued a report on hatchery

operations on the north coast of Washington (HSRG 2004). The report included population

viability ratings from managers of Olympic Peninsula winter and summer steelhead populations

(HSRG 2004). To assess population viability, they considered multiple factors such as age class

structure, spawner escapement, and proportion of hatchery-origin fish in natural spawning

populations (HSRG 2004). Each stock’s viability was rated as “critical,” “at risk,” or “healthy”

(HSRG 2004).

The report did not list any wild Olympic Peninsula steelhead populations as “healthy”

(HSRG 2004). The report identified the following populations’ viability as “critical”: Copalis River

winter steelhead, Goodman Creek winter steelhead, Hoh River summer steelhead, Moclips

River winter steelhead, Mosquito Creek winter steelhead, Kalaloch Creek winter steelhead,

Ozette River winter steelhead, Quillayute River system summer steelhead, and Sooes River

winter steelhead (HSRG 2004). The report rated the following populations as “at risk”: Hoh

River winter steelhead, Hoko River winter steelhead, Queets River winter steelhead, Quillayute

system winter steelhead, Quinault River summer steelhead, and Quinault River winter

steelhead (HSRG 2004). HSRG could not assess the viability of Queets River summer steelhead

(HSRG 2004).

G. Steelhead at Risk Report (2018)

In 2018, WDFW released the Steelhead at Risk Report, which assessed steelhead

populations statewide based on the available, albeit limited, data at the time (Cram et al. 2018).

Importantly, the last year of data the report covered was from 2013 and, therefore, it did not

include data on the significant declines in abundance that occurred from 2014 through 2018.

The authors noted that the “the lack of abundance, productivity, and diversity data was the

most common impediment to conducting wild steelhead status assessments statewide” (Cram

et al. 2018). The authors also explained they had limited productivity data. Cram et al. (2018)

acknowledged “very little is known about temporal and spatial patterns of freshwater

population productivity (smolts per spawner) and smolt to adult return rates (SAR) for Olympic

Peninsula DPS wild steelhead populations, and this is a substantial data gap” (Cram et al. 2018).

However, the authors observed that smolt-to-adult return ratios (SARs) for Olympic Peninsula

steelhead have declined significantly over time (Cram et al. 2018). Additionally, the authors

highlighted that between brood years 2005 and 2010, the population productivity of Olympic

Peninsula steelhead was lower than the productivity of the lower Columbia River, middle

Columbia River, Upper Columbia River, and Snake River steelhead DPSs (Cram et al. 2018).

When possible, Cram et al. (2018) generated a risk score based on short- and long-term

trends in run size, extinction risk, and proportion of time the population met its escapement

goal. Again, the report did not include escapement data for years 2014-2018 and, therefore, it

did not consider significant failures to meet escapement goals during those years. They found

that 11 (73%) of 15 winter steelhead populations displayed a decreasing trend in abundance

(Table 3), but the Lower Quinault River was the only population that met the high-risk standard

for long-term abundance. They also determined Goodman Creek and the Calawah River were

considered as high risk based on declines in short-term abundance, while Goodman Creek, Salt

Creek, and Clallam River populations were identified as having a high risk of extinction (Table

3). Of the 13 populations with defined escapement goals, Cram et al. (2018) identified six

populations at high risk for failure to meet escapement goals, but also noted that 54% of

populations had met escapement goals in at least 70% of last 10 years (2004 – 2013).

Table 3 (from Cram et al. 2018). Risk assessment results and ratings for Olympic Peninsula

Steelhead DPS populations. Red text indicates values that exceeded specific criterion for each

metric. Status relative to abundance goal represents the percentage of years the population

achieved its escapement goal. Abbreviations: win. = winter steelhead; sum. = summer

steelhead; insuf. = insufficient.

In addition to lacking summer steelhead data, the data was insufficient to determine

extinction risks for the following winter steelhead populations: Copalis River, Kalaloch Creek,

Lyre River, Mosquito Creek, Ozette River, Raft River, Sekiu River, Sail River, Sooes River, and

Wa’atch River populations (Cram et al. 2018).

According to the results for all four risk metrics, Cram et al. (2018) determined that one

population was at a high total risk, seven at moderate total risk, seven at low total risk (Table 3)

and 16 at undetermined risk.

X. ABUNDANCE AND POPULATION TRENDS

Based on Cram et al. (2018), McMillan et al. (2022), NOPLE (2015), run size and

escapement estimates by WDFW and tribes (co-managers), and an online WDFW data

repository (https://fortress.wa.gov/dfw/score/score/species/steelhead.jsp?species=Steelhead,

Accessed online 3/20/2022), we compiled a list of all the major and small watersheds known to

support wild steelhead by their respective Watershed Resource Inventory Area (WRIA) and

other factors, such as mean annual abundance, in cases where such data was available (Table

1). The populations are predominantly or solely the winter run life history, and the most

abundant populations are found in the Quillayute, Queets, Quinault, and Hoh Rivers, while the

remaining rivers and creeks support much smaller populations (Table 1). Data is almost totally

lacking for summer steelhead, and they are only thought to be present in the four largest

watersheds (Quillayute, Hoh, Queets, and Quinault Rivers). However, snorkel survey data

indicates that summer run populations are at critically low levels and close to extirpation (Table

1). Below we summarize information on the historic and current trends in abundance and

escapement of wild steelhead where such information is available.

A. Historical Abundance Winter and Summer Steelhead

Information on the historic abundance of wild winter steelhead is only available for the

four largest populations of wild winter steelhead in the Olympic Peninsula DPS, which include

the Quillayute, Hoh, Queets, and Quinault River populations (McMillan et al. 2022). There is

also some historical catch information for summer steelhead in each watershed, which we

reference later. We could not find historical data for the other, smaller populations of wild

winter steelhead. McMillan et al. (2022) analyzed old cannery, sport, and tribal catch records to

estimate historic run size and run timing for the Quillayute, Hoh, Queets, and Quinault River

populations. Information on historic releases of hatchery winter and summer steelhead are

available from Duda et al. (2018), which provides a thorough summary of all releases of

hatchery steelhead for watersheds draining from Olympic National Park.

Although there is no formal analysis of historical catch of summer steelhead in the four

largest watersheds known to support them (Quillayute, Hoh, Queets, and Quinault Rivers), we

include recreational and tribal catch data for a few years during the 1950’s when data was

occasionally available. We also include catch data from 1962-1977, which mostly consists of

data from sport record catch cards collected annually by WDFW (Summer steelhead catch data

is in excel file “Summer steelhead catch data for Quillayute_Hoh_Queets_Quinault River

populations” and based on data from .pdf file “Historic winter and summer steelhead tribal

catch by month Hoh, Quinault, Queets, Quillayute 40s-70s”). We ended the “historical” period

in 1977 because that was the first year that hatchery summer steelhead were released into two

tributaries of the Quillayute River system (Calawah and Sol Duc Rivers). It was not possible to

determine the relative proportion of hatchery and wild fish in the catch after 1977 because

hatchery summer steelhead were not outwardly marked with a fin clip until 1985. Regardless,

this data was available to Busby et al. (1996), but it was not evaluated in their status review.

WDFW and the treaty tribes (collectively, the “co-managers”) considered steelhead to

be “summer runs” if they were caught from May 1 through October 31, which is slightly

problematic because that suggests they were also potentially catching spawned out winter run

kelts in May and June that were emigrating back to the ocean (McMillan 2006). Kelts can be

distinguished from unspawned summer steelhead upon capture, but it is not known whether

such distinctions were made. Regardless, the overall catch of summer steelhead is quite small

in most years and inconsistent across months, which generally suggests that large numbers of

kelts were not reported in the catch.

We summarized cumulative summer steelhead catch by year and watershed and then

generated a very rough estimate of run size by assuming a harvest rate of 25% and calculating

lower and upper bounds at 15% and 35% harvest rates. We do not assume the estimates of

summer run are robust owing to the simple calculation. Nonetheless, we provide them because

they offer some insight into how catch could translate to abundance.

Below, we summarize what is known about the historical winter steelhead and summer

steelhead catch and run sizes. Overall, the estimates suggest the largest watersheds historically

contained abundant runs of winter and summer steelhead, with fresh steelhead likely entering

the rivers each month of the year (McMillan 2006).

Table 4 (modified from McMillan et al. 2022). Comparison of historical mean annual wild winter steelhead abundance estimates

based on cannery record data, expansion of historical commercial and recreational catch data, historical commercial fishery catch

per unit effort (CPUE), and accessible stream kilometers (SKM) of habitat, an ensemble historical (circa 1948 - 1960) mean estimate,

the contemporary (circa 1980 - 2017) mean estimate, and the percent decline of each population in relative to the ensemble

historical estimate. Percent decline is the difference between the ensemble historical abundance estimate and the contemporary

mean abundance. The percent decline reported for the most recent five-year period in McMillan et al. (2021) refers to 2013-2017

(except for Quinault which referred to 2009-2013). Since the estimate in McMillan et al. (2021) does not capture the most recent run

sizes, we updated the estimate so that the comparison between the ensemble estimate and the most recent five-year period for

each population refers to 2016-2020.

Population

Cannery

records

Historical

catch

Historical

CPUE

Accessible

habitat

(SKM)

Ensemble

historical

abundance

estimate

Contemporary

mean

abundance

Percent

decline

Most recent

5-yr mean

abundance

(~ 2016-2020)

Percent

decline for

most recent

5-yr

Quillayute 22,567 23,391 19,571 21,843 13,595 38 % 8,528 61 %

Hoh 15,923 14,160 10,431 13,505 4,206 69 % 2,880 79 %

Queets 32,659 19,875 13,553 12,144 15,191 7,648 50 % 4,658 69 %

Quinault 13,743 22,226 14,723 16,897 6,181 63 % 3,370 80 %

Ensemble historical estimate does not include cannery record estimate for Queets in 1923

1. WRIA 20 Populations

a. Hoh River

Using historical catch data, McMillan et al. (2022) estimated the average historical run

size from 1948 through 1960 to be 15,923 winter steelhead (lower and upper bounds for

historic mean: 9,023, 24,901) (Table 4). With that approach, the lowest and highest single-year

estimate for the Hoh River population was 7,118 and 24,684 winter steelhead in 1948 and

1956, respectively. Using historical catch per unit effort (CPUE) and total accessible habitat

produced average historical run size estimates of 14,160 and 10,431 steelhead, respectively

(Table 4). The average ensemble estimate for annual run size for all three approaches was

13,505 steelhead (Table 4).

Estimates of historic run size prior to the modern monitoring period are not available for

summer run steelhead. Because they were not a target species less data is available on historic

catch. We could only find two years of harvest data prior to 1962 (i.e., tribal catch in 1954 and

1957: data is in excel file “Summer steelhead catch data for Quillayute_Hoh_Queets_Quinault

River populations” and based on data from .pdf file “Historic winter and summer steelhead

tribal catch by month Hoh, Quinault, Queets, Quillayute 40s-70s”), after which sport catch

record card data is available in addition to a few years of tribal catch up to 1977. We consider

this timeframe to be the “historical” period because it is prior to the onset of releases of

hatchery summer steelhead that began in 1977. Mean catch of summer steelhead for the

period beginning in 1954 and ending in 1977 was 118 fish and ranged from a low of 38 fish to a

high of 274 fish (Table 5).

Assuming a 25% harvest rate and using bounds of 15% and 35% harvest rates, we

estimated a mean run size of 472 summer steelhead (Lower bound = 337, Upper bound =

1,179) with a low run size of 152 fish and a high of 1,096 (Table 6). The mean of the top-five run

sizes is 884 summer steelhead (Lower bound = 631, Upper bound = 1,473).

Although there is uncertainty associated with our estimations, wild summer steelhead

were formerly abundant in the Hoh River. Further, it is likely that the run sizes were larger if the

harvest rates are lower than we assume and not all the catch was reported, both of which are

plausible. Last, the earliest data point we found was from 1954, which is decades after the

onset of commercial and recreational fisheries and industrial logging practices. Therefore, it is

almost certain that wild summer steelhead were more abundant than we report here.

b. Quillayute System

Using historical catch, McMillan et al. (2022) estimated that the average historical

abundance for winter steelhead in the Quillayute River from 1948-1960 was 22,567 steelhead

(lower and upper bounds for the historical mean: 16,733, 31,591) (Table 4). Annual variability

within the historical period ranged from the lowest and highest single year estimates in the

Quillayute River of 6,702 and 34,757 winter steelhead in 1948 and 1951, respectively. Using

historical CPUE and total accessible habitat produced average historical run size estimates of

23,391 and 19,571 steelhead, respectively (Table 4). The average ensemble estimate for annual

run size for all three approaches was 21,843 steelhead.

Estimates of historic run size are not available for summer run steelhead. Because they

were not a target species less data is available on historic catch. However, there is evidence

that summer steelhead were present in each of the three largest tributaries, including the

Bogachiel, Sol Duc, and Calawah Rivers (McMillan 2006). We summarized all the catch data for

the tributaries as part of the Quillayute River system (Data is in excel file “Summer steelhead

catch data for Quillayute_Hoh_Queets_Quinault River populations” and based on data from

.pdf file “Historic winter and summer steelhead tribal catch by month Hoh, Quinault, Queets,

Quillayute 40s-70s”). Mean catch of summer steelhead in 1946 and from 1962-1977 was only

67 fish with annual catch ranging from 12 to 309 fish (Table 5), suggesting not many fishers

focused on catching summer steelhead in the Quillayute system.

Assuming a 25% harvest rate and using bounds of 15% and 35% harvest rates, we

estimated a mean run size of 268 summer steelhead (Lower bound = 191, Upper bound = 670)

with a low run size of 48 fish and a high of 848 (Table 6). The mean of the top-five run sizes is

478 summer steelhead (Lower bound = 342, Upper bound = 797). If the population was

harvested with the range of rates we proposed, then wild summer steelhead were once quite

abundant in the Quillayute River system.

2. WRIA 21 Populations

a. Quinault River

Using historical catch, McMillan et al. (2022) estimated the average historical

abundance for winter steelhead in the Quinault River from 1948-1960 was 13,743 steelhead

(lower and upper bounds for the historical mean: 9,345, 20,258) (Table 4). Annual variability

within the historical period ranged from the lowest and highest single year estimates in the

Quinault River of 7,475 and 30,332 winter steelhead in 1956 and 1952, respectively. Using

historical CPUE and total accessible habitat produced average historical run size estimates of

22,226 and 14,723 steelhead, respectively (Table 4). The average ensemble estimate for annual

run size for all three approaches was 16,897 steelhead.

We summarized all the catch data on wild summer steelhead for the Quinault River

system for 1954-1955, 1957, and 1962-1977 (Table 5, Raw data is in excel file “Summer

steelhead catch data for Quillayute_Hoh_Queets_Quinault River populations” and based on

data from .pdf file “Historic winter and summer steelhead tribal catch by month Hoh, Quinault,

Queets, Quillayute 40s-70s”). Mean catch of summer steelhead in 1954-55, 1957, and from

1962-1977 was 162 fish with annual catch ranging from 12 to 447 fish, and peak catch being

reported in 1954 (Table 5).

Assuming a 25% harvest rate and using bounds of 15% and 35% harvest rates, we

estimated a mean run size of 649 summer steelhead (Lower bound = 464, Upper bound =

1,624) with a low run size of 48 fish and a high of 1,788 (Table 6). The mean of the top-five run

sizes is 1,130 summer steelhead (Lower bound = 807, Upper bound = 1,884). This suggests

summer steelhead were very abundant in the Quinault River dating back to 1954. Further, as

with other populations, catch records were very low in several years, suggesting some of our

estimates are likely much lower than the true total run size.

Table 5. Catch of wild sumer steelhead in Quillayute, Hoh, Queets, and Quinault Rivers prior to onset of hatchery summer steelhead

releases in Quillayute system in 1977, with 1946-1957 representing tribal gillnet catch and 1962-1977 representing almost solely

recreational catch. Tribal catch data was only reported and included here for 1972 and 1975-1978 in Quillayute River system, 1975-

1976 in Hoh River, 1974-1977 in Queets River, and 1974-1976 in Quinault River. Bolded values represent peak catch for the period

of record for each population. Data is in excel file “Summer steelhead catch data for Quillayute_Hoh_Queets_Quinault River

populations” and based on data from .pdf file “Historic winter and summer steelhead tribal catch by month Hoh, Quinault, queets,

quillayute 40s-70s”),

Population 1946 1954 1955 1957 1962 1963 1964 1965 1966 1967 1968 1969 1970 1971 1972 1973 1974 1975 1976 1977

Quillayute

River

15 N/A N/A N/A 12 36 39 30 37 78 105 81 84 116 309 66 175 143 136 51

Hoh River N/A 133 N/A 39 59 106 126 112 160 137 106 105 125 142 274 169 188 66 92 72

Queets

River

N/A 373 29 136 111 143 230 220 266 225 299 239 217 337 345 267 315 385 171 124

Quinault

River

N/A 447 15 49 165 157 208 119 206 151 238 279 221 180 241 289 236 140 357 12

Table 6. Mean estimated run size of wild summer steelhead based on recreational and tribal catch data in Table 7, where we

estimated run size with an assumed 25% harvest rate and we estimated upper and lower bounds using 15% and 35% harvest rates,

respectively, for the top five run sizes for each population. Data from 1946-1957 is from tribal gillnet catch records. Data from 1962-

1977 almost solely consists of recreational catch, with tribal catch only being reported for 1972 and 1975-1978 in Quillayute River

system, 1975-1976 in Hoh River, 1974-1977 in Queets River, and 1974-1976 in Quinault River. Bolded values represent peak catch

for the period of record for each population. Calculations for each year’s run size, if not clear to the reader in this petition, can be

accessed in the excel file “Summer steelhead catch data for Quillayute_Hoh_Queets_Quinault River populations”

Population 1946 1954 1955 1957 1962 1963 1964 1965 1966 1967 1968 1969 1970 1971 1972 1973 1974 1975 1976 1977

Quillayute

River

N/A

144

156

120

148

312

420

(300-

700)

324

(231-

540)

336

(240-

560)

464

(331-

773)

848

(606–

1,413)

264

308

104

296

204

Hoh River

N/A

532

N/A

156

236

424

504

448

640

(457-

1,067)

548

(391-

913)

424

420

500

568

(406-

947)

1,096

(783-

1,827)

676

(483-

1,127)

520

152

360

288

Queets

River

N/A

1,492

(1,066-

2,487)

116

544

444

572

920

880

1,064

(760-

1,773)

900

1,196

(854-

1,993)

956

868

1,348

(963-

2,247)

1,380

(986-

2,300)

788

676

212

380

496

Quinault

River

N/A

1,788

(1,277-

2,980)

196

660

628

832

476

824

604

952

(680-

1,587)

1,116

(797-

1,860)

884

(631-

1,473)

720

964

(689-

1,607)

484

596

296

212

b. Queets River

Unlike the other populations, there are two historic data periods for Queets winter

steelhead: one from a cannery record in 1923 and another from tribal catch records from 1948-

1960. In 1923, 1,500 cases of wild winter steelhead were packed at the Queets River cannery,

with each case containing 48 cans at 1 lb per can, or 48 lb per case (McMillan et al. 2022). This

is equivalent to a total weight of canned winter steelhead of 72,000 lb. Based on a wastage rate

of 0.40 and a mean weight of 9.8 lb per Queets River winter steelhead, the number of wild

winter steelhead processed in 1923 is equivalent to 12,245 fish. Assuming a cannery

exploitation rate equivalent to the median contemporary exploitation rate (0.38), the total

1923 Queets return of wild winter steelhead is estimated to be 32,223 fish, with upper and

lower bounds of 43,732 and 27,829 fish, based on 75% and 25% quartiles of contemporary

exploitation rates (0.44, 0.28), respectively (McMillan et al. 2022).

Using historical catch, McMillan et al. (2022) estimated the average historical

abundance for winter steelhead in the Queets River from 1948-1960 was 19,875 steelhead

(lower and upper bounds for the historical mean: 13,025, 32,878) (Table 4). Annual variability

within the historical period ranged from the lowest and highest single year estimates in the

Queets River of 6,191 and 52,200 winter steelhead in 1960 and 1954, respectively. Using

historical CPUE and total accessible habitat produced average historical run size estimates of

13,553 and 12,144 steelhead, respectively (Table 4). During the peak estimate year in the

Queets River, which was the largest estimated run size amongst all years and populations, over

14,000 winter steelhead were harvested, including over 5,200 winter steelhead harvested in

the month of December alone. The average ensemble estimate for annual run size for all three

approaches was 15,191 steelhead.

We summarized catch data on wild summer steelhead for the Queets River system for

1954-1955, 1957, and 1962-1977 (Table 5: Raw data is in excel file “Summer steelhead catch

data for Quillayute_Hoh_Queets_Quinault River populations” and based on data from .pdf file

“Historic winter and summer steelhead tribal catch by month Hoh, Quinault, Queets, Quillayute

40s-70s”). Mean catch of summer steelhead in 1954-55, 1957, and from 1962-1977 was 200

fish with annual catch ranging from 29 to 373 fish, with peak catch being reported in 1954

(Table 5).

Assuming a 25% harvest rate and using bounds of 15% and 35% harvest rates, we

estimated a mean run size of 802 summer steelhead (Lower bound = 573, Upper bound =

2,004) with a low run size of 116 fish and a high of 1,492 (Table 6). The mean of the top-five run

sizes is 1,296 summer steelhead (Lower bound = 926, Upper bound = 2,160). Based on the catch

data wild summer steelhead were formerly abundant in the Queets River.

3. WRIA 19 Populations

As mentioned previously, we could not find quantifiable information on historic

abundance for populations in WRIA 19. However, there is likely some historic catch information

from WDFW punch-cards, and anecdotal references suggest that “the largest steelhead trout

populations were found in the Lyre, Pysht, and Hoko rivers. The Clallam and Sekiu rivers, as well

as Deep Creek also supported significant steelhead populations” (NOPLE 2015). Additionally,

each stream was fished for winter steelhead by Native American tribes that lived in the area,

including the Lower Elwha Klallam Tribe and the Makah Tribe.

B. Current Abundance and Population Trends (circa 1980-2020)

Data on run size and escapement is not available for all populations in the Olympic

Peninsula DPS (Table 1). The Steelhead at Risk Report for Washington State by Cram et al.

(2018) identified 31 populations of wild steelhead (24 winter run populations and 7 summer

run populations) in the DPS, which included delineating larger sub-basins within each major

watershed (e.g., counted Sol Duc, Bogachiel, Calawah, and Dickey River (tributaries in

Quillayute River watershed) as individual populations). We did not delineate the Upper and

Lower Quinault River into sub-populations per Cram et al. (2018). We otherwise delineated

populations the same as Cram et al. (2018). We identified 22 distinct watersheds, with the

Quillayute River system consisting of four sub-basins and the Queets River containing the

Clearwater River, resulting in 26 populations of winter steelhead (Table 1). The Hoh, Calawah,

Queets, and Quinault Rivers are known to support wild populations of summer steelhead,

though data is very limited because they are not monitored. There is also evidence of summer

steelhead in the Sol Duc, Bogachiel, and Clearwater Rivers (Table 1). This brings the total

number of populations to 30 for the entire DPS.

Cram et al. (2018) evaluated current abundance and population trends for winter

steelhead where data was available. Cram et al. (2018) reported that sufficient escapement

abundance data were available for 48% (15 of 31) of populations (Figure 2). We use Cram et al.

(2018), a report by NOPLE (2015), and a publication by McMillan et al. (2022) to summarize

existing information on the run size, escapement, and trends in run size (or escapement if only

escapement data is available). We also rely on McMillan et al. (2022) to compare contemporary

abundance (circa 1980 – 2018) to historical abundance (circa 1948-1960: Table 4) to illustrate

that contemporary populations, most of which are in decline (Cram et al. 2018, McMillan et al.

2022), had already experienced a high level of depletion before the onset of the modern

monitoring programs that began in the late-1970’s through the early-1980s (Table 4).

Figure 2 (from Cram et al. 2018). Abundance data for the Olympic Peninsula Steelhead DPS

populations available from years 1980 to 2013. The thick blue line represents the total

escapement (for the population or a portion of it, based on data availability). The thin grey line

represents total run size (escapement plus reported sport and tribal harvest plus, for Hoh River

population only, estimated non-tribal and tribal bycatch mortality). The dashed grey line is the

escapement goal. The light grey number is the percent change in abundance (increase or

decrease) over the time period with data.

In terms of numbers of populations, the Olympic Peninsula DPS is predominantly

composed of small populations of wild steelhead in small coastal creeks and streams (Table 1).

Data sets for most of those populations are less consistent compared to the larger, major

populations, which contribute the greatest proportion of fish to the cumulative total number of

wild steelhead in the DPS. We provide a synthesis of the location of each watershed (by WRIA),

the type of steelhead life history that is present, the period of record of monitoring, mean and

range for run size and escapement, and the status of the population trend – as either declining

or increasing, and the rate of decline (based on Cram et al. 2018, NOPLE 2015, and McMillan et

al. 2022). As we outline below, the data indicates that most populations of wild winter

steelhead are in decline, and comparatively, Cram et al. (2018) estimated the Olympic

Peninsula DPS had the second lowest proportion of populations with increasing trends (20%)

among all the steelhead population segments in Washington State (Figure 3).

Figure 3 (From Cram et al. 2018). Washington steelhead population escapement abundance

trends in each DPS in time periods from 1980-2013. N represents in the number of populations

in each DPS that had suitable abundance data for trend analysis and the percentage indicates

the proportion of populations that have increased in abundance since 1980.

Run size declines in wild winter steelhead have generally continued since the report by

Cram et al. (2018), with most populations reaching their lowest level of abundance in the past

several years (Figures 4-8). To account for the most recent data, we updated the most recent

five-year estimates by McMillan et al. (2022) to 2016-2020 instead of 2013-2017. Using that

adjustment, we estimate the Quillayute, Queets, Hoh, and Quinault River wild winter steelhead

populations have declined by 61%, 69%, 79%, and 81%, respectively, when comparing the most

recent five-year average run size to the historic estimates from the 1950s (Table 4). The

estimated decline for the Queets winter steelhead population increases to 86% when

compared to the 1923 cannery estimate in McMillan et al. (2022). Given the older cannery

record from the Queets River, declines are likely even greater than estimated in McMillan et al.

(2022) because their estimates only dated back to the 1950’s, which is after the initial boom in

logging and the operation of commercial salmon canneries at the mouths of the Quillayute,

Hoh, Queets, and Quinault Rivers in the early-1900’s.

The long-term, chronic declining trends coupled with a recent sharp downturn not

documented in Cram et al. (2018) puts the Olympic Peninsula Steelhead DPS at greater risk of

extinction than when Busby et al. (1996) conducted the last federal status review for the DPS

(ESU at the time). Further, as we outline below, the populations of wild summer steelhead are

almost all at critically low levels of abundance, something that was also not reported by Busby

et al. (1996), and could be facing extirpation in the near term if some are not already

functionally extinct.

1. Winter Steelhead

The four largest winter steelhead populations on the Olympic Peninsula have been on a

forty-year-long decline.

Figure 4 (From McMillan et al. 2022). Contemporary trends (circa 1980 – 2017) in wild winter

Steelhead abundance in the a) Quillayute, b) Hoh, c) Queets, and d) Quinault Rivers. Black lines

represent the best fit linear regression models.

a. WRIA 20 Populations

i. Hoh River

The Hoh River has one of the largest populations of wild winter steelhead in the Olympic

Peninsula DPS (Table 1). From 1980-2020 the mean annual winter steelhead run size of Hoh

River winter steelhead was 4,117 fish with annual run sizes ranging from 2,541 – 5,783 fish

(Table 1). The population was in a significant declining trend from 1980-2013 (Figure 2) and

from 1980-2017 (Figure 4), and it has experienced most of its lowest returns on record in the

past decade (Figure 5). Cram et al. (2018) estimated a 16% decline from 1978-2013, while

McMillan et al. (2022) estimated a decline of 37% from 1980-2017 (Table 1), which is equivalent

to a loss of 513 (95% CI: 320, 710) adults per decade (linear regression on log-transformed

abundance, F

1, 36

= 29.01, P < 0.001: McMillan et al. 2022). The decline has not reversed since

the analysis by McMillan et al. (2022) (Figure 5).

The Hoh River population of winter steelhead has failed to meet its escapement goal in

5 of the past 11 years (2010-2020), and it just barely made its escapement goal in the past two

years (Figure 5).

Figure 5. Annual run size and escapement of wild winter steelhead in the Hoh River from 1980-

2021 in relation to the escapement goal of 2,400 fish. Years when escapement falls below the

escapement goal indicate that the fishery did not meet its spawner target goal.

