United States

Department

of Agriculture

Forest Service

Rocky Mountain

Research Station

General Technical

Report RMRS-GTR-174

September 2006

Strategies for Conserving Native 

Salmonid Populations at Risk 

From Nonnative Fish Invasions: 

Tradeoffs in Using Barriers to Upstream Movement

Kurt D. Fausch

Bruce E. Rieman

Michael K. Young

Jason B. Dunham
You may order additional copies of this publication by sending your 

mailing information in label form through one of the following media. 

Please specify the publication title and number.

Telephone (970) 498-1392

FAX (970) 498-1122

E-mail rschneider@fs.fed.us

Web site http://www.fs.fed.us/rm

Mailing Address Publications Distribution

Rocky Mountain Research Station

240 West Prospect Road

Fort Collins, CO 80526

Fausch, Kurt D.; Rieman, Bruce E.; Young, Michael, K.; Dunham, Jason B. 2006. Strategies for 

conserving native salmonid populations at risk from nonnative fish invasions: tradeoffs 

in using barriers to upstream movement. Gen. Tech. Rep. RMRS-GTR-174. Fort Collins, CO: 

U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station. 44 p.

Abstract__________________________________________

Native salmonid populations in the inland West are often restricted to small isolated habitats at 

risk from invasion by nonnative salmonids. However, further isolating these populations using barriers to prevent invasions can increase their extinction risk. This monograph reviews the state of 

knowledge about this tradeoff between invasion and isolation. We present a conceptual framework 

to guide analysis, focusing on four main questions concerning conservation value, vulnerability 

to invasion, persistence given isolation, and priorities when conserving multiple populations. Two 

examples illustrate use of the framework, and a final section discusses opportunities for making 

strategic decisions when faced with the invasion-isolation tradeoff.

Keywords: barriers to fish movement, biological invasions, conservation biology, isolation 

management, native salmonids

Authors___________________________________________________

Kurt D. Fausch is a Professor in the Department of Fish, Wildlife, and Conservation Biology at Colorado 

State University, in Fort Collins, CO. He earned a B.S. from the University of Minnesota, and M.S. and 

Ph.D. degrees from Michigan State University. During the last 30 years his research has focused on 

conservation and management of salmonids and other stream fishes, throughout the West and several 

locations worldwide.

Bruce E. Rieman is a Research Fisheries Biologist with the Rocky Mountain Research Station Boise 

Aquatic Sciences Laboratory in Boise, ID. He holds B.S., M.S. and Ph.D. degrees from the University of 

Idaho. He has worked in native fish conservation and fisheries management and research in the Pacific 

Northwest for the past 32 years.

Michael K. Young is a Research Fisheries Biologist with the Rocky Mountain Research Station in 

Missoula, MT. He holds a B.S. and M.S. from the University of Montana and a Ph.D. from the University 

of Wyoming. His work focuses on the conservation biology of cutthroat trout.

Jason B. Dunham is an Aquatic Ecologist with the U.S. Geological Survey, Forest and Rangeland 

Ecosystem Science Center, in Corvallis, OR. He holds a B.S. from Oregon State University, an M.S. from 

Arizona State University, and Ph.D. from the University of Nevada-Reno. His work focuses on native 

and nonnative species in freshwaters.

Cover photos and images clockwise from top: A gabion barrier constructed on North Barrett Creek in the Snowy 

Range, Wyoming, to prevent upstream movement by nonnative trout (photo by G. T. Allison, U.S.D.A. Forest Service). Nonnative brook trout (top) and native greenback cutthroat trout (bottom) from Hidden Valley Creek, Rocky 

Mountain National Park, Colorado (photo by K. D. Fausch, Colorado State University). Culvert beneath forest road 

in Lost Creek, Idaho (photo by J. B. Dunham, U.S. Geological Survey).Map of the North Fork Coeur d’Alene River 

basin, Idaho, showing stream segments where nonnative brook trout have invaded (red) and potential brook trout 

habitat (yellow). See Figure 7 in Section VI for details.
Executive Summary

Native salmonid populations have declined throughout the 

world due to a host of human influences, including habitat degradation and loss, invasion of nonnative fishes, and overfishing. 

In many regions, suitable coldwater habitats for salmonids are 

found mainly in protected natural areas, so that these fishes are 

increasingly relegated to smaller and more isolated pieces of their 

former native ranges. Moreover, populations of native trout and 

charr in these relatively undisturbed habitats are often at further 

risk from invasion by nonnative salmonids from downstream. 

Faced with this dilemma, fisheries managers frequently consider 

using barriers to upstream movement to prevent invasion and 

displacement or hybridization. However, in doing so they face a 

tradeoff because isolating native salmonid populations in small 

headwater habitats may also increase their risk of extinction. 

Here we focus on native salmonids in the inland western U.S. 

(for example, cutthroat trout, Oncorhynchus clarkii; bull trout, 

Salvelinus confluentus), but this is an important problem for 

salmonids and other stream biota worldwide.

This monograph reviews the state of knowledge about the 

factors that affect this tradeoff between invasion and isolation, and presents a framework for analyzing it and prioritizing 

conservation actions. Barriers to prevent invasions can pose 

problems for salmonids because these fish often need to move 

to complete their life history. Although anadromous salmonids 

are known for their extensive migrations, freshwater trout, 

charr, grayling, and whitefish also show remarkable flexibility 

in life history and diverse movement behaviors. Movements 

allow fish living in patchy environments to maximize fitness by 

placing each life history stage in habitats that provide optimum 

growth and survival. Likewise, in environments that fluctuate 

seasonally or among years, diverse life histories, and movements allow fish to avoid harsh conditions, or recolonize habitats 

after catastrophes.

Isolation of fish populations using barriers can extirpate 

mobile life history types, restrict fish populations to habitats 

inadequate for long-term persistence, and prevent natural 

recolonization after catastrophes. Both direct and indirect evidence from field research indicate that isolated populations of 

cutthroat trout and bull trout are more likely to be extirpated in 

smaller watersheds, but studies have been conducted in only a 

few regions. These case studies suggest that these salmonids 

need approximately 10 km of suitable stream habitat to persist 

for 25 to 50 years and maintain genetic diversity, although the 

long-term fate and evolutionary potential of populations in such 

small watersheds is unknown. Much more research is needed 

to refine these estimates for all taxa, because the amount of 

habitat needed for persistence is likely to vary strongly with 

climate and basin characteristics.

Invasion by three commonly introduced salmonids (brook 

trout, Salvelinus fontinalis; brown trout, Salmo trutta; and rainbow trout, O. mykiss) can also extirpate inland native stream 

salmonids, by competition and predation or introgressive hybridization. For example, field research on brook trout in the 

inland western U.S. showed that they rapidly invade upstream 

in some regions, and displace or completely eliminate native 

cutthroat trout by reducing survival during the first two years of 

life. However, in other regions brook trout invasions appear to 

have stalled, perhaps limited by environmental factors. Likewise, 

nonnative rainbow trout often hybridize with native cutthroat 

trout, and nonnative brook trout with bull trout, but in certain 

cases environmental factors or isolating mechanisms apparently reduce risks of introgression. More research is needed to 

understand where each species will invade and displace native 

species versus where they will coexist or fail to invade.

Because the outcome of invasion and isolation can differ 

among species of native salmonids, and among regions or 

habitats, installing barriers is often a complex problem that must 

be analyzed carefully. Four key questions provide a framework 

for considering the use of barriers to prevent invasions and 

conserve native fishes:

1. Is a native salmonid population of important conservation 

value present?

2. Is the population vulnerable to invasion and displacement?

3. If the native salmonid population is isolated to prevent 

invasion, will it persist?

4. If there are multiple populations of value, which ones are 

priorities for conservation?

The first step in any conservation plan is to set clear management objectives by considering the values embodied in 

a native salmonid population. Three conservation values 

emerge from the literature: evolutionary, ecological, and socio-economic. Evolutionary value focuses on distinct species, 

races, and populations, such as those protected in the U.S. 

by the Endangered Species Act, many of which are adapted 

to specific environments. For example, in some regions small 

remnant populations of native cutthroat trout represent a rare 

and significant genetic resource. Ecological value includes 

important ecological processes and functions at the population, 

community, and ecosystem levels. For example, species like 

salmon and trout have strong effects on trophic webs in aquatic 

systems, which may also extend into the terrestrial ecosystem. 

Also of ecological value are species and populations that 

are self-sustaining, resilient, adaptable, and require minimal 

inputs of resources to maintain them. Socio-economic values 

include recreational and economic benefits from fishing and 

tourism. Although values may overlap, often it is not possible 

to conserve all of them simultaneously in any one location, so 

effective management will involve clearly defining priorities.

The second question to address for a population of conservation value is whether it is vulnerable to invasion and 

displacement by, or hybridization with, a nonnative salmonid. 

This depends on the ability of the invader to be transported to 

the basin or spread to the target location, establish a reproducing population, and dominate in interactions with the native 

salmonid. Currently, most transport of nonnative salmonids is 

by unauthorized introductions by the public. Spread from these 

initial invasions may be facilitated by long distance movements 

(that is, “jump dispersal”), or hampered by barriers to upstream 

movement like cascades and waterfalls. Headwater stocking 

facilitates downstream dispersal throughout entire drainage 

basins. Once nonnative salmonids arrive, they tend to establish populations where abiotic regimes of temperature, flow, 

and natural disturbances allow reproduction and recruitment. 

Nonnative salmonids may have impacts through competition, 

predation, hybridization, or spreading parasites or pathogens. 

Empirical evidence has demonstrated strong individual biotic 

interactions (competition or predation) with native salmonids, 

particularly during the first two years of life. Hybridization is 

also commonplace and spreading in some cases, but rare or 

localized in others. Recent evidence suggests that movement 

of nonnative trout may transport nonnative parasites (such 

i
as Myxobolus cerebralis, which causes whirling disease) 

upstream to contact native salmonids.

The third question for valuable populations vulnerable to 

displacement by a nonnative salmonid is whether the native 

fish population will persist if isolated using a barrier to prevent 

upstream invasion. Both direct and indirect field evidence indicate that larger interconnected patches sustain populations 

longer than do smaller isolated ones. Natural disturbances 

like fire, flood, drought, and freezing, as well as more subtle 

climate changes that affect stream flow and temperature, can 

drive small salmonid populations to extinction. Isolation can 

eliminate migratory life histories that confer resilience through 

large fecund adults, the opportunity for gene flow, and recolonization following catastrophic events. Isolation may also interact 

with habitat degradation and overfishing to reduce population 

persistence in small isolated stream segments.

The fourth question to address is which native salmonid 

populations of value are deemed the highest priorities for 

conservation, given limited management resources. Priorities 

should be based on the relative conservation values of each 

population, the risk of losing them, and the feasibility of reducing that risk through management. These values, risks, and 

opportunities for management may vary strongly across any 

area of interest, so a strategic, spatially explicit approach is 

useful. It is important to understand the natural spatial structure 

of native salmonid populations and their dynamics through 

time, rather than arbitrarily focusing on individual streams or 

other geopolitical units. Planning to conserve native salmonids 

requires maintaining broad representation of the remaining ecological diversity, ensuring there is redundancy (more than one 

population of each type) to buffer against loss, and considering 

the resilience and resistance of populations to environmental 

disturbances. An appropriate goal is to seek a set of populations 

that represent the diversity of possible conservation values, and 

that can interact, persist, and evolve to maintain those values 

through time.

Two examples from the Coeur d’Alene River, Idaho and 

the Little Snake River, Colorado-Wyoming, illustrate use of 

the framework in markedly different situations. In the Coeur 

d’Alene River, brook trout and rainbow trout have invaded 

some streams, but large amounts of relatively undisturbed 

habitat remain uninvaded. Here, there are opportunities to use 

barriers judiciously to protect genetically pure populations of 

westslope cutthroat trout (O. c. lewisi) with evolutionary values 

from further invasion, and to sustain key ecological and socioeconomic values by either retaining or restoring passage for 

migratory life history forms. In contrast, in the Little Snake River, 

nonnative trout have invaded almost completely so that many 

barriers have been installed in headwater streams to protect 

the evolutionary legacy of remnant Colorado River cutthroat 

trout (O. c. pleuriticus) populations. Unfortunately, there is little 

possibility for these populations to support the full spectrum of 

ecological and socio-economic values without large efforts to 

install barriers and eradicate nonnative trout from downstream 

segments. Because many of the remnant isolated populations 

are small, persistence of native cutthroat trout in the basin may 

depend on maintaining resilient populations through intensive 

habitat protection or restoration, and on reintroductions by humans if they are extirpated. Priorities could focus on removing 

nonnative salmonids to allow extending the larger populations 

downstream and connecting them with other populations.

We developed a decision-space diagram to allow biologists managing a particular basin and set of native salmonid 

populations of conservation value to consider the available 

options and make strategic decisions under different circumstances like those in the examples above. These decisions 

will be influenced primarily by: 1) the degree of isolation of the 

populations of interest and 2) the degree of invasion threat. If 

most remaining native populations are in headwater streams 

and are clearly vulnerable to invasion and displacement, then 

barriers will be required to protect them from rapid extinction. 

Given this, strategic decisions available include the number, 

size, quality, and spatial distribution of habitat patches; the 

representation and redundancy of populations needed to conserve remaining diversity; and effective protocols for removing 

nonnative salmonids, and translocating native salmonids to 

found new populations. The goal should be to develop larger 

and more connected habitat patches and to move the invasion 

threat farther away, even though in many cases this is not 

fully achievable at present. At the other end of the spectrum, 

if a basin has many large interconnected patches of suitable 

habitat with native salmonids and the invasion threat is distant 

or weak, then strategic decisions will involve preventing invasions by controlling the sources and monitoring their spread, 

preventing fish movement barriers and other types of habitat 

fragmentation, and maintaining natural ecological processes 

that provide complex habitats. Other regions of the decision 

space require different strategies for management, or a mix of 

strategies, and additional strategic decisions can be made about 

barrier placement or removal under different scenarios.

Finally, we emphasize that the appropriate use of barriers 

will depend not only on the current threats of invasion and 

isolation, but also may change through time as human values 

shift, as anthropogenic disturbances change, and as climate 

and natural disturbances are altered. Because we don’t fully 

understand the risks associated with barriers and invasion of 

non-native salmonids, most management should be considered an experiment and carefully evaluated to improve future 

efforts. Effective management will require careful monitoring 

of populations, evaluating the results of such management 

experiments (for example, the installation or removal of barriers), and investigating the basic processes of invasions and 

population persistence.

ii
Contents_____________________________________________

 Page

Executive Summary...............................................................................................................i

Introduction...........................................................................................................................1

How to Use this Publication..................................................................................................1

Section I: An Important Problem...........................................................................................2

Section II: Movement and Salmonid Life Histories...............................................................4

Section III: Effects of Isolation ..............................................................................................5

Direct Evidence...............................................................................................................6

Indirect Evidence ............................................................................................................8

Section IV: Effects of Invasions by Nonnative Salmonids ....................................................9

Displacement of Native Salmonids.................................................................................9

Hybridization, Pathogens, and Parasites......................................................................11

Section V: A Framework for Analyzing Tradeoffs................................................................12

Conservation Values.....................................................................................................12

Vulnerability to Invasion................................................................................................16

Isolation and Population Persistence............................................................................19

Considering Priorities among Multiple Populations ......................................................21

Section VI: Two Examples of Invasion and Isolation Tradeoffs ..........................................24

Coeur D’Alene River.....................................................................................................25

Little Snake River..........................................................................................................31

Section VII: Making Strategic Decisions.............................................................................34

Epilogue: What do We Need to Know? ..............................................................................36

References .........................................................................................................................37

Acknowledgments________________________________________

We thank K. Crooks, D. Simberloff, B. Nehring, A. Suarez, B. Noon, D. Peterson, R. Knight, and 

J. Savidge for information and discussion on the invasion-isolation tradeoff, and S. Dekome, E. Lider, 

M. Davis, J. Dupont, and N. Horner for providing information and review for the Coeur d’Alene River 

basin example. S. Barndt, M. Enk, A. Harig, F. Rahel, B. Riggers, K. Rogers, and B. Shepard provided 

constructive reviews that helped improve the manuscript. We also thank Kate Walker of Region 1 of 

the USDA Forest Service for help securing funding, Dona Horan for preparing the maps, K. Morita, 

G. Allison, R. Stuber, R. Thurow, and P. Valcarce for providing images, and R. Dritz for checking 

references. This synthesis was prepared during a sabbatical by K. Fausch at the Boise Forestry 

Sciences laboratory of the Rocky Mountain Research Station, which was generously supported 

by the USDA Forest Service (RJVA 04-JV-11222014-175, administered by B. Rieman), Trout 

Unlimited (administered by W. Fosburgh), and Colorado State University. Finally, we are indebted 

to the many fish biologists and ecologists whose dedicated research over many years provided 

the information needed to develop this synthesis.

iii
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 

Introduction_____________________

	 Native salmonid populations have declined throughout

theworld due to a host of human effects,including habitat

degradation and loss, invasion of nonnative fishes, and 

overfishing. In many regions, suitable coldwater habitats 

for salmonids are found mainly in protected natural 

areas, so that these fishes are increasingly relegated to 

small pieces of their former native ranges. Moreover, 

populations of native trout and charr in these relatively 

undisturbed habitats are often at further risk from invasion by nonnative salmonids, usually from downstream. 

Faced with this dilemma, fisheries managers frequently 

consider using barriersto upstreammovementto prevent

invasion and displacement. However, in doing so they 

face a tradeoff because isolating native salmonid populations in small headwater habitats may also increase 

their risk of extinction.

	 This monograph describes this widespread problem

for stream salmonids, reviews the state of knowledge

aboutthe factorsthat affect extinction risksfrominvasion

versus isolation, and presents a framework for analyzing 

this tradeoff and prioritizing conservation actions. We 

first briefly review the natural population structure and 

life history of stream salmonids, focusing on movements 

thatlinkpopulations acrosshabitatsdispersedthroughout

the riverscape. We then discuss the effects of isolating 

salmonid populations above barriers, which restricts

these movements and may drive populations extinct, 

and the effects of invading nonnative salmonids, which

may move upstream (or downstream in some cases) and 

displace native salmonids.

	 Second, we present a conceptual framework to guide

analysis of the tradeoff, focusing on four main questions 

concerning conservation value, vulnerability to invasion, 

persistence if isolated, and priorities when conservStrategies for Conserving Native

Salmonid Populations at Risk

From Nonnative Fish Invasions:

Tradeoffs in Using Barriers to Upstream Movement

Kurt D. Fausch

Bruce E. Rieman

Michael K. Young

Jason B. Dunham

ing multiple populations. We provide two examples to 

illustrate the use of the framework and show the key 

elements in the decision process, and end by discussing 

opportunitiesformaking strategic decisions when faced

with this tradeoff.

	 This paper is limited in scope to native and non-native

salmonids and the stream networks they inhabit or could 

invade, because salmonids dominate vertebrate communities in coldwater streams and are often the main focus

of management. However, the tradeoff is also relevant to 

other aquatic taxa, both vertebrate and invertebrate, so 

the framework we propose may serve as a foundation for 

considering similar issues for these groups.

How to Use this Publication________

	 Thisreport is written in seven majorsections, to allow

readers with different backgrounds to find information 

quickly. Some points are supplemented with additional 

material that appears in text boxes, and some important

concepts are presented first in the review of knowledge 

and used again in the framework for analyzing the 

tradeoff. Sections are arranged as follows:

	 •	 Section I describes why the tradeoff between invasion and isolation is an important problem for

salmonids worldwide.

	 •	 Section II summarizes the life history and population structure of stream salmonids, and describes

why movement may be critical for their persistence 

and long-term viability.

	 •	 Section III presents what is known about the effects

of isolation by movement barriers on persistence of 

stream salmonid populations.

	 •	 Section IV discusses the effects of invasion by 

nonnative salmonids on native salmonids.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

	 •	 Section V presents a conceptual framework for

analyzing the tradeoff between invasion and 

isolation.

	 •	 Section VI includes two examples with different

constraintsto illustrate the use ofthe framework and

show the key elements in the decision process.

	 •	 Section VII discusses opportunities for making

strategic decisions when conserving sets of native

trout populations in specific watersheds, given different degrees of isolation and invasion threat.

Section I: An Important Problem____

Habitat degradation and loss, invasion by nonnative 

salmonids, and overfishing have reduced the distribution of native salmonids in many regions throughout 

the world to small enclaves of headwater habitat where

they are vulnerable to extinction (Rieman and others 

1997b, Harig and others 2000, Kitano 2004). For example, in many watersheds of the eastern U.S., from 

New England to the southern Appalachian Mountains, 

native brook trout1

 (Salvelinus fontinalis) populations 

are now restricted to headwater streams that are fragmented by dams, heavily sedimented, and close to dense 

human populations (Hudy and others 2004). These 

same watersheds simultaneously have been invaded by 

nonnative rainbow trout (Oncorhynchus mykiss) and 

brown trout (Salmo trutta) that displace brook trout from 

downstream habitats (Larson and Moore 1985). Similar 

combinations of habitat fragmentation and nonnative fish 

invasions endanger native salmonids and other stream

organisms in other regions worldwide (see Box 1). These 

examples illustrate that, although nonnative salmonids

do not invade everywhere they are introduced (Fausch 

and others 2001, Adams and others 2002, Dunham 

and others 2002a), three widespread invaders (brook, 

brown, and rainbow trout) have established reproducing 

populations in many regions outside their native ranges 

(for example, see Lever 1996, Thurow and others 1997, 

Fuller and others 1999, Cambray 2003b), and often 

combine with habitat loss to limit native salmonids or

other native fishes to small headwater streams above 

movement barriers.

1 Fishes of the genus Salvelinus are charr, but here we use the accepted common names for brook trout and bull trout (S. confluentus). Throughout, we use the common names of fishes published 

by the American Fisheries Society (Nelson and others 2004).

Box 1. Examples of habitat fragmentation and 

invasion for stream fishes worldwide.

•	Native charr in northern Japan - Habitats of 

native Dolly Varden (Salvelinus malma) and whitespotted charr (Salvelinus leucomaenis) in Hokkaido 

Island, northern Japan, have been highly fragmented 

by channelization and erosion control dams constructed mainly during the last 35 years (Morita 

and Yamamoto 2002), and downstream populations 

are now being displaced by nonnative rainbow trout 

and brown trout (Takami and Aoyama 1999, Kitano 

2004). 

•	Native brown trout in Switzerland - Habitat for 

native brown trout in tributaries to the Rhine River 

has been fragmented by human activities over a long 

period. These populations now face invasion by 

rainbow trout, which are apparently better adapted 

to the altered flow regime (Peter and others 1998). 

•	Native stream fishes in the Southern Hemisphere 

- Native stream fishes in New Zealand, Australia, 

Venezuela, Chile, Sri Lanka, and South Africa are 

being decimated by habitat alteration and introduced 

brown trout and rainbow trout, and persist primarily 

in small headwater enclaves above waterfalls that

the nonnatives cannot ascend (Townsend and Crowl 

1991, Crowl and others 1992, Pethiyagoda 1994, Ruiz 

and Berra 1994, Pefaur and Sierra 1998, Pascaul and 

others 2002, Cambray 2003a, Gajardo and Laikre 

2003).

•	Other native stream biota - Native stream fishes 

are blocked from Great Lakes tributaries by barriers 

built to exclude nonnative sea lamprey (Petromyzon 

marinus; Smith and Jones 2005), and certain life 

stages of native amphibians and invertebrates that

use stream corridors in the Rocky Mountains are 

also blocked by barriers (Clarkin and others 2003, 

Adams and others 2005).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 

Dams, diversions, and culverts that create barriers to 

upstreammovement are among themost common agents

that fragment habitat. In the conterminous U.S., there 

are estimated to be about 77,000 dams at least 2 m high 

(6 feet) that store at least 62 X 103m

3

 (50 acre-feet) of 

water (U.S. Army Corps of Engineers 2006). This total 

does not include the huge number of smaller dams and

diversions for domestic water supply or micro-hydropower developments, nor the hundreds of thousands of

road culverts that create impassable barriers for fish 

(GAO 2001). In addition, after early settlers degraded 

downstream habitats and overexploited fish populations 

(Wohl 2001), they often translocated fish above natural 

barriers like waterfalls to establish new populations in

headwater habitats. Regardless of whether barriers to 

upstream movement are anthropogenic or natural, they 

fragment stream habitat in watersheds, simultaneously 

protecting native salmonid populations from impending invasions but potentially hastening their extinction 

from environmental catastrophes or small population

effects (Caughley 1994). They also disrupt movements 

of other fishes, amphibians, and invertebrates (Warren 

and Pardew 1998, Vaughan 2002, Clarkin and others 

2003, Adams and others 2005), and may have similar 

effects on these taxa.