Compared to the historic period (circa 1948-1960), mean annual abundance in the

contemporary period (1980-2017) has declined 69% from its mean annual abundance during

the historical period (circa 1948-1960) (Table 4). The decline jumps to 79% when comparing the

mean historical abundance to the most recent five-year average run size (Table 4).

We believe this should be considered a minimum level of depletion because the

approach assumes all caught fish were sold and reported in the catch. This is unlikely, as Wilcox

(1898) indicated for the Quinault River in 1895: “…quite a large number of salmon are taken by

Indians for their winter supply of food, and a small amount … was sold to buyers…” It could

therefore be anticipated that a considerable number of steelhead caught on the Queets River in

1923 were used by Native Americans for their own winter food supply rather than sold to the

cannery (McMillan 2006). In addition, the historical estimates in McMillan et al. (2022) were

based on fish populations that had already experienced decades of habitat changes and the

onset of canneries and commercial fisheries.

If this downward trajectory continues, run sizes will begin to consistently come in below

the escapement goals, eliminating the potential for fisheries and potentially leading to a level

where they can no longer effectively compensate (e.g., Ward et al. 2000; Atlas et al. 2015).

ii. Quillayute River System

The Quillayute River supports the largest population of wild winter steelhead in the

Olympic Peninsula DPS. Its mean annual run size from 1978-2020 was 13,064 steelhead (range

= 6,456 – 21,615) (Table 1). The Quillayute River system is largely comprised of three large

tributaries, including the Sol Duc, Calawah, and Bogachiel Rivers, and one smaller tributary, the

Dickey River (Table 1). We summarize all of these populations as part of the “Quillayute River”

system for purposes of simplicity and because each of these populations is showing relatively

similar patterns in annual abundance (Figure 2).

Unlike the Hoh, Queets, and Quinault Rivers, there was not a simple linear trend for wild

winter steelhead run size in the Quillayute River (Figure 4). This is because winter steelhead

returns increased from 8,761 fish in 1980 to a peak of 21,615 fish in 1996. Since the peak in

1996, annual returns of winter steelhead have declined at a rate of 5,533 fish per decade

(Figure 4). Run sizes for more recent years, which were not included in McMillan et al. (2022),

suggest the recent declining trend has continued (Figure 6a). As a result, winter steelhead run

sizes in recent years have been the lowest on record. Although the Quillayute River population

has fared slightly better than the others, the co-managers implemented restrictive regulations

in the 2020 and 2021 seasons due to poor returns, raising concerns about the long-term

sustainability of the population and the fisheries it supports.

Figure 6. Annual run size and escapement of wild winter steelhead in a) entire Quillayute River

watershed system, and annual escapement of wild winter steelhead in the four major individual

tributaries, which include the b) Dickey River, c) Sol Duc River, d) Calawah River, and e)

Bogachiel and Quillayute River proper. Black circles represent run size while open circles

represent escapement. The escapement goals for individual rivers within the Quillayute River

system are as follows: Calawah River (1,740 fish), Sol Duc River (2,910 fish), Bogachiel and

Quillayute River proper (1,127 fish), and Dickey River (123 fish), resulting in a system-wide goal

of 5,900 fish (Duda et al. 2018). Years when escapement falls below the escapement goal

indicate that the fishery did not meet its spawner target goal.

Mean annual abundance of wild winter steelhead in the Quillayute River from 1978-

2017 has declined by 38% compared to historic mean annual abundance from 1948-1960 (Table

4). The decline increases to 61% when compared to the mean annual abundance for the most

recent five-year period (Table 4). If the population continues to decline at its current rate

(1996-2020), run sizes in the coming decade to decades could – on average – come in lower

than the escapement goal, effectively eliminating the potential for future fisheries while

simultaneously increasing the risk of extinction.

iii. Other Populations

There are four other small populations of wild winter steelhead in WRIA 20, including

the Tsoo-Yess, Wa’atch, and Ozette Rivers, which are located north of the Quillayute River, and

Goodman Creek and Mosquito Creeks, which are located between the Quillayute and Hoh

Rivers.

Goodman Creek is the only stream for which we could find monitoring data on

escapement (Table 1). From 1995-2020 the mean annual escapement was 184 wild steelhead

(range of 45 – 374) and consequently, it has not met its escapement goal of 206 fish once in the

past decade ending in 2020 (Table 1). In fact, Goodman Creek has met its escapement goal only

once since 2004. Cram et al. (2018) estimated the average annual escapement has declined by

54% over its period of record, and the decline has continued through 2020. Data on run size

and escapement was otherwise not available.

b. WRIA 21 Populations

i. Quinault River

From 1978 to 2020 the mean annual winter steelhead run size was 5,883 fish (range =

2,179 – 9,726) in the Quinault River (Table 1). Unlike the Hoh and Quillayute Rivers, the

Quinault River is managed in two sections, including: (1) the lower part below the lake, which is

owned and managed by the Quinault Indian Nation, and (2) the upper section above the lake,

which is owned mostly by the Olympic National Park and its fish managed by WDFW (Figure 2,

from Cram et al. 2018). There is no escapement goal for the lower Quinault River, but there is

an escapement goal of 1,200 fish for the upper section (Cram et al. 2018).

There is a significant declining trend in Quinault River wild winter steelhead (Table 1).

Cram et al. (2018) estimated a decline of 69% for escapement of wild steelhead in the Lower

Quinault River and an increase of 24% in escapement for the Upper Quinault River (Figure 2).

McMillan et al. (2022) grouped the upper and lower river because run size cannot be estimated

for each component of the population separately, and they estimated a 44% decline (Table 1).

The McMillan et al. (2022) estimate equates to a loss of 1,052 (590 - 1,510) adults per decade

(Figure 4; F

1, 32

= 21.20, P < 0.001: McMillan et al. 2022). We have access to the full data set

from 1978-2020 (included in our data package) and the declining trend has continued, with the

smallest run sizes on record occurring in the most recent five-year period (Figure 7). The mean

escapement has been 3,107 fish with a range of 1,366 to 5,774 fish (Table 1). While it is unclear

whether the population would be meeting a biologically defensible escapement goal for the

whole watershed, the declining trend is concerning and more restrictive fishery practices,

including early closures, were implemented in 2020 and 2021 to limit impacts on the small run

sizes.

Figure 7. Annual run size and escapement of wild winter steelhead in the Quinault River from

1978-2021. There is no escapement goal for the entire Quinault population, as it is delineated

into Upper (above the lake) and Lower (below the lake) populations. Figure 2 illustrates the

trend by Cram et al. (2018) for the Upper and Lower Quinault populations.

Overall, McMillan et al. (2022) estimated that the contemporary mean annual

abundance of wild winter steelhead in the Quinault River from 1980-2013 declined by 63%

when compared to historic mean annual abundance from 1948-1960 (Table 4). The decline

increases to 80% when compared to the mean annual abundance for the most recent five-year

period (Table 4). If the decline rate of 1,052 adults per decade continues, Quinault River winter

steelhead will decline from their recent five-year average of 3,370 fish down to 2,318 fish

during the following decade. This concern is particularly heightened now because the dramatic

declines in 2019 and 2020 highlight how quickly the population could further deplete (Figure 7).

ii. Queets River

From 1980-2018 the mean annual winter steelhead run size was 7,648 fish (range =

4,200 – 13,309) in the Queets River (Table 1). Like the Hoh and Quinault Rivers, there is a

significant declining trend for winter steelhead in the Queets River from 1980-2017 (Figure 4),

which includes a 12% decline in the escapement trend for its major tributary, the Clearwater

River, from 1980-2013 (Figure 2). The extent of decline is greater in McMillan et al. (2022) (45%

decline) because Cram et al. (2018) (29% decline) only used data up until 2013, while McMillan

et al. (2022) included data up until 2018. The run size continued to decline after 2013 (Figure 4).

The decline reported by McMillan et al. (2022) equates to a loss of 1,220 (640, 1,820) adults per

decade (F

1, 36

= 19.56, P < 0.001).

Unlike the Hoh River population, the co-managers have not reached an agreement on

an escapement goal for the Queets River population, which has affected management for

several decades. The Queets Tribe uses a goal of 2,600 fish, while WDFW uses a goal of 4,200

fish (Cram et al. 2018). We show escapement in relation to the latter (Figure 8a) because it is

the only goal for which we can find scientific support (Gibbons et al. 1985). The Clearwater

River is a major tributary of the Queets watershed that has a separate escapement goal of

1,450 fish (Cram et al. 2018). Clearwater steelhead run size and escapement is included in the

overall data for the Queets River population, but Figure 8b depicts the escapement of wild

winter steelhead for just the Clearwater River. As a result of the sharp declines in run size in

recent years (Figure 8a), the Queets River fishery (as with the Quinault River) was closed to

protect small run sizes.

Figure 8. Annual run size and escapement of wild winter steelhead in the a), Queets River from

1980-2020 and b) escapement of wild winter steelhead in the Clearwater River, a major

tributary. Black circles represent run size, open circles represent escapement. In the Queets

River panel, the bold line depicts the 4,200 fish escapement goal (Escapement goal 1) that is

used by WDFW and is recognized by the National Park Service (a substantial portion of the

Queets River watershed lies within the Olympic National Park) and the 2,500 fish escapement

goal used by the tribe (Escapement goal 2). The bold line in the Clearwater River panel refers to

WDFW’s escapement goal of 1,450 fish. Years when escapement falls below the escapement

goal indicate that the fishery did not meet its spawner target goal.

The mean annual abundance of wild Queets River winter steelhead from 1980-2017 has

declined by 50% when compared to historic mean annual abundance from 1948-1960 (Table 4).

The decline increases to 69% when compared to the mean annual abundance for the most

recent five-year period (Table 4). However, as we mentioned previously, the declines could be

more substantial. For example, McMillan et al. (2022) analyzed one year of cannery data in

1923, when more than nominal cans of steelhead were reported. The run size estimate is

32,659 wild steelhead (Table 4), which would represent a decline of 86% compared to the most

recent five-year average run size. If the decline rate of 1,220 adults per decade continues, wild

winter steelhead abundance in the Queets River will decline from its recent five-year average of

4,658 fish down to an average of 3,438 fish during the following decade. Like the Hoh River, this

suggests that in just one decade run sizes would be coming in under the 4,200 fish escapement

goal set by WDFW and would be encroaching on the 2,400 fish goal set by the tribe.

Unfortunately, run sizes in recent years have declined rapidly, resulting in fishery

closures and suggesting that the population is on a chronic path that could lead to extinction

unless the 40-year trend is reversed.

iii. Other populations

There are four smaller populations of wild winter steelhead in WRIA 21, including

Kalaloch Creek, Moclips River, Raft River, and the Copalis River populations (Table 1). Among

those populations, we could find escapement data only for the Moclips River, which spans a

relatively short period that is not up to date. From 1988-2000, the mean escapement in the

Moclips River was 299 steelhead (range = 130 – 560), and based on that data set, Cram et al.

(2018) reported a 27% increasing trend in abundance (Table 1 and Figure 2). Given the limited

period of record and overall missing data, it is difficult to draw conclusions about the status of

these populations.

c. WRIA 19 Populations

There are ten small streams in WRIA 19 that support populations of wild winter

steelhead (Table 1). The Hoko River has the largest population with a mean annual escapement

of 566 fish (range = 193 – 990). The Hoko River has met an escapement goal of 400 fish in 80%

of the last ten years (Table 1). Based on the period of record, however, the Hoko River

population declined by 40% from 1985-2013 (Table 1) and that decline has not subsided in

recent years (WDFW Score website, Accessed 3/20/2022:

https://fortress.wa.gov/dfw/score/score/species/population_details.jsp?stockId=6357).

The Pysht River (and tributaries) and the Clallam River support the second and third

largest populations, respectively, and both populations are in long-term decline (Cram et al.

2018). Cram et al. (2018) estimate that winter steelhead abundance declined by 21% in the

Pysht River from 1995-2013 and 27% in the Clallam River from 1999-2013 (Table 1). More

recent data indicates these declines have continued through 2020 in the Pysht River (WDFW

Score website, Accessed 3/20/2022:

https://fortress.wa.gov/dfw/score/score/species/population_details.jsp?stockId=6343) and the

Clallam River populations (WDFW Score website, Accessed 3/20/2022:

https://fortress.wa.gov/dfw/score/score/species/population_details.jsp?stockId=6350).

Consequently, the Pysht River population has only made its escapement goal in 70% of the past

ten years (Table 1).

Salt Creek supports the fourth largest population in WRIA 19. The population has only

met its escapement goal once in the past decade (Table 1) and once overall since 2004. Cram et

al. (2018) estimated wild steelhead abundance declined by 43% from 1995-2013, and more

recent data suggests the decline has not reversed but may have stabilized (WDFW Score

website, Accessed 3/20/2022:

https://fortress.wa.gov/dfw/score/score/species/population_details.jsp?stockId=6329).

Escapement estimates are also available for populations in Deep Creek and the East and

West Twin Rivers from circa 1995 through 2020 (WDFW Score website, Accessed 3/20/2022:

https://fortress.wa.gov/dfw/score/score/species/population_details.jsp?stockId=6343). Using

this data set, a regional recovery document suggests that all three populations of wild

steelhead were in decline and there is a high level of concern about the depressed status of the

stocks (NOPLE 2015). Of particular concern, wild winter steelhead have only met the

escapement goal of 86 steelhead once in the past decade (Table 1) and in two of the past

twenty years. Even worse, the population in West Twin River has missed its escapement goal of

103 fish every year in the past decade (Table 1) and every year dating back to 2001. We could

not find any estimates of escapement for the Sekiu, Sail, and Lyre Rivers (Table 1).

2. Summer Steelhead

The co-managers do not monitor any populations of wild summer steelhead on the

Olympic Peninsula, nor do they have any established escapement goals for these populations.

The only contemporary data on abundance and distribution of adult summer steelhead comes

from snorkel surveys conducted by Olympic National Park (e.g., Brenkman et al. 2012), which

manages summer steelhead habitat and fisheries in the park, and McMillan (2022) (Table 7).

Those snorkel surveys were conducted in the Sol Duc River, Calawah River, SF Calawah River, NF

Calawah River, Sitkum River, Bogachiel River, SF Hoh River, and the EF and NF Quinault Rivers.

Snorkel surveys were generally conducted annually from September through October and,

occasionally early November, to ensure summer steelhead had time to migrate through the

watershed and reach the upper most habitats where they stage for several months before

spawning from winter through spring.

Table 7. Counts of summer steelhead via snorkel surveys in the Calawah, Sitkum, SF Calawah,

NF Calawah, SF Hoh, Sol Duc, EF Quinault, and NF Quinault Rivers, including year(s) of surveys,

mean abundance, and mean proportion of hatchery steelhead (Figures 9 – 11). Data from

McMillan (2022) is based on snorkel surveys conducted in early September through early

November, while data from Sam Brenkman et al. is based on snorkel surveys conducted in late

summer (Brenkman et al. 2012). Standard Deviation and range is provided for McMillan (2022),

while only range is provided for Brenkman et al. (2012).

Study and population Year (s)

Mean abundance

(SD and range)

Mean proportion

hatchery steelhead

(SD range)

McMillan 2022

Calawah River system

2002

89 (wild),

214 (hatchery)

N/A

Sitkum River

2002 – 2006,

2009 – 2021

53 (25.2, 19-105) 3% (3%, 0% – 10%)

SF Calawah River 2003 – 2006

18 (5.6, 10 – 23) 7% (6%, 0% – 13%)

NF Calawah River

2000 – 2006,

2009 – 2021

4 (4.1, 0 – 14) 33% (35%, 0% – 100%)

SF Hoh River

2000 – 2006,

2009 – 2014,

2016 – 2019

7 (3.9, 1 – 16) 41% (12%, 20% – 67%)

Brenkman et al. 2012

Bogachiel River 2005 – 2010 15 (8 – 26) 13% (0% – 23%)

Sol Duc River 2005 – 2010 6 (0 – 15) N/A

SF Hoh River 2005 – 2010 8 (4 – 12) 40% (14% – 76%)

EF Quinault River 2005 – 2010 17 (4 – 35) 16% (0% – 38%)

NF Quinault River

2005 – 2010

3 (0 – 8)

43% (0% – 100%)

Calawah River survey data from 2002 is a census of the entire watershed accessible to adult summer steelhead in

September and October.

Below we summarize the results of the surveys and when possible, compare them to

our historic estimates. The findings generally suggest that wild summer steelhead now only

exist at exceptionally low abundance (Brenkman et al. 2012; McMillan 2022) and are headed

towards extirpation if their status and trends are not rapidly improved.

a. WRIA 20 Populations

i. Hoh River

Snorkel surveys on the Hoh River have been conducted by the National Park Service

(Brenkman et al. 2012) and McMillan (2022). These surveys focused on several long sections

where summer steelhead were observed staging during previous years. The results are not a

census, but rather, an estimate of peak abundance in those important habitats.

The snorkel surveys in the SF Hoh River by Brenkman et al. (2012) sometimes covered a

longer section of stream than McMillan (2022), though not always, and in most years their

surveys overlapped spatially. Unfortunately, both sets of surveys indicate a very low abundance

of wild summer steelhead. Brenkman et al. (2012) data cover the period from 2005 – 2010

while surveys by McMillan (2022) ranged from 1999-2006, 2009-2014, and 2016-2019.

Brenkman et al (2012) reported a mean abundance of eight wild summer steelhead with peak

annual peak counts of 12 and 16 fish (Table 7). The longer time series by McMillan (2022)

suggests summer steelhead abundance has remained low, with a mean abundance of seven

wild summer steelhead (Figure 9). The lowest abundance estimates have almost all occurred

from 2016 to 2020 (McMillan 2022).

Figure 9 (From McMillan 2022). Box-whisker plot displaying mean (+) annual abundance of wild

and hatchery summer steelhead counted during snorkel surveys from late-summer through

early-fall in the SF Hoh River (1999-2006, 2009-2014, 2016-2019). Internal rectangle indicates

95% confidence interval for the mean. Tukey whiskers extend to data points that are less than

1.5 interquartile range away from 1

quartile.

Both surveyors also reported high levels of hatchery summer steelhead, including more

hatchery than wild adults in some years (Table 7). For example, the mean abundance of four

hatchery steelhead was not significantly lower than the mean estimate of seven wild steelhead

(Figure 9). Hatchery summer steelhead are not released into the Hoh River, but hatchery

summer steelhead could have strayed from their release locations in the Bogachiel or

Wynoochee Rivers, eventually ending up in the headwaters of the SF Hoh River.

Unfortunately, there is almost no information on summer run abundance in the

mainstem Hoh River, which likely supports most of the population. Nonetheless, there is also

little evidence to suggest that mainstem component is faring any better, even if more

abundant. Overall, the data indicate that SF Hoh River wild summer steelhead exist at low levels

of abundance. Considering our rough historic estimates of population size based on

recreational catch – a mean run size of 472 steelhead with a high of 1,096 fish (Table 6) – it is

likely that the SF Hoh River sub-component is critically depleted, and that the entire population

could be threatened with extinction.

ii. Quillayute System

Data on abundance of wild summer steelhead is limited to snorkel surveys conducted in

the Sol Duc, Bogachiel, Calawah, NF Calawah, SF Calawah, and Sitkum Rivers (Table 7). Surveys

by Brenkman et al. (2012) found a mean abundance of 15 wild summer steelhead in the upper

Bogachiel River from 2005 – 2010 and only six wild summer steelhead in the upper Sol Duc

River (Table 4). In some years no wild summer steelhead were observed in the Sol Duc River

surveys conducted by Brenkman et al. (2012).

McMillan (2022) conducted snorkel surveys in the NF Calawah River, Sitkum River, and

SF Calawah River (only 2003 – 2006) from circa 2000 – 2006 and from 2009 – 2021. These

locatoins are thought to be the predominant staging and spawning areas for summer steelhead

(McMillan 2006). Data is also available from a snorkel census in 2002 that covered 85% of all

habitats available to adult steelhead during the summer (e.g., McMillan et al. 2013). The diver

counted adult steelhead during the survey but did not publish the data in McMillan et al. 2013.

During the 2002 census, a total of 89 wild summer steelhead and 214 hatchery summer

runs were counted (Table 7). Snorkel surveys in the Sitkum River and NF Calawah River

produced a mean annual abundance of 53 and five wild summer steelhead, respectively (Figure

10a and 10b). Peak abundance in the Sitkum River snorkel counts was 105 wild summer runs in

2005, but its lowest estimates have dropped to 19 fish (Table 7) and counts were less than 30

adults in 2020 and 2021 (McMillan 2022). Surveys of sections of the upper SF Calawah from

2003-2006 suggest it also has summer steelhead staging and spawning habitat, though overall

abundance appears lower than in the Sitkum (McMillan 2022). Counts of wild summer

steelhead in the NF Calawah River have generally ranged between 0 – 10 fish (Figure 10b), with

the two peaks occurring in 2001 and 2002 (McMillan 2022). A drought in 2002 dewatered all of

the lower NF Calawah except for the lower-most 1km of stream, which killed thousands of

salmonids including all the adult steelhead, and since then, abundance of summer runs has

been at very low levels (McMillan 2022). In fact, zero adult steelhead were counted in the NF

Calawah from 2014-2021 except for 2015, when a single adult was observed (McMillan 2022).

The snorkel survey data has not been adjusted to account for potential sources of error in the

diver counts, and there is not a whole basin estimate for summer steelhead abundance. With

those caveats, the current counts still suggest the Quillayute River system population is severely

depleted in comparison to the historical catch estimates.

Figures 10 a-b (From McMillan 2022). Box-whisker plot displaying mean (+) annual abundance

of wild and hatchery summer steelhead counted during snorkel surveys from late-summer

through early-fall in the, a) Sitkum River (2003-2006, 2009-2019), and b) NF Calawah River

(2000-2006, 2009-2021). Internal rectangle indicates 95% confidence interval for the mean.

Tukey whiskers extend to data points that are less than 1.5 interquartile range away from

quartile.

b. WRIA 21 Populations

i. Quinault River

The only data we could find on the abundance of summer steelhead in the Quinault

River is from Brenkman et al. (2012). They conducted snorkel surveys in the EF and NF Quinault

Rivers from 2005 – 2010. The mean annual abundance of wild summer steelhead during those

surveys was 17 and three wild steelhead in the EF and NF Quinault Rivers, respectively (Table

7). The surveys suggest the Quinault River population of wild summer steelhead has

dramatically declined from the historic estimates we calculated based on recreational catch

from the 1950s through 1976, which show a mean run size of 649 summer steelhead with a

peak run size of 1,788 fish (Table 6). Although the current run size is unknown, it appears the

population of summer steelhead in the Quinault River is less than 1% of its historic abundance

and is close to extirpation.

ii. Queets River

The are no snorkel survey counts to determine the population trend of summer

steelhead in the Queets or Clearwater Rivers. McMillan (2006) estimated that, based on

available catch data, the run size of wild summer steelhead returning to the Queets and

Clearwater Rivers combined was no more than 100 fish. If the population is close to 100 fish,

that would suggest the population is now only a fraction of the historic estimate we calculated

based on catch, where the mean population size was 802 summer steelhead and the peak was

1,492 fish (Table 6). This uncertainty creates a great deal of concern for the status and

trajectory of the population moving forward.

XI. PRODUCTIVITY

In comparison to estimates of abundance, productivity data is very limited for the

Olympic Peninsula Steelhead DPS. Cram et al. (2018) reported on nine productivity data sets in

the Olympic Peninsula DPS from circa 1980-2013, eight of which focused on spawner-per-

spawner production (Figure 11) and one that focused on smolt-to-adult return survival (Figure

11). Hall et al. (2016) reported on the productivity of adults to juveniles in the Twin Rivers. We

reviewed the co-managers’ annual data sets and could only find productivity information for

three major populations, including 1) recruits-per-spawner data for the Quillayute River (1978-

2015), 2) Hoh River (2000-2016) populations of wild winter steelhead, and 3) a raw estimate of

number of adult returns per wild/natural winter steelhead smolt in the Queets River (1982-

2013). Cram et al. (2018) did not list which populations were included in their productivity

estimates, but it would seem likely that our data sets overlap.

Figure 11 (From Cram et al. 2018). For each DPS, population growth rates represented as the

log of the numbers of natural-origin adult spawners produced per spawner for all populations

with suitable data for brood years 1970 to 2010. The thick black line represents the average

value for each DPS, while the thin lines represent data from individual populations within

each DPS. The y-axis numbers in parentheses for the Puget Sound figure are the non-

transformed values for reference. A log productivity of 0, shown by the red lines, corresponds

to an untransformed productivity of 1 spawner per spawner (replacement). Data are available

for multiple populations in each DPS that occur in Washington, including: 7 populations in the

Puget Sound DPS; 8 in the Olympic Peninsula DPS; 5 in the Southwest Washington DPS, 8 in

the Lower Columbia River (LCR) DPS, 1 in the Mid-Colubmia River (MCR) DPS, 4 in the Upper

Columbia River (UCR) DPS, and 3 in the Snake River Basin (SRB) DPS.

There appears to be a slight decline in spawner-to-spawner recruitment from 1978-2010

for wild winter steelhead in the Olympic Peninsula DPS, and after the mid-1990’s, the

populations increasingly fail to replace themselves (Figure 11, Cram et al. 2018). Smolt-to-adult

return declined significantly over time for the single wild winter steelhead population evaluated

by Cram et al. (2018) (Figure 12).

Figure 12 (from Cram et al. 2018). For each DPS, smolt survival, estimated as smolt-to-adult

return (SAR) rates of natural-origin steelhead spawners from smolt outmigration year 1975 to

2011 (16 populations represented: 2 populations and 1 sub-population in Puget Sound; 4 sub-

populations and 1 population in Olympic Peninsula; 1 population each in Southwest

Washington, MCR, and UCR DPSs; 3 populations in the LCR DPS; and 2 populations in the SRB

DPS). The thick black line represents the average SAR value for each DPS, and the thin grey lines

represent data from individual populations therein.

Cram et al. (2018) indicated recruits-per-spawner productivity was not available for any

populations in the DPS. However, the co-manager data sets we analyzed referred to “recruits

per spawner” estimates for the Quillayute and Hoh Rivers. Perhaps this is just a minor

difference in terminology. We include data on recruits per spawner and extend the productivity

period of record for the Quillayute River population to 1978-2015 and the Hoh River population

to 1990-2016, both of which suggest a continued declining trend (Figure 13a).

Figure 13. Estimates of wild steelhead productivity, including a.) recruits per spawner for wild

winter steelhead in Quillayute (1978-2015 brood years) and Hoh Rivers (1990-2016 brood

years), and b.) smolt to adult return data for wild winter steelhead in Queets River from 1982-

2013. Dashed line in Panel a.) represents replacement of 1 recruit/spawner, while dashed in

Panel b.) represents mean smolt to adult return rate for Queets wild steelhead during period of

record, which is 0.15.

Of particular concern, productivity has been below replacement (recruits per spawner =

1.0) in four of the past ten years in both the Quillayute and the Hoh River populations, and the

Hoh has been below 1.0 in ten (50%) of the past twenty years (Figure 13a). There is no clear

trend in the Queets River population smolt-to-adult return data (Figure 13b), largely because of

two peaks in survival that seem abnormally high. Perhaps the co-managers adjust their

estimates or there were errors in the data sets that we had access to.

Cram et al. (2018) also reported that variation in freshwater per-capita productivity

among populations in the Olympic Peninsula DPS declined as abundance increased, as did Hall

et al. (2016) for the Twin Rivers. Indeed, the estimates of recruits per spawner for the

Quillayute (Figure 14a) and Hoh River populations (Figure 14b) reveal a similar pattern. As Cram

et al. (2018) described, the relationships suggest there are potential density dependent effects

on productivity (Walters et al. 2013; ISAB 2015). This could suggest habitat is limiting (Cram et

al. 2018). However, density-dependence effects on growth can be highest at low densities

(Lobon-Cervia 2007), and there are other factors to consider before we can conclusively

determine that productivity is being solely limited by habitat potential and that the populations

have the compensatory capacity to rebuild from their current low levels of abundance.

Figure 14a. Estimates of recruits per spawner for wild steelhead populations in the Quillayute

River.

Figure 14b. Estimates of recruits per spawner for wild steelhead populations in the Hoh River

(1990-2016).

The first factor to consider is the lack of data pointing to specific causes of productivity

declines. While there is suggestive evidence of density dependence in the Quillayute and Hoh

River populations based on the adult-to-adult recruits per spawner estimates, the factors

limiting these populations are unknown due to a lack of data on juveniles and smolts. There has

not been a systematic, long-term decline in marine survival on the Washington Coast (Kendall

et al. 2017; Welch et al. 2018), so it is possible that fishery, hatchery, and habitat effects have

combined to reduce the productivity of the populations. For example, the historic run timing of

wild winter steelhead has changed dramatically in the Quillayute, Hoh, and Queets River

populations with early-entering wild fish being replaced by hatchery fish (Figure 15; McMillan

et al. 2022), resulting in compressed migration timing that could increase density-dependent

effects by reducing spatial and temporal distribution and diversity of adults and juveniles.

Several studies on Atlantic salmon (Einum et al. 2006, 2008; Teichert et al. 2011; Finstad

et al. 2013) indicate that contraction in spatial distribution can lead to heightened density

dependence even without changes in habitat conditions. In this vein, declines in abundance

could have also resulted in reduced spatial distribution of spawners (Isaak and Thurow 2006),

which in turn could contribute to depensation (Atlas et al. 2015). There is also evidence that

diversity in run timing can help stagger the use of habitat by newly emerged offspring,

essentially increasing the productivity of the existing habitat (Gharett et al. 2013).

Additionally, the frequency of repeat spawners – which can be the most productive

individuals (Christie et al. 2018) – has apparently declined in the Queets and the Hoh River

populations, and they currently occur at low levels in the Quillayute River and are almost non-

existent in the Quinault River (Figure 16). Because life history diversity is critical to the

productivity and resilience of steelhead (Moore et al. 2014), it is plausible that productivity will

continue to decline because the original fabric of diversity has been greatly reduced. Although

this is postulation, the dearth of information outside the adult-to-adult data does not provide

the65 necessary insight to determine how and when population productivity is being regulated.

Figure 15 (from McMillan et al. 2022). Historical (circa 1955–1963) and contemporary (2000 -

2017) migration timing estimates based on CPUE of wild and hatchery winter steelhead (left

panels) in the (A–B) Quillayute, (C–D) Hoh, and (E–F) Queets rivers, and comparison of

cumulative run timing with estimates of dates at which 25% (q25), 50% (q50), and 75% (q75) of

the run had passed for historical and contemporary wild winter steelhead (right panels). Dark

gray lines and triangles represent contemporary hatchery returns; black lines and black circles

represent historical wild returns; and light gray lines and light gray circles represent

contemporary wild returns. Run timing estimation in the (G) Quinault River was limited to the

historical period because contemporary CPUE data were not available.

The second factor to consider is that productivity estimates for the major populations of

wild winter steelhead in Cram et al. (2018) ended in 2010 and the co-manager data sets we

accessed (ending in 2015) do not account for the recent declines in run size in the Quillayute,

Hoh, and Queets River populations. We also could not find estimates of productivity for the

Quinault River population (though we assume the co-managers do have estimates of

productivity). It will take several years for the productivity estimates to be updated so that we

can better understand the compensatory dynamics that have occurred over the past decade.

Petitioners are concerned that the populations are not compensating as effectively now

as they did 20-40 years ago for several reasons. For example, there are increasingly low levels

of productivity in the Quillayute River population (Figure 14a). Additionally, the peaks and

troughs in annual run size for the Hoh, Queets, and Quinault River populations have generally

declined each decade over the period of record (i.e., lower peaks, deeper troughs). These

factors have led to high levels of depletion in recent years (Figures 5-8),

The third factor to consider is that there is no data on the productivity of summer

steelhead stocks, nearly all of which are completely unstudied. This is a very large data gap.

Considering the declines in wild winter steelhead abundance and productivity, there is a high

level of risk that summer steelhead are facing an even more dire situation.

The fourth factor to consider is that there is tremendous uncertainty about productivity

for numerous smaller populations of wild winter steelhead that inhabit independent tributaries.

While these populations are indeed smaller, they collectively represent an important

component of diversity that spatially covers a much larger extent of the DPS than the four

largest, major populations. An HSRG (2004) report claims productivity is limiting for 15 ½ out of

16 wild steelhead populations (the Quinault River winter steelhead received two ratings), and

they associated the reduced productivity with degraded and “inadequate” habitat conditions.

And a report by NOPLE (2015) indicates the number of smolts produced in the following WRIA

19 populations had declined: Salt Creek (1995-2006), East Twin River (2001-2005); West Twin

River (2001-2005); Deep Creek (1998-2005). NOPLE (2015) did not provide smolt production

numbers for the Pysht, Clallam, or Hoko River populations.

In sum, the general assumption, based on fish data and habitat evaluations, is that wild

steelhead productivity is in long-term decline for most populations in the Olympic Peninsula

DPS. That trend, combined with depleted abundance and diversity, puts them at greater risk of

extinction in the coming decades.

XII. DIVERSITY

A robust population should maintain both genotypic and phenotypic diversity and have

distributions that are spatially and temporally diverse. The array of life history diversity of a

population is analogous to a “Portfolio” in the financial realm, where having different life

histories is like having a more diverse financial portfolio (Schindler et al. 2010). The diversity

helps buffer individuals and populations from environmental effects, the cumulative effects of

which improve the performance and resilience of populations with stronger and more diverse

portfolios (Schindler et al. 2010). Moore et al. (2014) applied this to steelhead populations in

British Columbia and found that life history diversity helped dampen fluctuations in population

abundance and increased their resilience to environmental factors.

Overall, there is very little data on the genetic diversity of wild steelhead on the Olympic

Peninsula or on the genetic relationships between the several apparently demographically

independent populations. However, there is evidence that life history diversity has been

reduced in several ways that could compromise their ability to remain productive and resilient

in the face of a changing climate. Important elements of steelhead life-history include: the

frequency and age-specific proportions of repeat spawners; proportions of smolt ages of each

sex; proportions of spawner ages of each sex; size-specific fecundity; and timing of migrations

to spawning habitat.

Repeat spawning is particularly important to the resilience of steelhead populations

(Hard et al. 2015; Gayeski et al. 2016). For example, available historical (Withler 1966) and

contemporary data from Kamchatka (Pavlov et al. 2001) indicate that the proportion of repeat

spawners in the annual returns of the area’s steelhead population historically exceeded 20-30%

of total spawners. This is also true for the Queets River, which had relatively high levels of

repeat spawners in the 1980s (Figure 16). Such relatively high levels of repeat spawning

suggests that this is a life-history characteristic to compensate for high and/or variable juvenile

marine survival rates. When this is the case, the proportion of repeat spawners can determine

whether a population can sustain levels of recruits per spawner above 1.0 (see Gayeski et al.

2016, pp. 92-93, and Table 1).

Figure 16. Proportion of repeat spawning individuals (repeats/total run size) by year for wild

winter steelhead populations in the Quillayute (1978-2021), Hoh (1993-2020), Queets (1980-

2019), and Quinault Rivers (1991-2020).

A. Summer Steelhead

The summer run life history represents an important component of the overall diversity

of the Olympic Peninsula Steelhead DPS. There is a heritable basis to the summer run life

history in steelhead (Prince et al. 2017; Fraike et al. 2021) that allows them to enter streams in

a highly immature state and stage for several months to up to a year in freshwater before

spawning. Entering in summer when temperatures are warmer also helps allow them ascend

waterfalls that winter runs cannot, providing them with access to habitats to headwater

habitats that would otherwise go unused. Thus, losing the summer life history component could

reduce, and likely already has, the spatial distribution of Olympic Peninsula steelhead. The

extinction of summer steelhead would also create large gaps in migration timing, hindering

each population’s ability to adapt to and keep pace with changing selective pressures

associated with climate change.

As noted by McMillan (2006), the situation for summer steelhead on the Olympic

Peninsula is dire:

“It is apparent that the Olympic Peninsula summer steelhead populations

examined are at the edge of extirpation. The Quinault population may be the

most dire, with estimated returns of less than 50 fish for the entire watershed

whose spawning destinations are further reduced in their split between the

North and East forks – potentially less than 25 fish destined for each. The

Clearwater population of the Queets system, and the Sol Duc and Bogachiel

populations of the Quileute system may be similarly low, with only 2-3 dozen fish

returning to each. In fact, the Quinault, Clearwater, Sol Duc, and Bogachiel

populations may already be functionally extinct.”

Snorkel data from Brenkman et al. (2012) and McMillan (2022) suggests the situation is

even worse now compared to McMillan (2006).

Considering the allelic variants for early migration have not arisen independently via

new mutations (Prince et al. 2017) and the life history is at critically low levels of abundance, it

is likely the genetic integrity of the populations is impaired. Consequently, there is great

concern wild summer steelhead populations are close to extinction and their demise could lead

to irreparable loss of the genetic basis for premature migration in the Quillayute, Hoh, Queets,

and Quinault River populations.

B. Early Returning Winter Steelhead

There is strong evidence that the migration timing of wild winter steelhead has changed

significantly from its historic norm (McMillan et al. 2022). Migration and run timing are strongly

heritable in salmonids such as steelhead (Carlson and Seamons 2008; Abadıa-Cardoso et al.

2013) and differences in migration and spawn timing represent adaptive responses that

maximize their reproductive fitness to local environmental conditions (Brannon et al. 2004).

However, there is concern that indirect and direct hatchery effects, combined with habitat

degradation, have selected against the early entering component (November – January) of the

wild winter steelhead (Bahls 2001; McMillan 2006). Fisheries can induce directional selection in

the timing of migration in salmon (Quinn et al. 2007), and in turn, shorten breeding seasons,

reduce phenotypic diversity, and lower population productivity (Tillotson and Quinn 2018).

Recently, McMillan et al. (2022) published a study that evaluated potential changes in

run timing of wild winter steelhead in the Quillayute, Hoh, and Queets Rivers. The author’s

comparison of tribal CPUE data for wild winter steelhead between circa 1948-1960 (historical

period) and 1980-2017 (contemporary period) revealed large changes in run timing (Figure 15,

McMillan et al. 2022). Overall, wild winter steelhead now enter later, and as a result, their

overall period of entry is compressed compared to the historical period (Figure 15, McMillan et

al. 2022).

Reconstruction of run timing provided evidence that wild winter steelhead runs began

significantly earlier in all three populations during the historical period (Figure 15). For example,

the percentage of wild winter steelhead migrating before January 1

(pJan1) during the

contemporary period was between 18% (14%-23%) and 43% (35%-50%) less than the historical

period in the Quillayute and Hoh Rivers, respectively (McMillan et al. 2022). The reduction of

earlier returning wild winter steelhead also corresponds with a significant shift in the date at

which half of the population had passed (q50) in the Quillayute and Hoh Rivers (Figure 15b and

16d). Estimated q50 occurred 25 (16-33 days) and 61 (47-71 days) days later during the

contemporary period in the Quillayute and Hoh Rivers, respectively. The magnitude of the shift

in q50 in the Queets River was similar to other populations, occurring 54 (-23-70 days) days

later during the contemporary period. Importantly, the timing of contemporary hatchery

steelhead returns directly overlaps with the historical early returning wild fish, which are now

greatly depleted (Figure 15).

The later entry timing during the contemporary period is causing run compression

(Figure 15). As a result, the number of days elapsed between when 25% and 75% of the run had

passed historically is now 16 (6-29 days), 26 (11-53 days), and 22 days (-4-83 days) shorter in

the Quillayute, Hoh, and Queets Rivers, respectively (McMillan et al. 2022).

McMillan et al. (2022) could not explicitly determine cause and effect, but they did offer

two hypotheses. First, spawning and rearing habitat used by early entering individuals may

have been sufficiently degraded to reduce the productivity of those fish (discussed in McMillan

et al. 2022). Second, declines in earlier migrating wild winter steelhead are also associated with

the introduction of hatchery winter steelhead (Cederholm and TU 1984; Bahls 2001).

Most hatchery programs on the western side of the Olympic Peninsula selectively bred

adults to return from November through early January, under the assumption that wild

steelhead were not abundant during that period (Crawford 1979; Cram et al. 2018). The

assumption was erroneous, however, because the return of the hatchery steelhead directly

overlaps with wild run timing in the historical period (Figure 15). This would have created

opportunities for interbreeding for wild steelhead that also spawned early. For example,

Seamons et al. (2012) found that a segregated winter steelhead hatchery program using the

early returning Chambers Creek stock failed to prevent interbreeding with wild winter

steelhead in Forks Creek, Washington. After only three generations of hatchery stocking, the

proportion of wild ancestry smolts and adults declined by 10-20% and up to 80% of naturally

produced winter steelhead were hatchery x wild hybrids (Seamons et al. 2012).

Further, the production and timing of hatchery winter steelhead set the stage for mixed-

stock in-river fisheries in which recreational and commercial fisheries targeting hatchery winter

steelhead also subjected co-mingling earlier-returning wild winter steelhead to high and

potentially unsustainable exploitation rates (Cederholm and TU 1984; Naish et al. 2007; Cram et

al. 2018). Contemporary fishing effort is highest from December through mid-January,

corresponding with the timing of hatchery winter steelhead returns and thereby providing the

potential for fisheries-induced directional selection against earlier migrating wild winter

steelhead (Quinn et al. 2007; Tillotson and Quinn 2018). Given the heritable component

underlying timing of migration, its plausible that reduced survival due to interbreeding with

hatchery fish combined with a variable fishing intensity that peaks from late-November through

early-January are the two most important factors responsible for the declines in early returning

wild steelhead. Regardless of the exact causes, unless action is taken to restore and protect

diversity, the fate of wild Olympic Peninsula steelhead rests on late-returning winter steelhead

that may not be able to keep pace with changes in spring and summer streamflow and

temperatures.

XIII. SPATIAL STRUCTURE

Ensuring that populations are well represented across diverse habitats helps to maintain

and enhance genetic and life history variability and population resilience (McElhany et al. 2000).

Additionally, ensuring wide geographic distribution across diverse climate and geographic

regions helps to minimize risk from catastrophes (e.g., droughts, floods, etc.; McElhany et al.

2000).

The Steelhead at Risk Report authors evaluated spatial structure of steelhead in the

Olympic Peninsula DPS using only previously documented or historical classifications (Cram et

al. 2018). The region is fortunate in that no habitat has been lost to large dams or barriers

(Cram et al. 2018). However, much of the landscape is covered with a dendritic network of

roads that were constructed for forest practices, many of which cross salmon-bearing streams

and have culverts (or lack culverts) that prevent upstream access (Smith 2000). We could not

determine how much habitat has been blocked by culverts and roads, but it is likely extensive

(Smith 2000). Considering that juvenile steelhead are distributed up to barriers in many small

creeks (McMillan and Starr 2008; McMillan et al. 2013), it is likely that numerous roads and

impassable culverts have truncated their spatial distribution.

It is also possible that distribution has been altered due to changes in run timing

(McMillan et al. 2022) due to spatial structure of spawning adults (McMillan et al. 2007). Early

returning steelhead often spawn in smaller streams or stream sections within a watershed

(Cederholm 1983; McMillan et al. 2007), perhaps because they are only accessible during

higher flows of early winter and such places may offer more secure spawning habitat during the

high flows of winter. Reduced abundance of those fish could alter the overall distribution of

steelhead. Similarly, the depletion of summer steelhead has potentially changed how and when

different life histories use parts of the watershed. While any changes in spatial distribution due

to the aforementioned effects are speculative, the lack of data and the importance of spatial

structure underscore the uncertainty associated with steelhead in this DPS.

Last, spatial structure of a DPS is also related to the patterns and frequency of natural

demographic exchange between apparently demographically independent populations. This

affects the extent to which the DPS and its component populations represent or exhibit

characteristics of metapopulations and/or exhibit source-sink dynamics and over what

temporal periods. This requires, among other things, knowledge of the genetic relationships

between populations, which appears to be largely lacking on Olympic Peninsula steelhead.

PART TWO

ENDANGERED SPECIES ACT LISTING FACTORS

I. PRESENT OR THREATENED DESTRUCTION, MODIFICATION, OR CURTAILMENT

OF ITS HABITAT OR RANGE

Olympic Peninsula steelhead habitat is threatened both inside and outside of Olympic

National Park (Klinger et al. 2008). Outside the park, steelhead habitat is degraded from

decades of destructive logging operations (McHenry et al. 1996). Timber harvesting has altered

hydrographs, increased water temperatures, and accelerated erosion. Extensive road building is

causing erosion and delivering fine sediments to streams. A widespread distribution of culverts

block access to spawning and rearing habitat and impede the downstream delivery of wood and

sediment (Smith 2000). These impacts continue to limit steelhead productivity despite

improved forest practices that were adopted in the 1990s.

“*** [T]he majority of streams on the Olympic Peninsula will not recover for well

over a century or possibly longer. The condition and land-use history on the

Olympic Peninsula are representative of the Pacific Northwest. The implications

for salmonid habitat, population, and resource value are not encouraging.

Degraded habitats will continue to produce at levels below their full capacity,

while other habitats will decline in quality.”

(McHenry et al. 1996). Climate change will only worsen these impacts on steelhead and their

habitat (Wade et al. 2013; Frissell et al. 2014).

Inside Olympic National Park, climate change is more than a threat – it is reality. Glaciers

that provide cold summer flows are shrinking and some have disappeared. By 2070, they may

all be gone (Fountain et al. 2022) and late spring snowpack is expected to decrease by 40-60%

(see Figure 17). Summer streamflows are decreasing, and peak winter streamflows are

increasing. Most models project drier summers, which will decrease summer and early fall

streamflows, especially in rain-dominant basins.

Figure 17. Trends in April Snowpack in the Western United States, 1955–2020, from:

https://www.epa.gov/climate-indicators/climate-change-indicators-snowpack.

Summer steelhead, which are nearly extinct, are particularly vulnerable to these habitat

changes (e.g., low summer flows during adult migration). Winter steelhead, which are declining,

are also vulnerable to these changes. These impacts will increase throughout the century.

A. Overview of Habitat Conditions on the Olympic Peninsula

1. WRIA 20 Habitat

Outside of Olympic National Park, steelhead habitat in WRIA 20 suffers from extensive

habitat degradation (Smith 2005, Klinger et al. 2008). A 2005 study conducted for the

Washington State Conservation Commission rated salmonid habitat in WRIA 20 as “poor-fair”

(Smith 2005). The study rated the following habitat attributes as poor: side-channel floodplain,

sediment quantity, sediment quality, bank/streambed/channel stability, instream large woody

debris (“LWD”), riparian, and water temperature (Smith 2005). The following attributes were

rated as fair: road density and hydro maturity high flows (Smith 2005). The only attribute rated

as “good” was pool habitat (Smith 2005).

All rivers in WRIA 20 where Olympic Peninsula steelhead occur have degraded habitat,

including but not limited to the following examples:

• The Sol Duc River has excessive sedimentation from landslides and high road densities,

poor LWD recruitment and riparian conditions, loss of wetland and off-channel habitat,

low and warm summer streamflows, loss of fog drip, blockages in tributaries, and loss of

cover and winter refuge habitat provided by debris jams (Smith 2000).

• The Bogachiel River has poor riparian and LWD conditions, an aggraded mainstem that

worsens as the river moves downstream, collapsing banks in the lower mainstem, fines

from exposed clay layers, and warm summer water temperatures (Smith 2000).

• The Calawah River experiences extensive landslides, high road densities, historic fire

and subsequent salvage logging impacts, excessive sedimentation, poor levels of LWD,

incised floodplains in the North Fork Calawah and South Fork Calawah Rivers (as well as

several tributaries), and low and warm summer streamflows (Smith 2000).

• The Hoh River suffers from significant habitat degradation outside of Olympic National

Park. Debris flows in the basin have been “common and devastating, resulting in

scoured, incised channels with few spawning gravels and LWD” (Smith 2000). There are

numerous areas with poor LWD and riparian conditions, passage blockages (in

tributaries), degraded water quality, few floodplain complexes, and fog drip loss due to

large conifer removal (Smith 2000). The Hoh River is also experiencing higher magnitude

flood events and lower summer flows (Piety et al. 2004; East et al. 2017; NIFC 2020)

• The Dickey River experiences sedimentation from roads and logging operations. The

Dickey River has poor riparian habitat, extensive substrate embeddedness, low and

warm summer flow, increased distribution of predators (e.g., northern pikeminnow),

passage blockages, reduced fog drip due to the removal of old conifers, altered

wetlands, decreasing LWD, and degraded channel and floodplain conditions in several

tributaries (Smith 2000). Low streamflows and high water temperatures are believed to

limit steelhead production in the river (Klinger et al. 2008).

• The Ozette River Basin has low levels of LWD, poor riparian conditions, warm water

temperatures, poor hydrologic maturity, invasive plant species, and a lack of marine-

derived nutrients (Smith 2000). The Ozette River basin has particularly high road

densities, which contribute to high sedimentation rates, a major limiting factor for the

area (Smith 2000; Klinger et al. 2008).

• The Wa’atch and Tsoo-yes Rivers suffer from numerous blockages throughout their

basins, lack LWD, and experience substantial sedimentation, mass wasting, high water

temperatures, and lack marine-derived nutrients (Smith 2000; Klinger et al. 2008).

• Goodman and Mosquito Creeks suffer from sedimentation, poor riparian conditions,

and, in the middle reach of Goodman Creek, low levels of LWD (Smith 2000).

2. WRIA 21 Habitat

Outside of Olympic National Park, steelhead habitat in WRIA 21 is degraded (QINLE

2011). Throughout most of the 20

century, the forest outside of the park was widely

harvested for timber, causing significant effects on the quantity and quality of salmon habitat

(Cederholm and Salo 1979; Cederholm et al. 1978; Smith and Caldwell 2001; McMillan 2006).

There is excessive sedimentation in the more intensely logged areas, especially in the

Clearwater subbasin (Cederholm et al. 1980; Klinger et al. 2008). Generally, water temperatures

and side-channel floodplain habitat are in poor condition (Smith 2005; Klinger et al. 2008).

Instream LWD, pool habitat, and riparian habitat are in fair condition (Smith 2005). Off-channel

habitat is limited in the Clearwater, Sams, and Salmon Rivers and Matheny Creek (Klinger et al.

2008). A glacier that fed the Quinault River melted away in 2011, contributing to lower summer

flows (Ahearn 2015). On the Queets River, peak and low flows are intensifying (QINLE 2011).

3. WRIA 19 Habitat

Steelhead habitat in WRIA 19 is degraded from decades of poor logging practices

(McHenry et al. 1996). Approximately 95% of the watershed’s old growth forest has been

converted into tree farms, and landslides triggered by logging operations have been common

(McHenry et al. 1996). For example, between 1950 and 1995, at least 330 logging-related

landslides occurred in the Hoko River basin alone (McHenry et al. 1996). Mass wasting events

have also occurred in the Clallam, Sekiu, and Pysht River basins (McHenry et al. 1996).

The basins in WRIA 19 suffer from floodplain development and alterations, loss of LWD,

estuary and nearshore alterations at stream mouths, degraded water quality, warm summer

stream temperatures, passage barriers, and conversion of riparian forests to non-forest uses

(McHenry et al. 1996; NOPLE 2015). Smith (2005) rated the following habitat attributes as being

in “poor” condition: side-channel floodplain habitat, sediment quantity, sediment quality,

bank/streambed/channel stability, instream LWD, riparian habitat, and hydro maturity high

flows. Water temperature varies from “poor” to “good” (McHenry et al. 1996, Smith 2005).

B. FOREST PRACTICES

Forestry has been, and continues to be, the primary land use outside of the park

(McHenry et al. 1998; Smith and Cadwell 2001; Copass 2016). Despite improvements to logging

practices on federal, state, and private land, the habitat of Olympic Peninsula steelhead

continues to suffer the consequences of past and ongoing logging operations throughout WRIA

19, 20, and 21 (McHenry et al. 1998; Smith 2000; NIFC 2020).

1. Timber Harvest on Federal and State Lands

a. Olympic National Forest

The Olympic National Forest spans 628,115 acres and surrounds most of Olympic

National Park. Logging in Olympic National Forest began in the 1920s and grew more intense in

the latter half of the 20th century (Halofsky et al. 2011). Extensive road building occurred

during the 1950s through the 1980s (Halofksy et al. 2011). Today, there are approximately

3,500 km of roads in the Olympic National Forest, which is roughly the driving distance between

Forks, Washington and Milwaukee, Wisconsin (Halosfky et al. 2011). Between the 1960s and

1990s, approximately half of the forest’s suitable land base was harvested (Halofksy et al.

2011). Prior to the adoption of the Northwest Forest Plan in 1994, timber operations involved

harvesting riparian trees, clearcutting, and broadcast burning (Halofksy et al. 2011). Stream

clearing and splash damming removed large wood from numerous stream channels (Halofksy et

al. 2011).

The adoption of the 1994 Northwest Forest Plan ushered in a new era of management,

which incorporates ecosystem management principles and specific protections for aquatic

ecosystems (Halofsky et al. 2011). However, despite improved forest practices, the effects of

logging continue to impair steelhead habitat in the Olympic National Forest. Logging roads and

associated channel crossings are still major issues for fish habitat quality (Halofsky et al. 2011).

The USFWS describes the sediment input from roads in the Olympic National Forest as

“chronic” (USFWS 2020). Although conifers have regenerated along some waterways, “many

riparian corridors have few conifers to provide large wood to streams” (Halofksy et. al 2011).

b. State Lands

The State of Washington began to overhaul its forest practice rules in the 1990s to

benefit salmonids and other species such as the marbled murrelet (Brachyramphus

marmoratus) and northern spotted owl (Strix occidentalis caurina). Prior to 1998, Washington’s

forestry rules lacked certain measures to protect salmonid habitat such as requiring timber

harvesters to leave trees in riparian zones for LWD recruitment (McHenry et al. 1998).

However, even with these improvements, the benefits to salmonid habitat are not clear.

“The effectiveness of currently implemented forest practices for minimizing

impacts remains uncertain. For example, incorrectly applied or inadequately

designed riparian management zones and incorrect stream typing classifications

are known problems that impair habitat protection strategies (Hansen 2001).

These practices result in loss of large woody material, fish passage impacts,

altered hydrology, water quality impacts, mass wasting (landslides), and elevated

stream temperatures (Naiman et al. 1998).”

(Cram et al. 2018).

The Olympic Experimental State Forest

The Olympic Experimental State Forest (OESF) was designated in 1992 and covers

approximately 270,000 acres of land (WDNR 2016). The OESF includes parcels of state-owned

trust land interspersed with private, federal, and tribal lands (Minkova et al. 2021). WDNR’s

goal is to manage the state trust lands to generate sustainable revenue for counties,

universities, and other trust beneficiaries, while also maintaining ecological values (WNDR

2016). Nearly half of the forest is young, with trees ranging between 20 and 39 years of age

(WDNR 2016).

Several major rivers occur in the OESF, including the Hoh, Queets, Quillayute,

Clearwater, Hoko, and Pysht Rivers (Minkova et al. 2021). WDNR does not manage streams in

the OESF to meet desired future conditions for salmonids (WDNR 2018). Instead, it seeks to

maintain or aid the restoration of riparian functions, water quantity, and water quality (WDNR

2018).

It is unclear whether the forest practices in OESF are significantly improving habitat for

steelhead. For example, Pollock et al. (2004) found that “within the OESF, a majority of streams

do not meet WDOE water quality standards for temperature, and that stream temperatures in

harvested basins are often (but not always) higher and more variable than stream

temperatures in unharvested basins” (Pollock et al. 2004). Additional study by Pollock et al.

(2009) found that the impact of harvest activities could not be fully mitigated by riparian

buffers alone. However, two studies by WDNR did not find similar water temperature

exceedances within OESF (Martens et al. 2019; Devine et al. 2022).

2. Logging Effects on Steelhead Habitat

Logging operations harm steelhead habitat by increasing water temperatures and

sedimentation in streams, removing large woody debris, reducing streamflows, and decreasing

habitat connectivity.

a. Temperature

Logging operations increase water temperatures in several ways: by removing upland

and riparian vegetation, which increases thermal radiation; decreasing groundwater inflow;

increasing sedimentation rates, which widen and shallowing river channels; and reducing large

woody debris recruitment (Cederholm and Salo 1979; Lynch et al. 1984; Frissell et al. 2014).

High water temperatures harm salmonids by causing direct mortality and metabolic distress,

altering migration and breeding behavior, increasing susceptibility to disease, and creating

thermal barriers that block access to habitat (Hicks 1999).

Several studies indicate that logging operations on the Olympic Peninsula increase

stream temperatures. Hatten and Conrad (1995) studied managed (i.e., logged) and

unmanaged stream reaches on tributaries to the Hoh, Queets, and Bogachiel Rivers and

Kalaloch Creek. They found that unharvested sites exceeded the state’s 16° C temperature

standard at an average of 1.8 days over a 39-day monitoring period (Hatten and Conrad 1995).

In contrast, the managed sites exceeded 16° C an average of 18.3 days during the same period

(Hatten and Conrad 1995). Because the harvested sites were representative of low elevation

harvesting sites, Hatten and Conrad (1995) assumed that most managed stream segments in

the western Olympic Peninsula did not meet water quality standards for temperature. Those

standards are intended to protect salmonids.