In this monograph, we focus on the tradeoff in extinction risks for native salmonids of the inland western

U.S. Native cutthroat trout (O. clarkii) were restricted 

to headwater streams by early mining, logging, grazing, overfishing, and water diversions (Gresswell 1988, 

Young 1995), and many of these remnant habitats were 

subsequently invaded by nonnative brook, brown, and 

rainbow trout that exclude or hybridize with cutthroat 

trout (Figure 1; Behnke 1992, Dunham and others 

Figure 1—Populations of native cutthroat trout 

(top; S. Emmons image) in the inland western 

U.S. have often been invaded by nonnative 

brook trout (bottom; K. Morita image) originally 

introduced in either downstream or upstream 

habitats.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

2002a). Likewise, bull trout and native rainbow trout in 

the Pacific Northwest and Canada face similar problems 

(Rieman and others 1997b, Thurow and others 1997). 

Although this tradeoff is just coming to light in many 

regions of the world, it has become a controversial topic

in the inland western U.S. There, fish biologists that 

seek to conserve native cutthroat trout and bull trout

disagree over the merits of building new barriers to

limit upstream invasions and preserve remaining native

salmonid populations versusremoving barriersto allow

movement of native fish and thereby enhance the chances 

for population persistence (Dunham and others 2002a, 

Peterson and others 2004, Shepard and others 2005).

	 Thisissue is alsoimportantbecause conservationbiologists in general face the same tradeoff – that managing

ecosystems to address habitat fragmentation precludes 

isolating them to prevent invasions. Some conservation 

biologists have proposed connecting fragments of habitat with corridors to reduce extinction risks (Beier and 

Noss 1998, Dobson and others 1999), whereas others 

have argued that increasing connectivity opens the door 

to nonnative species and diseases (Simberloff and Cox 

1987, Simberloff and others 1992, Hess 1994). However, 

there are few well-documented cases of the problem

or potential solutions, even in the extensive literature

on landscape connectivity (Bennett 1999, Soulé and 

Terborgh 1999, Crooks and Suarez in press).

Section II: Movement and Salmonid

Life Histories____________________

Barriers and invasions potentially pose problems for 

salmonids because these fishes often move to fulfill their 

life history. Within a fish’s lifetime, local environments 

often become marginal or inadequate for its needs, 

prompting movement to increase or maintain fitness 

(Thorpe 1994, Brannon and others 2004; see Box 2). A 

general model of life history developed for fishes (Harden 

Jones 1968), and refined for stream fishes (Schlosser and 

Angermeier 1995) and salmonids (Northcote 1997), 

holds that 1) movements link together habitats required 

forspawning, growth, and refuge fromharsh conditions,

and 2) these habitats are often dispersed throughout the 

riverscape (Schlosser 1991, Fausch and others 2002).

	 Although anadromous salmonids are well known for

their spectacular migrations to and from the ocean,

freshwater salmonids also show a great diversity of 

movement behaviors (for example, Meka and others 

2003; see Box 2). Salmonids show remarkable flexibility 

in expression of life-history types, including co-occurring (sympatric) populations of the same species with 

different life histories, alternative life histories within

a common gene pool, and phenotypic plasticity of a 

single genotype under different environmental conditions 

(Hendry and others 2004a, Hutchings 2004). Distinct 

forms that co-occur in the same watershed, or even the

same stream, allow a single species to exploit diverse

Box 2. Fish movement and migration.

Theory: Movement of organisms, both within and 

between generations, is a response to an environment

that is patchy in space or through time (Wiens 1976).

The theory of animal migration is based on the premise 

that organisms move to place each life-history stage 

in suitable habitat to maximize fitness (Northcote 

1978, 1992, Dingle 1996, Hendry and others 2004a, 

2004b). 

Migratory Life History Types: Fishes that migrate

within freshwater are termed potamodromous

(Gresswell 1997), but four additional terms are more 

commonly used to describe the diversity of movement 

behavior in freshwater fish (see below). Nevertheless, 

salmonids display great plasiticity and variation in 

movement behavior, which often defies this simple 

typology.

Lacustrine salmonids complete their life cycle 

entirely within lakes.

Adfluvial fish migrate from lakes into rivers to 

spawn, and juveniles later descend to lakes to grow. 

For example, “coaster” brook trout in tributaries of

Lake Superior (Huckins and others in press), and 

Lahontan cutthroat trout in Pyramid Lake, Nevada 

(Snyder 1917, Behnke 1992), display adfluvial life 

histories, spawning in tributaries but moving downstream to large lakes to grow to large adult sizes.

Fluvial fish complete their life cycle within riverine 

environments, but may use tributaries for spawning 

and juvenile rearing. For example, bull trout in the 

northern Rocky Mountains and Dolly Varden in rivers in northern Japan display fluvial life histories, 

spawning and rearing in tributary streams but moving 

downstream into larger rivers to grow to adulthood

(Rieman and McIntyre 1993, Rieman and Dunham 

2000, Koizumi and Maekawa 2004).

Resident salmonids complete their life cycle entirely 

within tributary streams, but may range widely seeking suitable habitatsforspawning, feeding, or refuges

(Gowan and others 1994, Fausch and Young 1995).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 

environments (Carl and Healey 1984, Bagliniere and 

others 1989, Brannon and others 2004) or persist in the 

face of variable or uncertain ones (Rieman and Clayton 

1997). Populations may include both resident and migratory individuals (Rieman and others 1994, Zimmerman 

and Reeves 2000), and the migratory forms may display 

substantial variation in the extent and timing of movements (Brannon and others 2004). The expression of 

different life-history types is apparently partly under 

genetic control, yet plastic in response to environmental 

conditions (Gislason and others 1999, Hendry and others 2004a), which can produce strong variation in life 

histories both within and among distinct environments

(Taylor 1990, Jonsson and Jonsson 1993, Yamamoto and 

others 1999). Recent evidence also suggests that when a 

species is exposed to a novel or changing environment,

new life histories (for example, resident vs. migratory 

forms, seasonal timing of migration) may emerge relatively quickly, in several decades or less (Näslund and 

others 1993, Quinn and others 2001, Hendry and others 

2004b, Riva Rossi and others 2004).

Movement is important not only in spatially patchy 

environments, but also in those that fluctuate through 

time so that fish are periodically reduced or extirpated in 

some habitat patches. Dispersal that contributes to gene 

flow, provides demographic support (that is, immigration 

that offsets mortality or increases population growth 

rate; see Stacey and Taper 1992, Hilderbrand 2003), 

or allows recolonization can be critical to persistence 

of these populations across generations. Subdivided 

populations that interact strongly through movements 

among them are termed metapopulations (Hanski 1999), 

although salmonid populations are unlikely to fit simple 

theoretical models (see Harrison and Taylor 1997, Dunham and Rieman 1999). Such metapopulation dynamics 

may be particularly important in watersheds where major 

disturbances like drought, flood, fire, and debris flows 

create and destroy habitat, leaving a mosaic of habitats 

in different successional stages (Reeves and others 

1995, Neville and others in press). Salmonid populations that have evolved in these dynamic environments 

may also have evolved diverse life histories, especially 

movement behaviors, that ensure long-term persistence

at larger scales (Thorpe 1994, Heino and Hanski 2001). 

For example, migratory fish may be critical to long-term 

persistence because returning adultsthat fail to home to

their natal watershed disperse and recolonize or support 

other habitats (Quinn 1993, Tallman and Healey 1994, 

Hendry and others 2004b). In other cases diversity in 

migratory timing and the occurrence of multiple age 

classes “at large” in downstream rivers, lakes, or oceans

during major disturbances can allow fish to quickly 

recolonize the disrupted habitats (Rieman and others 

1997a) or move to newly created ones (Leider 1989).

Small-stream salmonids may not exhibit the dramatic 

migrations of forms with access to larger habitats, but

movement is still a critical behavioral component. 

These fishes may move tens to thousands of meters 

daily, seasonally, or during their lifetime to seek suitable habitats. For example, foraging cutthroat trout in 

small streams have been observed to move over 150 

m during the diel cycle (Young and others 1997), and 

spawning runs of such fish have exceeded 2 km (Young 

1996). Movements in fall and winter, associated with 

declining water temperatures and ice formation, have

been of similar magnitude (Brown and Mackay 1995, 

Jakober and others 1998). Moreover, mobility is common 

among cutthroat trout of all ages in small streams, and

is not simply attributable to a few wandering individuals 

(Peterson and Fausch 2003a, Schmetterling and Adams 

2004).

Thus, movement plays important roles in stream salmonid populations, and yet remains poorly described 

(Gowan and others 1994, Rieman and Dunham 2000). 

The advent of advanced technologiessuch asradiotelemetry, PIT tags, and otolith microchemistry has enabled 

recognition of some aspects of trout movements (Rieman 

and others 1994, Fausch and Young 1995, Meka and 

others 2003, Wells and others 2003), but a full understanding awaits sampling that is continuous over scales

ofspace and time broad enough to track individuals of all

life histories throughout their life cycles (Baxter 2002, 

Fausch and others 2002). Such understanding remains a 

central issue for research because movement may be the 

key to population persistence in variable and changing 

environments (Rieman and Dunham 2000).

Section III: Effects of Isolation______

Given the potential for extensive movements by salmonids both within and among generations, it is clear that

barrierslike dams or culverts could extirpatemobile life

history types, restrict populations to habitats inadequate 

for long-term persistence, and prevent recolonization 

in the event of population extinction. Barriers might 

also disrupt movement of other aquatic organisms, and 

potentially change the structure of whole communities. However, even for salmonids, which are relatively 

well studied, evidence for negative effects of isolation

in headwater streams is still emerging. There is some 

direct evidence from studies of barriers in Japan and 

translocation of cutthroat trout in the inland western
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

U.S., and indirect evidence from patch occupancy in 

both countries. There is also indirect evidence from 

molecular genetics and population models, which we

summarize briefly.

Direct Evidence

Direct evidence for negative effects of isolation on 

headwaterstreamsalmonid populations comesfromtwo

studies where barriers were built, or fish were translocated 

above barriers, and the results measured. In both cases, 

investigators fit relationships between watershed area and 

species occurrence, allowing empirical estimates of the

watershed areas required for a specified probability of 

population persistence (for example, 0.50 or 0.90). We 

caution that these estimates may not be easily generalized to other regions, because watersheds that support 

habitat of sufficient size, length, or complexity to sustain 

native salmonid populations can differ markedly in 

area, depending on climate, geology, and other factors. 

In addition, different species, and even subspecies, of 

salmonids may have different life history characteristics that allow persistence despite isolation. Despite 

this complexity, these studies do show that watershed 

area is strongly associated with population persistence, 

and they provide first approximations of relevant size 

thresholds.

Morita and Yamamoto (2002) measured occurrence 

of whitespotted charr in 52 headwater fragments above 

erosion control dams (all >2 m high) in mountain streams 

of southern Hokkaido Island, Japan, and found that both 

watershed area and time since isolation had important

effects on population persistence (Figure 2). The authors 

inferred that all sites had charr populations before the

dams were built, because habitat was no different than

at 32 control sites that supported charr. Seventeen of 

the 52 populations had been extirpated in the 30 years 

since most barriers had been built, and 12 of the 35 

remaining populations were predicted to go extinct in

the next 50 years. Overall, they predicted that a minimum watershed area of 2.3 km2

 is needed to have a 50 

percent chance of sustaining a population for 50 years, 

and their figure shows that about 9 km2 is needed to

have a 90 percent chance. Southern Hokkaido receives 

much more precipitation than the inland western U.S. 

so much smaller watersheds sustain perennial streams. 

However, this study shows that useful relationships 

can be developed from historical records and relatively 

simple field measurements.

	 Further work on extant populations of these charr

revealed additional ecological effects of the barriers. 

Morita and others (2000) reported that growth rates of 

whitespotted charr juveniles in three rivers were higher

above dams than below, likely due to the lower spawning and fry density upstream. This resulted in more fish 

remaining resident than smolting, a common response to

increased juvenile growth rates (Thorpe and Metcalfe 

1998, Hendry and others 2004a). Moreover, growth 

rates of fish transplanted from below to above the dam 

in one of the streams also increased and smoltification 

decreased, indicating that the changes were due to phenotypic plasticity rather than rapid evolution.

In the second study, Harig and Fausch (2002) measured persistence of allopatric greenback cutthroat trout

(O. c. stomias) and Rio Grande cutthroat trout (O. c. 

virginalis) in segments of 28 streams (1.0 to 20.5 km 

long) in Colorado and New Mexico where they had been 

translocated above barriers during the previous 3 to 31 

years. The models they fit predicted that translocations 

in watersheds >14.7 km2 had a >50 percent chance of 

establishinga reproducingpopulation,whereasinsmaller

watersheds populations were more likely to die out or 

remain at low abundance. Watersheds of >33 km2were

predicted to have a >90 percent chance of establishing a 

reproducing population, a goal more in line with many 

conservation assessments. The empirical data used to 

develop the model showed that the 13 stream segments 

where cutthroat trout reproduced consistently were, on 

average, 7.3 km long (range: 1.8 to 20.5) and drained 

watersheds that averaged 23 km2, although 6 of the 

shortest also included lakes.

Virtually all of the historical stream populations of

these two subspecies of cutthroattroutwere isolated from

nonnative salmonid invasions above natural barriers. In 

most cases, they were probably translocated there by 

early settlers. This suggests that at least some populations of these subspecies can adapt to such habitats and

persist for decades. Indeed, Harig and Fausch (2002) 

found that 32 of the 70 historical populations persisted 

in stream fragments with watersheds smaller than 14.7 

km2. However, these may be at greater risk of eventual 

extinction due to insufficient habitat, an effect Tilman 

and others (1994) referred to as an “extinction debt” that 

is yet to be paid (see Hanski 1997 for review). Finally, 

we note that other studies have also reported loss of

stream fish populations above barriers, and reductions 

in fish species richness upstream versus downstream of 

barriers, in other temperate and tropical rivers (Winston 

and others 1991, Fausch and Bestgen 1997, Peter 1998, 

Schmutz and Jungwirth 1999, Pringle and others 2000). 

These observations lend further direct evidence for the

negative effects of isolation.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

Figure 2—Probability of whitespotted charr (top; K. Morita image) persistence above erosion control (sabo) dams (bottom; 

K. Fausch image) in southern Hokkaido Island, Japan, as a function of watershed area (from Morita and Yamamoto 2002; 

used with permission). Curves (center) were predicted from a logistic regression model of presence/absence as a function of 

watershed area, time since damming, and stream gradient for 52 sites above dams. The family of curves in the figure predicts 

the probability of population persistence in headwater streams that have been isolated for different time periods (given average 

gradient). Watersheds >10 km2

 have a high probability of sustaining these charr populations for 50 years, according to the 

model, but different salmonid species in different environments may have markedly different relationships.
 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

Indirect Evidence

	 Several lines of indirect evidence also indicate that

small watershed size and isolation reduce persistence of 

headwater salmonid populations. This evidence comes 

primarily from studies of patch occupancy (patches are 

defined as continuous networks of suitable stream habitat; 

Dunham and others 2002b), but other indirect evidence 

also comes from colonization after barrier removal, and 

genetics and population models (see Box 3). In the first 

study of patch occupancy for salmonids, Rieman and 

McIntyre (1995) reported that bull trout in the Boise River 

basin in Idaho had approximately a 50 percent chance 

of occurrence in patches >25 km2 (which support about 

5 km of stream habitat), and more than an 80 percent 

chance in patches >80 km2 (about 9 km of stream habitat). 

This suggested that populations in smaller patches were

more vulnerable to local extinction than those in larger

patches. In a second study, Dunham and others (1997) 

reported that Lahontan cutthroat trout (O. c. henshawi) 

Box 3. Additional indirect evidence on isolation.

Genetics: Isolation of small populations can cause loss of genetic diversity 

and random genetic drift (Rieman and Allendorf 2001). In turn, this can lead 

to a loss of phenotypic variation and evolutionary potential (Allendorf and 

Ryman 2002), and potentially to reduced effective population size (Waples 

2004) and inbreeding depression, both of which could hasten extinction of 

salmonid populations isolated above barriers (for example, Soulé and Mills 

1998, Frankham 2005). However, these mechanisms have not been demonstrated in headwater stream salmonids, and there has been only limited work 

to discriminate effects of habitat size and isolation. Examples include:

•	Yamamoto and others (2004) found lower genetic diversity and increased 

genetic drift in whitespotted charr above barriers versus below in three watersheds in Hokkaido.

•	Wofford and others (2005) reported that genetic diversity decreased and 

genetic drift increased for coastal cutthroat trout (O. c. clarki) isolated above 

more versus fewer barriers in a small headwater stream in Oregon.

•	Neville and others (in press) reviewed evidence for salmonids, and found 

that barriers were consistently associated with lower genetic diversity within 

populations. 

Population models: Population models have been used to 1) estimate total 

population sizes and compare them to theoretical guidelines, and 2) simulate 

the effects of isolation, population size, survival, and environmental variation 

on persistence and genetic variation.

•	 Rieman and Allendorf (2001) used individual-based models to estimate 

population sizes needed to sustain genetic variation in bull trout. They argued 

that few local populations would maintain genetic variation indefinitely 

without gene flow from surrounding populations. 

•	Morita and Yokota (2002) used an individual-based model to predict 

persistence of whitespotted charr populations. Persistence increased with 

larger populations and higher adult survival, but decreased substantially for 

small populations after 30 to 100 years in the model, suggesting that smaller 

populations are at a higher risk of eventual extinction.

•	Hilderbrand (2003) modeled inland cutthroat trout populations and found 

that persistence increased with increases in population size, immigration 

from adjacent source populations, and juvenile survival, and with decreases

in environmental variation.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 

in the Great Basin were more likely to occur in patches 

that were connected to other occupied patches along

stream networks. This suggested that populations were 

more likely to persist in less isolated patches. Further 

research in the Boise River basin showed that both larger 

patch size and smaller distance to occupied patches were 

important predictors of bull trout presence (Dunham and 

Rieman 1999). In both locations, disturbances like floods, 

droughts, fires, and debris flows can extirpate these two 

salmonids from whole stream segments (Rieman and 

Clayton 1997, Dunham and others 2002b), so movement 

is required to recolonize habitat fragments and maintain 

local populations.

	 Three recent studies have also shown the importance

of larger connected patches in supporting headwater

salmonid populations. Rich and others (2003) found 

that bull trout were more likely to be present in tributaries of Montana watersheds where there was a strong 

source population in the mainstem river downstream. 

Koizumi and Maekawa (2004) reported that Dolly 

Varden were more likely to occur in larger spring-fed 

tributaries of a Hokkaido river that were closer to other 

occupied tributaries than in smaller and more distant

ones. Young and others (2005) used data from extensive 

electrofishing surveys to calculate that streams of 3.4 to 

12.9 km in length were needed to sustain 500 to 5000 

age-1 and older greenback or Colorado cutthroat trout

(O. c. pleuriticus). These population size thresholds have 

been proposed as sufficient to maintain genetic diversity 

(n=500) and long-term evolutionary potential (n=5000; 

Allendorf and others 1997, Hilderbrand and Kershner 

2000, Young and Harig 2001), rather than simply to 

ensure relatively short-term persistence (that is, 50 to 

100 years). It is typical that half or fewer of the extant 

populations of native cutthroat trout in various regions

are found in streamslong enough to support populations

of these sizes (see Kruse and others 2001, Young and 

others 2002, Young and Guenther-Gloss 2004, Young 

and others 2005). Nevertheless, indirect evidence from 

field studies like these is particularly relevant because 

they encompassed the large spatial scales at which land 

managers must set priorities and make decisions about

resource management (Fausch and others 2002).

Once habitats are made available by removing barriers, recolonization of unoccupied habitat appears likely. 

The many examples of barrier removal followed by rapid 

and extensive upstream colonization in fishes (Hart and 

others 2002 and citations therein) serve as circumstantial evidence that renewed access to headwater habitats

above dams, particularly spawning areas, may increase 

the persistence of downstream populations (Smith and 

others 2000). Similarly, streams that have lost fish populations because of drought, severe floods, fire-related 

disturbances, or pollution are often quickly repopulated 

(Larimore and others 1959, Sigler and others 1983, 

Fausch and Bramblett 1991, Rieman and Clayton 1997, 

Roghair and others 2002), and habitats available for the 

first time (for example, following glacial retreat; Milner 

and Bailey 1989) typically develop fish populations in 

a predictable sequence. In some circumstances, fishes 

require extended periods to colonize newly available 

habitats downstream (Hepworth and others 1997), or 

may be more likely to colonize other habitats following 

displacement by floods (Leider 1989) or in response 

to high densities (Gowan and others 1994, Isaak and 

Thurow 2006).

Section IV: Effects of Invasions by

Nonnative Salmonids_____________

	 The other main factor in the tradeoff addressed here

is upstream invasion by nonnative trout that may reduce 

or eliminate native salmonids in headwater streams. As 

we described above, rainbow, brown, and brook trout

have been transported and released in most countries

and regions that have suitable coldwater habitat for salmonids (Lever 1996, Rahel 2000). In the western U.S., 

many of these introductions began in the 1870s with 

the arrival of large numbers of Euroamerican settlers,

who further spread these fishes by railroad, horseback, 

and on foot (Wiltzius 1985, Alvord 1991, Wiley 2003). 

State and federal agencies assumed responsibility for 

most stocking thereafter (Wiltzius 1985), which still 

continues in some parts of this region (Epifanio 2000). 

For example, brook trout were first released in the 11 

western states between 1872 and 1912 (MacCrimmon 

and Campbell 1969) and are still stocked in 7 of them 

(Dunham and others 2002a).

Displacement of Native Salmonids

Where barriers to upstream movement are lacking, 

nonnative salmonids like brook trout can move rapidly 

upstream and colonize habitat occupied by native salmonids (Gowan and Fausch 1996, Peterson and Fausch 

2003a). There have been multiple observations of rapid 

declinesin native salmonid populations after these nonnative trout invasions,such as declines of cutthroat trout

after nonnative brook trout and rainbow trout invasion

in western U.S. streams (Gresswell 1988, Behnke 1992, 

Young 1995). These effects have been studied in detail 

using both experimental and correlative approaches. 
10 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

At the scale of 1- to 10-km stream segments, several 

field removal experiments in which all age classes of 

the native and invaded populations were measured provide strong evidence of invader impacts. Peterson and 

others (2004) found that abundance of native Colorado 

River cutthroat trout fry and juveniles increased after 

four years of brook trout removal in a Colorado stream, 

compared to a control stream, because survival of age-0 

cutthroat trout increased 13 fold and that of age-1 fish 

doubled after brook trout removal. In contrast, no such 

increase occurred in another pair of streams at higher

elevation, because cutthroat trout recruitment often

failed, apparently due to cold temperatures (Coleman 

and Fausch 2006). Surprisingly, survival of adult cutthroat trout was unaffected by brook trout removal in 

either set of streams.

	 Several other nonnative trout removal experiments

conducted at the stream segment scale have also been

followed by increased native trout abundance. Moore 

and others (1983, 1986) removed nonnative rainbow 

trout from entire lengths of four streams above barriers

in Great Smoky Mountains National Park for 6 years. 

They reported increases in abundance of native brook 

trout compared to three control streams where abundance changed little. Shepard and others (2002) found 

that a westslope cutthroat trout (O. c. lewisi) population 

increased seven fold after removing brook trout from

above a barrier in a Montana stream for 8 years. The 

decade-long removal of brook trout with piscicides and

electrofishing from a stream in Crater Lake National 

Park, Oregon resulted in an almost three fold increase 

in abundance of native bull trout (Buktenica and others 

2001, Renner 2005). However, encouraging anglers to 

remove brook trout from an Alberta stream inhabited

by native bull trout and westslope cutthroat trout was 

ineffective and failed to reduce brook trout numbers

(Stelfox and others 2001).