Although riparian buffer standards have been implemented on federal and state

forestland, research conducted by Hatten and Conrad (1995) and Pollack et al. (2009) indicate

that buffers alone may not be sufficient to improve stream temperatures. For example, Hatten

and Conrad (1995) found that the most important variable affecting stream temperature is the

total amount of land logged in a basin. However, Devine et al. (2022) did not find similar water

temperature exceedances in OESF as Pollock et al (2009), which may indicate that increased

shade in riparian zones is benefiting stream temperature.

Logging has had significant impacts in WRIA 19 basins, which have mostly been

converted into tree farms (McHenry et al. 1996). For example, water temperatures as high as

69° F and 73.4° F have been recorded in the Hoko River (McHenry et al. 1996).

b. Hydrologic Flows

Timber harvest affects streamflow by increasing permeable surface areas, transporting

water directly to streams, removing vegetation that stores water, and altering snow retention

and melt rates (Davis and Schroeder 2009, as cited in WCSSP 2013). These impacts change the

timing and magnitude of peak and low streamflows (Davis and Schroeder 2009, as cited in

WCSSP 2013). Peak flows can scour eggs and flush juveniles out of rearing habitat (Davis and

Schroeder 2009, as cited in WCSSP 2013). Low flows can alter flow-related migration cues,

disconnect habit, and increase crowding, competition, and vulnerability to predation (Davis and

Schroeder 2009, as cited in WCSSP 2013).

A study on the Hoh River showed that logging significantly impacts peak and mean daily

streamflows at watershed, subbasin, and basin level in descending order (Achet 1997).

“If about 27% of the basin was harvested (extracting 46. 9 million m

of timber)

the impact on peak would increase by 44 to 51% at basin, 34 to 41% at sub-basin

and 55 to 141% at watershed level in medium and high hydrologic conditions. By

reducing the timber harvest by 26%, the corresponding reductions in peak at the

basin level, sub-basin level and watershed level were 9 to 10%, 7 to 9% and 16 to

18% respectively. Regarding mean daily flow, harvesting 27% of the basin would

result in increase in mean daily flow by 30 to 47% at the basin level, 22 to 36 % at

the at the sub-basin level and 77 to 137% at the watershed level in medium and

high hydrologic conditions. When the timber harvest was 34.6 million m3 the

corresponding reduction in impact on mean daily flow at basin, sub-basin and

watershed level would be 7 to 10%, 5 to 8% and 14 to 20%” (Achet 1997).

c. Sediment

Logging operations increase sedimentation through debris flows and surface runoff

from logging roads, which expand channel widths and reduce pool depth (Furniss et al. 1991;

Jones et al. 2000). Increased sedimentation affects salmonids throughout their freshwater life

stages by suffocating incubating eggs and fry, reducing macroinvertebrate composition/prey,

causing respiratory failure, reducing or blocking access to habitat, and warming stream

temperatures by lowering water depth (McHenry et al. 1996, USFWS 2020).

Sedimentation from timber harvest significantly impairs aquatic habitat on the Olympic

Peninsula. Logging operations have caused landslides throughout the area. The most severe

damage has occurred on the Quinault River, Queets River, Clearwater River, Hoh River, Calawah

River, Sol Duc River, Hoko River, Seiku River, Sooes River, and Deep Creek basins (McHenry et

al. 1996; Cederholm et al. 1980; WSCC 2001; USFWS 2020).

Landslides have been particularly destructive in the Hoh River basin (Logan et al. 1991).

For example, in the winter of 1988-1989, 105 landslides occurred on Huelsdonk Ridge above

the Hoh River, mobilizing an estimated 83,000 cubic yards of sediment that deposited in the

South Fork and mainstem Hoh River (Logan et al. 1991). Only two of the 105 landslides

occurred in entirely old growth stands (Logan et al. 1991).

Sedimentation from logging roads impacts all watersheds where Olympic Peninsula

steelhead occur. The Hoh River basin suffers from excessive sedimentation caused by logging

roads (USFWS 2020). The sub-basins with the largest percentage of sediment over background

levels include the South Fork of the Hoh River (156%), Owl Creek (286%) and Winfield Creek

(210%) (Marsh 2012).

d. Large Woody Debris

Logging removes wood from riparian, upland, and headwater riparian areas that could

be recruited to streams (Klinger et al. 2008). The loss of wood increases the magnitude of peak

flow events, gravel scour, and bank erosion. It also decreases channel sinuosity and stability as

well as the quantity of pools and off-channel features (East et al. 2017). As a result, steelhead in

streams with low levels of LWD have less cover, food, off-channel habitat, and pools for rearing

and refuge. Depletion of LWD has long-term consequences for stream productivity.

Outside of Olympic National Park, nearly all rivers and streams suffer from the loss of

LWD (Smith 2000; Piety et al. 2004). Based on GIS imagery, East et al. (2017) found that the

Quinault River has less LWD than the Hoh and Queets Rivers. The Quinault River had 1.9% LWD

cover by area compared to 2.5% on the Queets River and 2.7% on the Hoh River (East et al.

2017).

e. Habitat Connectivity

Logging roads and culverts decrease access to on-channel and off-channel habitat (Davis

and Schroeder 2009, as cited in WCSSP 2013). As a result, there is diminished spawning,

rearing, and foraging habitat; increased competition caused by crowding; and less access to

refuges during high flow events (Davis and Schroeder 2009, as cited in WCSSP 2013).

Undersized culverts also compromise habitat quality for varying downstream distances due to

their disruption of the downstream transport of water, wood, and sediment.

f. Roads

Outside of Olympic National Park, most areas in WRIAs 19-21 are not properly

functioning for salmonid habitat due to road density (NIFC 2020). According to NMFS,

watersheds with road densities greater than three miles per square mile of watershed are not

properly functioning for salmonid habitat (NMFS 1996). Based on this criterion, nearly all

forestland outside of the park is not properly functioning for salmon habitat (NIFC 2020). The

West Fork Dickey River watershed has the highest road density (5.5 miles per square mile)

(NIFC 2020). Other high-density watersheds include Elk Creek-Calawah, Crooked Creek, the East

Fork Dickey, the Lower Bogachiel River and Big River watersheds (NIFC 2020).

Roads threaten several of the larger watersheds as well. For example, road density

remains high in the Queets River watershed on both federal and private land (USFWS 2020).

Road density in the Clearwater drainage is 3.2 miles per square mile in the upper section and

3.7 miles per square mile in the lower section (USFWS 2020). There are approximately 2.5 miles

of road per square mile in the Queets River floodplain outside of the Olympic National Forest

(USFWS 2020).

Roads cause excessive sedimentation on the lower Hoh River (USFWS 2020). Private

landowners and the Washington Department of Natural Resources have extensively logged the

Hoh River watershed and logging roads cause excessive sedimentation (USFW 2020).

Despite lower road densities, the Quinault River is also impacted by roads (USFWS

2020). Logging roads built during the early 1900s spread throughout the basin and streambank

hardening is degrading habitat conditions (USFWS 2020).

Roads have also triggered mass wasting events. For example, the USFS (2012) found

that 49% of mass wasting events in the Sitkum and South Fork Calawah watershed were caused

by roads.

C. CLIMATE CHANGE

The evidence of climate change is irrefutable. According to the IPCC, “[i]t is unequivocal

that human influence has warmed the atmosphere, ocean and land” and that “[w]idespread

and rapid changes in the atmosphere, ocean, cryosphere and biosphere have occurred” (IPCC

2021). Each of the last four decades have been successively warmer than the last (IPCC 2021).

Under status quo emissions levels, average global temperatures are expected to rise 1.6° C

above 1850-1900 levels between 2021 and 2041. Under the very high GHG emissions scenario

(SSP5-8.5), global temperature will likely increase by 2°C between 2041 and 2060 (IPCC 2021).

Climate change will affect the climate and hydrology of the Pacific Northwest. The

region is projected to lose snowpack and glacier mass and incur frequent and extreme

hydrological conditions (Mantua et al. 2010; Halofsky et al. 2011). Most models indicate that

winters will be wetter and summers will be warmer and drier (Mote and Salathé 2010; Elsner et

al. 2010; Kunkel et al. 2013). Summer droughts will be more frequent and severe (CIG 2009;

Mantua et al. 2010; Halofsky et. al 2011).

The western Olympic Peninsula is anticipated to warm, although it may warm slightly

less than other areas of the Pacific Northwest due to the moderating effect of the Pacific Ocean

(Halofsky et al. 2011; Dalton et al. 2016). Annual precipitation is projected to increase during

the winter and spring and decrease during the summer (Dalton et al 2016). Snowpack is

projected to decline and streamflows are projected to increase in the winter and decrease in

the summer (Dalton et al. 2016). The Olympic Peninsula is also expected to experience sea level

rise, stronger and more frequent storms, increased erosion, warmer water temperatures, more

low and high flow events, and increased glacial melt (Miller et al. 2013).

Climate change will also significantly alter the marine environment. Under the RCP8.5

scenario, models indicate that more multi-year warming events will occur in the northeast

Pacific Ocean (marine heatwaves, MHW), which may cause profound effects on salmonid

habitat (Joh and Di Lorenzo 2017; Oliver et al. 2019; Cheung and Frolicher 2020).

“Widespread and extreme negative impacts on marine life and fisheries

associated with the 2014-2015 marine heat wave are well documented. If the

projected increases in the area, magnitude and frequency of extreme warm

events are realized, and they are superimposed upon a systematic

anthropogenic warming trend; this combination would likely cause profound

negative impacts on marine life and fisheries all along the west coast of North

America, particularly those in the Gulf of Alaska in the second half the 21st

century.”

(Joh and Di Lorenzo 2017).

These changes will impact Olympic Peninsula steelhead, especially the summer-run

component, which is nearly extinct. Climate change is expected to cause widespread declines in

the quantity and quality of habitat for Olympic Peninsula steelhead (Halofsky et al. 2011). It is

unknown whether these fish will adapt quickly enough to these changes.

“It remains an open question whether present-day salmonid fish populations on

the Olympic Peninsula can adapt (either through phenological, phenotypic, or

evolutionary responses) at rates required to deal with the combination of

anthropogenic climate change and other habitat and ecosystem changes that

will come in the next century (Crozier et al. 2008).” (Halfosky et al. 2011).

1. Projected Climate and Hydrological Changes

a. Air Temperature

The Pacific Northwest has not experienced the same magnitude of warming since the

period between glacial and interglacial periods (Dalton et al. 2016). Between 1900 and 2014,

the average air temperature in Washington rose by approximately 1.5°F (USFWS 2020; Mote

and Salathe 2010). This warming trend will continue.

“Regardless of the scenario, warming is projected to continue throughout the

century in the Pacific Northwest. For the 2050s (2041 to 2070) relative to

1950-1999, temperature is projected to rise +5.8°F (range: +3.1 to +8.5°F) for a

high greenhouse gas scenario (RCP8.5). Much greater warming is possible after

mid-century under the more aggressive scenarios (RCP 6.0 and 8.5) ***.”

Models using less aggressive emission scenarios also project increased warming over the course

of the 21

Century. For example, Elsner et al. (2010) used data from the IPCC Fourth

Assessment Report to project that air temperatures will increase 0.3 C per decade in the Pacific

Northwest. However, even this lower rate of warming could produce “profound changes in the

hydrology and environment of the Northwest” (Mote and Salathe 2010).

Modeling shows that summer air temperatures in the Pacific Northwest will rise

significantly. For example, multi-model averages of the A1B scenario indicate that average June

through August air temperatures will increase at the following rates over the century: 1.7° C

(0.43° C to 3.4° C) by the 2020s; 2.7° C (1.3° C to 5.1° C) by the 2040s; and 4.7° C (2.7° C to 8.1°

C) by the 2080s (Mantua et al. 2010). Multi-model averages of the B1 scenario project June

through August temperature to increase as follows: 1.2° C (0.18° C to 2.4° C) by the 2020s; 1.8°

C (0.2° C to 3.7° C) by the 2040s; and 2.9° C (1.3° C to 5.1° C) by the 2080s (Mantua et al. 2010).

Air temperatures on the Olympic Peninsula are projected to increase (USFWS 2020).

Under the business-as-usual emissions scenario, air temperatures in the Olympic National

Forest are projected to increase by 2° F to 5° F (approximately 1° C to 2.5° C) between 2016 and

2045 relative to late 20

Century temperatures (USFWS 2020). Air temperatures in Olympic

National Park are projected to increase by 2.93° F between 2016 and 2045 and by 5.85° F

between 2046 to 2075 (LCD 2022).

https://cig.uw.edu/learn/climate-change/

b. Glaciers & Snowpack

Climate change is melting the Olympic Peninsula’s glaciers (USFWS 2020) and reducing

snowpack (Figure 17). Between 1980 and 2015, glaciers on the Olympic Peninsula decreased by

34% (Riedel et al. 2015). During that period, thirty-five glaciers and 16 perennial snowfields

disappeared (Fountain et al. 2022), including the Anderson Glacier, which contributed to

streamflow to the Quinault River (Dalton et al. 2016; NIFC 2020). Other glaciers that provide

water to the Quinault, Queets, Hoh, and Bogachiel Rivers (McHenry et al. 1996) are receding at

rates higher than previously recorded (NIFC 2020). Between 1981 and 2015, the glaciers that

feed into the Hoh River decreased by 40% (NIFC 2020). Under the RCP 8.5 scenario, Fountain et

al. (2022) estimates that glaciers will largely disappear from the Olympic Peninsula by 2070.

Climate change is expected to reduce late spring snowpack on the Olympic Peninsula

(Halofsky et al. 2011). Parts of the Olympic National Forest historically maintained snowpack

until April 1

(USFWS 2020). By mid-century, most spring snowpack will only exist at high

elevations in Olympic National Park, which will also experience “dramatic reductions” in spring

snowpack (Elsner et al. 2010; USFWS 2020).

c. Precipitation

The majority of climate models project that spring and summer rain will decrease on the

Olympic Peninsula, with up to a 24% decrease in summer rainfall (USFWS 2020). At least one

model, however, shows slightly wetter summers (LCD 2022).

The majority of climate models project that fall and winter rain will increase on the

Olympic Peninsula (USFWS 2020). Precipitation during the months of December through

February is likely to increase by 4.5% to 5% on average and depending on location (Halofksy et

al. 2011). Across the Pacific Northwest, spring precipitation has already increased by about 2%-

5% per decade over the past century, a trend that will likely continue (Abatzoglou et al. 2014). A

multi-mean model using a high emissions scenario projects a 6.5% increase in spring

precipitation by 2041-2070 (Dalton et al. 2016).

The Pacific Northwest will also experience wetter storm events (Miller et al. 2013).

Some models indicate that storms will be wetter and stronger on the Olympic Peninsula

(Halofsky et al. 2011).

d. Streamflow

Climate change is altering the hydrology of watersheds on the Olympic Peninsula.

Winter peak flows are getting higher, and summer low flows are getting lower (NIFC 2020).

These conditions are predicted to become worse for the major steelhead producing watersheds

in the coming decades. As illustrated in Table 8, these changes are occurring throughout the

DPS and are likely to have significant effects on winter and summer steelhead. For example, as

we discussed for steelhead, the lower 9km of the NF Calawah River (tributary of Calawah River

in Quillayute River system) went completely dry in 2002 killing thousands of salmonids,

including all adult and juvenile steelhead (McMillan et al. 2013; McMillan 2002). The Calawah

River is a lower-elevation, rain-fed tributary and is susceptible to drought conditions (Smith

2000). Considering low summer flows are likely to become worse in the future (Wade et al.

2013), streams like the NF Calawah may become uninhabitable during the summer months.

This also helps explain why the Calawah River has experienced a large change in summer low

flows relative to other watersheds (Table 8).

Table 8: Changes in low and high streamflow in Calawah, Quinault, Hoh, and Hoko Rivers from

1976-2019. (Source: NIFC 2020)

Watershed

Low flow

decrease

High flow

increase

Calawah River -48% +5%

Quinault River -20% +7%

Hoh River -15% +18.4%

Hoko River -13% +17.8%

Summer streamflows will continue to decrease on the Olympic Peninsula (Mantua et al

2010; Wade et al. 2013; Beechie et al. 2013; Dalton et al. 2016). Using A1B and B1 warming

scenarios, Mantua et al. (2010) found that annual summer low flows on the Olympic Peninsula

will be approximately 5% to 45% lower by mid-century. According to Dalton et al. (2016),

average summer flows are projected to decline by 30% across the Quillayute, Hoh, Queets, and

Quinault River basins. Summer flows in headwater areas will likely become more ephemeral or

stop during the summer months, and the duration of low flow periods will increase significantly

in all but the most rain-dominant basins (i.e., Hoh and Queets River basins) (Halofsky et al.

2011). Increased winter flooding may exacerbate summer low flow conditions by increasing

porous sediment deposits (Halofsky et al. 2011).

Winter flooding is increasing and will continue to do so throughout the 21

century

(NIFC 2020, Mantua et al. 2010). Peak flows are commonly at or above flood stage on the

Bogachiel and Calawah Rivers (NIFC 2020). Average winter flows are projected to increase by at

least 30% over the majority of stream reaches in the Quillayute, Hoh, Queets, and Quinault

River basins (Dalton et al. 2016). Winter flow in stream reaches on the Quinault and Quillayute

Rivers are projected to increase by 40% (Dalton et al. 2016). Late fall and early winter flooding

is also projected to increase (Halofsky et al. 2011).

It is important to note that these summer and winter streamflow projections are based

on moderate (A1B) and low (B1) emission scenarios established by the IPCC. These scenarios

project similar amounts of emissions through the mid-21

century as the A2 scenario, which is a

high emissions scenario (Halofsky et al. 2011). The A2 scenario, however, projects higher

emissions during the latter half of the century (Halofsky et al. 2011).

Figure 18. Current and projected (year 2040) reductions in summer streamflow and increases in

winter streamflow in the Sol Duc, Dickey, Quillayute, Calawah, Bogachiel, Hoh, Queets, and

Quinault Rivers (Reeves et al. 2018).

e. Water Temperature

Water temperatures on the Olympic Peninsula are projected warm. Increasing stream

temperatures pose higher risks to the quality and quantity of steelhead habitat because

steelhead typically however, it is unlikely that they will exceed the thermal tolerances of

steelhead (USFS 2020; Miller et al. 2013; Mantua et al. 2010). Using A1B and B1 emission

scenarios, Mantua et al. (2010) projected that water temperatures will warm by 1° to 2° C (

Fig. 19. Current and projected 2040 summer water temperatures in the Sol Duc, Dickey,

Quillayute, Calawah, Bogachiel, Hoh, Queets, and Quinault Rivers (Reeves et al. 2018).

f. Other Water Quality Issues

Increased winter precipitation will likely increase runoff and landslides from logging

operations, which would increase sediment pollution in Olympic Peninsula rivers and streams

(Klinger et al. 2008; Halofsky et al. 2011).

2. Effect of Freshwater Habitat Changes on Olympic Peninsula Steelhead

Climate change will adversely modify freshwater salmonid habitat on the Olympic

Peninsula. Because of their longer freshwater residency, steelhead are more sensitive to these

freshwater habitat changes than other salmonids (Halofsky et al. 2011).

a. Low Summer and Early Fall Streamflow

Olympic Peninsula steelhead have high exposure to extreme low flows (Wade et al.

2013). Lower flows will decrease habitat availability, elevate water temperatures, and induce

thermal stress on salmonids (Crozier and Zabel 2006; Wade et al. 2013; Dalton et al. 2016;

Ohlberger et al. 2018). Reduced summer and fall streamflows in rain-dominated basins may

adversely affect summer steelhead migration (Holafsky et. al. 2011). For example, there will be

less cold water and fewer holding pools for migrating steelhead, which may be particularly

stressful for summer steelhead and lower their reproductive success (Dalton et al. 2016). Low

streamflows are already a limiting factor on the Calawah River (Phinney et al. 1975), particularly

the NF Calawah (McMillan 2002), and summer streamflows are showing a decreasing trend on

many Olympic Peninsula Rivers, such as the Hoh River (NIFC 2020). Climate change will

exacerbate these effects (NIFC 2020).

b. High Late Fall and Winter Streamflow

Olympic Peninsula steelhead have high exposure to increased winter streamflows

(Wade et al. 2013). Average winter flows (January – April) are expected to increase by at least

30% by 2040 (Dalton et al. 2016). Average winter flows are projected to increase by 31-40% in

the Hoh, Queets, and Quinault Rivers (Dalton et al. 2016). Increased winter flows could reduce

the survival of developing eggs, embryos, and juveniles (Dalton et al. 2016) This is a higher risk

in confined streams, which are more susceptible to scour (Halofsky et al. 2011). Summer

steelhead juveniles that emerge in winter in high gradient streams are at greater risk of

displacement (Dalton et al. 2016). Peak flows could also reduce the availability of slow-water

habitat for juveniles (Mantua et al. 2010; Halofsky et al. 2011), which may reduce parr-smolt

survival rates (Halofsky et al. 2011). Increased runoff will also likely increase sedimentation in

steelhead habitat (East et al. 2017).

c. Water Temperature

Based on models that use A1B and B1 scenarios, water temperatures are unlikely to

exceed lethal limits for steelhead, but they would increase to levels that reduce growth and

increase the risk of predation (Dalton et al. 2016). For example, temperatures at or above 15° C

can inhibit the smolt transformation process (Miller et al. 2013). As explained in the water

quality section, many stream reaches on the Olympic Peninsula already exceed 16° C and

warmer spring stream temperatures could impede smolting and force the populations to adapt

to smolt during an earlier time.

It is also possible that warmer temperatures could expand growing seasons and increase

food web productivity, which would benefit steelhead (Halofsky et al. 2011). However, this

benefit could be offset by increased flooding, lower summer flows, increase predation risks

(Dalton et al. 2016). Additional food availability may be also offset by increased competition

(Dalton et al. 2016).

3. Projected Changes in Marine Environment

Climate change is altering nearshore and offshore habitat of Olympic Peninsula

steelhead (Klinger et al. 2008; Miller et al. 2013). These changes include warming sea surface

temperatures and potential alterations in upwelling, hypoxia, and acidification (USFWS 2020).

The scope and intensity of these impacts are uncertain (Miller et al. 2013). For example, it is

possible that climate change could decrease upwelling, which would lower productivity (Klinger

et al. 2008). It could also increase upwelling, although warmer ocean temperatures could limit

productivity benefits (Miller et al. 2013). Increased sea surface temperatures will drive the

southern boundary of steelhead northward, potentially contracting the range of the species

(Abdul-Aziz et al. 2011). These changes will negatively affect salmon and steelhead survival

rates (e.g., Kilduff et al. 2015).

a. Sea Surface Temperature

Sea surface temperatures are projected to increase in the Pacific Northwest (Mote and

Salathe 2010; Miller et al. 2013; USFWS 2020). Mote and Salathe (2010) projected that ocean

surface temperatures will increase by approximately 1.2° C by 2050 relative to the 1970-1999

average temperature. Other models project that surface temperatures may increase by 1.2° C

to 3° C by mid to late century (USFWS 2020). These projected increases are substantially

outside 20

century variability (Mote and Salathe 2010). Marine heatwaves such as the 2014-

2015 “Blob” are also likely to reoccur more frequently (Oliver et al. 2019; USFWS 2020) as the

ocean heat content increases (Cheung and Frölicher 2020; von Schuckmann et al. 2020).

Fig. 20. Sea Surface Temperature Anomaly from August 1 - 31, 2019 (Image source:

https://earthobservatory.nasa.gov/images/145602/marine-heat-wave-returns-to-the-

northeast-pacific)

b. Upwelling

Climate change may affect upwelling, although there is much uncertainty about this

possibility. For example, models have produced mixed results on whether upwelling favorable

winds will change (Miller et al. 2013). There is some evidence indicating that upwelling patterns

may become shorter and more intense (USFWS 2020). Increasing surface temperatures may

influence the timing and magnitude of upwelling as well (Miller et al. 2013).

c. Acidification

The Pacific Northwest is vulnerable to ocean acidification (Miller et al. 2013). Increased

acidification threatens key species in the food web, including zooplankton, pteropods, crabs,

and krill (Busch et al. 2013; Mathis et al. 2015; Dalton et al. 2016).

d. Anoxic and Hypoxic Events

Anoxic and hypoxic events have occurred off the Washington coast (Klinger et al. 2018;

USFWS 2020;). As sea surface temperatures warm, dissolved oxygen levels are expected to

decrease (Miller et al. 2013).

4. Effect of Changed Marine Conditions on Olympic Peninsula Steelhead

NMFS anticipates that climate change will continue to limit ocean productivity for

salmonids (Ford 2022).

“Historically, ocean conditions cycled between periods of high and low

productivity. However, global climate change is likely to disrupt this pattern, in

general, leading to a preponderance of low productivity years, with an unknown

temporal distribution (Crozier et al. 2019). Recent (2015–19) ensemble ocean

indicator rankings include four of the worst seven years in the past 20, meaning

that an entire salmon or steelhead generation could have been subjected to

poor ocean productivity conditions.” (Ford 2022).

Warming will also cause steelhead ocean habitat to progressively decrease throughout

the century (Abdul-Aziz et al. 2011). Summer marine habitat is anticipated to contract 8-10% by

the 2020s, 15%-19% by the 2040s, and 24%-43 by the 2080s (Abdul-Aziz et al. 2011).

Although some forage fish may benefit from climate change, forage fish abundance is

likely to decrease (USFWS 2020), resulting in less prey for steelhead.

C. WATER QUALITY

Limited water quality monitoring makes it difficult to identify the status and trends of

water quality on the Olympic Peninsula (Klinger et al. 2008). However, the limited data that

does exist shows that there are many river and stream segments where Olympic Peninsula

steelhead occur that do not meet water quality standards (Klinger et al. 2008; WDOE 2016).

These water quality standards are intended to protect salmonids.

In WRIA 19, thirteen rivers and streams have segments that do not meet state

temperature standards and need Total Maximum Daily Loads (TMDLs). These waterbodies

include: Deep Creek, Hoko River, Little Hoko River, Clallam River, Sekiu River, Sekiu River (North

Fork), Sekiu River (South Fork), Pysht River, Pysht River (South Fork), Salt Creek, Lyre River, and

the West Twin River (WDOE 2016). Deep, Salt, and Bear Creeks are also on the 303(d) list for

dissolved oxygen (WDOE 2016).

Dozens of rivers and stream segments in WRIA 20 do not meet water temperature

standards (WDOE 2016). The following rivers and streams have at least one river segment that

does not meet temperature standards and are in need of a TMDL: Hoh River, Coal Creek,

Crooked Creek, Ozette River, Umbrella Creek, Willoughby Creek, Owl Creek, Split Creek, Maple

Creek, Anderson Creek, Line Creek, Alder Creek, Winfield Creek, Nolan Creek, Bogachiel River,

Coal Creek, Dickey River (mainstem, East Fork, Middle Fork, and West Fork) Elk Creek, Sol Duc

River, SF Calawah River, Sitkum River, Sooes River, Trout Creek, Big River, Fisher Creek

(McQuarry Creek), Ozette River, and Lake Creek (WDOE 2016). Monitoring by the Hoh Tribe

between 2006 and 2015 revealed that all but one major salmonid tributary to the Hoh River

had summer water temperatures that exceeded state water quality standards (NIFC 2020).

In addition to temperature exceedances, several rivers and tributaries in WRIA 20 have

at least one stream segment that does not meet pH or dissolved oxygen standards and need

TMDLs for these parameters (WDOE 2016). The following rivers and stream are on the 303(d)

list for pH: Coal Creek, Crooked Creek, Ozette River, Sol Duc River, Big River, and Palmquist

Creek (WDOE 2016). The following rivers and streams are on the 303(d) list for dissolved

oxygen: Big River, Coal Creek, Ozette River, Siwash Creek, South Creek, Umbrella Creek, Lake

Creek, and Bear Creek (WDOE 2016).

There is less water quality data for WRIA 21. Kalaloch Creek, Matheny Creek, and the

Sams River are on the 303(d) list for temperature (WDOE 2016).

D. ANTHROPOGENIC MIGRATION BARRIERS

When summarizing habitat conditions for Olympic Peninsula steelhead, Busby et al.

(1996) assumed that “minor blockages (such as impassable culverts) are likely throughout the

region." Today, we know that at least 1,214 human-made barriers exist in WRIA 19-21, many of

which are completely impassable (WDFW 2020). According to WDFW, there are 272, 470, and

473 barriers in WRIAs 19, 20, and 21, respectively (WDFW 2020). Although progress has been

made on removing and repairing culverts and other human-caused barriers, there are still

passage issues on many streams throughout the range of Olympic Peninsula steelhead.