	 Strong declines in native trout after nonnative trout

invasion have led investigators to propose competition

and predation as likely mechanisms. These mechanisms 

have been studied extensively in controlled laboratory 

and field experiments at small scales (see reviews in 

Fausch 1988, 1998, Dunham and others 2002a, Peterson and Fausch 2003b), but results are often difficult 

to generalize to larger spatial and longer time scales 

(Rieman and others 2006). An exception is experiments 

designed to explicitly consider environmental factors like 

temperature across a range of natural conditions, either

in the field (McHugh and Budy 2005) or by simulating 

conditions in the laboratory (Taniguchi and Nakano 

2000).

Finally, field sampling at various scales provides correlative evidence of the effects of nonnative trout invasions. For example, brook trout appear to have displaced 

cutthroat trout in a Nevada watershed (Dunham and others 1999), and bull trout in an Idaho watershed (Rieman 

and others 2006), where the native and nonnative species 

co-occurred. However, the effects were highly variable 

among streams. Rich and others (2003) showed that bull 

trout occurrence in small low-gradient streams below

barriers in a Montana watershed was strongly associated 

with complex habitat structure and strong neighboring

populationsthat could provide demographic support, but

was negatively associated with presence of brook trout 

(Figure 3). Townsend and Crowl (1991) and McIntosh 

(2000) reported strong negative correlations between 

brown trout and native galaxiid abundance in stream

segments below barriers throughout two New Zealand 

watersheds, and provided experimental evidence that

predation by large brown trout was the cause. At the 

scale of channel units, Morita and others (2004) used 

an extensive field survey to provide evidence that brown 

trout excluded nativewhitespotted charrfromindividual

pool-riffle sequences in a Hokkaido stream, whereas 

rainbow trout apparently excluded the charr from pools 

to adjacent riffles. These data at relatively small spatial 

scales support conclusions about the effects of these

nonnative trout invasions in Hokkaido derived from 

broad field surveys (Takami and Aoyama 1999, Kitano 

2004).

In contrast to examples where nonnative salmonids 

clearly displaced native trout, other cases show that nonnative salmonids apparently failed to invade all habitat, 

or had little effect on native trout when they did invade. 

Fausch (1989) reported that brook trout failed to invade 

certain high-gradient streams with cutthroat trout in

several Rocky Mountain regions (see also Rieman and 

others 1999, Paul and Post 2001). He suggested various 

hypotheses including limited upwelling groundwater 

required for egg incubation and winter floods that scour 

eggs or fry. Likewise, Adams and others (2002) found that 

brook trout invasion had apparently stalled in more than 

half the tributaries studied in a pristine Idaho watershed, 

based on comparison with historical data (see Section V 

for more details). Novinger and Rahel (2003) found no 

increase in native cutthroat trout in four small Wyoming 

streams isolated by barriers after 5 to 8 years of removing 

brook trout by electrofishing, suggesting that the brook 

trout had little effect. They attributed this primarily to 

lack of critical resources or abiotic conditions needed

for cutthroat trout recruitment and survival (see Harig 

and Fausch 2002, Sloat and others 2005, Coleman and 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 11

Fausch 2006), or to lack of demographic support from 

adult spawners. Fausch and others (2001) found that 

rainbow trout invasion in various regions worldwide

was most successful where flow regimes were similar to 

those in their native range, and reported that this widely 

introduced species was apparently unable to invade in at 

least one region where flooding was frequent during fry 

emergence. Although there has been no comprehensive 

analysis to date of locations where nonnative salmonids 

have failed to invade, and the environmental correlates,

such research is clearly needed to explain differences 

in invasion success and provide tools for managing the

invasion-isolation tradeoff.

Hybridization, Pathogens, and

Parasites

Hybridization is an additional process by which nonnative salmonids can adversely affect native salmonids 

(Allendorf and others 2001, 2004). Hybridization has 

obvious genetic influences, but ecological effects are also 

possible. Hybrids of many salmonid crosses are fertile 

(Suzuki and Fukuda 1971), which leads to backcrossing 

of hybrids with parental types and gene flow between 

genetically distinct populations, subspecies, or species, 

a condition called introgression (Rhymer and Simberloff 

1996). If backcrossing is extensive and hybrid individuals 

also mate among themselves, the resulting population

is termed a hybrid swarm. For example, introgressive 

hybridization of native cutthroat trout with rainbow trout 

is widespread throughout the western U.S. (Leary and 

others 1998, Peacock and Kirchoff 2004). Weigel and 

others (2003) reported that hybridization of westslope 

cutthroat trout with rainbow trout was more prevalent in

the downstream segments of an Idaho watershed, despite 

an apparent absence of migration barriers. They inferred 

that harsh abiotic factors upstreamlimited rainbowtrout

more than cutthroat trout. In contrast, Hitt and others 

(2003) reported that hybrids between the two species 

were spreading upstream throughout a large Montana 

watershed, and inferred that introgressionwould expand

because it appeared that hybridization was limited more 

by demographic factors like movement than abiotic factors that could limit spread.

Other cases suggest that hybridization can be restricted. 

For example, when brook trout invade and mate with

bull trout, hybrids and subsequent backcrosses appear to 

have reduced fertility or survival (Kanda 1998, Kanda 

and others 2002). Although such hybridization may not 

result in hybrid swarms (Kanda and others 2002), it does 

represent wasted reproductive effort by the native fish. 

Because invading brook trout can mature at an earlier

age and have higher reproductive potential (Power 1980, 

Kennedy and others 2003), hybridization could still lead 

to displacement of bull trout (Leary and others 1993, 

Rieman and others 2006). Recent evidence indicates 

that maintenance of large migratory bull trout may 

limit hybridization with smaller resident Dolly Varden 

Figure 3—Bull trout presence in Montana 

and Idaho watersheds is often positively 

associated with complex physical habitat 

and strong populations in neighboring 

watersheds. However, bull trout (fish at 

bottom) presence has also been found to 

be negatively associated with brook trout 

(top) in certain watersheds, suggesting that 

this nonnative salmonid either uses different 

habitats than bull trout or displaces them 

(K. Morita image).
12 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

through size-assortative mating (Redenbach and Taylor 

2003). Assortative mating and differential habitat use 

(B. Rieman and J. Dunham, personal observation) might 

similarly limit hybridization between large migratory 

bull trout and smaller resident brook trout, further

highlighting the potential importance of migratory life 

histories and stream interconnections in maintaining

the integrity of native salmonid populations.

Last, invading nonnative trout infected with pathogens 

or parasites may transmit them to native salmonid populations. For example, the myxozoan parasite Myxobolus 

cerebralisthatcauseswhirlingdiseasehasaresistantspore

stage that can remain latent for decades (El-Matbouli 

and others 1992), and is transmitted to trout fry through 

oligochaete (Tubifex) intermediate hosts. Nehring and 

Thompson (2003) and Nehring (2004) reported invasion of the parasite 25 km upstream in one Colorado 

watershed, and 4 km upstream in two smaller streams, 

apparently via rainbow trout and brown trout movement. 

In the first case, this movement brought the parasite to 

within about 3 km of a population of native Colorado 

River cutthroat trout. In contrast, Peterson and others 

(2004) found no evidence that brook trout invading native 

cutthroat trout habitat in four other Colorado streams

were infected with the parasite, but the distance to an

infected source population was unknown.

Section V: A Framework for

Analyzing Tradeoffs_ _____________

	 The information presented above makes clear that

fisheries managers often face a dilemma: nonnative 

trout invasions can reduce or extirpate native salmonid

populations, yet barriers to prevent these invasions can 

lead to extinction of native salmonids because of isolation. However, the outcomes of invasions and isolation 

differ among species and stocks of native and nonnative

salmonids due to their evolutionary history, and among 

locations due to habitat characteristics, environmental

variation, and time since isolation. This means that installation or removal of barriers to promote persistence of

native salmonid populationsis often a complex problem

that must be analyzed in the context of these ecological 

and evolutionary constraints. Such efforts will also be 

influenced by conservation goals and the available opportunities and management resources. Here, we propose 

a conceptual framework that explicitly considers these 

factors. We outline a process that incorporates the preceding synthesis of existing knowledge as well as the 

experience of field biologists. We illustrate the concept 

with examples from two basins in the Rocky Mountains 

that have markedly different settings and constraints, 

and end by considering an overall strategy for making 

decisions.

Four key questions provide a framework for considering the use or removal of barriers to conserve native

fishes in any stream or network of streams.

1. Is a native salmonid population of important conservation value present?

2. Is the population vulnerable to invasion and displacement by nonnative salmonids?

3. If the native salmonid population is isolated, will 

it persist?

4. If there are multiple populations of value, which 

ones are priorities for conservation?

Although these questions are relatively simple, each 

leads to more questions or dimensions that must be 

considered in any decision (Table 1; see Dunham and 

others 2002a), so we address each in turn.

Conservation Values

A first step in developing any native salmonid management plan is to consider the conservation values

associated with the populations of interest (Table 1). 

Without a clear concept of the relative values of different 

populationsthat could bemanaged, it isimpossible to set

specific management objectives. What constitutes value 

varies markedly among government and tribal agencies, 

those who study or manage native fish species, and the 

interested public. For natural resource management 

agencies, conservation of biodiversity is a widely held 

goal, and the conservation literature presents a diverse

set of guidelines toward that end (see Box 4).

We suggest that three general classes of conservation 

value emerge from the literature (for example, Angermeier and others 1993): evolutionary, ecological, and

socio-economic (Figure 4). Evolutionary conservation

values are considered explicitly by the U.S. Endangered Species Act (ESA), which calls for designating 

evolutionarily significant units and distinct population 

segments (Waples 1995). Here the focus is on distinct 

species, stocks, and populations that have evolved in,

and adapted to, unique and varied environments. Genetic and ecological divergence are important criteria

(Allendorf and others 1997) and the full representation 

of diversity at regional scales (104 to 105 km2) has been 

a focus of ESA-related management efforts to conserve

evolutionary legacies (Waples 1995, U.S. Fish and Wildlife Service 1998b). Some species such as westslope cutthroat trout and bull trout can show substantial genetic

and phenotypic divergence among tributaries within a 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 13

Table 1—Four questions to consider when evaluating the tradeoff of managing native 

salmonid populations using barriers to isolate them from invasions (after 

Dunham and others 2002a, S. Dunham unpublished). The answer to each 

question depends on further dimensions or considerations (right column) for a 

more complete analysis of the tradeoff.

Main questions Further dimensions or considerations

1. Is a native salmonid population of •	 Evolutionary values

important conservation value present?	 •	 Ecological values

•	 Socio-economic values

2. Is the population vulnerable to invasion •	 Transport and spread

and displacement by nonnative salmonids?	 •	 Establishment

•	 Effects 

	 	o	Displacement

	 	o	Coexistence

	 	o	Hybridization

	 	o	Transmit parasites or pathogens

3. If the native salmonid population is •	 “Small population” phenomena

isolated, will it persist? o	Loss of genetic variability

	 	o	Demographic stochasticity

	 	o	Environmental stochasticity

	 	o	Catastrophes

•	 Loss of migratory life histories

•	 Synergistic factors

4. If there are multiple populations of value, •	 Identify conservation units

which ones are priorities for conservation?	 •	 Prioritize populations

	 	o	Representation

	 	o	Redundancy

	 	o	Persistence

	 	o	Feasibility

•	 Consider viability

•	 Confront uncertainty

single river basin (102 to 103km2), arguing for conserving 

distinct populations at this scale as well (Allendorf and 

Leary 1988, Leary and others 1998, Quinn and others 

2001, Hendry and others 2004b). From this perspective, 

evolutionary conservation values are highest for distinct 

populations that represent a substantial or unique portion 

of extant genetic diversity, occupy unique or remnant 

habitats, or exhibit diverse or unusual life histories

(Allendorf and others 1997, Halupka and others 2003, 

Shepard and others 2005).

Ecological conservation values include important

ecological processes and functions at the population,

community, or ecosystem levels of organization. For 

example, species like salmon and trout can have strong

effects on trophic networks in aquatic systems (Simon 

and Townsend 2003), which can even cross the ecosystem 

boundary into adjacent terrestrial ecosystems (Spencer 

and others 1991, Willson and Halupka 1995, Baxter and 

others 2004, Koyama and others 2005). As already discussed, some populations may be the key to persistence 

of others through dispersal, gene flow, demographic 

support, and metapopulation-like processes (Schlosser 

and Angermeier 1995, Rieman and Dunham 2000). 

Thus, dispersal from stronghold or source populations

may allow the entire network to persist in the face of 

disturbance or changing environments. Similarly, the full 

expression of life-history diversity within populations 

may confer resilience to environmental disturbance (Rieman and Clayton 1997, Brannon and others 2004) and 

in some cases resistance to invasion. Ecological value 

can be far more varied than these simple examples, but

a key point is that the structure and function — that is, 

the composition and roles of individuals or populations

— are ecologically important. We may value essential 
14 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

Figure 4—Distribution of three conservation values in 

salmonid populations in natural and disrupted stream 

systems. In natural systems (top) all values may be 

represented in the same species distributions. In 

systems disrupted by human development and the 

introduction and invasion of non-native salmonids 

(bottom), all of these conservation values often cannot 

be supported in the same streams or fish communities. 

For example, isolation of genetically pure native trout 

populations may conserve primarily evolutionary 

values (solid line), but limited ecological functions and 

values (dashed line). There is currently much debate 

about whether hybrid trout retain significant ecological 

functions and values, but most are considered to 

have limited or no evolutionary value. Hatchery trout 

in a put-and-take fishery could have significant socioeconomic value but no evolutionary or ecological values. 

This simple classification is not mutually exclusive 

(for example, isolated pure trout may also have socioeconomic conservation values) and different values 

may overlap depending on the system and the species 

and forms involved. However, in disrupted systems 

the potential to represent all conservation values 

simultaneously is likely to be far more limited than in 

natural systems.

Box 4. Components of value for biodiversity.

Different authors have proposed different systems for 

considering the requirements for, and values of, sustaining biodiversity:

•	Noss (1990) and Groves (2003) proposed that sustaining biodiversity requires conserving three attributes of 

ecosystems: composition, structure, and function. Composition includes the unique elements such as species or 

genes. Structure represents the distribution, interconnections, and interactions of those elements. Function 

represents the key processes or ecological products or 

influences that result (for example, secondary production).

•	Noss and Cooperider (1994) described four explicit 

values of biodiversity: direct utilitarian (for example, 

harvestable protein), indirect utilitarian (for example, 

clean water), recreational/aesthetic (for example, fishing), and intrinsic spiritual/ethical aspects. 

•	Allendorf and others (1997) focused on biological 

values defined by the potential evolutionary and ecological consequences of extinction. Thus, extinction 

may result in loss of evolutionarily distinct, rare, and 

genetically diverse populations, as well as loss of 

members of unique communities or environments, or 

of “conservation umbrella” species. 

•	Callicot and Mumford (1997) outlined a classic debate in natural resource and biodiversity 

management: whether ecosystem sustainability might 

be met through either of two potentially distinct objectives - conserving ecosystem integrity or conserving

ecosystem health. Integrity refers to the native species 

complex, interactions, and diversity, whereas health 

describes a system that still retains its function and 

organization but not necessarily the native elements. 

The differences between the two have proven controversial. Although we can conserve ecological processes 

without the native community, functional ecosystems 

might still be substantially impoverished (Callicott 

1995, Groves 2003). Nevertheless, conservation of 

ecological processremains an important challenge that

should not be overlooked, even when there islittle hope

of conserving native diversity or integrity represented 

by “ghost” or “relict” species (Meyer 2004).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 15

ecological services that are provided, or the ability of an 

ecosystem and its elements to organize, reorganize, and 

respond to changes by rebounding, adapting, and evolving. From a management perspective, we value species, 

populations, and communities that are self-sustaining,

resilient, and adaptable because they don’t require continual input of resources and energy to maintain them. 

However, a key issue is that although evolutionary and 

ecological values sometimes overlap, they are not the 

same thing (Callicott 1995, Callicott and Mumford 1997). 

For example, a genetically introgressed cutthroat trout 

population could retain significant ecological value, but 

have little evolutionary value (Figure 4; Allendorf and 

others 2004). From this perspective, ecological conservation values are higher for populations that are large,

productive, and interconnected, and fill roles similar to 

those in the original ecosystem. Likewise, they are less 

for populationsthat are vulnerable to extinction without

extraordinary conservation effort, or those that threaten 

the persistence of ecosystem structure and function that 

we hope to maintain (for example, fish populations that 

are vectors of disease).

Socio-economic conservation values are perhaps the

most evident to fisheries managers who are anxious to 

conserve the recreational and economic benefits from 

fishing, tourism, and associated activities. These values 

may be more obvious because they can be quantified 

in economic terms, capture the public interest, and

often dictate the budgets and activities of management

agencies. They are also important because they can 

leverage conservation resources for other species and

populations that may hold significant evolutionary and 

ecological values. For example, conserving salmon may 

also protect lamprey (Petromyzontidae; Lee and others 

1997, Thurow and others 1997). Like the first two values, socio-economic values can also vary among local 

populations of conservation interest.

Recognizing distinct evolutionary, ecological, and socio-economic conservationvaluesfor aparticularpopulation or set of populations is important because they may 

afford very different opportunities for conservation and 

the implementation or removal of barriers. In some large 

wilderness rivers such as the Middle Fork Salmon River 

in Idaho or the South Fork Flathead River in Montana, 

undeveloped watersheds and native species complexes

clearly support substantial values in all three categories 

(Figure 5). But habitat and populations in most river 

systems throughout the region have been substantially 

altered, so that convergence of all three conservation

values in any single population is rare (Figure 4). For 

example, main-stem populations of cutthroat trout may 

be large and composed of individuals with mobile life

Figure 5—In relatively pristine wilderness river basins like the Middle Fork 

Salmon River, Idaho, shown here 

(Sulphur Creek, Frank Church River 

of No Return Wilderness; R. Thurow 

image), native salmonid assemblages 

can support substantial evolutionary, 

ecological, and socio-­economic values 

simultaneously. However, most river 

systems have been substantially altered, 

so convergence of all three conservation 

values in any single location is rare.
16 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

histories, but tend to be introgressed and co-occur with

nonnative salmonids,whereas headwater populations are

generally small and isolated yet offer the best prospects 

for being genetically pure. As suggested above, the mainstem populations may hold significant ecological value, 

but little evolutionary value, whereas the opposite may 

be true of the headwater populations. In these cases, 

the prioritization of conservation efforts is challenging 

because the tradeoffs are difficult or uncertain. Should 

we favor populations that represent remnant elements

of evolutionary history or those that have the potential 

to evolve and adapt in the future (Meyer 2004)? Should 

we favor locally hybridized populations that may retain 

some elements of local adaptation (Dowling and Childs 

1992, Peacock and Kirchoff 2004), or should we replace 

them with a genetically pure stock that represents many 

pooled populations (Leary and others 1998)? Should 

we maintain access for large migratory fish that support a key fishery even though local populations will be 

vulnerable to invasion and hybridization, or should we 

isolate the local populations to conserve genetic purity 

(Allendorf and others 2004)?

It is clear that all three conservation values are important depending on the context, and often fisheries 

managers will be faced with a mix of objectives and

possibilities. Opportunities to conserve evolutionary 

values, for example, may be very restricted and viewed 

as immediate priorities in a few distinct areas if they 

are to be conserved at all (see Shepard and others 2005). 

Therefore, conservation of native, wild locally-adapted 

populations with diverse life history forms will often be 

a primary goal. Yet as Meyer (2004) suggested, even in 

cases where we cannot conserve an intact evolutionary 

legacy, we may still be able to conserve some ecological processes and evolutionary potential. Toward this 

end, sustaining wild slightly-hybridized native trout 

populations that retain much evolutionary potential to 

fill important ecological roles will often be a second 

priority, followed by wild populations of non-native trout 

that support important ecological and socio-economic

values. Finally, in habitats that cannot sustain wild 

trout populations, usually due to lack of reproduction 

or recruitment, stocking hatchery fingerlings or adults 

may be desired to support valuable fisheries, so long as 

these do not threaten ecological or evolutionary values 

of other populations (for example, by introgressive 

hybridization, or transmitting parasites or pathogens). 

This hierarchical approach embodies the principle that

certain values are irreplaceable and should not be forfeited, but that inability to conserve one set of values 

does not eliminate the possibility of realizing others.

Vulnerability to Invasion

If a native salmonid population of important conservation value is present, the second question to address is 

whether it is vulnerable to invasion and displacement

by nonnative salmonids (Table 1). The invasion process 

has been divided into three or four basic steps (Vermeij 

1996, Kolar and Lodge 2001, Dunham and others 2002a), 

which can be summarized as transport, establishment, 

spread, and impacts. For simplicity, we discuss transport 

and spread together, since they have the same effect of 

bringing nonnative salmonids in contact with native

ones. Given this, assessing vulnerability to invasion 

requires asking three main questions about the stages 

of invasion, shown in Table 2.

Transport and spread—Whether nonnative salmonids are transported to specific locations by humans will 

depend on interest in them for recreation or food, and

distances to sources of fish for introduction. Although 

most agencies now evaluate fish stocking more carefully than in the past (Rahel 1997, 2000), unauthorized 

introductions by the public are burgeoning. For example, 

Rahel (2004) calculated that unauthorized introductions 

of nonnative fishes in five regions throughout the U.S. 

accounted for 90 percent of the new introductions during 1981 to 1999, compared to only 15 to 43 percent 

during all previous periods. Disgruntled anglers may 

translocate nonnative fishes back to their favorite fishing sites above barriers after they have been removed 

using chemicals or electrofishing (U.S. Fish and Wildlife 

Service 1998a, Rahel 2004). Established populations 

of nonnative salmonids close to native fish populations 

(for example, just downstream from accessible barriers, 

or a short distance by road) provide a source for such 

unauthorized introductions (Thompson and Rahel 1998, 

Dunham and others 2004, Rahel 2004, Munro and others 2005).

	 The spread of nonnative species without the aid of

humans will depend on distance from a source population and the time available for colonization, and on 

their tendency for long-distance movement, their rate of 

population growth, and the resistance of the corridor to

fish dispersal (With 2002, Hastings and others 2005). If 

invasions spread as an advancing front, they will arrive 

sooner when source populations are close. However, 

nonnative salmonids like brook trout and brown trout

can also move long distances (Young 1994, Gowan and 

Fausch 1996, Peterson and Fausch 2003a) and can have 

high intrinsic population growth rates (McFadden 1961, 

Power 1980, Kennedy and others 2003), both of which 

increase invasion rates (Kot and others 1996, Lewis 1997, 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 17

Hastings and others 2005). Therefore, trout invasions 

appear to advance through a combination of frequent 

short distance movements at the upstream margin

(Rodríguez 2002), combined with less frequent longdistance movements, termed “jump dispersal” (Peterson 

and Fausch 2003a). For example, direct measurements 

of brook trout movement by tagging and recapture in 

Colorado streams revealed that they ranged up to 2 to 

3 km within 100-day summer sampling periods, with a 

bias toward upstream movement (Riley and others 1992, 

Gowan and Fausch 1996). Brown trout movements of 

similar distances, and three that reached 23 to 96 km, 

were detected by telemetry in Wyoming streams (Young 

1994).

	 Stream segments with high-gradient reaches or

waterfalls may provide resistance to invasion from 

downstream, but source populations in headwater lakes

or stream reaches can spread downstream throughout

watersheds. Several syntheses of data from the Rocky 

Mountains suggested that brook trout are less successful at establishing populations in streams with gradients steeper than 4 to 7 percent (Fausch 1989, Rieman 

and others 1999). However, a mark-recapture study in 

Idaho showed that brook trout ascended slopes of 13 to 

22 percent, although immigrants in steeper channels 

were larger fish and moved more slowly (Adams and 

others 2000). In experimental flumes, adult brook trout 

>15 cm could jump over waterfalls up to about 75 cm 

high from plunge pools >40 cm deep (Kondratieff and 

Myrick 2006), but field results suggest that they may also 

be able to ascend higher barriers either by penetrating 

interstices at low flow (Thompson and Rahel 1998) or 

finding alternate routes at higher flow (Adams and others 

2000). In contrast, Adams and others (2001) found that 

brook trout dispersed downstreamfromheadwaterlakes

where they were stocked, through high-gradient channels 

up to 80 percent slope and over an 18-m high waterfall, 

potentially invading many reaches inaccessible from 

downstream. Likewise, Paul and Post (2001) reported 

that brook trout introduced in an Alberta watershed

spread downstream disproportionately from stocking 

locations to colonize reaches at lower elevations. There 

is also some evidence from northern Montana suggesting 

that hybrids of rainbow X cutthroat trout may disperse 

more widely than native cutthroat trout (Hitt and others 

2003), thereby increasing spatial spread. Rahel (2004) 

emphasized that fisheries managers should assume that 

any nonnative species that is introduced and becomes 

established in a basin will eventually colonize other 

suitable and accessible habitats.

Table 2—A hierarchy of three main questions to assess whether nonnative salmonids will invade and have 

negative effects on populations of native salmonids (after Vermeij 1996, Dunham and others 

2002a). The three main stages of invasion considered are influenced by additional factors shown 

on the right.

Main questions about the invasion process Factors influencing invasion process

1. Are invaders likely to be •	 Proximity of sources for unauthorized introductions 

transported or spread to the location •	 Distance to source population and time for invasion

that native salmonids inhabit? •	 Tendency for frequent or long distance dispersal 

•	 Population growth rate

•	 Resistance of the stream corridor to invasion

2. Can the invaders establish a •	 Environmental resistance - habitat suitability

reproducing population? o	Temperature regime

	 	o	Flow regime

	 	o	Stream size

	 	o	Gradient

	 	o	Productivity

•	 Propagule pressure 

•	 Biotic resistance from native biota

3. Will the invaders displace native •	 Biotic interactions and demographic rates may 

salmonids by competition, predation, change with abiotic factors (for example, 

introgression, or transmission of temperature and flow)

parasites or pathogens, or will they coexist? •	 Habitat modification may favor invaders

•	 Introgressive hybridization may be irreversible

•	 Parasites and pathogens may be dispersed by invaders
18 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

Establishment—All three main salmonid invaders 

have established reproducing populations in most U.S. 

states where they are not native (Rahel 2000), although 

they have not invaded all waters to which they have 

access (Fausch and others 2001, Adams and others 

2002, Dunham and others 2002a). Establishment of 

a nonnative salmonid population will depend on both

environmental resistance (habitat suitability) and “biotic 

resistance” by native species, as well as chance events 

(Moyle and Light 1996). Habitat suitability for native 

and nonnative salmonids in streams is often governed

by temperature regime (Meisner 1990, Rahel and Nibbelink 1999, Harig and Fausch 2002, Dunham and others 

2003a), flow regime (Jowett 1990, Jager and others 1999, 

Strange and Foin 1999, Fausch and others 2001), stream 

size (Rahel and Nibbelink 1999, Rieman and others 

1999), habitat factors correlated with gradient (Fausch 

1989, Adams 1999), and stream productivity (Kwak and 

Waters 1997). Because many of these factors influence 

fish at thresholds, measurements across a wide range of 

each variable often show that establishment is relatively 

predictable from simple physiological tolerances (see 

Box 5). In other cases, simultaneous consideration of 

a host of variables may be required (Moyle and Light 

1996, Marchetti and others 2004).

	 A main predictor of success of deliberate introductions or translocations of nonnative speciesis “propagule

pressure” (Kolar and Lodge 2001, Hilderbrand 2002, 

Marchetti and others 2004), which refers to the number of 

releases and the number of individuals per release. These 

data indicate that repeated stocking of nonnatives can

eventually increase likelihood of establishment. In part, 

this may be due to rare events, such as years with unusual 

weather, which favor nonnative species and allow them

to establish “beach head” populations and subsequently 

invade (Moller 1996, Moyle and Light 1996).

Intact assemblages of native fishes may also provide 

biotic resistance that prevents nonnative salmonidsfrom

invading, especially when the assemblages include more 

species. Current theory in invasion biology holds that 

assemblages with more species provide fewer “niche

opportunities” for invaders (Shea and Chesson 2002, 

Chase and Liebold 2003), defined as opportunities for 

invading species to increase from low densities due to

abundant unused resources or a lack of natural enemies. 

This theory may explain why brook trout in western 

North America are least successful at invading Pacific 

Northwest coastal watersheds that support intact runs

of salmon and anadromous trout (Oncorhynchus spp.) 

with many diverse life history types that use a wide 

variety of food and space resources.

Box 5. Examples of environmental resistance that 

limits invading salmonids.

Several examples indicate that abiotic factors like

low temperature and flooding regimes can limit 

establishment of nonnative salmonids, but each may 

interact with other factors.

Low temperature: Low temperature limits salmonid 

recruitment by delaying egg development and hatching which causes higher egg mortality (Stonecypher and others 1994), and by reducing fry growth 

and lipid storage which causes higher overwinter

mortality due to starvation (Biro and others 2004, 

Coleman and Fausch 2006). Low temperature also 

limits growth of adults, which reduces fecundity 

and recruitment.

•	Rahel and Nibbelink (1999) found that upstream 

limits of brown trout in a Wyoming watershed could 

be explained by a combination of cold temperature 

and small stream size. 

•	Adams (1999) reported that reduced fecundity at 

low temperatures could account for the upstream

limit of brook trout in a Montana stream.

•	Clark and others (2001) inferred that lower 

fecundity due to lower temperatures limited invading rainbow trout in the northern Appalachian

Mountains.

Flooding regime: Flooding during incubation or

emergence of fry from gravel redds may limit establishment of nonnative salmonids. 

•	Nehring and Anderson (1993) and Latterell and 

others (1998) found that recruitment of nonnative 

brook, brown, and rainbow trout in sets ofColorado

rivers was lowest in years when snowmelt runoff 

flows were highest, probably because fry emerge 

just before or during these flow peaks and are 

displaced.

•	Fausch and others (2001) reported that nonnative 

rainbow trout were least likely to establish reproducing populations in several regions worldwide

where flooding occurred during fry emergence in 

early summer, and most likely to establish in regions 

where flow regimes most closely matched those in 

their native range along the Pacific coast of North 

America. 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 19

Displacement, coexistence, and other effects—In 

many cases, after nonnative salmonids are transported 

orspread to a streamand establish a reproducing population, they reduce or displace native salmonids (Waters 

1983, Behnke 1992, Peterson and others 2004), but in 

other cases the invasion may stall or the two may coexist 

(Shepard 2004, Rieman and others 2006). Adams and 

others (2002) reported that upstream limits of brook 

trout distribution and recruitment advanced over a 25-

year period in fewer than half the streams resampled 

in an undisturbed Idaho watershed. Whether displacement or coexistence occurs may depend on abiotic 

factors like temperature that affect biotic interactions

among individuals at the local scale, or on temperature

and flow regimes that affect demographic responses 

at larger scales. For example, at the local scale, DeStaso and Rahel (1994) reported that brook trout were 

behaviorally dominant over Colorado River cutthroat 

trout at warmer temperatures in an artificial stream, 

whereas neither was dominant at colder temperatures

(but see Novinger 2000). However, it is more likely that 

the effects of temperature and other abiotic factors will

play out at riverscape scales by changing demographic 

rates. For example, low temperature may reduce growth 

or fecundity of nonnative salmonids and thereby limit 

their latitudinal distribution (Clark and others 2001) or 

upstream invasion (Adams 1999, Weigel and others 2003, 

Benjamin 2006). Alternatively, changes in environmental 

conditions that favor productivity of non-native species 

might allow them to invade new habitats and displace

native species. Moller and Van Kirk (2003) reported 

that nonnative rainbow trout have been more successful

at invading and displacing native Yellowstone cutthroat

trout (O. c. bouvieri) where flow regimes were modified to favor rainbow trout over cutthroat trout. Shepard 

(2004) suggested that brook trout were more likely to 

displacewestslope cutthroattroutwhere habitat had been

degraded by land management activities. Likewise, Olsen 

and Belk (2005) suggested that native fishes were less 

likely to coexist with non-native predators such as brown 

trout if habitat had been degraded so that off-channel

refugia were lost. These examples also suggest that judicious management of flow and temperature regimes, and 

physical habitat structure, to achieve more natural states 

might favor native species (Poff and others 1997, Fausch 

and others 2001), and allow some level of coexistence 

with nonnative species. This idea would benefit from 

careful study using an adaptive management approach 

at watershed scales (see Walters 1986).

Two other possible effects of invaders are hybridization 

and introduction of parasites or pathogens. Although 

invasions of nonnative salmonids can sometimes be

eradicated (Moore and others 1983, Knapp and Matthews 1998), especially if caught in the early stages 

(Simberloff 2003), neither hybridization nor introduction of parasites or pathogens to a native salmonid

population is usually reversible. Only in cases where 

hybrids have substantially lower survival or fertility, as 

perhaps for brook trout hybridizing with bull trout (see 

above), could hybridization be reversed. Introductions of 

parasites like Myxobolus cerebralis are also apparently 

irreversible once established, and may be transmitted 

farther by movements of invading salmonids upstream 

or downstream from infected source populations (see 

above; Nehring and Thompson 2003, Nehring 2004). 

However, whether conditions are favorable to persistence and expansion of pathogens may depend on local 

environmental conditions, and even past watershed

management (Anlauf 2005).

Isolation and Population Persistence

If a native salmonid population is vulnerable to a 

nonnative salmonid invading from downstream, one

management alternative is to construct a barrier to fish 

movement to prevent the invasion. However, then the 

influence of isolation caused by the barrier on persistence of the native salmonid population becomes an

important question (Table 1). Isolation or fragmentation 

of habitat has several important effects on native salmonids (Table 3; Rieman and Dunham 2000, Dunham 

and others 2003b), including declines in population and 

habitat size. Small populations can be more vulnerable 

to extinction because of a loss of genetic variability, 

random changes in demographic processes and rates

(demographic stochasticity), and environmental fluctuations (environmental stochasticity), collectively known 

as “small population” phenomena (Caughley 1994). 

Large-scale perturbations or “catastrophes” that severely 

reduce populations and habitats can be important for

both small and large populations, particularly if they are 

confined to a set of spatially restricted habitats that could 

be affected by the same disturbance, such as hurricanes, 

earthquakes, fires, floods, droughts, and temperature 

extremes (Propst and others 1992, Brown and others 

2001). Moreover, disturbances that pose little threat to 

larger, interconnected populations may become a threat 

when populations are highly fragmented (Dunham and 

others 2003b).

Isolation also eliminates migratory life histories, 

which confer resilience to local populations. Populations 

with migratory individuals are spread across many different habitats and may therefore be less vulnerable to 
20 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

disturbances influencing a single habitat. For example, 

short-term local extirpations of cutthroat trout by drought 

(Dunham and others 1997) and bull trout by wildfire 

(Rieman and others 1997a) were likely reversed by returns of migratory fish that were in habitats outside of 

the systems during the disturbance. Isolation or other 

influences that lead to loss of migratory life histories 

restrict a population spatially and thus increase the vulnerability to local disturbances. Isolation also reduces 

the opportunity for exchanges of individuals among 

local habitats, which may reduce the demographic stability of some populations (Hilderbrand 2003). Finally, 

migratory fish are typically larger and more fecund, 

which may provide important demographic support to 

existing populations, and possibly resistance to invading 

salmonids (Dunham and others 2002a).

Threats posed by isolation cannot be considered alone. 

Synergistic feedbacks among threats can lead to the 

demise of local populations (for example, an “extinction vortex,” Gilpin and Soulé 1986). Many factors can 

modify population responses of salmonid fishes to isolation and strongly influence local persistence, including 

angling, habitat degradation, and habitat and life history 

characteristics. The negative effects of angling on native 

trout populations can be substantial, even if regulations

restrict harvest (Post and others 2003), particularly for 

large migratory fish that are highly desirable to anglers. 

Habitat degradation often leads directly to habitat fragmentation in stream networks, and also reducessurvival

and population size making populations less resilient and 

more vulnerable to other threats. Habitat degradation can 

influence salmonids at a variety of spatial scales. Local 

factors that are strongly altered by land management, 

such as temperature, sediment, stream discharge, and

instream cover, are commonly cited as degrading habitat 

(Bjornn and Reiser 1991). At broader scales, the status of 

Table 3—Basic threats to the persistence of single, isolated salmonid populations, and 

guidelines for addressing each threat (see Rieman and McIntyre 1993, Allendorf 

and others 1997, McElhany and others 2000 for more detailed criteria). Criteria 

for multiple populations are given elsewhere (see Box 6).

Threat Guidelines

Loss of genetic variability Maintain enough adults to minimize loss of genetic 

variability through drift and inbreeding.

Loss of resilience Maintain or restore the condition of local habitats or 

reduce sources of mortality to compensate for lost 

fecundity and potential productivity due to loss of 

migratory adults.

Demographic stochasticity Maintain enough adults and year classes to minimize 

the probability that chance demographic events drive 

local populations extinct.

Environmental stochasticity Maintain a large enough population size and wide 

distribution within the isolated habitat to persist in the 

face of common environmental fluctuations.

Ensure that human influences do not result in habitat 

degradation or substantial changes to environmental 

regimes that reduce survival.

Catastrophes Maintain a large enough habitat and population size, and 

wide distribution, to ensure resilience in the face of rare, 

high-magnitude environmental changes.

Ensure that human influences do not alter the frequency 

and magnitude of natural catastrophes in a way that 

threatens remnant populations, or create new catastrophes 

(for example, toxic chemical spills, introductions of invasive 

species) that pose threats.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 21

salmonid populations has been consistently associated 

with the distribution of land management activities, including road density and management intensity (Thurow 

and others 1997, Baxter and others 1999). Habitat and 

life history characteristics can also influence the probability of salmonid population persistence in the face of 

isolation. For example, Koizumi and Maekawa (2004) 

observed that Dolly Varden charr were more likely to 

persist in smaller isolated habitats if they were strongly 

influenced by groundwater, which provided more stable 

and favorable conditions over time. Rieman and Dunham 

(2000) argued that cutthroat trout were more resistant to 

local extinction in small habitat patches than were bull

trout. These examples illustrate that other environmental 

conditions and species characteristics may mitigate or 

exacerbate the effects of isolation.

Considering Priorities among Multiple

Populations

	 To this point, ourframework has addressed the tradeoff

of invasion and isolation for single populations. Fisheries 

managers and conservation biologists have no lack of

problems to consider at this scale, but often move from

one project (that is, the implementation or removal of 

a barrier) to the next as the need or opportunity arises. 

Typically, this is done while struggling with limited 

financial, capital, and human resources. However, to 

be most effective it is important to avoid incremental

or “ad hoc” solutions and carefully prioritize use of 

limited resources (Pressey 1994, Allendorf and others 

1997, Groves 2003). This can be facilitated through a 

systematic assessment of risks (Francis and Shotton 

1997). Characterizing risk implies understanding: 1) 

the probability of an event (for example, a nonnative 

salmonid invasion), 2) the effects of the event if it actually occurs (for example, probability of ecological displacement or genetic introgression following invasion), 

and 3) the value that could be lost. Thus, a “threat” is 

a factor that could negatively impact a population, and 

a “risk” is the probability that a given factor will have

an impact on something we value.

	 Asimple example canillustrate these three components

of risk for native cutthroat trout. First, the probability of 

invasion may be higher for cutthroat trout populations 

closer to sources of invasion by brook trout, increasing 

the first component of risk. Second, the effects of invasion may be greater for some cutthroat trout populations 

because they are less resistant to invasion, either due to 

their weaker population status or because they live in 

habitats that favor brook trout, increasing the second

component of risk. Third, the conservation value of some 

cutthroat trout populations may be higher than others 

because they represent relatively rare pure genomes or 

life history types, increasing the third component of risk. 

In contrast, populations farther from invasion sources, 

more resistant to invasion, or of lower conservation

value would be at lower risk. Therefore, these would be 

lower priorities as candidates for constructing barriers

to prevent invasions. These are simple examples, and 

tradeoffs can become more complicated as multiple factors and populations, spatial structure, and the general

principles of conservation planning are considered.

The first step in planning is to identify the fundamental units of conservation and determine how they are 

arranged across spatial scales. Native salmonids often 

occur as collections of local populations distributed

across habitats that are nested within larger subbasins

(Rieman and Dunham 2000). This creates a hierarchy of 

spatial structure at scales ranging from local “patches”

of suitable habitats (Dunham and others 2002b) to patch 

networks, and even larger aggregations or regional

networks (Table 4). Moreover, some habitats may be 

potentially occupied but currently vacant. We emphasize 

that “scale” as used here should be based on how fish 

populations are organized across the landscape, and not 

on a fixed spatial scale or stream network structure. In 

many cases, the natural structure of local populations 

has been modified by human activities, such as barriers 

that fragment habitats. In other cases for rare taxa, only 

one local habitat exists, such as for the Paiute cutthroat 

trout (O. c. seleniris). Whatever the case, it is important 

to carefully consider the biology of the target species and 

how it interacts with the larger landscape, to define local 

populations, habitat patches, patch networks, or other

units for use in conservation planning (Dunham and 

others 2002b, Fausch and others 2002, Wiens 2002).

	 The second step is to establish priorities for management of the potential conservation units. Ultimately, 

managers will need to decide on the number, distribution, and characteristics of populations they will attempt 

to conserve and the actions required (that is, installing 

or removing barriers). The conservation literature is replete with proposals outlining goals for such efforts, but

the most common include representation, redundancy, 

persistence, and feasibility (Scott and Csuti 1997, Groves 

2003). Moreover, a formal decision process must also 

acknowledge that setting priorities is inherently uncertain 

(Ludwig and others 1993, McElhany and others 2000). In 

practice, these goals do not always translate neatly into 

specific guidelines (see Box 6), but we discuss each separately here to emphasize their individual importance.
22 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

Table 4—Potential relationships between conceptual scales at which natural populations of salmonid 

fishes are structured (Dunham and others 2002b) and units of conservation as defined in 

practice (following the Draft Bull Trout Recovery Plan; U.S. Fish and Wildlife Service 2002).

Draft Bull Trout

Conceptual Scale Recovery Plan Description

Patch Local Population A patch may be occupied or not at present, 

whereas local populations are based on

occupancy. Occupied patches and local

populations are characterized by frequent

(daily to seasonal) interactions among individuals.

Patch network Core Area Local aggregation of patches or local

or metapopulation populations characterized by less frequent

(annual to decadal) interactions.

Subbasin Recovery Unit Naturally discrete aggregations of patch

networks or core areas within larger drainage

basins that potentially interact over long time

scales (100s to 1000s of years).

Region Distinct Population Major biogeographic units that characterize

Segment distinct evolutionary lineages.

Representation refers to how well local populations

and habitats reflect the full suite of ecological and evolutionary conditions that are expressed within a given 

area (Lawler and others 2003, Ruckelshaus and others 

2003). In conservation planning, it is nearly always the 

case that only some of the “pieces” can be protected or 

restored. Accordingly, it is important to identify and 

protect those that provide the most complete representation of the original population characteristics (for 

example, allele frequencies, life history types, species 

occupancy or composition) and habitat conditions (for 

example, spatial distribution, successional stage), as well 

as the conservation values defined above.

Redundancy is a key component of conservation planning because no local population or habitat isimmune to

extinction or catastrophic events (Moyle and Sato 1991). 

Therefore, it is prudent to “not put all your eggs in one 

basket,” and instead to conservemultiple examples of all

the representative pieces. Although many have struggled 

to identify the minimum number and distribution of 

local populations necessary to insure persistence of a 

metapopulation (see Rieman and McIntyre 1993) that 

number will always depend on the local environments 

and events that are difficult to predict. In general, more 

populations will be necessary as individual populations 

become more vulnerable (for example, smaller), the 

potential for catastrophic disturbances increases, or the

potential for simultaneous disturbance across sets of

local populations increases (Hanski and Gilpin 1997). 

Replicate populations are especially important when 

all are isolated above barriers, to allow reintroduction

if local extinctions occur but habitats remain suitable

(Lubow 1996).

Persistence is a function ofthe resilience ofindividual

populations, as well as the larger network of populations (Table 4). Factors contributing to the persistence 

of individual populations in the face of isolation and

invasion were considered above, and have been addressed in detail for salmonids (Rieman and McIntyre 

1993, Allendorf and others 1997, McElhany and others 

2000). Populations that appear unlikely to persist may 

be low priorities for conservation actions. Persistence 

of multiple populations must usually be considered 

at several scales. For example, the dynamics of local 

habitats and populations may create a metapopulation, 

which may be nested within a larger collection of core 

areas that characterize a recovery unit, and ultimately 

a distinct population segment (DPS) or evolutionarily 

significant unit (ESU; Table 4). Whereas the spatial 

structure and temporal dynamics may be difficult to 

quantify precisely (Rieman and Dunham 2000, Fausch 

and others 2002), it is important to consider these factors because persistence of any single population may 

depend strongly on influences from adjacent populations 

or habitats. We strongly encourage readers to consult 

key references (Rieman and McIntyre 1993, Allendorf 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 23

Box 6. Guidelines for viability of multiple salmonid populations (partially excerpted and modified from 

McElhany and others 2000; see also Rieman and others 1993, Allendorf and others 1997).

•	 Some populations should exceed minimal guidelines for long-term viability. Larger and more productive 

(“resilient”) populations may be able to recover from a catastrophic event that would cause the extinction of smaller 

populations. Furthermore, such populations may serve as important sources of dispersing fish to less productive 

habitats. A collection of local populations that contains some populations with high abundance and productivity is 

less likely to go extinct in response to a single catastrophic event that affects all populations. It should be recognized 

that populations or habitats serving as sources and sinks may exchange roles over time.

•	 Populations that display diverse life-histories and phenotypes should be maintained. When local populations 

each have a fair degree of life history diversity (or other phenotypic diversity), a collection of local populations is 

less likely to go extinct as a result of correlated environmental catastrophes or changes in environmental conditions 

that occur too rapidly for an evolutionary response. In addition—assuming phenotypic diversity is caused at least 

in part by genetic diversity—maintaining diversity allows natural evolutionary processes to operate. Note that to 

protect genetic diversity, it may be necessary to maintain several populations with the same phenotype. 

•	 Habitat patches should not be destroyed faster than they are naturally created. Salmonid habitat is dynamic; 

suitable habitat is continually being created and destroyed by natural processes. Human activities should not decrease 

either the total area of habitat OR the number of habitat patches. 

•	 Maintain some habitat patches that appear to be suitable or marginally suitable, but currently contain no 

fish. In the dynamics of natural populations, there may be time lags between the appearance of empty but suitable 

habitat (by whatever process) and the colonization of that habitat. If human activity is allowed to render habitat 

unsuitable when no fish are present, the population as a whole may not be sustainable over the long term.

•	 Natural rates of straying among subpopulations should not be substantially increased or decreased by 

human actions. This guideline means that habitat patches should be close enough together to allow dispersal and

expansion of the population into under-used patches during times when productivity is high. Also, dispersal should 

not be much greater than natural levels because increased dispersal may reduce a population’s viability if fish wander 

into unsuitable habitat or interbreed with genetically unrelated fish.

•	 Some populations should be geographically widespread. Spatially correlated environmental catastrophes are 

less likely to drive a widespread collection of local populations to global extinction. 

•	 Some populations should be geographically close to each other. On long temporal scales, having populations 

geographically close to one another facilitates connectivity among existing populations. Thus, a viable collection 

of local populations requires both widespread AND spatially close populations.

•	 Populations should not all share common catastrophic risks. A collection of local populations that do not 

share common catastrophic risks is less likely to be driven to extinction by correlated environmental catastrophes. 

Maintaining geographically widespread populations is one way to reduce risk associated with correlated catastrophes, but factors unrelated to spatial proximity may be important. For example, geographically distant populations 

with similar vulnerability to broad-scale climate influences may share common catastrophic risks.