The Quinault Indian Nation, Quileute Indian Nation, Hoh Tribe, Makah Nation, and other

western Washington tribes won a major legal victory for salmonids that requires the State of

Washington to repair or replace state-owned culverts throughout the Olympic Peninsula and

elsewhere. United States v. Washington, 853 F.3d 946 (9th Cir. 2017) aff 138 S.Ct. 1832 (2018).

However, the Washington State Department of Transportation has less than 12% of the funding

it needs to complete these actions by the court-imposed deadline (NIFC 2020). If the state

continues at its current rate, it will not complete its work until 2034 (NIFC 2020).

State and private landowners have made progress repairing culverts on state and large

private forest roads under the Road Maintenance and Abandonment Plan (NIFC 2020).

Although the plan was supposed to be completed in 2021, approximately 85% of the work is

done in the coastal region (NIFC 2020). However, there are hundreds of non-RMAP culverts on

private, county, and federal land throughout WRIAs 19, 20, and 21 (NIFC 2020). Roughly half of

the non-RMAP culverts are not passable (NIFC 2020).

II. OVERUTILIZATION FOR COMMERCIAL AND RECREATIONAL PURPOSES

Wild winter steelhead populations in the Olympic Peninsula DPS have a long history of

supporting intensive recreational and commercial fisheries until the last few years when

fisheries were altered and eventually closed due to conservation concern over unprecedented

low run sizes (WDFW blog post: https://wdfw.medium.com/changes-to-the-coastal-steelhead-

season-67131dd05ba7). The chronic declines in run size combined with altered run-timing

(McMillan et al. 2022), low levels of repeat spawning, and now, dramatic freshwater and

marine effects linked to climate change, suggest that the four major populations of wild winter

steelhead are being overutilized in fisheries. Cram et al. (2018) was able to estimate harvest

rates for 33% of the populations in the Olympic Peninsula DPS up until 2013 and found a mean

annual harvest rate of 25.6%, which was the highest harvest rate among all steelhead DPSs in

Washington State. We updated the harvest estimates for the Quillayute, Hoh, Queets, and

Quinault Rivers through 2020 (Table 1) and discuss those results for each population specifically

in the Harvest Impacts section.

These harvest fisheries put Olympic Peninsula steelhead at risk. WDFW’s Steelhead at

Risk report did not evaluate “population-specific impacts” within the DPS because the “harvest

rates are difficult to interpret without accompanying demographic analyses, since risks posed

by harvest depend on a population’s productivity” (Cram et al. 20198). WDFW acknowledges

there are substantial data gaps regarding the productivity of Olympic Peninsula steelhead

(Cram et al. 2018). Nonetheless, WDFW and the Treaty tribes manage Olympic Peninsula

steelhead under a maximum sustainable harvest regime that does not include the level of detail

necessary to responsibly manage harvest or maintain the persistence of the species (Gibbons et

al. 1985; Burge et al. 2006). And summer steelhead are not managed or monitored but are

subject to indirect harvest in fisheries.

A. Escapement Goals

The framework for the modern winter steelhead fishery was established following the

1974 “Boldt Decision” (U.S. v. Washington, 384 F.Supp. 312 (1974)), which reaffirmed the

tribe’s Treaty rights to fish (Clark 1985). Shortly after the Boldt Decision was issued, WDFW

conducted a study based on a series of snorkel surveys to estimate steelhead parr production in

populations across western Washington, including the Quillayute River and its tributaries,

which they then used to generate models and estimate a range of escapement goals under

varying assumptions (Gibbons et al. 1985). While the effort was laudable at the time, there are

several limitations, some of which include sample size and representation (e.g., the juvenile

snorkel surveys only occurred in a few rivers), survey counts that were not adjusted for diver

error (likely resulting in underestimates), and failure to account for the depletion of early-timed

steelhead (McMillan et al. 2022) or the effects of spatial distribution (Finstad et al. 2013; Atlas

et al. 2015). Further, the escapement goals do not account for potential mining of certain life

histories in species that display a high level of life history diversity, which in turn can reduce

productivity and resilience (Ricker 1963). Last, there has not been any effort to evaluate or

develop escapement goals or management targets for populations of summer steelhead.

Despite these limitations, the range of escapement goals estimated for each population

of winter steelhead were not unreasonable for the time. In fact, the highest estimates could

prove to be sufficiently conservative if the major questions are addressed through a sensitivity

analysis and aspects of diversity, such as run timing and iteroparity, were accounted for.

Unfortunately, WDFW and the co-managers disagreed over the escapement goal science and

ultimately, the parr production models were rejected in favor of Maximum Sustained Harvest

(MSH) and as a result, the agreed upon escapement goals were either among the lowest or

were lower than the lowest estimates in Gibbons et al. (1985). The disagreement continues to

this day, and as such, there is no agreed upon escapement goal for the Queets River or the

Lower Quinault River (Cram et al. 2018). This is a great concern given the declines of wild winter

steelhead specifically in those populations, but also overall because wild steelhead in the

Olympic Peninsula DPS experience the highest harvest rates of any steelhead populations in

Washington (Cram et al. 2018) and likely the world. The high harvest rates and declining

population trends raise serious questions about whether the management targets are sufficient

to sustain the winter steelhead populations and the fisheries they provide into an uncertain

future.

A previous assessment of winter steelhead management in western Washington argued

that a strong reliance on an MSH approach jeopardizes the health and resilience of a diverse

species such as steelhead (Burge et al. 2006):

“It has become abundantly clear that MSH theory and harvest models have not

provided adequate protection for wild steelhead in the 20th and 21st centuries

because they are too simplistic and allow high harvest rates that are

unsustainable. These models do not annually or temporally account or plan for

environmental variation, management error, the role of genetic and life history

diversity in stock resilience and productivity, or the rebuilding of depleted

tributary or seasonal runs. Rather, the models are rigid numbers-based

equations and provide management tools that were developed to provide

maximum harvests from a population without adequate considerations for long

term stock health or the sustainability of annual fisheries.”

Limited information and poor data quality on escapement levels can contribute to

population decline (Knudsen 2000). A 2018 report from the National Park Service noted that

some escapement estimates of salmonids on the Olympic Peninsula “are currently derived from

surveys that occur in limited index reaches or are based on one or a few surveys during the

peak of spawning” (Duda et al. 2018). The authors also noted there is a need “to fully

understand the extent of bycatch and incidental mortality of Pacific salmonids in recreational

and commercial fisheries in Olympic Peninsula Rivers” (Duda et al. 2018).

Given what we have outlined above, current escapement goals for the four major

Olympic Peninsula populations are likely far too low and should be revised based on

consideration of all available data. To ensure adequate escapement, the escapement goals

should be revised to incorporate precautionary estimates of parr and smolt densities required

to secure the resilience of each population given current and expected future variability in early

marine survival. Estimates should be based on spawner age-composition and repeat spawner

frequency targets with the primary objective of attaining minimum egg deposition sufficient to

attain target mean parr or smolt densities in rearing habitats (see Gayeski et al. 2016 for some

likely target parr density values). And the goals should begin to account for life history diversity

that underpins the evolutionary strengths of steelhead.

In addition, any harvest – whether tribal commercial, ceremonial, subsistence, or

recreational – should be reduced far below current comanager estimates of maximum

sustainable yield/harvest (MSY). Recent declines in productivity indicate that MSY harvest rates

are likely incompatible with population persistence or recovery. Petitioners obtained

population data sets from WDFW for the Hoh River and Quillayute River winter steelhead

populations (WDFW 2022d) These data sets contained run reconstructions that provided time

series of spawners and maiden recruits spanning more than 25 years. Gayeski conducted a

preliminary Bayesian estimation of the Ricker spawner-recruit model with time-varying

productivity parameter (alpha) and constant capacity parameter (beta) on both data sets. The

analyses were conducted in the program Stan. A copy of this estimation is included with this

petition (See Appx. A).

Gayeski’s estimates are shown in Figures 21 and 22. Both populations display a strong

declining trend in alpha and the unfished equilibrium abundance (EQ). The Hoh River

population displays a steady decline since 2003; the Quileute River population shows a steady

decline beginning in 2007. The average value of alpha for the Hoh River population for the

recent period of decline (2003 to 2013) is 2.51; the average value of EQ is 2,297. The average

value of alpha for the Quillayute River population for the recent period of decline (2007 to

2015) is 2.16; the average value of EQ is 9,058. For comparison, the peak values of alpha and EQ

for the Hoh River population during the period spanned by the data (1987 to 2013) occurred at

the beginning of the time series, where alpha averaged between 3.6 and 3.7, and EQ between

3,200 and 3,300. Peak values for the Quillayute River population during the period spanned by

the data (1978 to 2015) occurred in the early 1990s, where alpha averaged between 4.2 and

4.5, and EQ averaged between 17,000 and 18,000. These are biologically very significant and

worrying declines. Further, values of both parameters in the past three years of record for each

population are significantly lower than the recent average values.

Figure 21. Time series of time-varying Ricker productivity parameter alpha and unfished

equilibrium abundance for fixed beta parameter from a Bayesian stock recruit analysis of Hoh

winter-run steelhead.

1.5

1.7

1.9

2.1

2.3

2.5

2.7

2.9

3.1

3.3

3.5

3.7

3.9

1500

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1900

2000

2100

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2300

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1986 1990 1994 1998 2002 2006 2010 2014

Equilibrium Spawner Abundance

Brood Year

Hoh winter run steelhead BYs 1987 to 2013, left axis = EQ;

right axis = Alpha

time-varying Ricker alpha with constant Beta = 2523.

EQ Alpha

Figure 22. Time series of time-varying Ricker productivity parameter alpha and unfished

equilibrium abundance for fixed beta parameter from a Bayesian stock recruit analysis of

Quileute winter steelhead.

It is very probable that the other two large winter steelhead populations (i.e., Queets

and Quinault River populations) and numerous smaller populations have experienced similar

declines. These declines are cause for alarm.

B. Harvest Impacts

Wild winter steelhead populations in the Olympic Peninsula DPS are fished for and

harvested in recreational and tribal fisheries at the highest levels of any steelhead populations

in Washington State (Cram et al. 2018). Harvest can affect abundance, diversity, productivity,

and spatial structure of salmon and steelhead populations, which in turn interact to shape the

short- and long-term viability of the exploited populations (McElhaney et al. 2000). There are

two main sources of harvest, including impacts from tribal gillnets and from recreational catch

and release fishing.

Tribal commercial harvest of wild steelhead annually occurs in the Quinault, Queets,

Hoh, and Quillayute River systems, and up until 2016, recreational anglers were permitted to

kill wild steelhead on eight rivers on the Olympic Peninsula (Cram et al. 2018). Tribal harvest

includes all fish that were reported, including commercial and subsistence catch.

1.5

1.7

1.9

2.1

2.3

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18500

1978 1982 1986 1990 1994 1998 2002 2006 2010 2014

Equilibrium Spawner Abundance

Brood Year

Quileute winter run steelhead BYs 1978 to 2015, left axis = EQ; right

axis = Alpha time-varying Ricker alpha with constant Beta = 11985.

EQ Alpha

The recreational fishery is now catch and release only for wild steelhead, and although

catch and release generally has low mortality rates (2-10%: Hooton 1986; Barnhardt and Taylor

1996; Nelson et al. 2005), recreational fishing pressure and success has increased to the point

where anglers are now catching and releasing the entire escapement, on average, more than

one time (Bentley 2017; WDFW 2022c). Prior to 2016, WDFW determined recreational harvest

using punch cards. Now that the fishery has shifted to catch and release, WDFW relies on creel

surveys to estimate angler encounter rates with wild winter steelhead and estimate a 10%

mortality to fish that are caught and released (Bentley 2017; Cram et al. 2018).

It is likely that harvest rates in the Quillayute, Hoh, Queets, and Quinault Rivers are

underestimates. For example, we could not find estimates of gillnet dropout rates for wild

winter steelhead in the tribal commercial fisheries. Injuries sustained by fish that escape gillnets

can dramatically reduce their reproductive success if they survive to spawn (Baker et al. 2013).

It is unknown how many redds are counted each year that were constructed by fish with lower

fitness due to such injuries.

In addition, on average, every wild winter steelhead that escaped to spawn in the Hoh

River in 2014 was caught and released, on average, 1.4 times by anglers (Bentley 2017). WDFW

has several years of raw creel survey data, but the Hoh River in 2014 was the only stream and

year for which they expanded those counts to estimate total encounter rates. The raw creel

data suggests high encounter rates in other streams, with possibly even higher encounter rates

in the Bogachiel and Sol Duc Rivers (WDFW 2022c, see Fishing Regulations section below).

Although the 10% mortality rate that WDFW applies to recreational encounters is relatively

high and conservative, those estimates are based on a fish being caught one time (Hooton

1987; Hooton 2001; Nelson et al. 2005; Twardek et al. 2018). The mortality rate for fish caught

multiple times is unknown, but fish that enter early and remain in the system longer have a

higher frequency of being caught more than once (Hooton 1987; Hooton 2001). Further,

emerging research on Atlantic salmon indicates there can be significant sub-lethal impacts that

can alter the migration and reduce the reduce fitness of adults (up to ~ 20-25%) that are caught

and released (Richard et al. 2013; Bouchard et al. 2021; Papatheodoulou et al. 2021).

Potential effects from net dropouts and high recreational encounter rates, combined

with the high harvest levels, raise concerns whether current estimates of harvest fully account

for all impacts.

Below we review what is currently known about harvest rates for the four major wild

winter steelhead in the Olympic Peninsula DPS and discuss harvest in relation escapement

goals, how it has most likely contributed to changes in run timing, and how it is potentially

affecting size and age and iteroparity.

We also summarize harvest records for summer steelhead and, when possible, we

distinguish between wild and hatchery fish, but that is a challenge with the catch data. See

(WDFW 2022e). From 1978-1985 returning adults were not marked with an adipose fin clip, and

thus, only total catch of summer steelhead is reported for tribal and recreational fisheries.

Thereafter, recreational catch consistently delineates hatchery and wild summer steelhead

from 1985-2016, while the tribal catch records do not – except for the Quillayute River system

from 2001-2020. Retention of wild summer steelhead by recreational anglers was banned by

WDFW and the National Park Service in 1992, and consequently, from 1993 onward all

recreational fisheries for wild summer runs were catch and release. Even though contemporary

catch of wild summer steelhead is not high by tribes and anglers must employ catch and

release, impacts could be more deleterious than they seem given the very depleted status of

the stocks and the high levels of hatchery fish (see Hatchery Impacts section below).

1. Quillayute River System

Annual harvest rates of wild winter steelhead from 1978-2020 ranged from 10% to 55%

in the Quillayute River system (Figure 23a), with a mean annual harvest rate of 28%, which is

the lowest harvest rate of the four major populations (Table 1). Compared to the mean annual

run size of 13,064 wild winter steelhead, the mean harvest rate equates to 3,724 fish per year.

Total wild winter steelhead harvested per year ranged from 1,166 to 7,561 fish. Harvest rates

have generally declined the past few years (Figure 23a) in response to smaller run sizes and

shortened fishing seasons. The proportion of hatchery to wild winter steelhead in the

commercial fishery was 2.1:1, and the proportion of hatchery to wild winter steelhead in the

recreational fishery was 3.3:1 (Duda et al. 2018). The mean annual harvest rate of hatchery

winter steelhead was 64% (range 32% – 86% of total run size; SD 12%) (Duda et al. 2018).

Figure 23. Annual harvest rates of wild winter steelhead populations in the a) Quillayute, b)

Hoh, c) Queets, and d) Quinault Rivers. Based on co-manager data sets covering period circa

1980-2020.

Harvest rates of wild summer steelhead from 1978-2020 were not available because co-

managers do not monitor summer runs. We do, however, have catch data for summer

steelhead, though wild and hatchery fish could not be distinguished prior to 1986, the year the

first adults returned with adipose fin-clips. Hatchery summer runs were first planted in the

Calawah and Sol Duc Rivers in 1977 and were stopped in the Sol Duc in 2012 (Figure 24), and

according to the hatchery records we reviewed, the Quillayute River watershed is the only

system to have received direct plants of hatchery summer steelhead (Duda et al. 2018).

Figure 24. Number of hatchery summer steelhead smolts released into the Sol Duc and Calawah

Rivers, which are tributaries to the Quillayute River watershed.

We separated the Quillayute River catch data into recreational (1978-2016) and tribal

Treaty catch (1978-2020) because the recreational catch distinguishes between hatchery and

wild fish in all years possible while we could only find that delineation for tribal catch from

2001-2020. As mentioned previously, historical catch of wild summer steelhead was very low

and inconsistent in the Quillayute prior to hatchery releases in 1977, but from 1978 to 1985

there was a spike in harvest of adult summer steelhead, many of which were likely hatchery

(Figure 25). After fin-clipped hatchery adults started to return in 1986, the catch of wild fish

declined and remained low while the harvest of hatchery fish increased (Figure 25). Retention

of wild summer steelhead by recreational anglers was banned by WDFW and the National Park

Service in 1992, and consequently, from 1993 onward all recreational fisheries for wild summer

runs were catch and release. However, WDFW records show some wild summer steelhead

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harvest after 1992, and we are not sure why, though it is possible the data represents fish that

were illegally harvested (Figure 25). Regardless, based on the punch card data, 2006 was the

last year a wild summer steelhead was reported as harvested by recreational anglers in the

Quillayute River system (Figure 25). Mean annual recreational harvest from 1986-2006 was 54

wild summer steelhead (range = 0 – 239) and from 1986-2016 mean annual recreational

harvest was 673 hatchery summer steelhead (range = 119 – 1,974).

Figure 25. Number of summer steelhead harvested by recreational anglers from 1962-2016 in

the four largest populations in the Olympic Peninsula DPS. The black line denotes wild

steelhead with an adipose fin. The grey line denotes unknown steelhead because harvest

occurred during a period when hatchery fish were introduced, but not outwardly marked, so

wild and hatchery could not be distinguished. The dashed black line denotes hatchery

steelhead. Retention of wild summer steelhead by recreational anglers was banned from 1993

onward by WDFW and the National Park Service.

We found two sets of tribal catch data with slightly different annual catch numbers. The

first ranges from 1978-2016 and does not distinguish between hatchery and wild summer

steelhead. The second data set ranges from 2001-2020 and, although it does not contain catch

records for all years, it does delineate hatchery and wild steelhead in years when catch records

were available (Figure 26 and Figure 27). The data sets are compared in Figure 27, which shows

the data from the 1978-2016 and 2001-2020.

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Overall, the tribal harvest of summer steelhead in the Quillayute River is generally

higher than other populations, except for the Quinault River (Figure 26). Mean annual tribal

catch from 1978-2016 was 341 summer steelhead of unknown origin (range = 43 – 1,120), while

mean annual tribal catch from 2001-2020 was 107 wild summer steelhead (range = 24 - 244)

and 308 hatchery summer steelhead (range = 28 – 839). Given the lack of monitoring and

escapement goals, there should be concern because the reported harvest of wild steelhead is

high (Figure 24) compared to the snorkel counts of staging adults (Table 7) from circa 2000-

2010.

Figure 26. Number of summer steelhead harvested by tribal fishers in Quillayute, Hoh, Queets,

and Quinault Rivers, with occasional data available for some populations from 1946-1975 and

generally consistent annual catch data available from 1976-2017. Data were provided by

WDFW and did not include information on which fish were hatchery or which were wild. Based

on releases of hatchery smolts in Quillayute River system in 1977, 1978 would be the first year

of potential adult returns. So, data thereafter is conflated by presence of hatchery fish, which

could represent a substantial component of the catch – even in watersheds outside the

Quillayute River as evidenced by sport catch data in Hoh and Queets Rivers.

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Figure 27. Number of summer steelhead harvested in Quillayute River by tribal fishers, which

includes two disparate data sets: data set 1, which only provides total numbers of summer

steelhead harvested and does not delineate between hatchery and wild; and data set 2, which

delineates total harvest by hatchery and wild. We provide both data sets because they

contained different harvest numbers.

2. Hoh River

The mean annual harvest rate from 1980-2020 was 35% for Hoh River wild winter

steelhead (Table 1), with a range of 7% to 55% per year (Figure 23b). This equates to a harvest

of 1,361 wild winter steelhead per year, with a range of 261 – 2,800, compared to a mean

annual run size of 4,118 steelhead. Harvest rates have generally declined the past few years

(Figure 23b) in conjunction with declining run sizes and shortened fishing seasons. The ratio of

hatchery to wild winter steelhead in the commercial fishery was 2:1, and it was 3.4:1 in the

recreational fishery. (Duda et al. 2018).

As with the other populations, we delineate the recreational harvest of summer

steelhead in the Hoh River population into three periods, including: (1) the historical period

discussed in a previous section dating from circa-1950s through 1977; (2) the post hatchery

release period when wild and hatchery steelhead could not be distinguished (1978-1985); and

(3) the period from 1985 to the present when wild and hatchery fish could be distinguished

(Figure 25). Although the Hoh River did not receive hatchery releases, the harvest of summer

steelhead sharply peaked from 1978-1985 like the Quillayute River did. Since 1985, the harvest

of wild summer steelhead declined to zero in 2007, while the harvest of hatchery summer runs

has remained relatively high (Figure 25). Mean annual recreational catch from 1978-1985 was

459 steelhead (range = 96 - 711). The last year a wild summer steelhead was reported as

harvested by recreational anglers was 2006, the same as the Quillayute River system, though

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retention of wild summer runs was banned by managers in 1993. Mean recreational harvest

from 1986-2006 was 52 wild summer steelhead (range = 0 - 257) and, interestingly, 204

hatchery summer steelhead (range = 33 - 433). Given the consistent hatchery catch by

recreational anglers over the period of record in the Hoh River, the spike in catch from 1978-

1985 was potentially associated with hatchery steelhead straying from the Quillayute River.

Tribal harvest of summer steelhead has been variable since 1978 and was not

delineated by hatchery and wild fish (Figure 26). The mean annual tribal catch from 1978-2016

was 207 summer steelhead of unknown origin, with a peak of 937 fish and a low of 10. Overall,

the reported catch levels, as with the Quillayute River system, seem high compared to the very

low abundance of wild summer steelhead reported in snorkel surveys in the SF Hoh River (Table

4). This is even more concerning due to the lack of monitoring and escapement and fishery

goals.

3. Queets River

The mean annual harvest rate from 1978-2020 was 35% for Queets River wild winter

steelhead (Table 1), with a range of 10% to 55% per year (Figure 23c). This equates to a harvest

of 2,672 wild winter steelhead per year, with a range of 575 – 6,291, compared to a mean

annual run size of 7,384 steelhead. There could be a slight declining trend in harvest rate over

the period of record (Figure 23c). The ratio of hatchery to wild winter steelhead in the

commercial fishery was 0.7:1 and 4.7:1 in the recreational fishery. (Duda et al. 2018). Since

1976, the mean annual commercial harvest of hatchery steelhead has been 1,472 fish (range

298 – 3,308; SD 766) (Duda et al. 2018).

Recreational catch of wild summer steelhead in the Queets River was highest during the

historical period, and then showed a similar peak in catch as the Quillayute and Hoh Rivers from

1978-1985, followed by a sharp decline in catch thereafter until recreational harvest of wild

summer runs was eliminated in 1993 though some summer steelhead were reported as

harvested from 2001-2004 (Figure 25). Like the Hoh River, the Queets River does not receive

hatchery summer steelhead smolts, and yet, they have been caught each year by anglers since

hatchery fish were first marked (Figure 25). Mean annual recreational catch from 1978-1985

was 201 steelhead (range = 100 - 310), compared to 53 wild summer steelhead (range = 0 - 153)

from 1986-1992 and 2001-2004 and 30 hatchery summer steelhead (range = 8 - 81) from 1986-

2016.

Tribal catch is not delineated by hatchery and wild. Zero catch was reported in 1984 and

only two fish were reported in 1985. Except for these years, relatively low levels of catch have

been reported (Figure 26). Mean annual tribal catch from 1978-2016 was 102 summer

steelhead (range = 2 – 469). Unfortunately, there is almost no data on the abundance of wild

summer steelhead in the Queets River, so it is very difficult to draw inferences about potential

harvest impacts. Nonetheless, wild summer steelhead were formerly quite abundant as we

described earlier in the Historic Abundance section, and it is likely they are now highly

depleted. Consequently, we have grave concern about this population because it is not

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monitored, there are no escapement goals, and there has been no effort to evaluate status and

trends to inform management and conservation.

4. Quinault River

Annual harvest rates of wild winter steelhead from 1978-2020 ranged from 15% to 65%

in the Quinault River (Figure 23d), with a mean of 46% (Table 1). That is the highest mean

harvest rate among the four largest populations in the DPS (Table 1). It equates to 2,832 fish

per year, with a range of 814 – 6,356, compared to a mean annual run size of 5,968 wild winter

steelhead. Since 1977, there has been a mean annual commercial harvest of 7,089 hatchery

winter steelhead (SD = 3,956) with a range of 1,068 – 15,979 fish, and a mean annual

recreational harvest of 692 hatchery winter steelhead (SD = 418) and a range of 61 – 1,656 fish

(Duda et al. 2018). The ratio of hatchery to wild winter steelhead in the commercial fishery was

2.9:1 and in the recreational fishery it was 3.0:1 (Duda et al. 2018).

The pattern and extent of annual recreational catch of summer steelhead in the

Quinault River differs from other watersheds in that it did not experience the peak in catch

during the period when hatchery fish were unmarked and overall, catch remained very low

until some hatchery steelhead started to show up in the catch in the late-1990’s (Figure 25).

Overall, mean annual recreational catch from 1978-1985 was 15 steelhead (range = 8 - 34),

compared to 8 wild summer steelhead (range = 0 - 18) from 1986-1992 and 2001-2002 and 15

hatchery summer steelhead (range = 0 - 44) from 1986-2016.

Tribal catch of summer steelhead from 1976-2017 is higher than all other populations

(Figure 26). The mean annual catch was 753 steelhead, with a range of 191 to 2,345 fish.

However, given the very low snorkel counts of wild adult summer steelhead in the upper EF and

NF Quinault Rivers (see Abundance section), it is unclear whether true wild summer runs are

being harvested at very high rates or whether some of the fish are winter run kelts. It is also

unclear whether hatchery summer runs are contributing. A combination of all three factors

could possibly explain the very high catch levels reported by the tribe.

We delineated catch by month. The highest mean catch was in May, which thereafter

declined through September until catch slightly increased in October (Figure 28). May is an

important month of spawning for Olympic Peninsula steelhead (Busby et al. 1996), so some (or

many) of those fish could be kelts. Kelts are likely being captured in June, considering spawn

timing in other nearby populations (e.g., McMillan et al. 2007). Perhaps some of the fish are

also hatchery summer steelhead from releases in streams further south in Grays Harbor, such

as the Wynoochee River. Whatever the case, it seems imperative to determine the life histories

and source of the fish, because snorkel surveys in the upper watershed suggest the population

is greatly depleted and potentially close to extinction.

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C. Recent failures to Meet Escapement Goals

A report by Cram et al. (2018) evaluated the number of years the harvested wild winter

steelhead populations met their escapement goals from 2004-2013 (Table 3). We updated

those estimates to cover the most recent ten-year period (2011-2020) for all populations and

tributary populations (Table 1). There is an overall problem with many populations failing to

meet their escapement goals, particularly in recent years (Table 1). For example, among the

wild steelhead populations that experience harvest, the Hoh River has only met its escapement

goal of 2,400 fish in six of the past ten years (Table 1) and only 50% of the years dating back to

2003 (Figure 5).

The Queets River population has only met its WDFW and National Park Service

escapement goal of 4,200 fish in 30% of the last ten years (Table 1), and run sizes are now

coming in below the escapement goal (Figure 8a).

The Clearwater River, the major tributary of the Queets, has met its escapement goal of

1,450 fish only 50% of the years in the past decade (Table 1), with recent escapements being

the lowest during the period of record (Figure 8b).

There is no escapement goal for the entire Quinault River population, but its run sizes

have collapsed to the point where it certainly would be below any reasonable estimate for

spawner goals (Figure 7).

The Quillayute River system overall has remained the strongest among the populations,

meeting its escapement goal of 5,900 fish for the entire watershed in nine out of the past ten

years (Table 1) and only missing the goal in two of the past twenty years (Figure 6a). Among its

major tributaries, the Dickey (Figure 6b) and Calawah Rivers (Figure 6d) have met their

escapement goal each of the past ten years (Table 1), and they are the only two tributaries that

do not have a long history of receiving releases of hatchery winter steelhead. In contrast, the

Bogachiel/Quillayute (Figure 6e) and Sol Duc Rivers (Figure 6c) have only met their escapement

goal in 60% and 70% of the last ten years, respectively (Table 1).

As mentioned in the abundance and trends section, nearly all the smaller, independent

populations of wild winter steelhead have failed to meet their escapement goals during much

of – if not all – the past decade (Table 1). Most concerning, some of the populations have not

achieved their escapement goals for extensive periods of time dating back 20 years or more.

D. Demographic, Genetic, and Ecological Risks Caused by Harvest

Harvest poses demographic and genetic risk to wild salmon and steelhead (ISAB 2005;

Hard et al. 2008) and can alter attributes including, but not limited to, size and age at maturity,

migration timing, spatial distribution, and life history diversity, which, in turn, interact to shape

the productivity and resilience of wild salmon and steelhead populations (McElhaney et al.

2000; ISAB 2005; Hard et al. 2008). Given the high exploitation rates on the four largest

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populations of Olympic Peninsula wild winter steelhead, it is possible that chronic, high levels of

harvest have imposed a variety of impacts. Unfortunately, research on such topics is almost

entirely lacking for the DPS, except for evidence of change in run timing of wild winter

steelhead (McMillan et al. 2022).

Cram et al. (2018) noted the risk of fishery selection on run timing since the number of

fishing days per week for treaty fisheries is highest early in the season, when the fishery targets

greater harvest of hatchery adults. Historical recreational fisheries were structured similarly,

with a focus on harvesting early returning hatchery adults that were intermixed with wild

adults. The variable historical and contemporary fishing rates indeed appear to have

contributed to the substantial depletion of early returning wild steelhead (Figure 15). As a

result, run timing is more compressed than it was historically.

The results challenge the findings in Busby et al. (1996), which described hatchery and

wild winter steelhead as being temporally segregated because of differences in run timing. It

also underscores the importance of the Shifting Baseline Syndrome, where managers and

scientists come to accept the current level of abundance and diversity as the norm (Pauly

1995). For example, in the Quileute Tribal language, the month of January is defined as the

“time of steelhead running” and February is defined as the “time of steelhead spawning”

(Frachtenberg 1916). Thus, based on quantitative data and traditional ecological knowledge, it

is clear there has been a large change in run timing that has not been accounted for or

addressed in previous reviews of the Olympic Peninsula DPS (e.g., Busby et al. 1996).

The change in run timing could impact the population in several ways. Per McMillan et

al. (2022):

“Changes in run timing can shorten breeding seasons, reduce phenotypic

diversity, and lower population productivity (Tillotson and Quinn 2018). More

protracted anadromous fish migrations, like those that we estimated to occur

historically for wild Winter Steelhead, can allow fish to temporally stagger the

use of spawning habitat (Gharrett et al. 2013), thereby reducing density-

dependent effects on juvenile survival and increasing local habitat capacity

(Chandler and Bjornn 1988).

Migration timing is also often associated with the spatial structure of breeding

locations in anadromous fish populations (e.g., Everest 1973; Stewart et al. 2002;

Beacham et al. 2012). On the [Olympic Peninsula], McMillan et al. (2007)

observed spatial correlations with spawn timing in the Quillayute River, where

spawning higher in the stream network in smaller stream channels occurred

about a month earlier than in the lowermost mainstem spawning reaches.

Previous observations by Cederholm (1984) also found that earlier returning

Queets River wild Winter Steelhead were more likely to spawn in smaller

tributaries, while mainstem spawning tended to occur several weeks later in the

season. Importantly, Cederholm (1984) observed this occurrence in low

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elevation tributary streams, so the pattern of earlier spawn timing in tributaries

is not necessarily limited to higher elevation headwaters as might be inferred

just from the results of McMillan et al. (2007), but rather it appears to be a

pattern associated with stream size.”

In addition, migration timing underpins a population’s adaptive capacity to keep pace

with shifting climatic conditions, such as changing streamflow and temperature regimes

(Manhard et al. 2017; Reed et al. 2011; Austin et al. 2020). For example, winter steelhead tend

to migrate and spawn earlier in the season in warmer streams in more southerly habitats

(Busby et al. 1996), presumably because the higher streamflows provide better access and

earlier spawning is needed to ensure offspring emerge before the onset of summer baseflows.

Climate change models predict water temperature regimes on the Olympic Peninsula (see

Climate Change section) will become more similar to those in more southerly climates (Wade et

al. 2013). This suggests that early migrating life histories will become increasingly important to

population resilience. However, rebuilding earlier-timed wild winter steelhead migrations will

be challenging without addressing the current fishery structure and hatchery practices (see

Hatchery section below).

Harvest could also be impacting diversity though selection on size and age and selection

against iteroparity. Harvest can alter size and age in salmonids (Hard et al. 2008), and Cram et

al. (2018) reported that scale samples indicated gillnetting was selectively capturing older adult

wild steelhead, while recreational fisheries were capturing more younger fish (Cram et al.

2018). It is unknown how these effects have manifested over the decades, but it is possible that

both fisheries have truncated the overall distribution of size and age in the populations

experiencing high levels of harvest.

Steelhead are iteroparous, and individuals that repeat spawn are significantly more

productive on their second attempt than their first and forgo early reproduction to devote

additional energy to continued survival (Christie et al. 2018). Repeat spawning fish may also

come back larger and with more eggs (Christie et al. 2018; Gayeski, personal observation on

recaptured tagged steelhead in the Sopichnaya River, Kamchatka 2005). In this context, kelts

migrating back to the ocean after spawning their first time are only likely to contribute to

reproduction if they are allowed to spawn a second time. Harvesting those individuals is

analogous to removing the most productive life history from the next generation of spawners.

Kelts can appear in the system beginning as early as late-January, but they are most common

after the peak of spawning from April through May. During that time kelts are harvested in

gillnet fisheries and captured by recreational anglers in catch and release fisheries.

Although the steelhead fishery is closed in May, there have traditionally been

recreational and tribal Spring/Summer Chinook Salmon fisheries in each of the rivers that

begins in April and runs through May and into June. Those fisheries could have significant

impacts on kelts. McMillan (2006) hypothesized a portion of the relatively high numbers of wild

steelhead harvested by Treaty fisheries in May are likely kelts – up to 1,800 fish in May in some

years – despite being considered “summer runs” by co-managers. This seems possible,

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especially in the Quinault River where the catch numbers of steelhead in May are particularly

high (Figure 28). There is also an intensive sport fishery in the Quillayute and Sol Duc Rivers for

hatchery spring/summer Chinook, although it is unclear how many kelts are caught, it could be

substantial because most anglers are using bait. And while sport fisheries for spring Chinook

salmon are now closed on the Hoh and Queets Rivers, historically they were also open to

recreational angling at different periods. Because kelts start to feed after spawning, they are

also prone to being caught by recreational anglers and, considering the low lipid levels and poor

condition of kelts (Penney and Moffitt 2014), any further expenditure of energy – such as being

hooked and caught – could result in delayed mortality.

In addition, as noted by Gayeski et al. 2016, steelhead kelts (if unstressed by post-

spawning capture in river fisheries) have higher survival rates to repeat spawning in the

following one or two years than smolts that must survive several years in the marine

environment in order to become maiden spawners. Thus, repeat spawners provide a

considerable boost to the overall productivity and juvenile capacity of the population than

smolts that survive to spawn for the first time. This is particularly critical to the recovery of

depressed populations.

Ultimately, the sharp declines in iteroparity in the Queets River population since 1980

and low levels of iteroparity in other wild winter steelhead populations (Figure 16) have likely

contributed to the chronic declines over the past twenty years or more. The depleted early

returning wild steelhead combined with potential changes in size and age and reductions in

iteroparity raise serious questions about the sustainability of the fishery model and highlights

the need to incorporate run timing and repeat spawning into steelhead management and

fisheries.

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Figure 28. Proportion of repeat spawning individuals (repeats/total run size) by year for wild

winter steelhead populations in the Quillayute (1978-2021), Hoh (1993-2020), Queets (1980-

2019), and Quinault Rivers (1991-2020).

III. DISEASE AND PREDATION

A. Disease

The M genogroup of the infectious hematopoietic necrosis virus (IHNV), which is

harmful to steelhead, has been previously detected in seven hatchery locations where wild

Olympic Peninsula steelhead occur (Breyta et al. 2013). Between 2007 and 2011, there were

two distinct waves of the virus that occurred in multiple watersheds, including the Hoh, Queets,

Quinault, and Quillayute Rivers (Breyta et al. 2013). Most fish detected with the virus were

hatchery fish, however, wild fish are less commonly sampled (Breyta et al. 2013). The virus was

detected in wild fish in the Hoh and Quinault Rivers (Breyta et al. 2013). According to Breyta et

al. (2013), “it is not clear what the future burden of IHNV in coastal steelhead trout might be.”

B. Predation

There is an increased distribution of predators in the Dickey River, likely because of

warming water temperatures (Smith 2000). It is likely that predation risks will increase in

Olympic Peninsula rivers as summer streamflows decrease and water temperatures increase

(Dalton et al. 2016).

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IV. INADEQUACY OF EXISTING REGULATORY MECHANISMS

A. FEDERAL

1. National Forest Management Act & Northwest Forest Plan

The National Forest Management Act (NFMA) requires the U.S. Forest Service (USFS) to

manage fish and wildlife habitat in the Olympic National Forest to “maintain viable populations

of existing native and desired non-native vertebrate species.” 36 C.F.R. § 219.19. In 1990, the

USFS adopted a Land and Resource Management Plan (“LRMP") for the Olympic National Forest

(USFS 1990). The USFS claimed that implementation of the plan would increase fish production

potential by more than 10% by the end of first 10 years of implementation (USFS 1990). It also

claimed that it would correspond with 1,200,000 additional anadromous smolts and that

anadromous fish production would increase by 25% above then current levels due to decreased

sedimentation and habitat enhancement projects (USFS 1990).

In 1998, the LRMP was amended to include management changes consistent with the

Northwest Forest Plan (1994) (Hoffman 1998). The NFP includes an Aquatic Conservation

Strategy (ACS), which establishes measures intended to restore and maintain ecological

processes of aquatic and riparian habitat (Thomas et al. 1993; Reeves et al. 2006). Among other

things, the ACS requires the USFS to “maintain and restore the sediment regime under which

aquatic ecosystems evolved” (USDA 1994). To date, the ACS has not maintained or restored the

sediment regime under which Olympic Peninsula steelhead evolved, and Petitioners did not

locate evidence to suggest that the LRMP and the improvements made by the ACS have

increased anadromous fish production by 25% or more over 1990 levels.

The ACS also established riparian reserves, which place primary emphasis on protecting

fish habitat. The size of the buffers depends on whether the streams are designated as fish- or

non-fish-bearing. Riparian buffers on fish-bearing streams are two site-potential tree heights (>

200 years old) or 300 ft, whichever is greater. Riparian buffers for non-fish bearing streams are

one site-potential tree height or 100 ft, whichever is greater (Wilhere and Quinn 2018). The

buffers are intended to help ensure that species meet certain viability standards (Wilhere and

Quinn 2018). When developing the ACS, the USFS assumed that if a species has at least an 80%

likelihood of a stable, well-distributed population over 100 years, it is viable (Wilhere and Quinn

2018).

The USFS is also guided by an Olympic National Forest Strategic Plan, which “integrates

aquatics, wildlife, silviculture, and fire, helping to identify priority areas for management

activities such as habitat restoration, road decommissioning, forest thinning, and fuel reduction

treatments” (Halofsky et al. 2011). It also has a Road Management Strategy, which it

developed, in part, to meet ACS standards (Halofsky et al. 2011). The road strategy sets

priorities for road maintenance, upgrading, and decommissioning based on several factors,

including aquatic risk and high-value watershed goals (Halofsky et al. 2011).

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The NFP established the Aquatic and Riparian Ecosystem Monitoring Program (AREMP)

to monitor whether implementation of the ACS is improving watershed conditions (Gaines et al.

2022). Unfortunately, the program has been hindered by insufficient funding and changes to

monitoring protocols, which make it challenging to measure the impacts of the ACS on aquatic

resources (Gaines et al. 2022).

Roads

The ACS has not been effective in reducing road density or improving other road-related

factors that affect fish (Frissell et al. 2014). According to Frissell et al. “[t]he magnitude of

existing road impacts on watersheds and streams in the [NFP] may equal or exceed the effect of

all other activities combined” (Frissell 2014).

Roads present major risks to fish and aquatic resources in the Olympic National Forest.

In 2020, the USFWS stated that “forest roads in the Olympic National Forest have been a

chronic source of sediments for decades” (USFWS 2020). Although the USFS has

decommissioned 435 miles of roads in the Olympic National Forest since 1990 (Halofsky et al.

2011), hundreds of road miles still present significant risks to aquatic resources (USFS 2015).

Fifty-one percent (1,032 miles) of all roads in the Olympic National Forest present high aquatic

risks (USFS 2015). Thirty-three percent (651 miles) of the roads in the national forest are rated

as presenting medium aquatic risks (USFS 2015). As a result, only 17% (338 miles) of roads in

Olympic National Forest present low aquatic risk (USFS 2015). Nearly one-third of the Olympic

National Forest’s roads are proposed for decommissioning (Halofsky et al 2011).

In addition to falling behind on road decommissioning, the Olympic National Forest has

not received the funding necessary to bring all of the roads up to current standards (USFWS

2020). Additionally, some roads that have been previously maintained need additional work

(USFWS 2020).

Riparian Revegetation

Although there has been some revegetation in riparian corridors in the Olympic National

Forest, many corridors continue to have few conifers because of historic logging practices

(Halofsky et al. 2011). Reestablishing conifers in these corridors would help restore sources of

large woody debris (Halofsky et al. 2011). However, these projects require costly, long-term

commitments and, therefore, they have not been a high priority for forest managers (Halofsky

et al. 2011).

Additionally, treatments in riparian areas can disturb soils and decrease effectiveness in

retaining sediment and nutrients (Frissell et al. 2014). Thinning in riparian areas can diminish

summer flows because of increased water demand by regrowth of vegetation (Frissell et al.

2014).

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Climate Change

It is unclear if the USFS has reviewed the ACS to determine whether any science-based

changes to the strategy are necessary in response to climate change (Frissell 2014). Current

forest plans are not flexible enough to adapt to climate change challenges (Gaines et al. 2022).

To strengthen watershed resiliency to climate change, the ACS could be revised to include

additional steps to reduce non-climatic stressors (Gaines et al. 2022). For example, reducing the

impact of roads would foster greater ecosystem resiliency to climate change (Gaines et al.

2022). To date, the ACS has not been updated to specify steps to reduce non-climate-related

stressors on watersheds as a method to increase resiliency to climate change.

2. Endangered Species Act

Several populations of Olympic Peninsula steelhead occur in areas where ESA-related

conservation and recovery efforts are underway for other species, including bull trout (USFWS

2020). As a result, Olympic Peninsula steelhead may benefit from the habitat protections

afforded to bull trout (USFWS 2020). The status of Olympic Peninsula steelhead, however,

demonstrates that these benefits are not sufficient to halt or reverse their decline. Olympic

Peninsula steelhead need their own protection under the ESA.

a. Critical Habitat for Bull Trout

In 2005, the USFWS designated a series of waterbodies as critical habitat for coastal

populations of bull trout. 70 Fed. Reg. 56212, 56304-56306 (Sept. 26, 2005). In 2010, the

USFWS updated the designations for 32 critical habitat units (CHUs), including the Olympic

Peninsula Unit (75 Fed. Reg. 63898 (Oct. 18, 2010)). The USFWS designated 121 waterbodies as

critical habitat in the Olympic Peninsula Unit, excluding certain geographic areas covered by the

Washington State Forest Practices Plan (HCP) or covered by tribal plans. Id., at 63968-69369.

Critical habitat for bull trout overlaps with multiple areas where Olympic Peninsula

steelhead occur (75 Fed. Reg. 63875-63978) Specifically, Olympic Peninsula steelhead occur in

the following critical habitat for bull trout: Clearwater River, Copalis River, Hoh River, Kalaloch

Creek, Moclips River, Mosquito Creek, Quinault River, Queets River, Raft River, Salmon River,

South Fork Hoh River, and Tshletshy Creek. As a result, Olympic Peninsula steelhead in these

rivers and streams may benefit from critical habitat protection afforded to bull trout.

b. Biological Opinion on Forest Management Activities on the Olympic National

Forest

On June 24, 2020, the USFWS issued a biological opinion addressing Forest Management

Activities on the Olympic National Forest (USFWS 2020). The biological opinion evaluates the

potential effects of management actions on ESA-listed bull trout, Northern Spotted Owl, and

Marbled Murrelet. It requires certain conservation measures for projects that occur within bull

trout core area watersheds or in designated critical habitat for bull trout. Several populations of

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Olympic Peninsula steelhead occur in bull trout core area watersheds and critical habitat and

therefore, may indirectly benefit from these conservation measures. These measures include

the following: steps to prevent erosion and enable large wood recruitment; limits on

commercial thinning and road maintenance activities; prohibitions on new culvert installations

or culvert replacements; no-cut buffers restrictions; and road standards (USFWS 2020).

However, despite these conservation measures, the USFWS anticipates that adverse

impacts will continue to occur in bull trout habitat (USFWS 2020). Olympic Peninsula steelhead

that occur in the same areas as bull trout will likely incur some of these impacts as well. For

example, new and temporary road construction, existing road reconstruction, road repairs, log

hauling, road grading/blading, and drainage maintenance is anticipated to cause indirect

adverse effects to bull trout. These effects include increased fine sedimentation, altered

watershed hydrology, reduced water quality, and increased substrate embeddedness (USFWS

2020). The USFWS expects that the Queets and Quinault Core Areas, which overlap with the

habitat of Olympic Peninsula steelhead, will continue to be depressed by poor water quality

(e.g., up to 408 tons of sediment per year in the Queets River) increased substrate

embeddedness, and altered flow regimes caused by logging operations (USFWS 2020).

c. Habitat Conservation Plans

Olympic Peninsula steelhead may benefit from the Washington Department of Natural

Resources Trust Lands Habitat Conservation Plan (DNR HCP) and the Forest Practices Habitat

Conservation Plan (FPHCP). However, these plans have not been fully implemented and there

are conflicting data on whether forest practices are improving aquatic resources. These plans

are discussed in the state forest management section.

3. Clean Water Act

a. Overview

The Clean Water Act (CWA), 33 U.S.C. §§ 1251-1387, is the principal federal law

regulating water quality in U.S. surface waters. The CWA seeks “to restore and maintain the

chemical, physical, and biological integrity of the Nation’s waters.” 33 U.S.C. § 1251(a). The Act

establishes a goal to eliminate all discharges of pollutants into navigable waters by 1985. 33

U.S.C. §§ 1251(a)(1). It also sets an interim goal of achieving, wherever attainable, “water

quality which provides for the protection and propagation of fish, shellfish, and wildlife and

provides for recreation in and on the water” by July 1, 1983. Id., at 1251(a)(2).

The U.S. Environmental Protection Agency (EPA) is the primary federal agency

responsible for administering the CWA. EPA implements water pollution control programs,

establishes pollution control technology standards for industries, and reviews water quality

standards set by states. The U.S. Army Corps of Engineers (Corps) regulates the discharge of

certain materials (e.g., fill) in the waters of the United States.

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The CWA requires states to establish water quality standards to protect public health

and welfare, enhance water quality, and serve the purposes of the Act. States follow a two-step

process when setting water quality standards. First, they establish “designated uses” for

individual waterbodies, such as the protection and propagation of fish. Next, they set allowable

levels of pollutants to protect those uses. 40 C.F.R. §§ 131.10, 131.11. Washington has EPA-

approved water quality standards; however, it is not meeting them in many rivers and streams

where Olympic Peninsula steelhead occur (WDOE 2016; NIFWC 2020).

The CWA requires states to list “impaired waters,” which include segments of

waterbodies that do not meet water quality standards. The states must develop Total

Maximum Daily Loads (TMDLs) that set the maximum amounts of pollutants that may enter

impaired waters without violating water quality standards. The state may distribute those

amounts or “loads” to various sources of point and nonpoint sources of pollution. In effect, a

TMDL operates as a pollution budget for an impaired waterbody. Nonpoint source pollution

drives water quality exceedances in the rivers and streams where Olympic Peninsula steelhead

occur.

Washington has not developed, and EPA has not approved, any TMDLs for any

waterbodies where Olympic Peninsula steelhead occur.

b. Nonpoint Source Pollution

Nonpoint source pollution is the “leading cause of water pollution across the nation and

in Washington.” Wash. Dep’t of Ecology, Publ’n No. 040-010-009, Enforcement Report on Policy

and Trends, 49 (2004). Agriculture causes most nonpoint source pollution; however, that

industry is exempt from certain CWA regulations.

A limited number of logging-related activities are subject to the CWA. 40 C.F.R. §

1227.27. The EPA’s silvicultural rule defines silvicultural point sources to include “any

discernable, confined and discrete conveyance related to *** log sorting *** or log storage

facilities which are operated in connection with silvicultural activities and from which pollutants

are discharged in waters of the United States.” Id. § 1227.27(b)(1). The rule defines log sorting

and log storage facilities to mean “facilities whose discharge results from the holding of

unprocessed wood, for example, logs or roundwood with bark or after removal of bark held in

self-contained bodies of water (mill ponds or log ponds) or stored on land where water is

applied intentionally on the logs (wet decking). Id. § 1227.27(b)(3).

The silviculture rule does not apply to major sources of water pollution caused by

logging operations. Among other activities, the rule exempts harvesting operations, surface

drainage, and road construction and maintenance from which there is natural runoff. 40 C.F.R.

at § 1227(b)(1). In 2013, the U.S. Supreme Court upheld the EPA’s decision not to regulate

stormwater runoff from logging roads. Nw. Envtl. Def. Ctr. v. Decker, 133 S. Ct 1326, 1338

(2013). The following year, Congress amended the CWA to effectively prohibit the EPA from

requiring NPDES permits for discharges resulting from several silviculture-related activities,

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including surface drainage, road construction, and road maintenance. 33 U.S.C. § 1342(l).

Therefore, the Clean Water Act fails to adequately protect Olympic Peninsula steelhead.

Many river and stream miles occupied by Olympic Peninsula steelhead will experience

water quality impacts caused by logging operations for the foreseeable future (McHenry et al.

1996). The synergistic effect of timber harvesting and heavier rainfall caused by climate change

will impair water quality in rivers and streams where Olympic Peninsula steelhead occur.

4. National Environmental Policy Act

The National Environmental Policy Act (NEPA), 42 U.S.C. §§4321 et seq., requires federal

agencies to identify and evaluate the impacts of “major Federal actions significantly affecting

the quality of the human environment.” Thus, the U.S. Forest Service, U.S. National Park

Service, Bureau of Indian Affairs, and other federal agencies whose major actions may impact

Olympic Peninsula steelhead are subject to NEPA. Major federal actions include those actions

“subject to federal control and responsibility.” 40 C.F.R. §1508.1(q). To determine whether the

impacts of a proposed action are “significant” federal agencies must assess the “potentially

affected environment and degree of the effects.” Id. at § 1501.3(b). Federal agencies must

consider the short-term and long-term effects, beneficial and adverse effects, effects on public

health and safety, and effects that would violate laws that protect the environment. Id.

Federal agencies must prepare environmental impact statements (EISs) for proposed

actions with significant impacts. 42 U.S.C. § 4332. An EIS must assess the following topics: (1)

the environmental impacts of the proposal; (2) unavoidable adverse effects; (3) alternatives to

the proposed action; (4) the relationship between the short-term uses of the environment and

maintenance of long-term productivity; and (5) any irretrievable commitment of resources

involved if the proposed action is implemented. Id. An EIS must also provide a detailed

statement regarding the scope of its assessment, which must cover the following subjects: (1)

connected or similar actions; (2) a reasonable number of alternatives to the proposed action,

including no action alternatives and other reasonable alternatives, as well as mitigation

measures; and (3) effects. 40 C.F.R. §§ 1501.9, 1502.14. Federal agencies must release draft

EISs for comments from other agencies and the public. 42 U.S.C. § 4332(2)(C).

After undertaking all these analyses and procedures, NEPA does not require federal

agencies to select the alternatives with the least impacts on the environment; it simply requires

federal agencies to take a “hard look” at them. Robertson v. Methow Valley Citizens Council,

490 U.S. 332, 350 (1989). Therefore, the USFS and other federal agencies whose actions may

impact Olympic Peninsula steelhead are not required to choose the alternatives that best

protects the species. If they follow the procedural requirements of NEPA, agencies may select

any alternatives, including those that harm Olympic Peninsula steelhead. Therefore, NEPA is not

sufficient to protect Olympic Peninsula steelhead. The DPS requires protection under the ESA.

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5. Governmental Failure to Adequately Address Climate Change

Climate change impacts in freshwater and marine habitats pose significant threats to

the survival of Pacific salmonids. A critical feature of the contemporary and future threats of

climate change is the failure of the US government to address these threats through the

adoption of “green” energy policies and actions to drastically reduce reliance on fossil fuel

energy generation. These actions are necessary to begin a rapid reduction in the amounts of

greenhouse gases that are annually released into the atmosphere.

It is widely recognized in the peer-reviewed scientific literature that the concentration

of atmospheric carbon dioxide (CO2) is the primary driver of climate warming. The

concentration of CO2 in the atmosphere is currently 412.5 ppm [NOAA Climate.gov

https://www.climate.gov/news-features/understanding-climate/climate-change-atmospheric-

carbon-dioxide (accessed June 1, 2022)]. This level is the major contributor to the current global

energy imbalance. In 2020, von Schuckmann et al. estimated the current global energy

imbalance to be 0.87 +/0.12 watts-per-square meter (W m

-2

) of the earth’s surface. “The

amount of CO2 in the atmosphere would need to be reduced from 410 to 353 ppm to increase

heat radiation to space by 0.87Wm

-2

, bringing Earth back towards energy balance” (von

Schuckmann et al, 2020, p. 2014).

Considering both current policy and the fragmented political climate in the US today,

there is a high probability that such a timely reduction in atmospheric CO2 will not occur in time

to avoid significant and continued (and perhaps even acceleration of) global warming within the

next few decades. This further emphasizes the realism of making climate change predictions

based on the IPCC’s “business-as-usual”, RCP8.5 scenario, which Cheung and Frölicher (2020)

note leads to global atmospheric surface warming … of 3.2 C by 2081 – 2100 relative to

preindustrial” (Cheung and Frölicher (2020). Further, as explained by von Schuckmann et al.

(2020), the current large global energy imbalance coupled to the imbalance that has existed for

the past several decades means that considerable warming of the atmosphere as a result of the

legacy of heat already stored in marine waters is already locked in for several decades. This

would still be true even if the current imbalance were to disappear tomorrow. Therefore, the

current failure to adopt green energy policies that would rapidly reduce the global energy

imbalance should be understood as a fundamental component of the failure of existing

regulatory mechanisms to protect salmon and steelhead and the many components of

salmonid ecosystems on which they depend.

6. National Park Service Organic Act and Other Regulatory Mechanisms

The habitat in Olympic National Park has been protected for over 100 years, and it is in

relatively pristine condition (Halofsky et al. 2011). In 1988, Congress designated 95 percent of

the park as the Olympic Wilderness. The Olympic Wilderness, which was renamed the Daniel J.

Evans Wilderness, covers 1,370 square-miles, including 48 miles on the Washington coastline.

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a. National Park Service Organic Act

Portions of several Olympic Peninsula steelhead populations spawn and rear inside the

Olympic National Park, which provides relatively pristine aquatic habitat conditions compared

to areas outside of the park. The park’s enabling legislation and the National Park Service

Organic Act of 1916 (Organic Act) protect these resources. Specifically, the Organic Act requires

the NPS “to conserve the scenery and the natural and historic objects and wildlife therein and

to provide for the enjoyment of the same in such manner and by such means as will leave them

unimpaired for the enjoyment of future generations.” 16 U.S.C. § 1.

b. Fishing Regulations

Olympic Peninsula National Park has its own fisheries management program. The

program has three goals: (1) “manage aquatic resources as an important part of the park

ecosystem;” (2) “preserve and restore native fishes and their habitats;” and (3) “provide

recreational fishing opportunities for the enjoyment of park visitors, consistent with the first

two objectives.” Except for saltwater areas within park, a Washington State fishing license is not

required to fish in Olympic National Park. 36 C.F.R. § 2.3(b). However, a Washington State catch

record card is required to fish for steelhead.

All wild steelhead must be released in Olympic National Park. Except during limited

periods in select areas, bait fishing is generally prohibited in the park in order to reduce catch

and release mortality on steelhead. Studies show that catch-and-release fishing mortality is less

than 2% when fish are caught on flies and less than 5-7% overall with most gear types, except

for bait, which can have much higher rates of mortality – for both juvenile and adult steelhead

– and when fish are caught in elevated water temperatures (Warner and Johnson 1978; Hooton

1987; Pauley and Thomas 1993; Schisler and Bergersen 1996; Taylor and Barnhart 1997;

Hooton 2001; Nelson et al. 2005; Twardek et al. 2018). However, as mentioned earlier, the

regulations do not account for potential sublethal impacts that can reduce fitness of individuals

that survive catch and release (up to ~ 20-25% reduction in fitness: Richard et al. 2013;

Bouchard et al. 2021; Papatheodoulou et al. 2021).