•	 Evaluations of status should take into account uncertainty about among-population processes. Our understanding of spatial and temporal process and interactions among local populations is very limited. Because most 

collections of local populations of salmonid fishes are believed to have been historically self-sustaining, the historical 

number and distribution of populations serves as a useful “default” goal, unless better evidence is available.
24 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

and others 1997, McElhany and others 2000) to develop 

individual applications. Although throughout this monograph we address persistence of populations, we realize 

that the broader goal of viability may more often be the 

conservation target of interest (see Box 7).

Feasibility meansthat conservation actionsshould be

cost effective, sustainable, and socially and environmentally acceptable. If all else is equal, projects leading to 

the most efficient implementation or removal of barriers 

should be those that will produce the greatest benefit (for 

example, reduction in risk) for the least cost. Cost must 

be considered not only for the immediate project, but 

also for long-term maintenance. Constructing barriers or 

replacing impassible road culverts with ones that allow

fish passage is costly, so only a few can be implemented 

each year. Efforts to control or eradicate nonnative 

salmonids have proven expensive and are not always 

successful (Dunham and others 2002a), so thoughtful 

evaluation of the cost-effectiveness of these efforts will

be important. Constructing or removing barriers can 

have other environmental effects as well. Barriers to 

non-native fish are also barriers to other aquatic organisms that may need to move to persist and maintain 

productive aquatic and riparian communities. Barriers 

also disrupt flows of water and materials (for example, 

sediment) that shape upstream and downstream habitats 

or affect floodplains and human infrastructure.

Finally, it is clear that conservation planning decisions 

are inherently uncertain. For example, our understanding 

is very limited about where brook trout will invade, how 

long isolated populations of native fishes will persist, 

and how these differ among different environments and

species or subspecies. This creates genuine disagreement among biologists who have experience in different

regions. Although these limitations can be frustrating, 

we will never have all the answers and decisions must

be made, so confronting uncertainty must become part 

of the management process (Ludwig and others 1993, 

Francis and Shotton 1997, McElhany and others 2000). 

Ludwig and others (1993) suggested that managers should 

favor decisions that are robust to uncertainty (that is, 

the outcome is likely to be favorable regardless of our 

knowledge). When this is impossible, it is important to 

hedge, probe the system to learn, and favor reversible 

actions. For example, hedging could include maintaining redundant populations and constructing barriers in

some systems but not others to learn more about what 

happens, through careful monitoring. Some barriers 

may also be viewed as short-term, reversible actions to 

prevent impending invasions,while learningmore about

the threats of isolation. We are uncertain when the effects 

of barriers become irreversible because, for example,

migratory forms are lost. Recent work suggests that the 

evolution or re-expression of migratory life histories 

and colonization of newly accessible habitats can occur 

quickly (Hendry and Stearns 2004, Quinn 2005). However, until research or management experiments clearly 

demonstrate that the effects of isolation can be reversed,

any intentional isolation must be viewed as a calculated 

risk, balanced against the threat of invasion.

Section VI: Two Examples of

Invasion and Isolation Tradeoffs____

We use the Coeur d’Alene River basin in northern 

Idaho and the Little Snake River basin in southern 

Wyoming and northern Colorado (Table 5) as examples 

of relatively complex and simple systems, respectively, 

to show how our conceptual framework can be used to

Box 7. Persistence vs. viability.

Often the terms “persistence” and “viability” are 

used interchangeably, but here we recognize viability 

as a larger conservation objective. The basic impetus 

for conservation planning is not simply to guarantee 

persistence of a species, but to ensure that natural

ecological and evolutionary processes are allowed 

to continue and perhaps change through time. For 

a single species, this broader view of maintaining

process, not just persistence, is referred to here as

“viability” (see McElhany and others 2000). In the 

prioritization process, this may equate to conservation of both evolutionary and ecological values 

simultaneously. For example, evolutionary values 

can be associated with genetically pure but isolated 

populations that may persist for some time, but if 

those populations cannot evolve and adapt with

changing environments they may not be viable in 

the long term. Populations that are likely to persist 

and remain viable represent a higher overall value

and logically a higher priority in any assessment of 

risk. In short, persistence is generally viewed as a 

necessary but not sufficient objective for attaining 

full conservation of a species.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 25

consider the tradeoffs in managing invasions and isolation. There are many factors that affect conservation of 

native salmonids, including the history and extent of 

landmanagement, location of natural and anthropogenic

barriers to fish movement, introduction and invasion of 

nonnative salmonids, and goals of fisheries managers. 

These issues vary markedly across the western U.S., and 

it is beyond our scope to explore them in great detail 

here. Our goal is to use the framework outlined above 

to consider the opportunities and priorities that emerge

in each case. However, we caution that these are not 

intended to be specific proposals for management.

Coeur D’Alene River

Background—The upper Coeur d’Alene River basin 

in northern Idaho (Figure 6) has two major branches, 

the North Fork and Little North Fork Coeur d’Alene 

rivers, each with numerous 2nd- to 4th-order tributaries. 

The basin has relatively low topographic relief (approximately 600 m), but is influenced by a mixture of climatic 

processes. The hydrograph is predominantly driven by 

melting snow, but the Pacific maritime influence and 

the relatively low elevation of much of the basin result 

in some winter rain. Rain-on-snow floods during winter 

Table 5—Aspects of the physical characteristics, trout populations, and management issues in the 

Coeur d’Alene River basin, Idaho, and Little Snake River basin, Colorado-Wyoming.

Characteristic Coeur d’Alene River Little Snake River

Basin area (ha) 232,112 288,344

Elevation range (m) 600 to 2,100 1,900 to 3,300

Period of peak precipitation (form) November to January March to April 

(snow and rain) (snow)

Primary land management activities mining, logging grazing, water development

Cutthroat trout subspecies westslope cutthroat trout Colorado River cutthroat trout

Life histories present resident, fluvial, adfluvial resident

Cutthroat trout distribution (km) 1,063 160 to 190

Other sympatric native salmonids bull trout1 mountain whitefish1

mountain whitefish

Sympatric nonnative salmonids brook trout, rainbow trout none2

Brook trout distribution (km) 92 unknown but widespread

and much greater than 

cutthroat trout

Ranking of cutthroat trout socio-economic > evolutionary > ecological > 

population values ecological > evolutionary socio-economic

Management emphases maintain fishery for create and enlarge remnant

fluvial and adfluvial fish cutthroat trout populations

Management tactics restrict harvest, improve remove brook trout, install

habitat, remove barriers, barriers, translocate 

install barriers, stock cutthroat trout

sterile rainbow trout

1

Neither cutthroat trout subspecies currently coexists with these salmonids.

2

Colorado River cutthroat trout largely persist only above barriers preventing invasions of nonnative trout, but individuals occasionally drift downstream over barriers to form sympatric assemblages.
26 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

storms are more common than in much of the range of

westslope cutthroat trout. There are no headwater lakes 

in the basin that serve as sources for downstream invasion or modify the hydrologic and thermal regimes of 

downstream habitats.

Disruptive management and development have been 

extensive. Hard rock mining, dredging, log drives down 

streams, clearcut logging, extensive road building, and

overfishing have degraded habitats and depressed trout 

populations, especially in the lower half of the basin. 

In addition, road crossings have created barriers to 

fish movements in many tributaries. Native salmonids 

included westslope cutthroat trout, mountain whitefish 

(Prosopium williamsoni), and bull trout, although bull 

trout are now presumed to be extirpated. Because of its 

unusual geomorphic and climatic characteristics (for 

example, heterogeneous geology and frequent winter 

flooding), the Coeur d’Alene River basin could represent 

a unique evolutionary template for westslope cutthroat 

trout (for example, Allendorf and others 1997). Resident 

and migratory trout occur throughout the basin, including fluvial and adfluvial fish that reach 350 to 450 mm 

(14 to 18 inches). Brook trout and rainbow trout were 

introduced between 1900 and 1950, and both have become established in parts of the system. Hybrid swarms 

of cutthroat trout X rainbow trout are common in the 

lower tributaries, but hybrids declined or were absent 

entirely in samples upstream.

Figure 6—Predicted distribution of spawning and rearing habitat for westslope cutthroat trout (bold stream 

segments), and road culverts believed to be complete barriers to all fish passage (dots). Spawning and rearing 

habitat was predicted from associations between young-of-the-year cutthroat trout and stream size (Dunnigan 

1997, Abbott 2000), gradient, and confinement (Moore and Gregory 1988, Lentz 1998), and from direct field 

observations of spawning adults (J. Dupont, Idaho Fish and Game, personal communication). Barriers were 

identified by the Panhandle National Forests stream crossing inventory. The dotted line encompasses the area 

believed to support cutthroat trout populations that have not been hybridized by rainbow trout. Some segments 

above barriers outside this area may also retain pure cutthroat trout. One headwater tributary in the northeast 

portion of the basin has hybridized trout and so was excluded from this area.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 27

Restrictive angling regulations (that is, size limits, gear 

restrictions, and catch-and-release) were first imposed on 

portions of the basin in the 1970s. Hatchery rainbow trout 

have been stocked in the main-stem rivers for decades,

but less so in recent years. Now, only sterile fish are 

planted to reduce the risk of further hybridization. Habitat 

management has focused on road obliteration, channel

reconstruction, and barrier removal. The trout populations 

have responded and the upper river has recently become 

a renowned fly fishing destination for wild trout.

Conservation values—Evolutionary, ecological, and 

socio-economic conservation values are all considered

important in the Coeur d’Alene River basin, but the last 

two are dominant management objectives. A productive, diverse, and resilient population of trout with a

large migratory component is an emphasis of natural 

resource managers. The sport fishery for large cutthroat 

trout in the main-stem rivers plays a significant role 

in the economy of local communities, particularly as 

traditional extractive industries (for example, timber 

harvest and mining) are replaced by tourism.

Conservation of evolutionary value (which might be 

compromised by hybridization) has been a secondary 

concern, primarily because hybridized trout that may 

be morphologically indistinct from genetically pure cutthroat trout are widespread and have been considered

part of the formally recognized taxonomic group (U.S. 

Fish and Wildlife Service 2003, Campton and Kaeding 

2005). There is still controversy on this issue (Allendorf 

and others 2004, 2005). Little work has explored the different ecological roles of cutthroat trout, rainbow trout,

or their hybrids, but abundant populations of the latter 

two can become established in habitats once occupied

by cutthroat trout. A general assumption seems to be 

that rainbow trout or their hybrids may persist, evolve 

ecological roles, and support fisheries that are surrogates 

for native cutthroat trout, with little loss of the associated ecological functions and values (see Quist and 

Hubert 2004). However, although native and non-native 

trout are likely to have similar roles at some levels of 

ecological organization (McGrath 2004), in other cases 

non-native trout have very different effects (Baxter and 

others 2004). Little research has been done to clarify 

the ultimate ecological implications of introgression or

displacement of cutthroat trout by other salmonids, so 

some undefined risk remains.

	 To illustrate the fundamental units of conservation and

the distribution of local populations needed to maintain

all conservation values,wemapped thepotential distribution of spawning and rearing habitat for cutthroat trout. 

We assumed that natal habitats define local populations 

(Table 4), and represent the critical habitats vulnerable 

to invasion and isolation. We used simple models based 

on landscape and channel characteristics associated

with the distribution of juvenile fish (Dunnigan 1997, 

Abbott 2000) and observations of spawning areas by 

local fisheries biologists. We concluded that stream segments potentially important to conservation of ecological 

and socio-economic values occur in tributary systems 

throughout the basin (Figure 6).

Segments potentially important for maintaining evolutionary value are more restricted. Extensive hybridization 

is common in the lower tributaries, but pure populations

are found in the upper portions of both the Little North 

Fork and North Fork subbasins (B. Rieman, unpublished 

data). Therefore, it seems likely that genetically pure 

populations of westslope cutthroat trout important for

conserving evolutionary values also support socioeconomic and ecological values in the upper part of

the basin. Road crossings have also created barriers to 

movements of trout in many tributaries (Figure 6), and 

these may have temporarily prevented the invasion of 

rainbow trout and the extent of hybridization. Genetically 

pure populations of cutthroat trout may persist above 

some barriers, but surveys are needed to confirm this.

Vulnerability to invasion—Brook trout are well 

established in some tributaries of the Little North Fork 

and the lower North Fork Coeur d’Alene rivers. To 

consider habitats most vulnerable to future invasion

and displacement of westslope cutthroat trout, we used

an approach similar to that described above to map

potential habitat for brook trout (Figure 7). Our predictions suggest that brook trout could ultimately occupy 

and potentially displace cutthroat trout from portions of 

many tributaries throughout the basin. If the invasion 

proceeded to the ultimate limitsit could leave habitatfor

cutthroat trout highly fragmented. A complete invasion 

would threaten all three conservation values associated

with the existing trout populations, although brook trout

might still fill some of the ecological roles played by 

salmonids in these streams (see McGrath 2004). The 

most vulnerable habitats are tributaries adjacent to, or

habitats upstream of, current populations of brook trout

where propagule pressure should be highest. However, 

brook trout remain patchily distributed, even though 

they have been established in some tributaries for at 

least 60 years (Maclay 1940). Brook trout and cutthroat 

trout also appear to coexist in some streams. Whether 

the brook trout invasion is stalled or advancing very 

slowly is unknown, so it also is not clear that invasion 

and displacement is imminent even in sites close to

potential source populations.
28 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

The threat of invasion by rainbow trout and their hybrids appears more immediate than invasion by brook 

trout. Rainbow trout or their hybrids could ultimately 

invade throughout the system even without further 

stocking of fertile rainbow trout because hybridization 

may facilitate dispersal (Hitt and others 2003, Allendorf 

and others 2004). Thus, conservation of evolutionary 

values associated with pure populations of cutthroat

trout in the headwaters or any accessible tributaries 

throughout the basin is threatened. The distribution of 

rainbow trout genes may be constrained by elevation or 

thermal gradients (Weigel and others 2003) so the risks 

of hybridization are probably greater for populations 

closer to the advancing front. In contrast, conservation of ecological and socio-economic values is of less

concern because hybridized trout may retain many of 

these values.

Persistence with isolation—Westslope cutthroat trout 

have persisted in isolated stream segments throughout

the region, but little information is available to directly 

estimate the probability of persistence as a function of 

isolation time, habitat condition, or network size. Without 

data for this taxon we can only assume westslope cutthroat trout face threats similar to those outlined in the

preceding discussions for other salmonids. Consequently, 

we anticipate that stream networks with less than about

10 km of connected suitable habitat will be at increased 

risk of local extinction from stochastic, demographic,

and genetic processes. The risks will increase as network 

size declines, habitat is degraded, or the potential for 

catastrophic disturbance increases. Populations that are 

isolated will also lose any demographic resilience associated with migratory life histories unless the isolated 

network is large enough to support more productive

Figure 7—Known (red stream segments) and potential (yellow segments) distribution of brook trout in the 

upper Coeur d’Alene River basin. Potential brook trout habitat was predicted based on associations between 

brook trout distributions and stream size and channel gradient (Fausch 1989, Rieman and others 1999, Rich 

and others 2003, Petty and others 2005). However, brook trout could invade farther than indicated here 

because the models are simple approximations and some fish may disperse beyond their primary habitats. The 

brook trout distribution in the Coeur d’Alene River basin does not appear to have advanced in recent decades, 

so other factors like temperature or winter flooding may limit the distribution to less than predicted here.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 29

main-stem habitats that serve as foraging areas supporting faster growing individuals. Moreover, the basin 

is vulnerable to winter rain-on-snow floods associated 

with slow-moving,moistwinterstorms,which can create

catastrophic disturbances especially in the aftermath of 

large fires. These events could drive small populations 

extinct, so any strategy depending on intentional isolation 

should consider spatial replication across broad scales

to reduce the potential for simultaneous extinctions.

Potential priorities and actions—Priorities for management in this basin might include both removing and

constructing barriers as well as improving habitats to

enhance the resilience of local populations. Because there 

is still opportunity to provide some representation and 

diversity of all primary conservation values, a mix of 

strategies could be important. In general, constructing 

barriers could be used to conserve evolutionary values 

in remaining genetically pure populations, whereas removing barriers could be used to conserve and expand

the potential distribution of large migratory trout that 

represent important ecological and socio-economic

values.

There may be an unusual (albeit costly) opportunity to 

conserve all three conservation values simultaneously. 

The fishery and resilience of the remaining cutthroat 

trout populations depend strongly on migratory life 

histories and the connection of tributary and main-stem 

habitats. However, these connections could allow both 

brook trout and rainbow trout to expand throughout

the system. The brook trout invasion appears most 

imminent in the Little North Fork subbasin, whereas 

the risks are probably lower in the North Fork where 

the distance from potential sources is much greater. 

In contrast, the rainbow trout invasion may extend 

throughout the system. Large barriers might be used to 

conserve evolutionary as well as ecological and socioeconomic values if they could isolate tributary-river 

networks where pure populations and migratory life 

histories still persist (Figure 8). The best opportunity 

for this appears to be in the headwaters of both river

subbasins where pure populations and multiple tributary 

networks are linked to the main-stem rivers. Intentional 

barriers could be installed first in the headwaters of the 

Little North Fork Coeur d’Alene River because brook 

trout and rainbow X cutthroat trout hybrids are present 

immediately downstream (Figures 6, 7, and 9). If this 

proved successful and invasions continued to advance

in the North Fork, large barriers could be installed with

the benefit of the experience gained in the first project. 

Although large barriers like low-head main-stem dams

and diversions are possible, they may not be feasible. 

Large barriers would be expensive, would disrupt important ranging movements of large fish and potentially 

other species in the upper rivers (J. Dupont, Idaho Fish 

Figure 8—Pure populations of large fluvial westslope cutthroat trout (P. Valcarce, Limnophoto, image) might be 

conserved in the upper North Fork Coeur d’Alene River basin by using barriers to prevent upstream invasion 

and hybridization by nonnative rainbow trout. If this were successful, the native cutthroat trout would support all 

three types of conservation values.
30 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

and Game Department, Coeur d’Alene, unpublished 

data), and could have negative effects on hydrologic and 

geomorphic processes. Barrier management of this scale 

would require significant commitment of resources and 

environmental compromises.

More modest alternatives might include strategic use of 

barriers to conserve representative genetic diversity and 

evolutionary values in smaller stream networks throughout the basin. Existing impassable culverts have isolated 

several relatively large (5 to 10 km) habitat networks 

across the system (Figure 9). If remnant populations 

above these barriers have not been hybridized or invaded 

by brook trout, culverts could be replaced by intentional 

barriers to protect these populations indefinitely. Good 

barrier sites must be geomorphically stable and economically feasible, and the isolated stream network should 

support a productive population and be large enough to

ensure long-term persistence. Larger networks and more 

intact habitats would be higher priorities.

Figure 9—Opportunities and potential priorities for installing and removing barriers that isolate native 

cutthroat trout in stream networks. The double parallel lines represent sites where large barriers might 

be used to isolate relatively large stream networks that could protect socio-economic, evolutionary, and 

some ecological values simultaneously (but see cautions about feasibility in the text). The four parallel 

lines represent a similar site that might be a high priority because brook trout and rainbow X cutthroat 

trout hybrids occur in nearby tributaries downstream. The black arrows point to existing culvert barriers 

that are candidates either for conversion to intentional barriers to conserve remnant genetic diversity, or 

for removal to expand connectivity and restore ecological and socio-economic values. All other barriers 

above intentional barriers could be removed to expand the isolated network. The benefits of intentional 

isolation would depend on the genetic integrity of the population above the existing barrier, the condition 

of the habitat, and the ability to create a spatially diverse collection of isolated networks throughout the 

basin. It would also depend on the feasibility of creating the barrier. Other existing barriers that are high 

in the headwaters of tributaries are lower priorities for either permanent barriers or removal because 

there is limited potential habitat above the site. Such small areas would either be vulnerable to local 

extinction if isolated or contribute relatively little to the broader network if removed.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 31

Because we are considering only evolutionary values 

at this stage, representation, redundancy, and spreading of risk also become important criteria for selecting

sites for barriers. However, there has been no detailed 

genetic inventory to guide management. Genetic similarity tends to decline with distance among salmonid 

populations (Quinn 2005), and environmental conditions influencing local adaptation probably do also. If 

the goal is to represent as much of the genetic diversity 

and evolutionary legacy for westslope cutthroat trout 

as possible, the logical objective would be to conserve

populations widely distributed throughout the system. 

Broad representation would also minimize the threat 

of simultaneous extinctions from large catastrophic

disturbances.

	 Barriers other than those selected above could be

removed to extend connectivity and conserve or restore 

ecological and evolutionary values in other areas of the 

basin. Barriers that currently isolate small fragments of 

potential cutthroat trout habitatmight be lower priorities

for removal than those that could expand the network,

regardless of the values in question (Figure 9).

Little Snake River

Background—Much different circumstances characterize the Little Snake River basin in southern 

Wyoming and northern Colorado (Table 5, Figure 10). 

The climate is typical of the central Rocky Mountains. 

The first snowfall is in September or October, annual 

Figure 10—Headwaters of the Little Snake River, showing the current estimated 

distribution of populations of Colorado River cutthroat trout as bold stream segments 

(after Hirsch and others 2006). In many cases stream segments shown as occupied are 

fragmented by natural or artificial barriers, or the true distribution of native cutthroat trout 

is incompletely known. Estimates of the lowermost extent of trout habitat in the Little 

Snake River main stem in 1955 (Eiserman 1958) and in the Muddy Creek headwaters are 

denoted by solid red bars.
32 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

precipitation peaks in March and April in the form of 

snow, and there may be heavy rains associated with 

monsoonal storms in July and August (Knight 1994). 

Peak discharge is caused by snowmelt runoff in June, 

with occasional localized stormflows in late summer. 

Elevations range from 1900 m at the mouth of Muddy 

Creek near Baggs, Wyoming, to over 3300 m in the 

headwaters. Streams pass through three major vegetation types: sagebrush grasslands with riparian gallery 

forests at the lowest elevations, extensive aspen stands

at mid-elevation sites, and coniferous forests higher in

the basin.

Land ownership and use are typical of many watersheds in the western U.S. Lower-elevation valleys are 

privately owned and higher-elevation lands are in the 

public domain. Historically this region served as a hub of 

grazing activity. For example, in the early 1900s over 10 

million sheep annually were pastured on summer range 

or herded on stock driveways on the Wyoming side alone 

(Thybony and others 1985). Livestock grazing continues 

at much lower intensity, although sedimentation, riparian alteration, and irrigation diversions associated with

agriculture have altered fish habitat quality and tributary 

access for decades (Eiserman 1958). Resource extraction 

on public lands, such as hard rock mining and timber

harvest, is less than it was historically, but recreational 

use of public lands hasincreasedwith population growth

in much of the mountain West, and publicly-built and 

user-created roads and trails are extensive. An exception is a large portion of the uppermost North Fork

basin (Figure 10), which lies within the Huston Park 

Wilderness Area. Finally, the entire North Fork basin 

has undergone water development. Flows of all perennial 

tributaries pass through water collection structures that

block upstream migrations of fish and fragment some 

populations of cutthroat trout. In addition, portions of 

the high spring flows (and probably juvenile fish) are 

diverted into a pipeline transferring water east underneath the Continental Divide. Water development is also 

projected or underway in adjacent Wyoming basins, 

several of which harbor populations of cutthroat trout.

Native fishes are relatively few. Colorado River cutthroat trout once ranged throughout the basin, sympatric with mottled sculpin (Cottus bairdi) and mountain 

whitefish in the colder upper reaches, and perhaps with 

some members of the now-rare Colorado River fauna, 

such as Colorado pikeminnow (Ptychocheilus lucius), 

in the warmer downstream reaches of the main stem. 

Eiserman (1958) considered the Little Snake River near 

the mouth of Willow Creek and the town of Dixon as the 

downstream extent of trout (Figure 10). Below this point 

the river was regarded as too warm and silty to provide 

trout habitat in summer, as were many of the lower 

portions of Muddy Creek. Brook, brown, and rainbow 

trout were first introduced in the 1930s (Eiserman 1958) 

and currently dominate fish communities in coldwater 

portions of the basin, whereas cutthroat trout are relegated almost solely to short stream segments protected 

from upstream invasions by barriers. A waterfall near 

its mouth protected the North Fork Little Snake River 

population, one of the largest of this subspecies, from

early invasions. Later ones were thwarted by some of 

the first barriers built (in the mid-1970s) specifically to 

prevent upstream invasions of nonnative fishes. The main 

stem has a mix of species, including nonnative channel

catfish (Ictalurus punctatus), but does not appear to 

retain any native salmonids.