The National Park Service has taken emergency actions to protect vulnerable steelhead.

For example, in February 2021, the NPS closed recreational fishing for steelhead on the Queets

River inside Olympic National Park in response to record low returns (LaBossiere 2021).

c. Roads & Structures

Historically, maintenance and repair of Olympic National Park roads that are adjacent to

the Sol Duc, Hoh, Queets, and Quinault Rivers have caused major impacts on fish and aquatic

life (Halofsky et al. 2011). Today, the National Park Service takes steps to reduce these impacts.

For example, when feasible, the National Park Service relocates roads and other facilities from

floodplains to other areas (Halofsky et al. 2011). It also limits construction of new facilities

within floodplains to protect fish habitat (Halofsky et al. 2011).

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B. STATE

1. Washington Department of Fish and Wildlife

WDFW management of Olympic Peninsula steelhead is guided by the Statewide

Steelhead Management Plan, Hatchery and Fishery Reform Policy (C-3624), and harvest

management plans with the Quinault, Quileute, Hoh, and Makah Tribes (WDFW 2008, WDFW

2021, Hoh Tribe and WDFW 2020; Quileute Tribe 2020; Quinault Dept. of Fisheries 2021). The

continued decline of Olympic Peninsula steelhead with these plans in place shows they are

insufficient to adequately manage these populations. Further, the Washington Fish and Wildlife

Commission recently replaced its previous hatchery policy, C-3619, with its new hatchery

policy, C-3624, which further weakens the standards governing the allowable proportion of

hatchery-origin spawners (pHOS) on spawning grounds. The adoption of C-3624 represents a

step backwards in addressing the harms posed by hatcheries.

a. Statewide Steelhead Management Plan (2008)

The 2008 Statewide Steelhead Management Plan describes strategies intended to help

restore and maintain the abundance, distribution, diversity and long-term productivity of the

state’s wild steelhead and their habitats to assure healthy stocks (WDFW 2008). To achieve this

goal, the plan establishes a series of policies to guide natural production, habitat protection and

restoration, fishery management, artificial production, regulatory compliance, monitoring and

evaluation, adaptive management, research, and education (WDFW 2008).

The plan provides that “[s]teelhead management shall place the highest priority on the

protection of wild steelhead stocks to maintain and restore stocks to healthy levels.” (WDFW

2008). To achieve this goal, in part, the policy calls for providing sufficient wild steelhead

spawners, which requires setting escapement objectives (WDFW 2008). The plan acknowledges

that setting informed escapement objectives requires an understanding of stock population

dynamics, habitat conditions, and stock status (WDFW 2008). WDFW maintains a list, which

was last updated in 2002, that describes the status of each stock as “unknown,” “depressed,”

“critical,” or “healthy.” If a stock’s status is unknown, WDFW is supposed to “apply a

precautionary strategy by implementing low risk fishery and hatchery management regimes.”

(WDFW 2008). If the status is “depressed,” “critical,” or ESA-listed, WDFW should promote a

trend of increasing wild fish numbers. (WDFW 2008). If the status is “healthy,” WDFW should

maintain wild steelhead escapement objectives at or above MSH levels (WDFW 2008).

Regardless of whether the nearly two-decades-old “healthy” status designations still

apply to certain Olympic Peninsula steelhead stocks - which they do not (e.g., McMillan et al.

2022) - recent (2010-2021) escapement numbers indicate that WDFW is not maintaining

escapement levels at or above MSH levels. For example, WDFW did not maintain Queets River

wild winter steelhead escapement at or above MSH levels during the last eight out of nine years

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and have often failed to do so in the Hoh River (WDFW 2022a). Further, they do not monitor or

manage wild summer steelhead.

Additionally, the Statewide Steelhead Management plan requires WDFW fish and

wildlife managers to establish a network of wild stock gene banks across the state. These gene

banks are intended to be located where steelhead were “largely protected from the effects of

hatchery programs”. This policy came about from a 2004 review of Washington hatchery

facilities by the Hatchery Scientific Research Group, recognizing the need to protect populations

from the threat of hatchery effects (HSRG, 2004).

The plan calls for establishing one wild stock gene bank for each major population group

(“MPG”) within a larger conservation unit, such as a DPS (WDFW 2008). The plan defines

“MPG” to mean:

“[a] group of populations within a larger conservation unit such as a DPS or ESU

that share genetic, life-history, or ecological characteristics that are sufficiently

distinct from those of other groups of populations to make conservation or

recovery of the group essential for the conservation or recovery of the larger

conservation unit.”

(WDFW 2008). This definition implies that more than one MPG should occur within a

conservation unit, as an MPG is a subcomponent of a DPS or similar conservation unit.

Fourteen years after the plan was adopted, there is still only one wild steelhead gene

bank on the Olympic Peninsula: the Sol Duc Wild Steelhead Gene Bank, which WDFW

established in 2013. As a result, since the Snider Creek hatchery was closed in accordance with

the Sol Duc gene bank designation. Although this has benefitted Sol Duc steelhead, WDFW

should be establishing additional gene banks on the Olympic Peninsula.

Considering the differences between the WRIA 19-21 populations, WDFW should have

established more gene banks. There are characteristics that make these groups of populations

sufficiently distinct from the group of populations represented by the Sol Duc gene bank

designation. For example, WRIA 19 populations ecologically differ from WRIA 20 populations,

which includes the Sol Duc population. WRIA 19 populations occur in rain-dominant basins that

discharge into the Strait of Juan de Fuca. WRIA 20 populations occur rain-dominant basins –

several of which are glacially influenced – that discharge into the Pacific Ocean. The WRIA 21

populations also have characteristics that may make them sufficiently distinct from WRIA 20

populations. Despite these differences, WDFW has not announced plans to designate additional

gene banks for any other major population groups within the Olympic Peninsula DPS. Although

Cram et al. (2018) recommended establishing additional gene banks in other DPSs, it did not

recommend establishing more gene banks within the Olympic Peninsula DPS. Based on Cram et

al. (2018), it does not appear that WDFW intends to establish additional gene banks for the

Olympic Peninsula steelhead DPS.

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b. Hatchery and Harvest Reform Policy

In the years following NMFS’s 1996 status assessment of Olympic Peninsula steelhead

(Busby et al. 1996), WDFW developed two hatchery reform policies. The original Hatchery and

Fishery Reform Policy (C-3619) included several guidelines that, had they been implemented,

could have reduced hatchery threats to Olympic Peninsula steelhead. Recently, WDFW adopted

a new policy that rolls back these guidelines. As a result, WDFW’s hatchery policies have failed

to protect Olympic Peninsula steelhead.

On November 6, 2009, the Washington Fish & Wildlife Commission (WFWC) adopted a

Hatchery and Fishery Reform Policy that focused on the “scientific and systematic redesign of

hatchery programs to recover wild salmon and steelhead and support sustainable fisheries.”

(Policy C-3619). The policy directed WDFW to follow eleven policy guidelines (Policy C-3619).

The first three guidelines were the following instructions:

1. Use the principles, standards, and recommendations of the HSRG to

guide WDFW’s hatcheries, and to enable adaptive management based on

a structured monitoring, evaluation, and research program.

2. Prioritize improved broodstock management to reduce the genetic and

ecological impacts of hatchery fish.

3. Develop watershed-specific action plans that implement hatchery reform

as part of a comprehensive strategy for meeting conservation and

harvest goals. (C-3619).

The 2009 policy was reviewed through the State Environmental Policy Act and within

that review committed WDFW to follow a “phased approach” where subsequent reform efforts

would be developed and implemented for each hatchery facility, known as Hatchery Action

Implementation Plans (WDFW, 2009). To the best of our knowledge these plans were never

developed, submitted for state environmental review, or implemented.

In 2018, the Commission suspended these three guidelines for salmon species and

directed WDFW to review the hatchery and fishery reform policy and the department’s

performance of its strategies. Subsequently, WDFW produced two reports: (1) A Review of

Hatchery Reform Science in Washington State (Anderson et al. 2020), including independent

review by the Washington Academy of Natural Sciences; and, (2) WDFW Hatchery and Fishery

Reform Policy Implementation Assessment: Final Report, 2009-2019 (Murdoch and Marston

2020).

i. Review of Hatchery Reform Science in Washington State (Anderson et al. 2020)

As directed by the FWC through the C-3619 review process, WDFW and the Washington

State Academy of Sciences conducted a Review of Hatchery Reform Science in Washington

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State to identify advances in hatchery reform science that have occurred since C-3619 was

adopted in 2009. This thorough WDFW-produced and independently reviewed report provides

the following key conclusions:

1. “The HSRG principles of reducing pHOS and increasing pNOB to achieve fitness

gains in wild populations are well-founded and should be fundamental goals in

any hatchery reform management action.

2. Excessive hatchery program size requires more careful scrutiny and scientific

justification because it affects virtually every aspect of hatchery risks.

3. Hatcheries have potential for large magnitude ecological impacts on natural

populations that are not well understood, not typically evaluated and not

measured.

4. Hatchery risks include fishery risks, ecological risks and genetic risks. Fisheries

targeting abundant hatchery runs can unintentionally increase mortality of co-

mingled natural populations.

5. Research on ecological [HxW] interactions lags far behind the attention devoted

to genetic risks of hatcheries. Importantly, research suggests the potential for

ecological interactions in marine environments shared between multiple

hatchery and natural populations, yet very little is known about the likelihood or

magnitude of population-scale ecological impacts of hatcheries.”

6. Studies comparing the number of offspring produced by hatchery-origin fish and

natural-origin fish when both groups spawn in the wild (relative reproductive

success, RRS) have demonstrated a general pattern of lower reproductive

success of hatchery-origin fish.

7. In WDFW’s hatchery system, a focus on efficiency and maximizing abundance

prevents widespread implementation of risk reduction measures.

8. We recommend a more rigorous, consistent and intentional evaluation of

cumulative hatchery effects across multiple hatchery programs operating within

a geographic region.”

9. WDFW invests considerable effort into population monitoring, yet this

information does not often achieve its potential as a hatchery evaluation tool

because analysis, reporting, and synthesis are typically underfunded.

Furthermore, for many hatchery programs, the absence of a clear framework for

application of monitoring data in decision making precludes clearly articulated

risk tolerance thresholds.”

(Anderson et al. 2020).

ii. WDFW Hatchery and Fishery Reform Policy Implementation Assessment (Murdoch

and Marston 2020)

As directed by the WFWC through the C-3619 review process, WDFW also conducted a

Hatchery and Fishery Reform Policy Implementation Assessment (2020) to evaluate the policy’s

effectiveness at protecting wild salmon and steelhead populations (Murdoch and Marston

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2020). However, for reasons described in the report beginning on page 3, WDFW found that the

monitoring data necessary to answer that question were unavailable or inadequate (Murdoch

and Marston 2020). Consequently, WDFW re-focused the assessment to evaluate whether and

to what extent the agency had implemented the fishery and hatchery reform actions mandated

in the 2009 policy C-3619. Among the many findings in this report, the following WDFW

conclusions are particularly relevant to this petition: A lack of funding was a common reason

that prevented implementation of some guidelines; a lack of comprehensive statewide

monitoring and evaluation program are areas of special concern; and defining program success

and collecting and analyzing data to adaptively manage our programs are critical missing

components.

WDFW’s review also found that the department had made little to no progress

implementing seven HSRG recommendations for the Bogachiel Hatchery summer and winter

steelhead programs, as required under the first guideline of the 2009 Hatchery and Fishery

Reform Policy (Murdoch and Marston 2020). Those recommendations included steps intended

to set clear goals for conservation, ensure hatcheries are meeting management goals, and

minimize risks to natural populations (Murdoch and Marston 2020). As discussed in the

hatchery section of this petition, the winter and summer steelhead programs at the Bogachiel

Hatchery did not meet pHOS goals in 2009 (WDFW 2022b). Data for other years are not

available on WDFW’s online database (WDFW 2022b).

On April 9, 2021, the Commission adopted a revised Hatchery Policy (C-3624),

superseding the previous policy (C-3619) (WFWC 2021). The revised policy abandons

commitments to follow HSRG guidelines, and it sets forth new direction for WDFW to follow

when managing salmon and steelhead hatcheries. Among other things, the policy directs

WDFW to prepare Hatchery Management Plans for every WDFW salmon and steelhead

hatchery program (WFWC 2021). These HMPs must be based on the best available science

regarding the risks of hatchery production on wild salmon and steelhead. However, the HMP

provisions must “*** reflect a balance between minimizing genetic and ecological risks to

coincident wild populations and providing for the ecological and societal benefits of hatchery

propagated salmon and steelhead ***.” (WFWC 2021).

The Washington Fish and Wildlife Commission replaced C-3619 with C-3624 without

State Environmental Policy Act review, which is the subject of ongoing litigation. Wild Fish

Conservancy et al. v. Washington Dep’t. of Fish & Wildlife et al., King County Superior Court

Docket No. 21-2-13546-0 SEA.

Even after adoption, the new C-3624 hatchery policy is behind schedule on creating the

evaluation protocols for the yet-to-be-developed hatchery management plans. Despite the

known, ongoing harms that the steelhead hatchery programs are causing to the Olympic

Peninsula Steelhead DPS, no changes to relevant hatchery management plans have been

proposed. Additionally, previous phased state environmental reviews are being abandoned,

starting the decade-long process all over again. It is unclear if, or when, WDFW will release a

HMP for its Bogachiel steelhead hatchery operations.

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c. Harvest Management Plans

The Quinault, Hoh, Quileute, and Makah Tribes have treaty rights to receive shares of

salmon and steelhead as determined in U.S. v. Washington, 384 F. Supp. 312, aff’d 520 F.2d 676

Cir. 1975), cert. denied 423 U.S. 1086. Each year, WDFW and the tribes prepare harvest

management plans that set guidelines for tribal fisheries, seek to promote wild steelhead

conservation, establish monitoring protocols (e.g., sample catches for size, sex, and age), and

commit the parties to working together to evaluate fishery mortalities that occur in tribal and

non-tribal fisheries (e.g., catch and release mortality, net drop out, and marine mammal

predation) (Hoh Tribe and WDFW 2020; Quileute Tribe 2020; Quinault Dept. of Fisheries 2021).

Despite the monitoring protocols in harvest management plans, it is unclear how many

steelhead are harvested each year in tribal fisheries, as indicated in Cram et al. (2018).

“Methods used to estimate treaty harvest and non-retention mortality are not

well-documented and do not currently contain estimates of uncertainty. This is a

problem particularly in areas where harvest rates are relatively high (e.g.,

Olympic Peninsula, Grays Harbor, upstream of Bonneville Dam on the Columbia

River). Work with tribal co-managers is needed to document methods used to

estimate treaty harvest, to test assumptions related to estimations, and to

report estimates of uncertainty.”

(Cram et al. 2018). Cram et al. (2018) did not explain why previous work with tribal co-

managers has failed to fill this data gap.

d. Monitoring

WDFW does not sufficiently monitor Olympic Peninsula winter steelhead, and it does

not monitor Olympic Peninsula summer steelhead at all. The monitoring recommendations in

Cram et al. (2018) acknowledge that there are significant data gaps. For example, Cram et al.

(2018) recommends that WDFW “initiate” monitoring of wild summer-run steelhead in the

Olympic Peninsula DPS. Cram et al. (2018) also recommends initiating robust population-scale

monitoring in one of more Olympic Peninsula steelhead populations and expanding life cycle

monitoring of the DPS.

Likewise, WDFW Hatchery and Fishery Reform Policy Implementation Assessment

(Murdoch and Marston 2020) found that a lack of comprehensive statewide monitoring and

evaluation program are areas of special concern. Defining program success and collecting and

analyzing data to adaptively manage WDFW hatchery programs are critical missing components

despite previous WDFW plans that required this data to be collected.

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2. Washington Forest Practices Act

The Washington Forest Practices Act (FPA) regulates timber harvest on state and private

lands. Wash. Rev. Code §§ 76.09.010-.935. Among other goals, the FPA seeks to “recognize

both the public and private interest in the profitable growing and harvesting of timber,” and

“provide for regulation of forest practices so as to avoid unnecessary duplication in such rules,”

and “achieve compliance with all applicable requirements of federal and state law with respect

to nonpoint sources of water pollution from forest practices.” Wash. Rev. Code §

76.09.010(2)(c), (e), (g). The Act established a Forest Practices Board charged with adopting

forest practice regulations, and it established a permit system operated by DNR that covers

certain forest practices. Id. §§ 76.090.030, 76.09.050, 76.09.020(7).

a. The State Trust Lands Habitat Conservation Plan

WDNR manages state trust forestlands under the State Trust Lands Habitat

Conservation Plan (the “DNR HCP”) (WDNR 1997). The OESF is one of nine planning units under

the DNR HCP (WDNR 1997). The DNR HCP includes a Riparian Forest Restoration Strategy that

aims to restore salmonid habitat (WDNR 1997). Under that strategy, riparian buffers differ in

width based on particular stream needs and disturbance history (WDNR 2016). Generally, the

average riparian buffers, which are measured horizontally from the outer edge of the 100-year

floodplain, are 150 feet on type 1 and 2 streams and 100 feet on type 3 and 4 streams (WDRN

2016). These buffers are intended to minimize disturbance to unstable banks and adjacent

hillslopes and maintain key biological and physical functions. This approach is not designed to

achieve a desired future condition for salmonid habitat, but rather to “maintain or aid

restoration of riparian functions important to salmonid habitat” (WNDR 2016). In addition to

riparian buffers, the HCP includes other habitat protections, including road building

requirements and wetland protections (WDNR 2016).

Improved management standards have not increased large woody debris in streams in

the OESF. Martens et al. (2019) found that large woody debris was either stable at reduced

levels or declining, and that using passive restoration alone is unlikely to increase salmonid

productivity.

“Our results add to the current scientific literature that has found passive

restoration of salmonid habitat in the Pacific Northwest is a slow process, which

could take an additional 12–70 years for riparian forests and over 50 years for

instream wood to accumulate (McHenry et al. 1998; Connolly and Hall 1999;

Kaylor et al. 2017).”

(Marten et al. 2019).

The most recent monitoring report for the OESF also found that the majority of LWD is

in decay, and that historic logging practices continue to interrupt the supply of new LWD to

streams (Devine et al. 2022).

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Some research indicates that riparian buffers alone will not be sufficient to meet water

quality standards for temperature in the OESF (Pollack et al. 2009). A 2009 study collected

water temperature data from 42 subbasins on DNR lands, including 22 readings on tributaries

to the mainstem of the Hoh River, 10 on the South Fork of the Hoh River, nine on the

Clearwater River, and one small tributary to the Bogachiel River (Pollock et al. 2009). The study

found that 17 out of 40 streams had at least one seven-day Average Daily Maximum (7DADM)

temperature that exceeded 16.0°C, and thus did not meet water quality standards for

temperature in core salmon rearing and spawning habitat (Pollock et al. 2009). One of six

subbasins with 25-50% harvest, nine of eighteen subbasins with 50-75% harvest, and seven of

nine subbasins with >75% harvest had 7DADM exceedances (Pollock et al. 2009). None of the

temperature exceedances occurred in unharvested basins (Pollock et al. 2009).

Pollock et al. (2009) explained that several mechanisms related to logging could be

responsible for increased stream temperatures, including widening and shallowing of stream

channels, widening above-channel canopy openings, decreasing large woody debris and

alluvium, and increasing debris flow frequency (Pollock et al. 2009). If these mechanisms are

causing these changes, the authors concluded that “*** reestablishment of riparian forests

alone will not be sufficient to return stream temperature regimes to natural conditions”

(Pollock et al. 2009).

Other researchers have found different temperature results. For example, Martens et al.

(2019) found that 7-day average water temperatures for DNR-managed watersheds were below

the 16°C temperature standard. Martens et al. (2019) found that summer stream temperatures

have decreased and riparian canopy cover has increased (Martens et al. 2019). Devine et al.

(2022) reported that the average 7-day maximum water temperatures on DNR managed lands

was 14.4°C, with only 16 exceedances of 16°C out of 329 observations of the 7-day maximum.

Devine et al (2022) explained that the cool maritime climate and increased shading influenced

water temperatures.

b. Washington State Forest Practices Habitat Conservation Plan

The Washington State Forest Practices HCP (“FPHCP”) applies to private lands on the

Olympic Peninsula. Although the FPHCP includes habitat protections that benefit aquatic

species, NMFS and USFWS noted that forestry activities could still potentially adversely affect

aquatic habitat by increasing temperature pollution and sedimentation and decreasing large

wood recruitment (NMFS and FWS 2006). Additionally, legal questions have been raised

regarding whether the HCP’s Clean Water Act assurances violate Washington’s antidegradation

policy and undermine the TMDL program (Steifel 2013). Federal agencies, including NMFS, have

also raised concerns about water temperature, riparian function, and Clean Water Act

assurances.

There are ongoing issues with water typing classifications as well. Water typing is

fundamental to effectively protect riparian buffers from forest practices, and it can be used in

other management contexts such as fish population assessments and harvest management.

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Within the context of forest practices, the FPHCP riparian buffer width rules are based on a

water type classification system intended to identify Type F (fish-habitat) stream reaches and

Type N (non fish-habitat) reaches. The state, however, has been operating under Interim Water

Type Rules (WAC 222-16-031) since 1999, which inappropriately allows streams to be classified

as Type N based on the results of a single point-in-time electrofishing survey.

NMFS and USFWS have warned WDNR that its water typing practices are not consistent

with the FPHCP. In 2015, NMFS and USFWS sent a letter to Washington’s Commission of Public

Lands, describing WDNR’s failure to follow the FPHCP.

“Since the HCP was signed, too many water type determinations that result in

permanent modifications continue not to use physical stream characteristics or

models specified by the FPHCP. Water typing decisions based on a

demonstration of “fish absence” are not consistent with the permanent

approaches agreed to in the FPHC.”

(Letter from Kim Kratz, Assistant Regional Administrator, NMFS, and Eric V. Rickerson, State

Supervisor, USFWS, to Peter Goldmark, Commissioner of Public Lands, DNR (July 2, 2015) (copy

provided). As a result of WDNR’s water typing practices, NMFS and USFWS do not believe fish

habitat has been correctly identified in a “substantial number of instances” Id. Therefore, the

Services requested that WDNR consider its water type modifications as temporary rather than

permanent until “water typing is better aligned with the original HCP requirements.” Id.

In addition to water typing concerns, NMFS and USFWS expressed concerns about

WDNR’s compliance reporting. Id. Specifically, the Services stated that “sample sizes are

decreasing and reported patterns are disconcerting.” Id. It is unclear if these compliance

monitoring issues have been resolved to the Services’ satisfaction.

WDNR is still behind schedule on fulfilling its commitment to “rapidly and iteratively

improv[e] the water typing map and model” (NMFS and USFWS 2015). Although WDNR has

committed to develop and adopt permanent water type rules that more accurately identify fish

habitat as defined in WAC 222-16-010, including recoverable fish habitat, this still has not

occurred despite being 16 years into the 50-year habitat conservation plan (WDNR, 2019).

2. Water Pollution Control Act

Washington has established water quality standards for surface waters throughout the

state. The purposes of these standards are to protect public health and public enjoyment of

Washington’s surface waters and to “protect[] *** fish, shellfish, and wildlife ***” pursuant to

the Washington Pollution Control Act. WAC 173-201A-010(1).

The WPCA prohibits the Washington Department of Ecology (WDOE) from establishing a

permit system for nonpoint source pollution caused by forest practices. RCW § 90.48.420(3). It

also prohibits WDOE from fining or penalizing nonpoint sources of pollution so long as the

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polluters follow forest practice rules. Id. WDOE is allowed to review forest practice rules, and

WDOE and DNR are required to agree that the rules will allow water quality standards to be

met. RCW § 90.48.420(3).

At least one author has suggested that the incidental take permit (ITP) associated with

the FPHCP may violate Washington’s antidegradation policy, which seeks to “restore and

maintain the highest possible quality of the surface waters of Washington.” (Stiefel 2013). The

policy provides that “[n]o degradation may be allowed that would interfere with, or become

injurious to, existing or designated uses ***.” Wash. Admin. Code 173-201A-200. Because the

ITP authorizes activities that impair fish and aquatic life uses, the ITP expressly violates

Washington’s antidegradation policy (Stiefel 2013).

3. State Environmental Policy Act

The State Environmental Policy Act (SEPA) requires state and county agencies to identify

environmental impacts of proposed projects before committing to a particular course of action.

RCW Ch. 43.21C. Under SEPA, state and county agencies must evaluate these impacts, consider

alternatives and mitigation measures, and involve the public in decision making. Like NEPA,

SEPA is a procedural law; it does not require agencies to select the alternative that causes the

least harm to the environment.

4. Fishing regulations

The Olympic Peninsula DPS supports the most popular and intensive recreational

fisheries for wild winter steelhead in Washington State (Bentley 2017; Cram et al. 2018), largely

because most rivers in Puget Sound are now closed to steelheading in winter (Burge et al.

2006). With little other opportunity, many fishing guides have relocated to the Olympic

Peninsula and as such, each year thousands of anglers travel to fish famed rivers like the Sol

Duc, Bogachiel, Hoh, Queets, and Quinault Rivers from November through April, which has

resulted in very high encounter rates with wild steelhead (e.g., Bentley 2017). Some other

smaller populations, including but not limited to the Hoko River, Clallam River, and Sekiu River,

also support fisheries, though not to nearly the same extent as the largest populations. Further,

nearly all the rivers and streams – both large and small – in the DPS are open to recreational

fishing during summer (June – October) when anglers may encounter summer steelhead.

Below we discuss how fishery regulations and assumptions by WDFW have potentially helped

and hindered the productivity and resilience of wild steelhead.

WDFW has taken steps with fishery regulations to reduce impacts on wild steelhead,

including eliminating retention of wild summer steelhead (we could not determine the date

when retention was eliminated by WDFW and National Park) and eliminating retention of wild

winter steelhead in 2015. Further, in response to the sharp declines in run size the past few

years WDFW has shortened or closed fishing seasons for wild winter steelhead and reduced

angler effectiveness by eliminating fishing from boats and banning the use of bait

(https://wdfw.medium.com/frequently-asked-questions-march-2022-coastal-steelhead-

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closure-364cfa62826f). Each of these actions may have slowed the decline of steelhead, but

these actions, as a whole, have not seemed to reverse the plight of wild steelhead.

However, a lack of prior due diligence and regulatory action by WDFW has also likely

contributed to the recent declines and closures. The largest watersheds on the Olympic

Peninsula experience high levels of angling pressure that has resulted in very high encounter

rates with wild winter steelhead (Bentley 2017). Bentley (2017) evaluated the creel data

collected by WDFW to estimate total angler effort and total number of wild steelhead caught

and released in the Hoh River in 2014-2015. That analysis indicated WDFW’s prior method

(referred to hereafter as the “old expansion method”) for expanding creel surveys

underestimated the total angler effort and catch by a large amount (Bentley 2017). For

example, using the old expansion method WDFW estimated 2,580 wild steelhead were caught

and released by recreational anglers in 2014-2015, but that number jumped to 4,580 steelhead

using the improved model by Bentley (2017). That equates to a 1.77x increase. Total

escapement that year was 3,171 steelhead, meaning that every steelhead that escaped to

spawn was caught, on average, 1.4 times by anglers (Table 9).

Table 9. Estimated number of wild winter steelhead caught and released (CnR) by recreational

anglers in 2014-2015 in relation to escapement goal using the old expansion method by WDFW

(old method) and using the new model developed by WDFW (Bentley et al. 2017) in the Hoh

River and Quillayute River system excluding the Dickey River, which does not get creel surveys.

Bentley et al. (2017) only applied their model to the Hoh River in 2014-2015 and based on that

analysis, the old method underestimated catch by 1.77 times. We applied that multiplier to

estimates of catch using the old method in the Bogachiel, Calawah, and Sol Duc Rivers (summed

as the “Quillayute system”) for the 2014-2015 season to generate a rough estimate of potential

recreational catch using assumptions in Bentley et al. (2017).

Population Escapement

Old Method

Total wild

steelhead CnR

Bentley et al.