Presently, from 160 to 190 km of up to 38 streams 

are believed to contain 15 to 20 distinct populations of 

Colorado River cutthroat trout (Figure 10; Young and 

others 1996, Hirsch and others 2006). These populations display different degrees of hybridization with 

rainbow trout or Yellowstone cutthroat trout, but a few

populations appearto have retained their genetic integrity. Some populations in Colorado may have resulted 

from early stocking of Colorado River cutthroat trout 

originating from Trappers Lake, an adfluvial population in the headwaters of the White River basin in 

northwestern Colorado. No lakes in the Little Snake 

River basin were known to contain this subspecies.

Conservation values—Colorado River cutthroat 

trout have been the subject of intensive management in

this basin (Speas and others 1994), primarily for their 

evolutionary values. Their populations are rare enough 

that each receives individual attention, and the focus of

management is to maintain all existing populations. We 

infer from the extensive tributary network (for example, 

Battle, Savery, and Slater creeks) that fluvial cutthroat 

trout once existed. If so, their loss has reduced or eliminated ecological valuessuch as demographic support of

resident cutthroat trout populations, transfer of nutrients

upstream, and providing food for terrestrial predators

like bears (Koel and others 2005). Socio-economic 

values contribute little to the current management of

Colorado River cutthroat trout. Angling for cutthroat 

trout is an incidental activity, primarily because their 

populations are scattered, most streams are small and

difficult to access by automobile, and adult fish do not 

reach large sizes because of the short growing season 

at high elevations. Moreover, regulations prohibiting or 

limiting harvest have been in place for decades.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 33

Vulnerability to invasion—Following decades of 

stocking (Eiserman 1958, Wiltzius 1985), brook trout 

and rainbow trout have become widely distributed in 

most available habitats, and there is little evidence of

environmental or biotic resistance to their continued

spread should additional waters become accessible. For 

example, within four years of an illegal introduction 

just above the artificial barrier in the main-stem North 

Fork Little Snake River, brook trout had spread to three 

tributaries and ascended 8 km up the main stem despite 

the robust population of Colorado River cutthroat trout 

already present (M. Young, personal observation). In 

addition, perennial streams lacking fish are exceptionally rare, and presumably contain geological or artificial 

barriers, or habitat unsuitable for trout.

Persistence with isolation—Nearly all remaining 

populations of Colorado River cutthroat trout persist 

only above barriers (Figure 11), yet individuals in some 

of these populations migrate hundreds or thousands of

meters to spawn (Young 1996) and move extensively 

for foraging (Young and others 1997). Nevertheless 

only four populations occupy more than 10 km of connected habitat and probably contain enough individuals 

to afford security from the long-term loss of genetic 

variation (Allendorf and others 1997, Young and others 

2005). Most populations inhabit stream segments less 

than 3 km (often due to natural or artificial barriers that 

fragment habitat), likely consist of a few hundred fish at 

most (Young and others 2005), and are at a heightened 

risk of extirpation from an array of natural disturbances, 

such as drought or post-wildfire debris torrents. Similar 

populations ofthissubspecies, and salmonids elsewhere

in the inland West, have suffered this fate (M. Fowden, 

Wyoming Game and Fish Department, personal communication; Brown and others 2001, Dunham and others 

2003b).

Potential priorities and actions—Currently, most 

populations of Colorado River cutthroat trout in this 

basin inhabit short headwater stream segments and are

probably at an elevated risk of extirpation. The genetic 

structure of these populations has not been studied, so

the most conservative approach would be to attempt to

Figure 11—In the Little Snake River watershed, Colorado River cutthroat trout persist only in headwater 

streams, above barriers to upstream invasion by nonnative brook, brown, and rainbow trout (M. Young image).
34 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

conserve all populations to maximize redundancy and 

representation of whatever genetic diversity remains. 

From this perspective, the smallest populations are at

the greatest risk. The probability of persistence of all 

populations regardless of size could be enhanced by extending their distribution downstreamto suitable habitats

currently occupied by nonnative species once the latter 

have been removed. This approach could be particularly 

valuable if it connects populationsto additionalstreams

that decrease the risk of population loss from localized 

catastrophic events. In addition, there may be opportunities to restore connectivity by providing fish passage 

around water diversions, and to achieve redundancy by 

introducing genetically pure populations into additional, 

nearby waters that were previously occupied by cutthroat 

trout. Such efforts are underway (B. Wengert, Wyoming 

Game and Fish Department, and D. Brauch, Colorado 

Division of Wildlife, personal communication), as are 

attempts to identify waters where such actions could 

be undertaken (Hirsch and others 2006). Nevertheless, 

attempts to establish populations in locations with

access to additional streams or to refugia that permit

natural recolonization has been, and will be, challenging. 

Overcoming water management issues, resolving private 

land-public land conflicts, and locating sites suitable for 

barrier installation render this a long-term, site-specific 

solution unless other methods for controlling nonnative

fishes become available. The feasibility of any of these 

actions will likely guide the ultimate selection of sites 

for further work.

Where suitable habitat exists but cutthroat trout 

populations are absent, the primary conservation actions involve repeated chemical treatment or intensive

electrofishing to remove nonnative species and installing 

artificial barriers to prevent reinvasions. This work has 

been challenging. Occasionally, treatments have not 

killed all nonnative fish present, forcing biologists to 

repeat treatment of some streams or abandon reclamation in others. Some barriers were passable to upstream 

migrating brook trout, or became so after damage by 

high flows. In other cases brook trout were introduced 

above these structures by anglers, apparently to retaliate for restricted harvest of cutthroat trout or the loss

of individually prized fisheries for nonnative trout. In 

these latter cases, preservation of populations may rely 

more on road closures, persistent public outreach, and a

regular law enforcement presence than on fish population management.

Finally, the relatively few indigenous populations of 

Colorado River cutthroat trout remaining have served as 

a source for fish stocked elsewhere in the basin, either 

by direct transfers or indirectly as broodstock used in 

hatcheries or off-site ponds. Although the use of lakes 

as broodstock refugia has been successful elsewhere

(K. Rogers, Colorado Division of Wildlife, personal 

communication), it has been repeatedly attempted with 

limited success in this area. Use of hatchery or semi-wild 

broodstocks argues for adopting genetic management

plans to ensure such stocks do not deviate from wild

populations (Mobrand and others 2005).

Section VII: Making Strategic

Decisions_______________________

The framework we presented in Section V outlines 

important questions to consider when managing the 

invasion-isolation tradeoff. For biologists faced with 

a particular basin and set of native salmonid populations that represent important conservation values, the

next step is to consider the available options and make

strategic decisions, before setting priorities for action. 

A lesson from the two examples presented above is that

the best solutions for conserving native salmonids are

not the same in all regions. For example, in the Coeur 

d’Alene River basin the main issues are whether nonnative salmonids will invade and displace or hybridize 

with native cutthroat trout, and how much habitat will

be required to sustain important migratory life history 

types if they are isolated. In contrast, in the Little Snake 

River basin the primary issues are locating any remaining 

pure populations,replicating themin headwaterstreams

above barriers after removing nonnative salmonids, and

improving the probability of persistence of these many 

isolated populations.

Despite these fundamental differences among regions, 

the two examples illustrate that fisheries managers do 

have opportunities to make strategic decisions under

these different circumstances. For any given basin and 

particular set of salmonid populations of conservation

value, these decisions will be influenced primarily 

by two conditions: 1) the degree of isolation of the 

populations of interest, and 2) the degree of invasion 

threat (Figure 12). For example, if most native salmonid 

populationsin the basin are in small isolated patches, the

invasion of nonnative salmonids is either imminent or

relatively complete, and the invaders have strong negative effects (as in the Little Snake River basin; upper left 

corner of Figure 12), then barriers will be required to 

protect most populations from rapid extinction. Given 

this, the strategic decisions available include: 1) the 

size, habitat quality, and number of patches needed to 

ensure persistence of a minimum number of the isolated
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 35

populations through time; 2) the representation of as much 

genetic, morphological, behavioral, and habitat diversity 

as possible; 3) the spatial distribution and redundancy of 

populations to avoid extinction of evolutionary lineages 

by correlated catastrophic disturbances; and 4) working 

out protocols for effective removal or control of nonnative trout, and subsequent translocations of native trout. 

Despite these strong constraints on management options, 

the goal of these decisions is to move the set of native

salmonid populations toward the lower right corner of

this decision space where the invasion threat is farther

away and patches are larger and more connected, even 

though in many cases this goal is not fully achievable 

at present.

	 At the opposite end of the spectrum, if a manager is

fortunate to have many large interconnected patches with 

native salmonids, and the invasion threat is far away or the 

effects of the invader are weak, then strategic decisions

will involve: 1) preventing invasions, 2) preventing fish 

movement barriers or land uses that fragment habitat, 3) 

monitoring the spread of any nonnative salmonids in the 

basin, and 4) ensuring that natural processes continue 

to provide the habitat heterogeneity that sustains the 

native populations (Figure 12; lower right corner). For 

example, a key strategy might be to minimize sources 

of nonnative fishes that could foster unauthorized introductions (Rahel 2004) and educate the public about 

their dangers to native biota (Cambray 2003a, 2003b). 

Likewise, under this scenario managers have more opportunity to ensure representation and redundancy of 

populations, and to experiment and monitorto learn and

thereby reduce future uncertainty.

Figure 12—A conceptual diagram of the opportunities for strategic decisions when managing the 

invasion – isolation tradeoff for native salmonid populations of conservation value. Examples of 

strategic decisions to maximize conservation of remaining populations in several regions of the 

decision space are shown. 
36 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

Some managers will find themselves in other regions 

of this decision space, where different strategies are

required. For example, a basin may have little invasion 

threat but many culverts or diversions that fragment 

headwater stream habitats (Figure 12; upper right 

corner). In this situation, a strategic decision would be 

to minimize habitat degradation (and improve habitat) 

in these small patches to prevent further extinctions,

and increase connectivity by removing barriers where 

feasible to maximize resilience of these populations. In 

contrast, other managers may find that their basin has 

large patches of habitat isolated by natural barriers, but 

that nonnative trout were stocked above them and the

invasion is nearly complete (lower left corner). In this 

case they may prioritize some of these basins for the 

difficult task of removing nonnative trout, which may 

require building temporary barriers to prevent upstream 

invasion while successive treatments are used to extend

habitat downstream for native salmonids. This will require strong public relations and a long-term commitment 

of agency funding and time. Overall, an understanding 

of where their native salmonid populations lie in this

decision space will help managers develop strategic

decisions for other possible scenarios.

When barriers are used to prevent invasions, human 

translocation of nonnative salmonids above them may 

also be a problem, requiring further strategic decisions. 

If native salmonids are restricted to many small patches 

above barriers (Figure 12, upper left), managers could 

construct several barriers at intervals (for example, 1 

km) near the downstream end and monitor the buffer zone as insurance against reintroductions, but the 

stream fragments are usually too short to justify other 

measures. In contrast, in a large basin of intact and 

relatively remote habitat where strong invaders are removed in stages (lower left), a strategic decision would 

be to place the downstream barrier in an inaccessible

location to minimize human translocations. If temporary 

barriers were constructed for the project, the materials

could be left on site to allow replacing barriers quickly 

if translocations occur at a downstream barrier.

Finally, where constraints are strongest (Figure 12; 

upper left corner), it is likely that managers will select a 

mixed strategy and hedge their bets against uncertainty. 

This may include conserving many small populations 

thatpersist abovebarriers,translocatingamongreplicates

when they go extinct (Lubow 1996, Hilderbrand 2002), 

and working to protect or develop larger populations in

remote or protected basins (see Shepard and others 2005 

for an example). In the worst case, if barriers on the large 

populations are breached by human or natural agents 

and invasions proceed quickly, fish from the smaller 

replicate populations can be used to refound them.

Epilogue: What do We Need to

Know?

	 Conserving native salmonids at risk from invasions

in the western U.S. is a widespread, long-term, and 

expensive problem, so investment in better information

to improve decisions is likely to pay off. In developing 

this synthesis we found important uncertainties about 

the key processes that drive the tradeoff between invasion and isolation. These often emerged as differences 

in opinion among the authors and among reviewers,

based on their data and experience in different regions

and with different species and subspecies of salmonids. 

From these discussions, we found three main questions 

in need of research to inform better management:

	 •	 What are rules of thumb for stream length or watershed area needed to sustain different salmonid species 

and subspecies in isolated populations above barriers? — It is clear that the size of the stream network 

can strongly affect the resilience, persistence, and 

viability of isolated salmonid populations. However, 

these effects also vary strongly with environmental 

conditions and the evolutionary history of species or 

populations. Basic field data on species occurrence 

and an inventory of natural and anthropogenic barriers can be used to develop models to estimate risk of

extinction with time (see Section III). Rapid progress 

could be made by using existing data and collaborating with biologists already engaged in inventory and 

monitoring programs throughout the region.

	 •	 What environmental, ecological, and evolutionary factors control where the main nonnative salmonids will 

invade? — The extent and effects of nonnative invasions 

appear to vary substantially across the interior West. 

In many regions, it is not clear whether invasions are 

continuing orstalled, andwhetherthe process depends

on changing climate and even evolution of the species

involved. Management of invasions, and the effective use 

of barriers and limited management resources, would

benefit from a better understanding of where and how 

invasion and displacement is likely to progress.

	 •	 Does life history diversity of native salmonids buffer 

invasion and allow coexistence? If so, how? — Conventional wisdom and life history theory suggest that 

migratory forms could be more resilient and resistant 

to invasion than resident life history types, due to differences in demographic processes and habitat use. 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 37

However, little is known about this topic so more specific 

hypotheses and observations need to be developed, 

followed by empirical analysis and simulation.

	 Although some initialresearch has been conducted on

each of these questions for specific species or environments (Dunham and Rieman 1999, Adams and others 

2002, Harig and Fausch 2002), more is needed to allow 

accurate predictions on which to base sound management. 

Likewise, many other questions need to be answered 

for a more complete understanding of the tradeoff (see 

Dunham and others 2002a, Peterson and Fausch 2003b, 

Dunham and others 2004) and the feasibility of different 

management options. In addition, a careful monitoring 

program followed by widespread reporting of the results 

would rapidly advance our understanding of the need 

for, and effects of, barriers used to manage populations

of native trout throughout the western U.S

References______________________

Abbott, A. M. 2000. Land management and flood effects on the 

distribution and abundance of cutthroat trout in the Coeur

d’Alene River basin, Idaho. Master’s thesis. University of Idaho, 

Moscow.

Adams, S. B. 1999. Mechanisms limiting a vertebrate invasion: brook 

trout in mountain streams of the northwestern USA. Doctoral 

dissertation. University of Montana, Missoula.

Adams, S. B., C. A. Frissell, and B. E. Rieman. 2000. Movements 

of nonnative brook trout in relation to stream channel slope. 

Transactions of the American Fisheries Society 129:623-638.

Adams, S. B., C. A. Frissell, and B. E. Rieman. 2001. Geography 

of invasion in mountain streams: consequences of headwater lake 

fish introductions. Ecosystems 2001:296-307.

Adams, S. B., C. A. Frissell, and B. E. Rieman. 2002. Changes in 

distribution of nonnative brook trout in an Idaho drainage after 

two decades. Transactions of the American Fisheries Society 

131:561-568.

Adams, S. B., D. A. Schmetterling, and M. K. Young. 2005. Instream 

movements by boreal toads (Bufo boreas boreas). Herpetological 

Review 36:27-33.

Allendorf, F. W., D. W. Bayles, D. L. Bottom, K. P. Currens, C. A. 

Frissell, D. Hankin, J. A. Lichatowich, W. Nehlsen, P. C. Trotter, 

and T. H. Williams. 1997. Prioritizing Pacific salmon stocks for 

conservation. Conservation Biology 11:140-152.

Allendorf, F. W., R. F. Leary, P. Spruell, and J. K. Wenburg. 2001. 

The problems with hybrids: setting conservation guidelines. 

Trends in Ecology and Evolution 16:613-622.

Allendorf, F. W., R. F. Leary, N. P. Hitt, K. L. Knudsen, L. L. Lundquist, and P. Spruell. 2004. Intercrosses and the U.S. Endangered 

Species Act: should hybridized populations be included as westslope cutthroat trout? Conservation Biology 18:1203-1213.

Allendorf, F. W., and R. F. Leary. 1988. Conservation and distribution of genetic variation in a polytypic species: the cutthroat 

trout. Conservation Biology 2:170-184.

Allendorf, F. W., R. F. Leary, N. P. Hitt, K. L. Knudsen, M. C. Boyer, 

and P. Spruell. 2005. Cutthroat trout hybridization and the U.S. 

Endangered Species Act: one species, two policies. Conservation 

Biology 19:1326-1328.

Allendorf, F. W., and N. Ryman. 2002. The role of genetics in 

population viability analysis. Pages 50-85 in D. R. McCullough 

and S. R. Beissinger, editors. Population viability analysis. University of Chicago Press, Chicago, Illinois.

Alvord, W. 1991. A history of Montana’s fisheries management 

1890-1985. Fisheries Division, Montana Department of Fish, 

Wildlife, and Parks, Helena.

Angermeier, P. L., R. J. Neves, and J. W. Kauffman. 1993. Protocol 

to rank value of biotic resources in Virginia streams. Rivers 

4:20-29.

Anlauf, K. J. 2005. Understanding the ecology of whirling disease 

through modeling of the tubificid host. Master’s thesis. University 

of Idaho, Moscow.

Bagliniere, J. L., G. Maisses, P. Y. Lebail, and A. Nihouarn. 1989. 

Population dynamics of brown trout Salmo trutta L. in a tributary in Brittany, France: spawning and juveniles. Journal of Fish 

Biology 34:97-110.

Baxter, C. V. 2002. Fish movement and assemblage dynamics in 

a Pacific Northwest riverscape. Doctoral dissertation. Oregon 

State University, Corvallis.

Baxter, C. V., K. D. Fausch, M. Murakami, and P. L. Chapman. 

2004. Fish invasion restructures stream and forest food webs by 

interrupting reciprocal prey subsidies. Ecology 85:2656-2663.

Baxter, C. V., C. A. Frissell, and F. R. Hauer. 1999. Geomorphology, logging roads, and the distribution of bull trout spawning 

in a forested river basin: implications for management and 

conservation. Transactions of the American Fisheries Society 

128:854-867.

Behnke, R. J. 1992. Native trout of western North America. American 

Fisheries Society Monograph 6, Bethesda, Maryland.

Beier, P., and R. F. Noss. 1998. Do habitat corridors provide connectivity? Conservation Biology 12:1241-1252.

Benjamin, J. R. 2006. Invasion by nonnative brook trout in Panther 

Creek, Idaho: roles of habitat quality, connectivity, and biotic resistance. Master’s thesis. Boise State University, Boise, Idaho.

Bennett, A. F. 1999. Linkages in the landscape: the role of corridors 

and connectivity in wildlife conservation. IUCN-The World 

Conservation Union, Gland, Switzerland.

Biro, P. A., A. E. Morton, J. R. Post, and E. A. Parkinson. 2004. 

Over-winter lipid depletion and mortality of age-0 rainbow 

trout (Oncorhynchus mykiss). Canadian Journal of Fisheries and 

Aquatic Sciences 61:1513-1519.

Bjornn, T. C., and D. W. Reiser. 1991. Habitat requirements of 

salmonids in streams. American Fisheries Society Special Publication 19:83-138.

Brannon, E. L., M. S. Powell, T. P. Quinn, and A. Talbot. 2004. 

Population structure of Columbia River Basin Chinook salmon 

and steelhead trout. Reviews in Fisheries Science 12:99-232.

Brown, D. K., A. A. Echelle, D. L. Propst, J. E. Brooks, and W. L. 

Fisher. 2001. Catastrophic wildfire and number of populations as 

factors influencing risk of extinction for Gila trout (Oncorhynchus 

gilae). Western North American Naturalist 61:139-148.

Brown, R. S., and W. C. Mackay. 1995. Fall and winter movements 

of and habitat use by cutthroat trout in the Ram River, Alberta. 

Transactions of the American Fisheries Society 124:873-885.

Buktenica, M. W., B. D. Mahoney, S. F. Girdner, and G. L. Larson. 

2001. Response of a resident bull trout population to nine years of 

brook trout removal, Crater Lake National Park, OR. Pages 127-

132 in D. Schill, S. Moore, P. Byorth, and B. Hamre, editors. Wild 

Trout VII Symposium. Trout Unlimited, Washington, DC.

Callicott, J. B. 1995. Conservation ethics at the crossroads. American 

Fisheries Society Symposium 17:3-7.

Callicott, J. B., and K. Mumford. 1997. Ecological sustainability as 

a conservation concept. Conservation Biology 11:32-40.

Cambray, J. A. 2003a. Impact on indigenous species biodiversity 

caused by the globalisation of alien recreational freshwater 

fisheries. Hydrobiologia 500:217-230.
38 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

Cambray, J. A. 2003b. The global impact of alien trout species – a 

review; with reference to their impact in South Africa. African 

Journal of Aquatic Science 28:61-67.

Campton, D. E., and L. R. Kaeding. 2005. Westslope cutthroat trout, 

hybridization, and the U.S. Endangered Species Act. Conservation Biology 19:1323-1325.

Carl, L. M., and M. C. Healey. 1984. Differences in enzyme frequency and body morphology among three juvenile life history 

types of chinook salmon (Oncorhynchus tshawytscha) in the 

Nanaimo River, British Columbia. Canadian Journal of Fisheries 

and Aquatic Sciences 41:1070-1077.

Caughley, G. 1994. Directions in conservation biology. Journal of 

Animal Ecology 63:215-244.

Chase, J. M., and M. A. Liebold. 2003. Ecological niches: linking 

classical and contemporary approaches. University of Chicago 

Press, Chicago, Illinois.

Clark, M. E., K. A. Rose, D. A. Levine, and W. W. Hargrove. 

2001. Predicting climate change effects on Appalachian trout: 

combining GIS and individual-based modeling. Ecological Applications 11:161-178.

Clarkin, K., A. Connor, M. J. Furniss, B. Gubernick, M. Love, K. 

Moynan, and S. W. Musser. 2003. National inventory and assessment procedure for identifying barriers to aquatic organism 

passage at road-stream crossings. USDA Forest Service, National 

Technology and Development Program, Report 7700-Transportation Management, San Dimas Technology and Development 

Center, San Dimas, California. Available: http://www.stream.

fs.fed.us/publications/PDFs/NIAP.pdf. (accessed April 2006).

Coleman, M. A., and K. D. Fausch. 2006. Causes of recruitment 

bottlenecks in translocated cutthroat trout populations: investigation of low temperature effects. Final Report to Colorado 

Division of Wildlife, USDI Central Utah Project Completion 

Act Office, USDA Forest Service, and National Park Service. 

Department of Fishery and Wildlife Biology, Colorado State 

University, Fort Collins.

Crooks, J. A., and A. V. Suarez. in press. Hyperconnectivity, invasive species, and the breakdown of barriers to dispersal. Pages 

000-000 in K. R. Crooks, and M. Sanjayan, editors. Connectivity 

conservation. Cambridge University Press, Cambridge, U.K.

Crowl, T. A., C. R. Townsend, and A. R. McIntosh. 1992. The impact 

of introduced brown and rainbow trout on native fish: the case of 

Australasia. Reviews in Fish Biology and Fisheries 2:217-241.

De Staso, J., III, and F. J. Rahel. 1994. Influence of water temperature on interactions between juvenile Colorado River cutthroat 

trout and brook trout in a laboratory stream. Transactions of the 

American Fisheries Society 123:289-297.

Dingle, H. 1996. Migration: the biology of life on the move. Oxford 

University Press, Oxford, U.K.

Dobson, A., K. Ralls, M. Foster, M. E. Soulé, D. Simberloff, D. 

Doak, J. A. Estes, L. S. Mills, D. Mattson, R. Dirzo, H. Arita, 

S. Ryan, E. A. Norse, R. F. Noss, and D. Johns. 1999. Corridors: 

reconnecting fragmented landscapes. Pages 129-170 in M. E. 

Soulé and J. Terborgh, editors. Continental conservation: scientific foundations of regional reserve networks. Island Press, 

Washington, DC.

Dowling, T. E., and M. R. Childs. 1992. Impact of hybridization on 

a threatened trout of the southwestern United States. Conservation Biology 6:355-364.

Dunham, J. B., S. B. Adams, R. E. Schroeter, and D. C. Novinger. 

2002a. Alien invasions in aquatic ecosystems: toward an understanding of brook trout invasions and their potential impacts on

inland cutthroat trout in western North America. Reviews in Fish 

Biology and Fisheries 12:373-391.

Dunham, J. B., M. M. Peacock, B. E. Rieman, R. E. Schroeter, and 

G. L. Vinyard. 1999. Local and geographic variability in the distribution of stream-living Lahontan cutthroat trout. Transactions 

of the American Fisheries Society 128:875-889.

Dunham, J. B., D. S. Pilliod, and M. K. Young. 2004. Assessing 

the consequences of nonnative trout in headwater ecosystems in 

western North America. Fisheries (Bethesda) 29(6):18-26.

Dunham, J. B., and B. E. Rieman. 1999. Metapopulation structure of 

bull trout: influences of physical, biotic, and geometrical landscape 

characteristics. Ecological Applications 9:642-655.

Dunham, J. B., B. E. Rieman, and J. T. Peterson. 2002b. Patch-based 

models of species occurrence: lessons from salmonid fishes in 

streams. Pages 327-334 in J. M. Scott, P. J. Heglund, M. Morrison, M. Raphael, J. Haufler, and B. Wall, editors. Predicting 

species occurrences: issues of scale and accuracy. Island Press, 

Covelo, CA.

Dunham, J. B., R. Schroeter, and B. Rieman. 2003a. Influence of 

maximum water temperature on occurrence of Lahontan cutthroat trout within streams. North American Journal of Fisheries 

Management 23:1042-1049.

Dunham, J. B., G. L. Vinyard, and B. E. Rieman. 1997. Habitat fragmentation and extinction risk of Lahontan cutthroat trout. North 

American Journal of Fisheries Management 17:1126-1133.

Dunham, J. B., M. K. Young, R. E. Gresswell, and B. E. Rieman. 

2003b. Effects of fire on fish populations: landscape perspectives 

on persistence of native fishes and nonnative fish invasions. Forest 

Ecology and Management 178:183-196.

Dunnigan, J. L. 1997. The spatial distribution of westslope cutthroat trout in the Coeur d’Alene River Basin. Master’s Thesis. 

University of Idaho, Moscow.

Eiserman, F. 1958. A fisheries survey of the Little Snake River 

drainage. Fisheries Technical Report 6, Wyoming Game and 

Fish Commission, Cheyenne.

El-Matbouli, M., M. T. Fischer, and R. W. Hoffman. 1992. Present 

knowledge on the life cycle, taxonomy, pathology, and therapy 

of some Myxosporea spp. important in freshwater fish. Annual 

Review of Fish Diseases 1992:367-402.

Epifanio, J. 2000. The status of coldwater fishery management in 

the United States. Fisheries (Bethesda) 25(7):13-27.

Fausch, K. D. 1988. Tests of competition between native and introduced salmonids in streams: what have we learned? Canadian 

Journal of Fisheries and Aquatic Sciences 45:2238-2246.

Fausch, K. D. 1989. Do gradient and temperature affect distributions 

of, and interactions between, brook charr (Salvelinus fontinalis) 

and other resident salmonids in streams? Physiology and Ecology 

Japan Special Vol. 1:303-322.

Fausch, K. D. 1998. Interspecific competition and juvenile Atlantic 

salmon (Salmo salar): on testing effects and evaluating the evidence across scales. Canadian Journal of Fisheries and Aquatic 

Sciences 55:218-231.

Fausch, K. D., and K. R. Bestgen. 1997. Ecology of fishes indigenous 

to the central and southwestern Great Plains. Pages 131-166 in F. 

L. Knopf and F. B. Samson, editors. Ecology and conservation 

of Great Plains vertebrates. Springer-Verlag, New York.

Fausch, K. D., and R. G. Bramblett. 1991. Disturbance and fish 

communities in intermittent tributaries of a western Great Plains 

river. Copeia 1991:659-674.

Fausch, K. D., Y. Taniguchi, S. Nakano, G. D. Grossman, and C. R. 

Townsend. 2001. Flood disturbance regimes influence rainbow 

trout invasion success among five Holarctic regions. Ecological 

Applications 11:1438-1455.

Fausch, K. D., C. E. Torgersen, C. V. Baxter, and H. W. Li. 2002. 

Landscapes to riverscapes: bridging the gap between research 

and conservation of stream fishes. BioScience 52:483-498.

Fausch, K. D., and M. K. Young. 1995. Evolutionarily significant 

units and movement of resident stream fishes: a cautionary tale. 

American Fisheries Society Symposium 17:360-370.

Francis, R. I. C. C., and R. Shotton. 1997. “Risk” in fisheries management: a review. Canadian Journal of Fisheries and Aquatic 

Sciences 54:1699-1715.

Frankham, R. 2005. Genetics and extinction. Biological Conservation 126:131-140.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 39

Fuller, P. L., L. G. Nico, and J. D. Williams. 1999. Nonindigenous 

fishes introduced into inland waters of the United States. American 

Fisheries Society, Bethesda, Maryland.

Gajardo, G., and L. Laikre. 2003. Chilean aquaculture boom is based 

on exotic salmon resources: a conservation paradox. Conservation 

Biology 17:1173-1174.

GAO (U.S. General Accounting Office). 2001. Restoring fish 

passage through culverts on Forest Service and BLM lands 

in Oregon and Washington could take decades. GAO-02-136, 

Washington, DC.

Gilpin, M. E., and M. E. Soulé. 1986. Minimum viable populations: 

processes of species extinction. Pages 19-34 in M. E. Soulé, editor. Conservation biology: the science of scarcity and diversity. 

Sinauer Associates, Sunderland, Massachusetts.

Gislason, D., M. Ferguson, S. Skulason, and S. S. Snorrason. 1999. 

Rapid and coupled phenotypic and genetic divergence in Icelandic 

Arctic char (Salvelinus alpinus). Canadian Journal of Fisheries 

and Aquatic Sciences 56:2229-2234.

Gowan, C., and K. D. Fausch. 1996. Mobile brook trout in two 

high-elevation Colorado streams: re-evaluating the concept of 

restricted movement. Canadian Journal of Fisheries and Aquatic 

Sciences 53:1370-1381.

Gowan, C., M. K. Young, K. D. Fausch, and S. C. Riley. 1994. 

Restricted movement in resident stream salmonids: a paradigm 

lost? Canadian Journal of Fisheries and Aquatic Sciences 

51:2626-2637.

Gresswell, R. E., editor. 1988. Status and management of interior 

stocks of cutthroat trout. American Fisheries Society Symposium 

4, Bethesda, Maryland.

Gresswell, R. E. 1997. Introduction to ecology and management of 

potamodromous salmonids. North American Journal of Fisheries 

Management 17:1027-1028.

Groves, C. R. 2003. Drafting a conservation blueprint: a practitioner’s 

guide to planning for biodiversity. Island Press, Washington, DC.

Halupka, K. C., M. F. Willson, M. D. Bryant, F. H. Everest, and 

A. J. Gharrett. 2003. Conservation of population diversity of 

Pacific Salmon in southeast Alaska. North American Journal of 

Fisheries Management 23:1057-1086.

Hanski, I. A. 1997. Metapopulation dynamics: from concepts and 

observations to predictive models. Pages 69-91 in I. A. Hanski 

and M. E. Gilpin, editors. Metapopulation biology: ecology, 

genetics, and evolution. Academic Press, New York.

Hanski, I. A., and M. E. Gilpin. 1997. Metapopulation biology: ecology, 

genetics, and evolution. Academic Press, San Diego, California.

Hanski, I. L. 1999. Metapopulation ecology. Oxford University 

Press, Oxford, U.K.

Harden Jones, F. R. 1968. Fish migration. Edward Arnold, London, 

U.K.

Harig, A. L., and K. D. Fausch. 2002. Minimum habitat requirements for establishing translocated cutthroat trout populations. 

Ecological Applications 12:535-551.

Harig, A. L., K. D. Fausch, and M. K. Young. 2000. Factors influencing success of greenback cutthroat trout translocations. North 

American Journal of Fisheries Management 20:994-1004.

Harrison, S., and A. D. Taylor. 1997. Empirical evidence for metapopulation dynamics. Pages 27-42 in I. A. Hanski and M. E. 

Gilpin, editors. Metapopulation biology: ecology, genetics, and 

evolution. Academic Press, New York.

Hart, D. D., T. E. Johnson, K. L. Bushaw-Newton, R. J. Horwitz, 

A. T. Bednarek, D. F. Charles, D. A. Kreeger, and D. J. Velinsky. 

2002. Dam removal: challenges and opportunities for ecological 

research and river restoration. BioScience 52:669-681.

Hastings, A., K. Cuddington, K. F. Davies, C. J. Dugaw, S. Elmendorf, A. Freestone, S. Harrison, M. Holland, J. Lambrinos, 

U. Malvadkar, B. A. Melbourne, K. Moore, C. Taylor, and D. 

Thomson. 2005. The spatial spread of invasions: new developments in theory and evidence. Ecology Letters 8:91-101.

Heino, M., and I. Hanski. 2001. Evolution of migration rate. American Naturalist 157:495-511.

Hendry, A. P., T. Bohlin, B. Jonsson, and O. K. Berg. 2004a. To 

sea or not to sea? Anadromy versus non-anadromy in salmonids. 

Pages 92-125 in A. P. Hendry and S. C. Stearns, editors. Evolution illuminated: salmon and their relatives. Oxford University 

Press, New York.

Hendry, A. P., V. Castric, M. T. Kinnison, and T. P. Quinn. 2004b. 

The evolution of philopatry and dispersal: homing versus straying in salmonids. Pages 52-91 in A. P. Hendry and S. C. Stearns, 

editors. Evolution illuminated: salmon and their relatives. Oxford 

University Press, New York.

Hendry, A. P., and S. C. Stearns, editors. 2004. Evolution illuminated: salmon and their relatives. Oxford University Press, 

Oxford, U.K.

Hepworth, D. K., M. J. Ottenbacher, and L. N. Berg. 1997. Distribution and abundance of native Bonneville cutthroat trout

(Oncorhynchus clarki utah) in southwestern Utah. Great Basin 

Naturalist 57:11-20.

Hess, G. 1994. Conservation corridors and contagious disease: a 

cautionary note. Conservation Biology 8:256-262.

Hilderbrand, R. H. 2002. Simulating supplementation strategies 

for restoring and maintaining stream resident cutthroat trout

populations. North American Journal of Fisheries Management 

22:879-887.

Hilderbrand, R. H. 2003. The roles of carrying capacity, immigration, and population synchrony on persistence of stream-resident 

cutthroat trout. Biological Conservation 110:257-266.

Hilderbrand, R. H., and J. L. Kershner. 2000. Conserving inland 

cutthroat trout in small streams: how much stream is enough? 

North American Journal of Fisheries Management 20:513-520.

Hirsch, C. L., S. E. Albeke, and T. P. Nesler. 2006. Range-wide 

status of Colorado River cutthroat trout. (Oncorhynchus clarkii 

pleuriticus): 2005. Colorado River cutthroat trout Conservation 

Team Report. Colorado Division of Wildlife, Fort Collins.

Hitt, N. P., C. A. Frissell, C. C. Muhlfeld, and F. W. Allendorf. 

2003. Spread of hybridization between native westslope cutthroat 

trout, Oncorhynchus clarki lewisi, and nonnative rainbow trout,

Oncorhynchus mykiss. Canadian Journal of Fisheries and Aquatic 

Sciences 60:1440-1451.

Huckins, C. J., E. Baker, K. D. Fausch, and J. B. K. Leonard. in 

press. Ecology and life history of coaster brook trout Salvelinus fontinalis and potential bottlenecks in their rehabilitation. 

Transactions of the American Fisheries Society.

Hudy, M., T. M. Thieling, and J. K. Whalen. 2004. A large-scale 

risk assessment of the biotic integrity of native brook trout watersheds. Pages 93-101 in S. E. Moore, R. F. Carline, and J. Dillon 

editors.Wild Trout VIII Symposium, Trout Unlimited, U.S. Fish 

and Wildlife Service, and U.S. Geological Survey Biological 

Resources Division, Washington, DC.

Hutchings, J. A. 2004. Norms of reaction and phenotypic plasticity 

in salmonid life histories. Pages 154-174 in A. P. Hendry and 

S. C. Stearns, editors. Evolution illuminated: salmon and their 

relatives. Oxford University Press, New York.

Isaak, D. J., and R. F. Thurow. 2006. Network-scale spatial and temporal variation in Chinook salmon (Oncorhynchus tshawytscha) 

redd distributions: patterns inferred from spatially continuous 

replicate surveys. Canadian Journal of Fisheries and Aquatic 

Sciences 63:285-296.

Jager, H. I., W. V. Winkle, and B. D. Holcomb. 1999. Would hydrologic climate changes in Sierra Nevada streams influence trout 

persistence? Transactions of the American Fisheries Society 

128:222-240.

Jakober, M. J., T. E. McMahon, R. F. Thurow, and C. G. Clancey. 

1998. Role of stream ice on fall and winter movements and habitat 

use by bull trout and cutthroat trout in Montana headwater streams. 

Transactions of the American Fisheries Society 127:223-235.
40 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

Jonsson, B., and N. Jonsson. 1993. Partial migration: niche shift 

versus sexual maturation in fishes. Reviews in Fish Biology and 

Fisheries 3:348-365.

Jowett, I. G. 1990. Factors related to the distribution and abundance of brown and rainbow trout in New Zealand clear-water 

rivers. New Zealand Journal of Marine and Freshwater Research 

24:429-440.

Kanda, N. 1998. Genetics and conservation of bull trout: comparison 

of population genetic structure among different genetic markers and hybridization with brook trout. Doctoral dissertation. 

University of Montana, Missoula.

Kanda, N., R. F. Leary, and F. W. Allendorf. 2002. Evidence of 

introgressive hybridization between bull trout and brook trout. 

Transactions of the American Fisheries Society 131:772-782.

Kennedy, B. M., D. P. Peterson, and K. D. Fausch. 2003. Different 

life histories of brook trout populations invading mid-elevation

and high-elevation cutthroat trout streams in Colorado. Western 

North American Naturalist 63:215-223.

Kitano, S. 2004. Ecological impacts of rainbow, brown and brook 

trout in Japanese inland waters. Global Environmental Research 

8:41-50.

Knapp, R. A., and K. R. Matthews. 1998. Eradication of nonnative 

fish by gill netting from a small mountain lake in California. 

Restoration Ecology 6:207-213.

Knight, D. H. 1994. Mountains and plains: the ecology of Wyoming 

landscapes. Yale University Press, New Haven, Connecticut.

Koel, T. M., P. E. Bigelow, P. D. Doepke, B. D. Ertel, and D. L. 

Mahoney. 2005. Nonnative lake trout result in Yellowstone cutthroat trout decline and impacts to bears and anglers. Fisheries 

(Bethesda) 30(11):10-19.

Koizumi, I., and K. Maekawa. 2004. Metapopulation structure of 

stream-dwelling Dolly Varden charr inferred from patterns of 

occurrence in the Sorachi River basin, Hokkaido, Japan. Freshwater Biology 49:973-981.

Kolar, C. S., and D. M. Lodge. 2001. Progress in invasion biology: 

predicting invaders. Trends in Ecology and Evolution 16:199-

204.

Kondratieff, M. C., and C. A. Myrick. 2006. How high can brook 

trout jump?: a laboratory evaluation of brook trout jumping 

performance. Transactions of the American Fisheries Society 

135:361-370.

Kot, M., M. A. Lewis, and P. v. d. Driessche. 1996. Dispersal data 

and the spread of invading organisms. Ecology 77:2027-2042.

Koyama, A., K. Kavanagh, and A. Robinson. 2005. Marine nitrogen in central Idaho riparian forests: evidence from stable 

isotopes. Canadian Journal of Fisheries and Aquatic Sciences 

62:518-526.

Kruse, C. G., W. A. Hubert, and F. J. Rahel. 2001. An assessment 

of headwater isolation as a conservation strategy for cutthroat 

trout in the Absaroka Mountains of Wyoming. Northwest Science 75:1-11.

Kwak, T. J., and T. F. Waters. 1997. Trout production dynamics and 

water quality in Minnesota streams. Transactions of the American 

Fisheries Society 126:35-48.

Larimore, R. W., W. F. Childers, and C. Heckrotte. 1959. Destruction 

and re-establishment of stream fish and invertebrates affected 

by drought. Transactions of the American Fisheries Society 

88:261-285.

Larson, G. L., and S. E. Moore. 1985. Encroachment of exotic 

rainbow trout into stream populations of native brook trout in the

southern Appalachian Mountains. Transactions of the American 

Fisheries Society 114:195-203.

Latterell, J. J., K. D. Fausch, C. Gowan, and S. C. Riley. 1998. 

Relationship of trout recruitment to snowmelt runoff flows and 

adult trout abundance in six Colorado mountain streams. Rivers 

6:240-250.

Lawler, J. J., D. White, and L. L. Master. 2003. Integrating representation and vulnerability: two approaches for prioritizing areas 

for conservation. Ecological Applications 13:1762-1772.

Leary, R. F., F. W. Allendorf, and S. H. Forbes. 1993. Conservation 

genetics of bull trout in the Columbia and Klamath river drainages. Conservation Biology 7:856-865.

Leary, R. F., B. B. Shepard, B. W. Sanborn, W. P. Dwyer, J. A. Brammer, R. A. Oswald, A. Tews, D. Kampwerth, M. Enk, R. Wagner, 

and L. Kaeding. 1998. Recommendations from the Westslope 

Cutthroat Trout Technical Committee for the conservation of the

westslope cutthroat trout in the Upper Missouri River drainage. Upper Missouri Westslope Cutthroat Trout Committee, Final Report. 

Montana Department of Fish, Wildlife, and Parks, Helena.

Lee, D., J. Sedell, B. Rieman, R. Thurow, and J. Williams. 1997. 

Assessment of the condition of aquatic ecosystems in the Interior 

Columbia River Basin. Chapter 4. Eastside Ecosystem Management Project. USDA Forest Service, PNW-GTR-405, Pacific 

Northwest Research Station, Portland, Oregon.

Leider, S. A. 1989. Increased straying by adult steelhead trout, 

Salmo gairdneri, following the 1980 eruption of Mount St. Helens. 

Environmental Biology of Fishes 24:219-229.

Lentz, H. S. 1998. Distribution and habitat of recently emerged 

young-of-the-year trout. Master’s thesis. University of Montana, 

Missoula.

Lever, C. 1996. Naturalized fishes of the world. Academic Press, 

San Diego, California.

Lewis, M. A. 1997. Variability, patchiness, and jump dispersal in 

the spread of an invading population. Pages 46-69 in D. Tilman 

and P. Karieva, editors. Spatial ecology: the role of space in 

population dynamics and interspecific interactions. Princeton 

University Press, Princeton, NJ.

Lubow, B. C. 1996. Optimal translocation strategies for enhancing 

stochastic metapopulation viability. Ecological Applications 

6:1268-1280.

Ludwig, D., R. Hilborn, and C. Walters. 1993. Uncertainty, resource 

exploitation, and conservation: lessons from history. Science 

260:17-36.

MacCrimmon, H. R., and J. S. Campbell. 1969. World distribution 

of brook trout, Salvelinus fontinalis. Journal of the Fisheries 

Research Board of Canada 26:1699-1725.

Maclay, D. J. 1940. Tentative fish management plan: Coeur d’ Alene 

National Forest. USDA Forest Service, Panhandle National Forests, Coeur d’Alene, Idaho.

Marchetti, M. P., P. B. Moyle, and R. Levine. 2004. Alien fishes in 

California watersheds: characteristics of successful and failed 

invaders. Ecological Applications 14:587-596.

McElhany, P., M. H. Ruckelshaus, M. J. Ford, T. C. Wainwright, 

and E. P. Bjorkstedt. 2000. Viable salmonid populations and the 

recovery of Evolutionarily Significant Units. U.S. Department of 

Commerce, NOAA Technical Memorandum NMFS-NWFSC-42, 

Seattle, Washington.

McFadden, J. T. 1961. A population study of the brook trout, Salvelinus fontinalis. Wildlife Monographs 7:1-73.

McGrath, C. 2004. Trophic roles of native greenback cutthroat 

trout and non-native brook trout in montane streams of Colorado. 

Doctoral dissertation. University of Colorado, Boulder.

McHugh, P., and P. Budy. 2005. An experimental evaluation of 

competitive and thermal effects on brown trout (Salmo trutta) 

and Bonneville cutthroat trout (Oncorhynchus clarkii utah) 

performance along an altitudinal gradient. Canadian Journal of 

Fisheries and Aquatic Sciences 62:2784-2795.

McIntosh, A. R. 2000. Habitat- and size-related variations in 

exotic trout impacts on native galaxiid fishes in New Zealand 

streams. Canadian Journal of Fisheries and Aquatic Sciences 

57:2140-2151.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 41

Meisner, J. D. 1990. Potential effect of climatic warming on the 

southern margins of the native range of brook trout, Salvelinus 

fontinalis. Canadian Journal of Fisheries and Aquatic Sciences 

47:1065-1070.

Meka, J. M., E. E. Knudsen, D. C. Douglas, and R. B. Benter. 2003. 

Variable migratory patterns of different adult rainbow trout life 

history types in a southwest Alaska watershed. Transactions of 

the American Fisheries Society 132:717-732.

Meyer, S. M. 2004. End of the wild. Boston Review: April/May 

2004. Available: http://bostonreview.net/BR29.2/meyer.html 

(accessed December 2004).

Milner, A. M., and R. G. Bailey. 1989. Salmonid colonization of 

new streams in Glacier Bay National Park, Alaska. Aquaculture 

and Fisheries Management 20:179-192.

Mobrand, L. E., J. Barr, L. Blankenship, D. E. Campton, T. T. 

P. Evelyn, T. A. Flagg, C. V. W. Mahnken, L. W. Seeb, P. R. 

Seidel, and W. W. Smoker. 2005. Hatchery reform in Washington state: principles and emerging issues. Fisheries (Bethesda) 

30(6):11-23.

Moller, H. 1996. Lessons for invasion theory from social insects. 

Biological Conservation 78:125-142.

Moller, S., and R. V. Kirk. 2003. Hydrologic alteration and its effect on 

trout recruitment in the South Fork Snake River. Project Completion Report, Idaho Department of Fish and Game, Boise.

Moore, K. V., and S. V. Gregory. 1988. Geomorphic and riparian 

influences on the distribution and abundance of salmonids in a 

Cascade mountain stream. USDA Forest Service, PSW-1101988, 

Portland, Oregon.

Moore, S. E., G. L. Larson, and B. Ridley. 1986. Population control 

of exotic rainbow trout in streams of a natural area park. Environmental Management 10:215-219.

Moore, S. E., B. Ridley, and G. L. Larson. 1983. Standing crops 

of brook trout concurrent with removal of rainbow trout from

selected streams in Great Smoky Mountains National Park. North 

American Journal of Fisheries Management 3:72-80.

Morita, K., J. Tsuboi, and H. Matsuda. 2004. The impact of exotic 

trout on native charr in a Japanese stream. Journal of Applied 

Ecology 41:962-972.

Morita, K., and S. Yamamoto. 2002. Effects of habitat fragmentation by damming on the persistence of stream-dwelling charr 

populations. Conservation Biology 16:1318-1323.

Morita, K., S. Yamamoto, and N. Hoshino. 2000. Extreme life 

history change of white-spotted charr (Salvelinus leucomaenis) 

after damming. Canadian Journal of Fisheries and Aquatic Sciences 57:1300-1306.

Morita, K., and A. Yokota. 2002. Population viability of streamresident salmonids after habitat fragmentation: a case study with 

white-spotted charr (Salvelinus leucomaenis) by an individual 

based model. Ecological Modelling 155:85-94.