(2017)

1.77 multiplier

Percent of escapement

CnR

Old

method

1.77

multiplier

Hoh 3171 2580 4580 0.8 1.4

Quillayute system 7914 5277 9340 0.7 1.2

Calawah 3081 576 1019 0.2 0.3

Bogachiel 1315 1893 3350 1.4 2.5

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Sol Duc 3229 2808 4970 0.9 1.5

Bentley (2017) did not evaluate any other years or streams but did suggest re-evaluating

the prior year’s assumptions using the new method. Unfortunately, WDFW has not done that

so far, and therefore, they do not have a rigorous estimate of total encounter rates outside of

the single year analyzed by Bentley (2017). To help better understand the level of encounters in

the neighboring Quillayute River system, we applied the 1.77x correction to the old expansion

totals generated by WDFW for the Bogachiel, Calawah, and Sol Duc Rivers, and then summed

those for the entire system (Table 9). To be clear, Bentley (2017) did not come up with a

“correction factor” that can be applied to other populations. Rather, we did this only as a

means of roughly gaging how much WDFW has underestimated total numbers of fish caught

and released. The encounter rates rose sharply when we applied an expansion, but even the old

method suggested very high encounter rates in the Bogachiel and Sol Duc Rivers, which are

more easily accessed than the Calawah River (Table 9). If those estimates are in the general

ballpark, then encounter rates in the Bogachiel and Sol Duc River are, like the Hoh River, very

high.

It is difficult to estimate what effects such encounter rates could have. WDFW assumes

a fairly conservative 10% mortality rate with each encounter, which is higher than the typical 3-

5% rate found in many studies (Hooton 1987; Taylor and Barnhart 1997; Hooton 2001; Nelson

et al. 2005; Twardek et al. 2018). Nonetheless, they don’t account for fish being caught multiple

times, and they do not account for potential sublethal impacts associated with catch and

release. As discussed earlier, emerging research on Atlantic salmon indicates there can be

significant sub-lethal impacts that can alter the migration and reduce the reduce fitness of adult

females (up to ~ 20-25%) that are caught and released (Richard et al. 2013; Bouchard et al.

2021; Papatheodoulou et al. 2021). If sublethal impacts are similar in steelhead, then those

impacts, combined with the high encounter rates could be far more detrimental to productivity

than a simple assumption of a 10% mortality.

There are two other areas of notable concern. First, the fishery for summer steelhead is

not monitored and while WDFW has enacted emergency regulations for other stocks, including

closures, to protect depleted run sizes of wild winter steelhead (WDFW emergency regulations,

accessed online: https://wdfw.wa.gov/news/state-announces-full-closure-coastal-steelhead-

fishing-support-conservation-following) and summer coho salmon (WDFW 2021 emergency

regulations, accessed online:

https://wdfw.wa.gov/sites/default/files/about/regulations/filings/2021/2021_nof_statewide_r

ec_ces_final_7_1_2021.pdf), they have done nothing to help conserve wild summer steelhead.

Although WDFW does not monitor summer steelhead, they have been aware of the snorkel

counts in Brenkman et al. (2012) for well over a decade. Even low encounter rates by

recreational anglers could have impacts on the struggling populations, particularly because

mortality increases when water temperatures are warmer and sublethal impacts are not

accounted for. For instance, Barnhardt and Taylor (1996) found mortality greatly increased

when water temperatures reached and exceeded 21°C. While it is rare for the Hoh and Queets

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Rivers to reach 21°C during summer, such temperatures are fairly common in sections of the

Calawah (John McMillan, The Conservation Angler, unpublished instantaneous water

temperature data 2000-2006; 2009-2021) and Sol Duc Rivers (Sol Duc stream and temperature

gage: https://apps.ecology.wa.gov/ContinuousFlowAndWQ/StationDetails?sta=20A070).

Second, it is unclear how recreational angling is impacting juvenile steelhead. Juvenile

steelhead are found throughout a variety of main-stem and tributary habitats (McHenry et al.

1996; Smith 2000; McMillan and Starr 2008; McMillan et al. 2013), most of which are open to

angling during the summer months. It is possible that many juvenile steelhead are captured

each season by anglers, though at this time we are nor sure to what extent and how such

encounters could impact survival. Perhaps more worrisome is the potential for high encounter

rates on older, larger steelhead parr and steelhead smolts that are found in the main-stem

rivers during the winter steelhead season – from January – April. If juveniles are killed during

summer there is potential for compensation by other individuals, but if smolts are killed during

their journey downstream, there is almost no chance for compensation. Personal fishing

experiences and observations (John McMillan, The Conservation Angler) indicate many juvenile

steelhead are captured on steelhead gear in winter, and smolts are commonly captured during

the spring when they migrate to the ocean. Because they are non-target fish and the gear is for

larger fish, lethal hookings appear more common and smolts are often handled without care.

Given the estimated smolt to adult survival of Olympic Peninsula steelhead, a river like the Hoh

may only be producing 30,000-40,000 smolts per year. Anglers reported catching and releasing

over 4,500 adults in the Hoh River in 2014-2015 (Bentley 2017). Anglers likely encounter more

smolts than adults, perhaps substantially so because they are feeding. Understanding these

impacts, although challenging, seems important because of the sheer number of anglers and

their potential for catching and releasing large numbers of juvenile steelhead. At this point,

given the declines in the populations, all smolts are important to productivity and resilience.

V. OTHER NATURAL OR ANTHROPOGENIC FACTORS

A. Hatcheries

In 1996, NMFS acknowledged the genetic risk that hatcheries pose to wild steelhead on

the Olympic Peninsula (Busby et al. 1996). At the time, around 40,000 summer steelhead,

primarily out-of-DPS Skamania stock, were annually released in the Quillayute River Basin

(Crawford 1979; WDF et al. 1993); 840,000 winter steelhead, primarily from

Bogachiel/Chambers/Cook Creek stock, were annually planted; (WDF et al. 1993; WDFW

1994a); and winter steelhead of mixed origin were being released into the Quinault River

(Busby et al. 1996). The proportion of natural spawning winter steelhead composed of

hatchery-origin recruits (pHOS) ranged from 16% in Quillayute River to 44% in Quinault River

(Busby et al. 1996).

In the twenty-five years following the 1996 Status Review, additional studies have

discussed the genetic and ecological harm hatchery steelhead programs cause to wild steelhead

populations (e.g., Goodman 2005, Araki et al. 2008; Araki et al. 2009; Araki and Schmid 2010;

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Chilcote 2003; Naish et al. 2007; Chilcote et al. 2011; Kostow et al. 2003; Kostow and Zhou

2006; Kostow 2009). Considering these risks, state, and federal agencies, as well as the

Hatchery Scientific Review Group, have raised concerns about the effects of steelhead hatchery

production on the Olympic Peninsula (HSRG 2004, 2015; USFWS 2009; Duda et al. 2018; Cram

et al. 2018). And research by McMillan et al. (2022) indicates hatchery fish, both directly and

indirectly, have likely contributed to the depletion of early returning wild winter steelhead.

Below we summarize what is known about the extent of hatchery operations in the DPS,

straying of hatchery steelhead, and potential ecological and genetic impacts posed by hatchery

fish.

1. Hatchery Operations on the Olympic Peninsula

Hatchery steelhead have been released into Olympic Peninsula rivers for over a century

(Duda et al. 2018). Between 1933 and 2014, a total of 9.2 million winter steelhead and 1.6

million summer steelhead were released into the Quillayute River system (Duda et al. 2018).

Between 1958 and 2014, a total of 4.1 million winter steelhead and 0.9 million summer

steelhead were released into the Hoh River (Duda at al. 2018). And, between 1962 and 2010, a

total of 3.01 million juvenile winter-run steelhead were planted in WRIA 19 streams (NOPLE

2015).

The largest releases have occurred in WRIA 21, primarily on the Quinault and Queets

Rivers (Table 1). From 1915 to 2014, 26.2 million winter steelhead were released into the

Quinault River, and from 1974 to 2014, 5.2 million winter steelhead were released into the

Queets (Duda et al. 2018). The Quinault River program continues to release close to half-a-

million steelhead every year: 395,606 (2015), 452,482 (2016), 487,190 (2017), and 475,488

(2018) (QDF and WDFW 2021).

As of 2013, eighteen hatchery release operations (on-site and off-site) were ongoing on

the following rivers: Calawah River (summer: Figure 24, and winter steelhead), Clallam River,

Goodman Creek, Hoh River, Hoko River, Lower Quinault River, Lyre River, Pysht/Independents,

Queets River, Quillayute/Bogachiel River, Sail River, Sekiu River, Sooes/Waatch Rivers, Sol Duc

Rivers (terminated all hatchery steelhead releases after 2013), and Upper Quinault River (Table

1; Cram et al. 2018). Between 2000 and 2008, these programs released an average of 1,383,024

hatchery steelhead per year (Cram et al. 2018.) As a result of some off-site release program

reductions that occurred between 2009 and 2013, the average annual release has reduced to

1,072,781 (Cram et al. 2018). Population-specific data from Cram et al. (2018) are provided in

Table 1.

These programs are operated for harvest augmentation and mostly use early returning

Chambers Creek winter stock and early returning Skamania summer stock, neither of which are

native to the Olympic Peninsula DPS (Cram et al. 2018). It is unclear if the Quinault programs

are segregated (Cram et al. 2018). In 2013, an integrated hatchery program using early

returning wild broodstock was initiated on the Bogachiel River, and summer and winter

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steelhead hatchery releases were terminated on the Sol Duc River (Table 1 and Figure 24; Cram

et al. 2018)

a. Straying

Hatchery steelhead released within, and outside of, the Olympic Peninsula DPS

boundaries are known to stray into the DPS’s river and streams (Cederholm 1993; Houston and

Contor 1984; Phelps et al. 1997; McMillan 2006; USFWS 2009; Brenkman et al. 2012; Kassler et

al. 2011; Cram et al. 2018). WDFW acknowledges this threat to the genetic integrity of Olympic

Peninsula steelhead (Cram et al. 2018). As a result of past and current hatchery practices, the

Sitkum River summer steelhead population may be the only population of Olympic Peninsula

steelhead that has escaped large levels of hatchery straying (Table 7). Almost no information is

available for straying of hatchery winter steelhead, if only because it is difficult to discern which

populations the fish came from, and because in places like the Queets River, the hatchery

steelhead are not marked. The lack of marking for hatchery winter steelhead in the Queets, and

one of the hatchery stocks in the Quinault River, is of great concern because it is possible those

fish and their feral offspring are masking (Quinones et al. 2013; Willmes et al. 2018) potentially

greater declines in wild winter steelhead than is currently thought. Not marking hatchery fish

also undermines state selective fishing regulations that are intended to protect wild fish from

harvest while reducing pHOS by encouraging harvest of hatchery fish.

Snorkel survey data suggests there is a substantial amount of straying by hatchery

summer steelhead into watersheds that do not have any releases of hatchery summer

steelhead (Table 7). For example, surveys by Brenkman et al. (2012) from 2005-2010 found a

mean percent origin hatchery adult of 40% in the SF Hoh River, 16% in the EF Quinault River,

and 43% in the NF Quinault River (Table 7). In some years, there were so few wild summer

steelhead and so many hatchery summer runs that the annual proportion of hatchery adults

ranged up to 76% and 100% in the SF Hoh and NF Quinault Rivers (Table 7). McMillan (2022)

reported similar results as Brenkman et al. (2012) in the SF Hoh River, although over a longer

period (Figure 9), with a mean percent hatchery steelhead of 41% and an annual peak of 67%

(Table 7). Although it is possible some hatchery steelhead moved downstream in the Quillayute

River system and may not have spawned in the upper watershed, the surveys were purposely

conducted late in the season to minimize that behavior (McMillan 2022). Regardless, if we

translate the proportion of hatchery adults in late summer and early fall into pHOS, the

estimates easily exceed the HSRSG’s recommended pHOS limits of 5%-10% for segregated

programs (WCSSP 2013; HSRG 2015.

Snorkel survey data also suggests hatchery steelhead are straying within the Quillayute

River system outside of the release location in the lower Calawah River. Brenkman et al. (2012)

reported a mean proportion of hatchery steelhead of 13% in the upper Bogachiel River (Table

7), while McMillan (2022) reported a mean of 33% in the NF Calawah River with an annual

range of 0% up to 100% (Table 7). In contrast, relatively few hatchery steelhead have been

documented in the Sitkum River and upper SF Calawah River, where most of the spawning is

believed to occur (Table 7). The most likely explanation is the series of numerous waterfalls that

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must be ascended to reach the upper Sitkum and SF Calawah Rivers (McMillan 2006; McMillan

2022). Most steelhead ascend the falls from late summer through early fall after the first

freshets in October (McMillan 2022). It is likely that many summer steelhead are not capable of

ascending the falls and others are not able to capably time their passage with the appropriate

weather window.

As of 2013, weirs and other adult traps intended to remove hatchery steelhead were

absent on multiple Olympic Peninsula rivers (Cram et al. 2018). Specifically, as of 2013, there

were no weirs or traps on the Calawah River, Goodman Creek, Hoh River, Lyre River,

Pysht/Independents, Sail River, or Sekiu River (Cram et al. 2018). It is unclear if weirs or traps

were in place on the lower or upper sections of the Quinault River in 2013 (Cram et al. 2018); it

is unclear if they are there now.

Even where weirs and adult traps are in place, they may not sufficiently protect wild

populations (Seamons et al. 2012). They may not be complete barriers and they do not prevent

hatchery fish from spawning elsewhere in a watershed (Quinn 1993; Dittman et al. 2010).

Overall, straying hatchery summer steelhead combined with low abundance of wild

summer steelhead would appear to represent a great risk to the genetic integrity and diversity

of the wild populations, particularly in the Hoh, Quinault, and NF Calawah Rivers. Further, there

is almost no information on how many of those hatchery summer steelhead remain in the

system and spawn with wild winter runs. The uncertainty and potential for strong negative

hatchery impacts underscores the need for an ESA listing for the Olympic Peninsula steelhead

DPS.

b. Genetic and Ecological Risks

Concerns regarding interbreeding between hatchery and wild steelhead on the Olympic

Peninsula have been raised for nearly thirty years, if not longer. In the early 1980’s, Cederholm

(1983) raised concerns about the large number of hatchery smolts released in the Quinault,

Queets, Hoh, and Quillayute Rivers. Cederholm (1983) noted there was a high degree of within

and between river straying and that there may be interbreeding with wild stocks, which could

change the “long-term, spawning, timing, growth and survival of wild fish.” (Cederholm 1983).

The record shows Cederholm was correct, and in fact, subsequent snorkel surveys (mentioned

in previous section) indicate this is also a significant problem for wild summer steelhead.

NMFS expressed similar concerns in 1996 (Busby et al. 1996). In its stock assessment,

the BRT mentioned there was “widespread production of hatchery steelhead” within the DPS,

which presented genetic risks despite management efforts to minimize genetic introgression

(Busby et al. 1996).

In 2004, the Hatchery Scientific Review Group (HSRG) raised several concerns regarding

interbreeding on the Olympic Peninsula (HSRG 2004). Specifically, the HSRG identified

interbreeding between wild and hatchery steelhead as a concern for Quillayute River winter

134

and summer steelhead populations and the Hoh River winter steelhead population, and it

recommended the termination of multiple off-site release programs (HSRG 2004). Several of

those off-site release programs have been terminated to avoid genetic and ecological risks. For

example, the Quinault Tribe discontinued transfers Quinault NFH winter steelhead to the Hoh

River and Cook Creek. (https://www.fws.gov/Quinaultnfh/Hatchery.cfm).

The Washington Coast Sustainable Salmon Plan (2013) indicates that previous

proportions of hatchery fish spawning with natural origin spawners likely exceeds the HSRG’s

recommended pHOS limits of 5%-10% for segregated programs (WCSSP 2013; HSRG 2017). The

report includes the following estimated percentages of natural winter steelhead spawners by

river system: Sooes/Waatch: 50-74%; Ozette: 75-94%; Quil/Bogie: 75-94%; Dickey: 95-100%; Sol

Duc: 75-94%; Calawah: 75-94%; Goodman Creek: 50-74%; Mosquito Creek: 75-94%; Hoh River:

75-94%; Klalaloch Creek 95-100%; Queets River: 75-94%; Clearwater River 95-100%; Raft River:

75-94%; Quinault/Lake Quinault (75-94%); Quinault River (50-74%); Moclips River (95-100%);

and Copalis: (95-100%) (WCSSP 2013). Based on these estimates, genetic risk would be

increased on all systems except Kalaloch Creek and the Dickey, Moclips, and Copalis Rivers.

In 2008, WDFW reported that introgression resulting from the release of Chambers

Creek winter steelhead may have occurred in the Pysht/Independent, Hoko River, and Sol Duc

River populations (WDFW 2008). The report describes Chambers Creek winter and Skamania

River summer steelhead as “pos[ing] substantial risk to both the among-population diversity

and the fitness of natural steelhead populations” (WDFW 2008). Because of introgression, the

report estimated that the Pysht/Independents, Hoko, and Sol Duc winter steelhead populations

had “high” reductions in diversity (WDFW 2008).

“A limited amount of information was available to evaluate changes in spatial

structure and diversity in the Olympic Peninsula region. Most notably, we

compared the 1993 genetic characteristics of the Pysht Winter and Hoko Winter

steelhead populations with samples collected in 1973. The analysis indicated a

5.5-14.5% gene flow (modal value of 9.5%) from hatchery-origin, Chambers type

stock to the Hoko natural population; a 12-75% gene flow (modal value of

26.5%) to the Pysht natural population; and a 2.5-6% (modal value of 4%) to the

Sol Duc winter natural population (see Chapter 4, Artificial Production).”

(WDFW 2008). WDFW did not review gene flow samples from other wild steelhead populations

for the 2008 report. Fortunately, in 2012, WDFW stopped releasing hatchery winter and

summer steelhead in the Sol Duc River (Cram et al. 2018), but to date, we are not aware of data

on potential hatchery effects before and after cessation of hatchery releases.

In 2009, WDFW reported pHOS measurements related to the operation of its hatchery

program on the Bogachiel River (See

https://fortress.wa.gov/dfw/score/score/hatcheries/hatchery_details.jsp?hatchery=Bogachiel%

20Hatchery). WDFW did not meet its 5% and 10% pHOS goals for winter and summer

135

steelhead, respectively. WDFW reported a 9% pHOS rate for winter steelhead and a 23% pHOS

rate for summer steelhead (WDFW Cons. 2021).

Ecological impacts of hatchery programs may occur through a variety of pathways, and

the effects – ranging from predation to competition for food and space – have been reviewed

extensively (Einum and Fleming 2001; Weber and Fausch 2003; Kostow 2009; Tatara and

Berejikian 2012; Rand et al. 2012). Given the high pHOS levels in many of the largest

populations, and the ability of those fish to produce some feral offspring and interbreed with

wild fish, there is heightened potential for those feral and hybrid juveniles to compete with wild

steelhead for limited food and habitat, and negatively influence declining and depleted

populations of wild steelhead.

Outside of freshwater, there is a growing body of evidence that large-scale releases of

hatchery pink and chum salmon are negatively impacting growth and survival of other

salmonids in the North Pacific (Levin et al. 2001; Ruggerone et al. 2010; Ruggerone and Connors

2015; Ruggerone and Irvine 2018). A recent presentation on the subject found hatchery pink

salmon negatively impacted the growth and survival of steelhead in the Thompson River and

Chinook salmon in Puget Sound and the Columbia River (Ruggerone et al. 2021). It is therefore

possible that wild steelhead in the Olympic Peninsula DPS, particularly populations with older

fish that migrate further into the North Pacific, are also being impacted by competition with

pink and chum salmon.

c. Hatchery Fishery Structure and Erosion of Adaptive Capacity

As we described earlier, Cram et al. (2018) noted the risk of fishery selection on run

timing in wild winter steelhead since the number of fishing days per week for treaty fisheries is

highest early in the season to target greater harvest on hatchery adults. However, it should be

noted that the recreational fisheries operated in the same manner for decades before hatchery

steelhead were marked, and as such, both wild and hatchery steelhead were harvested at high

rates (McMillan 2006). Further, while it is now catch and release only for wild winter steelhead,

individuals that enter earlier in the season and remain in the system for longer periods are

more likely to be caught multiple times, and such effects are unknown (Hooton and Lirette

1986). Owing to this combination of factors, and others, it appears the hatchery fishery

structure has contributed to the depletion of early returning wild steelhead and as a result, run

timing is more compressed than it was historically (McMillan et al. 2022).

This is a concern because run timing is highly heritable in salmonids and it represents a

mechanism through which fish can adapt to changes in stream flow and water temperature

(Manhard et al. 2017; Tillotson and Quinn 2018). As outlined in this document previously,

stream flows and water temperatures are predicted to become flashier and warmer,

respectively (Wade et al. 2013). Winter steelhead enter and spawn earlier in the winter and

spring in more southerly areas of their native range (Busby et al. 1996), which suggests wild

steelhead on the Olympic Peninsula will need to rebuild the front end of their run timing to

keep pace with climate change.

136

Environmental conditions on the Olympic Peninsula have changed in the past forty years

and they will continue to shift in relation to climate effects. For example, early winter peak

flows are expected to become more intense and summer streamflows and temperatures will

occur earlier and for longer durations (Wade et al. 2013). To take advantage of higher

streamflows earlier in the winter and to ensure they get their juveniles out of the gravel before

the onset of summer base flows and peak temperatures, steelhead will need to increasingly

enter and spawn earlier. This will not be possible if the current hatchery management and

fishery framework remains in place, because those impacts are essentially blocking the

potential for adaptations in migration timing. Hence, the population could become increasingly

less productive if the fishery and hatchery practices continue to select for later entering and

spawning adults.

Of course, run timing is only one trait that is likely to be under selection as climate

effects unfold. Genetics are a complex field and numerous other traits will also likely need to

change.

“Many traits appear to have responded to recent climate change, apparently

without genetic adaptation. However, to keep pace with climate change, genetic

adaptation may be necessary in the long-run; thus maintaining genetic diversity

within DPSs and species as a whole is a high priority for salmon conservation.”

Crozier et al. (2019). See also (Pitman et al. 2020).

Although steelhead are highly adaptable, they have their limits (Wade et al. 2013) and it

is uncertain whether Olympic Peninsula steelhead can adapt quickly enough to climate change

(Halofesky et al. 2011). This concern is particularly heightened for early returning wild winter

steelhead and wild summer runs, the latter of which may no longer have the necessary genetic

and phenotypic diversity to sustain fitness and productivity in the coming years and decades.

B. Ocean Conditions

Correlations exist between recurring, decadal-scale variability (including the Pacific

Decadal Oscillation and the El Nino Southern Oscillation) and salmonid abundance in the Pacific

Northwest (Stout et al. 2012). When ocean conditions are unfavorable for salmonids, their

abundance declines. Historically, salmonids have persisted through shifts in ocean productivity.

However, it is uncertain whether salmonids can withstand these shifts in marine conditions on

top of climate change, habitat loss, and other increasing threats.

C. Loss of Salmon Nutrients

The loss of marine-derived nutrients from other salmonids in Olympic Peninsula rivers

and streams is likely limiting steelhead productivity (Minkova et al. 2021; Bilby et al. 1996,

1998; NOPLE 2015; McMillan 2006; Halofsky et al. 2011). For example, McMillan (2006)

137

estimated that the combined loss of pink, chum, and coho in the Quillayute River system has

resulted in 87,411 - 201,761 fewer salmon as compared to early 20th century numbers, a

significant reduction in marine nutrients. As a result of losing these nutrients, streams are less

able to produce steelhead smolts that spend an average of two years instream before

outmigrating (McMillan 2006). The same is likely true for all other Olympic Peninsula rivers that

historically supported significant numbers of salmonids.

EFFECTIVENESS OF CONSERVATION ACTIVITIES BY STATES AND OTHER PARTIES

State, local, and tribal governments, as well as nonprofit organizations, have invested

significant resources in recovering steelhead habitat on the Olympic Peninsula. For example,

the Wild Salmon Center, Western Rivers Conservancy, The Nature Conservancy, and WDNR

have purchased thousands of acres of habitat along the Hoh River to protect migration, rearing,

and spawning habitat (McMillan 2006; McMillan and Starr 2008). This is in addition to many

other projects to restore habitat, which can be reviewed on the Washington Governor’s Salmon

Recovery Office website: https://srp.rco.wa.gov/projectmap?mlayer=projects.

WDNR administers a Family Forest Fish Passage Program that assists private landowners

with fish passage improvements. Over seventeen years, the program has completed 67 projects

on the Olympic Peninsula (WDNR 2020). WDNR also administers a River and Open Space

Program that purchases habitat for conservation purposes (WDNR 2020). Since the program’s

inception, it has funded 16 conservation easements that protect 1,043 acres, although it is

unclear if any of those easements occur within Olympic Peninsula steelhead habitat.

Since 1999, the Pacific Coast Salmon Recovery Fund has distributed $12.6 million in

funding to habitat projects throughout Washington’s coast. This federal spending leveraged an

additional $33 million in federal and state funding for these projects.

Although habitat has likely improved in the limited areas where habitat restoration has

occurred, the results – as measured by increased abundance, productivity, spatial distribution,

and diversity – are unclear. For example, Bilby et al. (2022) reported that habitat restoration in

Deep Creek watershed produced mixed results in response to habitat restoration. Additionally,

Bilby et al. (2022) indicates that land use practices continue to interfere with restoration

efforts.

“Historical and ongoing land-use practices (road construction, logging) continue

to influence the effectiveness of the restoration activities. Mass-wasting events

and avulsions can dramatically change or negate the effects of restoration.

Equilibration of the systems will take decades or longer to occur. Riparian

recovery in particular will take centuries.”

(Bilby et al 2022).

138

To make restoration projects effective, there need to be enough fish to utilize restored

habitat. As McMillan (2006) explained, “habitat purchases made to recreate functioning salmon

and steelhead ecosystems are rendered ineffective without sufficient numbers of the key

species that drive them.”

REQUEST FOR CRITICAL HABITAT DESIGNATION

The Petitioners request the designation of critical habitat for Olympic Peninsula

steelhead concurrent with listing. Critical habitat should encompass all known and potential

freshwater spawning and rearing areas, migratory routes, estuarine habitats, riparian habitats

and buffers, and essential near-shore ocean habitats.

Acknowledgments

The following individuals prepared this petition:

John McMillan, Science Director, The Conservation Angler

Rob Kirschner, Legal and Policy Director, The Conservation Angler

Nick Gayeski, Fisheries Scientist, Wild Fish Conservancy

Conrad Gowell, Biologist, Wild Fish Conservancy

139

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APPENDIX A

Summary of Bayesian analysis of Hoh and Quillayute wild winter-run steelhead spawner-recruit

analyses from which Figures 21 and 22 were created

Nick Gayeski

Wild Fish Conservancy

June 17, 2022

Spawner and recruit data for each river were provided in the two attached Excel files received

by Wild Fish Conservancy (WFC) from WDFW in May 2022. See Gayeski Bayesian Analysis data

folder provided in submitted materials.

Sheet “Recruits Per Spawner 1” of the file “Quilayutte_Winter_Steelhead_data.xlsx, columns C

and G (rows 6 to 43) contained spawner and recruit data for brood years 1978 to 2015

calculated by WDFW staff and was taken at face value. The corresponding data file used in the

Stan analysis is named “QuilBY78-15.txt”.

The data set for Hoh wild winter run steelhead received by WFC (“Hoh Stlhd broodyear return

by age.xlsx”) did not include recruits calculated from the cohort analysis (worksheet “Cohort

Analysis”) so I calculated maiden recruits from the cohort analysis data (Gayeski 2022). This is

contained in workheet “R and S” of the file “Hoh Stlhd broodyear return by

age_NG_June_2022.xlsx” columns BI and BJ (rows 6 to 32). The corresponding data file used in

the Stan analysis is named “HohBY87-13.txt”.

The text files “Printfit_Hoh steelhead_RickerRW May 9 2022.txt” and “Printfit_Quil

steelhead_RickerRW May 9 2022.txt” contain the summary output of the Bayes analyses

conducted in program Stan (Gayeski 2022a)

The Ricker random walk in alpha model treats the natural logarithm of the Ricker alpha as

undergoing a random walk with the capacity parameter beta treated as a constant (estimated

from the data as part of the fitting of the model). In fitting each of the two data sets, broad

uniform prior distributions were placed on beta (between 1000 and 20000), sigmaw, the

standard error of the annual random walk (between 0 and 2.0), and sigma, the population-and-

time (brood year)-specific process error of the R/S relationship. A normal prior in natural log

space was placed on log alpha (mean = 1.5078, standard deviation = 0.3486), which confines

alpha to the interval [1.8, 11].

Four chains of 10000 replicates each were run, the first 5000 samples of which were treated as

“warmup” and the last 5000 retained as samples from the posterior distribution of the

parameters, for a total of 20,000 posterior samples of each parameter. Both models converged

rapidly. Trace plots of log alpha, beta, sigmaw, and sigma showed coherent sampling of the

parameter posteriors, and Rhat values were all equal to 1.00+ indicating convergence.