Moyle, P. B., and T. Light. 1996. Biological invasions of freshwater: 

empirical rules and assembly theory. Biological Conservation 

78:149-161.

Moyle, P. B., and G. M. Sato. 1991. On the design of preserves to 

protect native fishes. Pages 155-169 in W. L. Minckley and J. E. 

Deacon, editors. Battle against extinction: native fish management 

in the American West. University of Arizona Press, Tucson.

Munro, A. R., T. E. McMahon, and J. R. Ruzycki. 2005. Natural 

chemical markers identify source and date of introduction of 

an exotic species: lake trout (Salvelinus namaycush) in Yellowstone Lake. Canadian Journal of Fisheries and Aquatic Sciences 

62:79-87.

Näslund, I., G. Millbrink, L. O. Eriksson, and S. Holmgren. 1993. 

Importance of habitat productivity differences, competition, and 

predation for the migratory behaviour of Arctic charr. Oikos 

66:538-546.

Nehring, R. B. 2004. Stream fisheries investigations. Federal Aid 

in Fish and Wildlife Restoration, Job Progress Report, Colorado 

Division of Wildlife, Fort Collins.

Nehring, R. B., and R. M. Anderson. 1993. Determination of population-limiting criticalsalmonid habitatsinColorado streams using

the Physical Habitat Simulation system. Rivers 4:1-19.

Nehring, R. B., and K. Thompson. 2003. Whirling disease investigations. Federal Aid in Fish and Wildlife Restoration, Final Progress 

Report, Colorado Division of Wildlife, Fort Collins.

Nelson, J. S., E. J. Crossman, H. Espinosa-Pérez, L. T. Findley, C. 

R. Gilbert, R. N. Lea, J. D. Williams, editors. 2004. Common 

and scientific names of fishes from the United States, Canada, 

and Mexico, 6th Edition. American Fisheries Society, Bethesda, 

Maryland.

Neville, H., J. Dunham, and M. Peacock. in press. Assessing connectivity in salmonid fishes with DNA microsatellite markers. Pages 

000-000 in K. R. Crooks and M. Sanjayan, editors. Connectivity 

conservation. Cambridge University Press, Cambridge, U.K.

Northcote, T. G. 1978. Migratory strategies and production in 

freshwater fishes. Pages 326-359 in S. D. Gerking, editor. Ecology 

of freshwater fish production. Blackwell Scientific Publications, 

Oxford, UK.

Northcote, T. G. 1992. Migration and residency in stream salmonids 

- some ecological considerations and evolutionary consequences. 

Nordic Journal of Freshwater Research 67:5-17.

Northcote, T. G. 1997. Potamodromy in salmonids - living and 

moving in the fast lane. North American Journal of Fisheries 

Management 17:1029-1045.

Noss, R. F. 1990. Indicators for monitoring biodiversity: a hierarchical approach. Conservation Biology 4:355-364.

Noss, R. F., and A. Y. Cooperrider. 1994. Saving Nature’s legacy: 

protecting and restoring biodiversity. Island Press, Washington, 

DC.

Novinger, D. C. 2000. Reversals in competitive ability: do cutthroat 

trout have a thermal refuge from competition with brook trout?

Doctoral dissertation. University of Wyoming, Laramie.

Novinger, D. C., and F. J. Rahel. 2003. Isolation management with 

artificial barriers as a conservation strategy for cutthroat trout in 

headwater streams. Conservation Biology 17:772-781.

Olsen, D. G., and M. C. Belk. 2005. Relationship of diurnal habitat 

use of native stream fishes of the eastern Great Basin to presence 

of introduced salmonids. Western North American Naturalist 

65:501-506.

Pascaul, M., P. Macchi, J. Urbanski, F. Marcos, C. R. Rossi, M. 

Novara, and P. Dell’Arciprete. 2002. Evaluating potential effects 

of exotic freshwater fish from incomplete species presence-absence 

data. Biological Invasions 4:101-113.

Paul, A. J., and J. R. Post. 2001. Spatial distribution of native and 

nonnative salmonids in streams of the eastern slopes of the Canadian Rocky Mountains. Transactions of the American Fisheries 

Society 130:417-430.

Peacock, M. M., and V. Kirchoff. 2004. Assessing the conservation value of hybridized cutthroat trout populations in the Quinn 

River Drainage, Nevada. Transactions of the American Fisheries 

Society 133:309-325.

Pefaur, J. E., and N. M. Sierra. 1998. Distribution and density of the 

trout Oncorhynchus mykiss (Salmoniformes: Salmonidae) in the 

Venezuelan Andes. Revista de Biologia Tropical 46:775-782.

Peter, A. 1998. Interruption of the river continuum by barriers and 

the consequences for migratory fish. Pages 99-112 in M. Jungwirth, S. Schmutz, and S. Weiss, editors. Fish migration and fish 

bypasses. Fishing News Books, Oxford, U.K.

Peter, A., E. Staub, C. Ruhlé, and T. Kindle. 1998. Interactions 

between brown and rainbow trout in the Alpine Rhine valley and 

its effects on their management. Schweiz Fischereiwissenschaft 

98:5-10. (in German)

Peterson, D. P., and K. D. Fausch. 2003a. Upstream movement by 

nonnative brook trout (Salvelinus fontinalis) promotes invasion 

of native cutthroat trout (Oncorhynchus clarki) habitat. Canadian 

Journal of Fisheries and Aquatic Sciences 60:1502-1516.
42 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

Peterson, D. P., and K. D. Fausch. 2003b. Testing population-level 

mechanisms of invasion by a mobile vertebrate: a simple conceptual framework for salmonids in streams. Biological Invasions 

5:239-259.

Peterson, D. P., K. D. Fausch, and G. C. White. 2004. Population 

ecology of an invasion: effects of brook trout on native cutthroat 

trout. Ecological Applications 14:754-772.

Pethiyagoda, R. 1994. Threats to the indigenous freshwater fishes 

of Sri Lanka and remarks on their conservation. Hydrobiologia 

285:189-201.

Petty, J. T., P. J. Lamothe, and P. M. Mazik. 2005. Spatial and seasonal dynamics of brook trout populations inhabiting a central 

Appalachian watershed. Transactions of the American Fisheries 

Society 134:572-587.

Poff, N. L., J. D. Allan, M. B. Bain, J. R. Karr, K. L. Prestegaard, B. 

D. Richter, R. E. Sparks, and J. C. Stromberg. 1997. The natural 

flow regime. BioScience 47:769-784.

Post, J. R., C. J. Mushens, A. J. Paul, and M. Sullivan. 2003. Assessment of alternatemanagementstrategiesforsustaining recreational fisheries: model development and application to bull trout. 

North American Journal of Fisheries Management 23:22-34.

Power, G. 1980. The brook charr, Salvelinus fontinalis. Pages 

141-203 in E. K. Balon, editor. Charrs: salmonid fishes of the 

genus Salvelinus. Dr. W. Junk Publishers, The Hague, The 

Netherlands.

Pressey, R. L. 1994. Ad hoc reservations: forward or backward 

steps in developing representative reserve systems? Conservation 

Biology 8:662-668.

Pringle, C. M., M. C. Freeman, and B. J. Freeman. 2000. Regional effects of hydrologic alterations on riverine macrobiota 

in the New World: tropical-temperate comparison. BioScience 

50:807-823.

Propst, D. L., J. A. Stefferud, and P. R. Turner. 1992. Conservation and status of Gila trout, Oncorhynchus gilae. Southwestern 

Naturalist 37:117-125.

Quinn, T. P. 1993. A review of homing and straying of wild and 

hatchery produced salmon. Fisheries Research 18:29-44.

Quinn, T. P. 2005. The behavior and ecology of Pacific salmon and 

trout. American Fisheries Society, Bethesda, Maryland.

Quinn, T. P., M. T. Kinnison, and M. J. Unwin. 2001. Evolution of 

Chinook salmon (Oncorhynchus tshawytscha) populations in New 

Zealand: pattern, rate, and process. Genetica 112-113:493-513.

Quist, M., and W. A. Hubert. 2004. Bioinvasive species and the 

preservation of cutthroat trout in the western United States: 

ecological, social, and economic issues. Environmental Science 

and Policy 7:303-313.

Rahel, F. J. 1997. From Johnny Appleseed to Dr. Frankenstein: 

changing values and the legacy of fisheries management. Fisheries (Bethesda) 22(8):8-9.

Rahel, F. J. 2000. Homogenization of fish faunas across the United 

States. Science 288:854-856.

Rahel, F. J. 2004. Unauthorized fish introductions: fisheries management of the people, for the people, or by the people? American 

Fisheries Society Symposium 44:431-443.

Rahel, F. J., and N. P. Nibbelink. 1999. Spatial patterns in relations 

among brown trout (Salmo trutta) distribution, summer air temperature, and stream size in Rocky Mountain streams. Canadian 

Journal of Fisheries and Aquatic Sciences 56:43-51.

Redenbach, Z., and E. B. Taylor. 2003. Evidence for bimodal hybrid zones between two species of char (Pisces: Salvelinus) in 

northwestern North America. Journal of Evolutionary Biology 

16:1135-1148.

Reeves, G. H., L. E. Benda, K. M. Burnett, P. A. Bisson, and J. 

R. Sedell. 1995. A disturbance-based ecosystem approach to 

maintaining and restoring freshwater habitats of evolutionarily 

significant units of anadromous salmonids in the Pacific Northwest. American Fisheries Society Symposium 17:334-349.

Renner, D. 2005. Distribution and habitat use of bull trout following 

the removal of nonnative brook trout. Master’s thesis. Oregon 

State University, Corvallis.

Rhymer, J. M., and D. Simberloff. 1996. Extinction by hybridization 

and introgression. Annual Review of Ecology and Systematics 

27:83-109.

Rich, C. F. J., T. E. McMahon, B. E. Rieman, and W. L. Thompson. 

2003. Local-habitat, watershed, and biotic features associated 

with bull trout occurrence in Montana streams. Transactions of 

the American Fisheries Society 132:1053-1064.

Rieman, B., and J. Clayton. 1997. Wildfire and native fish: issues 

of forest health and conservation of sensitive species. Fisheries 

(Bethesda) 22(11):6-15.

Rieman, B., J. Dunham, and J. Peterson. 1999. Development of a 

database to support a multi-scale analysis of the distribution of 

westslope cutthroat trout. USDA Forest Service, Rocky Mountain 

Research Station, Boise, Idaho.

Rieman, B. E., and F. W. Allendorf. 2001. Effective population size 

and genetic conservation criteria for bull trout. North American 

Journal of Fisheries Management 21:756-764.

Rieman, B. E., and J. B. Dunham. 2000. Metapopulations and 

salmonids: a synthesis of life history patterns and empirical 

observations. Ecology of Freshwater Fish 9:51-64.

Rieman, B. E., D. C. Lee, D. Chandler, and D. Myers. 1997a. Does 

wildfire threaten extinction for salmonids? Responses of redband 

trout and bull trout following recent fires in the Boise River Basin, 

Idaho. Pages 45-57 in J. Greenlee, editor. Proceedings of the conference on wildfire and Threatened and Endangered species and 

habitats, November 13-15, 1995, Coeur d’Alene Idaho. International 

Association of Wildland Fire, Fairfield, Washington.

Rieman, B. E., D. C. Lee, and R. F. Thurow. 1997b. Distribution, 

status, and likely future trends of bull trout within the Columbia 

River and Klamath River basins. North American Journal of 

Fisheries Management 17:1111-1125.

Rieman, B. E., and J. D. McIntyre. 1993. Demographic and habitat 

requirements for conservation of bull trout. USDA Forest Service, Intermoutain Research Station, General Technical Report 

INT-302, Ogden, UT.

Rieman, B. E., and J. D. McIntyre. 1995. Occurrence of bull trout in 

naturally fragmented habitat patches of varied size. Transactions 

of the American Fisheries Society 124:285-296.

Rieman, B. E., D. L. Myers, and R. L. Nielsen. 1994. The use of 

otolith microchemistry to discriminate Oncorhynchus nerka of

resident and anadromous origin. Canadian Journal of Fisheries 

and Aquatic Sciences 51:68-77.

Rieman, B. E., J. T. Peterson, and D. L. Myers. 2006. Have brook 

trout (Salvelinus fontinalis) displaced bull trout (Salvelinus confluentus) along longitudinal gradients in central Idaho streams? 

Canadian Journal of Fisheries and Aquatic Sciences 63:63-78.

Riley, S. C., K. D. Fausch, and C. Gowan. 1992. Movement of brook 

trout (Salvelinus fontinalis) in four small subalpine streams in 

northern Colorado. Ecology of Freshwater Fish 1:112-122.

Riva Rossi, C. M., E. P. Lessa, and M. A. Pascual. 2004. The 

origin of introduced rainbow trout (Oncorhynchus mykiss) in 

the Santa Cruz River, Patagonia, Argentina, as inferred from 

mitchondrial DNA. Canadian Journal of Fisheries and Aquatic 

Sciences 61:1095-1101.

Rodríguez, M. A. 2002. Restricted movement in stream fish: the 

paradigm is incomplete, not lost. Ecology 83:1-13.

Roghair, C. N., C. A. Dollof, and M. K. Underwood. 2002. Response 

of a brook trout population and instream habitat to a catastrophic

flood and debris flow. Transactions of the American Fisheries 

Society 131:718-730.

Ruckelshaus, M., P. McElhany, and M. J. Ford. 2003. Recovering 

species of conservation concern - are populations expendable?

Pages 305-329 in P. Kareiva and S. A. Levin, editors. The importance of species: perspectives on expendability and triage. 

Princeton University Press, Princeton, New Jersey.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006 43

Ruiz, V. H., and T. M. Berra. 1994. Fishes of the high Biobio River 

of south-central Chile with notes on diet and speculations on

the origin of the ichthyofauna. Ichthyological Exploration of 

Freshwaters 5:5-18.

Schlosser, I. J. 1991. Stream fish ecology: a landscape perspective. 

BioScience 41:704-712.

Schlosser, I. J., and P. L. Angermeier. 1995. Spatial variation in 

demographic processes in lotic fishes: conceptual models, empirical evidence, and implications for conservation. American 

Fisheries Society Symposium 17:360-370.

Schmetterling, D. A., and S. B. Adams. 2004. Summer movements 

within the fish community of a small montane stream. North 

American Journal of Fisheries Management 24:1163-1172.

Schmutz, S., and M. Jungwirth. 1999. Fish as indicators of large 

river connectivity: the Danube and its tributaries. Archiv für 

Hydrobiologia 115/3:329-348.

Scott, J. M., and B. Csuti. 1997. Noah worked two jobs. Conservation Biology 11:1255-1257.

Shea, K., and P. Chesson. 2002. Community ecology theory as 

a framework for biological invasions. Trends in Ecology and 

Evolution 17:170-176.

Shepard, B. B. 2004. Factors that may be influencing nonnative brook 

troutinvasion and their displacement of nativewestslope cutthroat

trout in three adjacent southwestern Montana streams. North 

American Journal of Fisheries Management 24:1088-1100.

Shepard, B. B., B. E. May, and W. Urie. 2005. Status and conservation 

of westslope cutthroat trout within the western United States. North 

American Journal of Fisheries Management 25:1426-1440.

Shepard, B. B., R. Spoon, and L. Nelson. 2002. A native westslope 

cutthroat trout population responds positively after brook trout 

removal and habitat restoration. Intermountain Journal of Sciences 8:191-211.

Sigler, W. F., W. T. Helm, P. A. Kucera, S. Vigg, and G. W. Workman. 1983. Life history of the Lahontan cutthroat trout, Salmo 

clarki henshawi, in Pyramid Lake, Nevada. The Great Basin 

Naturalist 43:1-29.

Simberloff, D. 2003. How much information on population biology 

is needed to manage introduced species? Conservation Biology 

17:83-92.

Simberloff, D., and J. Cox. 1987. Consequences and costs of conservation corridors. Conservation Biology 1:63-71.

Simberloff, D., J. A. Farr, J. Cox, and D. W. Mehlman. 1992. 

Movement corridors: conservation bargains or poor investments? 

Conservation Biology 6:493-504.

Simon, K. S., and C. R. Townsend. 2003. Impacts of freshwater 

invaders at different levels of ecological organization, with emphasis on salmonids and ecosystem consequences. Freshwater 

Biology 48: 982-994.

Sloat, M. R., B. B. Shepard, and R. G. White. 2005. Influence of 

stream temperature on the spatial distribution of westslope cutthroat trout growth potential within the Madison River basin, 

Montana. North American Journal of Fisheries Management 

25:225-237.

Smith, K. L., and M. L. Jones. 2005. Watershed-level sampling 

effort requirements for determining riverine fish species composition. Canadian Journal of Fisheries and Aquatic Sciences 

62:1580-1588.

Smith, L. W., E. Dittmer, M. Prevost, and D. R. Burt. 2000. Breaching of a small irrigation dam in Oregon: a case history. North 

American Journal of Fisheries Management 20:205-219.

Snyder, J. O. 1917. Fishes of the Lahontan system of Nevada and 

northeastern California. U.S. Bureau of Fisheries Bulletin 

35:33-86.

Soulé, M. E., and L. S. Mills. 1998. No need to isolate genetics. 

Science 282:1658-1659.

Soulé, M. E., and J. Terborgh. 1999. Continental conservation: 

scientific foundations of regional reserve networks. Island Press, 

Washington, DC.

Speas, C., M. Fowden, M. Gorges, T. Rinkes, G. Eaglin, and B. 

Wengert. 1994. Conservation plan for Colorado River cutthroat 

trout (Oncorhynchus clarki pleuriticus) for the Little Snake 

River drainage in southeastern Wyoming. U.S. Forest Service, 

Medicine Bow National Forest, Laramie, Wyoming.

Spencer, C. N., B. R. McClelland, and J. A. Stanford. 1991. Shrimp 

introduction, salmon collapse, and eagle displacement: cascading interactions in the food web of a large aquatic ecosystem. 

BioScience 41:14-21.

Stacey, P. B., and M. Taper. 1992. Environmental variation and 

the persistence of small populations. Ecological Applications 

2:18-29.

Stelfox, J. D., D. M. Baayens, A. J. Paul, and G. E. Shumaker. 

2001. Quirk Creek brook trout suppression project. Pages 37-46 

in M. K. Brewin, A. J. Paul, and M. Monita, editors. Bull Trout 

II Conference Proceedings. Trout Unlimited Canada, Calgary, 

Alberta.

Stonecypher, R. W., W. A. Hubert, and W. A. Gern. 1994. Effect of 

reduced incubation temperatures on survival of trout embryos. 

The Progressive Fish-Culturist 56:180-184.

Strange, E. M., and T. C. Foin. 1999. Interaction of physical and 

biological processes in the assembly of stream fish communities. 

Pages 311-337 in E. Weiher and P. Keddy, editors. Ecological 

assembly rules: perspectives, advances, retreats. Cambridge 

University Press, Cambridge, UK.

Suzuki, R., and Y. Fukuda. 1971. Survival potential of F1 hybrids 

among salmonid fishes. Bulletin of the Freshwater Fisheries 

Laboratory of Tokyo 21:69-83.

Takami, T., and T. Aoyama. 1999. Distributions of rainbow and 

brown trouts in Hokkaido, northern Japan. Wildlife Conservation 

Japan 4:41-48. (In Japanese)

Tallman, R. F., and M. C. Healey. 1994. Homing, straying, and gene 

flow among seasonally separated populations of chum salmon 

(Oncorhynchus keta). Canadian Journal of Fisheries and Aquatic 

Sciences 51:577-588.

Taniguchi, Y., and S. Nakano. 2000. Condition-specific competition: implications for the altitudinal distribution of stream fishes. 

Ecology 81:2027-2039.

Taylor, E. B. 1990. Environmental correlates of life-history variation in juvenile chinook salmon, Oncorhynchus tshawytscha 

(Walbaum). Journal of Fish Biology 37:1-17.

Thompson, P. D., and F. J. Rahel. 1998. Evaluation of artificial barriers in small Rocky Mountain streams for preventing upstream 

movement of brook trout. North American Journal of Fisheries 

Management 18:206-210.

Thorpe, J. E. 1994. Salmonid flexibility: responses to environmental extremes. Transactions of the American Fisheries Society 

123:606-612.

Thorpe, J. E., and N. B. Metcalfe. 1998. Is smolting a positive or a 

negative developmental decision? Aquaculture 168:95-103.

Thurow, R. F., D. C. Lee, and B. E. Rieman. 1997. Distribution and 

status of seven native salmonids in the interior Columbia River 

Basin and portions of the Klamath River and Great Basins. North 

American Journal of Fisheries Management 17:1094-1110.

Thybony, S., R. G. Rosenberg, and E. M. Rosenberg. 1985. The 

Medicine Bows: Wyoming’s mountain country. Caxton Printers, 

Caldwell, Idaho.

Tilman, D., R. M. May, C. L. Lehman, and M. A. Nowak. 1994. 

Habitat destruction and the extinction debt. Nature 371:65-66.

Townsend, C. R., and T. A. Crowl. 1991. Fragmented population 

structure in a native New Zealand fish - an effect of introduced 

brown trout. Oikos 61:347-354.

U.S. Army Corps of Engineers. 2006. National inventory of dams. 

Available: http://crunch.tec.army.mil/nid/webpages/nid.cfm. 

(accessed April 2006)

U.S. Fish and Wildlife Service. 1998a. Greenback cutthroat trout recovery plan. U.S. Fish and Wildlife Service, Denver, Colorado.
44 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-174. 2006

U.S. Fish and Wildlife Service. 1998b. Proposal to list the CoastalPuget Sound, Jarbidge River, and St. Mary-Belly River population 

segments of bull trout as Threatened species. Federal Register 

63:31693-31710.

U.S. Fish and Wildlife Service. 2002. Bull trout (Salvelinus confluentus) draft recovery plan. Available: http://pacific.fws.gov/

bulltrout/colkla/recovery/Default.htm. (accessed April 2006).

U.S. Fish and Wildlife Service. 2003. Endangered and threatened 

wildlife and plants: reconsidered finding for an amended petition 

to list the westslope cutthroat trout as threatened throughout its

range. Federal Register 68:46989-47009.

Vaughan, D. M. 2002. Potential impact of road-stream crossings 

(culverts) on the upstream passage of aquatic macroinvertebrates. 

in USDA Forest Service Report, San Dimas Technology and 

Development Center, San Dimas, California. Available on line: 

http://fsweb.sdtdc.wo.fs.fed.us/programs/eng/aopxing/Products/

Products.shtml.

Vermeij, G. J. 1996. An agenda for invasion biology. Biological 

Conservation 78:3-9.

Walters, C. J. 1986. Adaptive management of renewable resources. 

MacMillan Publishing Company, New York.

Waples, R. S. 1995. Evolutionarily significant units and the conservation of biological diversity under the Endangered Species Act. 

American Fisheries Society Symposium 17:8-27.

Waples, R. S. 2004. Salmonid insights into effective population 

size. Pages 295-314 in A. P. Hendry and S. C. Stearns, editors. 

Evolution illuminated: salmon and their relatives. Oxford University Press, Oxford, UK.

Warren, M. L., and M. G. Pardew. 1998. Road crossings as barriers 

to small-stream fish movement. Transactions of the American 

Fisheries Society 127:637-644.

Waters, T. F. 1983. Replacement of brook trout by brown trout over 

15 years in a Minnesota stream: production and abundance. 

Transactions of the American Fisheries Society 112:137-146.

Weigel, D. E., J. T. Peterson, and P. Spruell. 2003. Introgressive 

hybridization between native cutthroat trout and introduced 

rainbow trout. Ecological Applications 13:38-50.

Wells, B. K., B. E. Rieman, J. L. Clayton, D. L. Horan, and C. M. 

Jones. 2003. Relationships between water, otolith, and scale 

chemistries of westslope cutthroat trout from the Coeur d’Alene 

River, Idaho: the potential application of hard-part chemistry to 

describe movements in fresh water. Transactions of the American 

Fisheries Society 132:409-424.

Wiens, J. A. 1976. Population responses to patchy environments. 

Annual Review of Ecology and Systematics 7:81-120.

Wiens, J. A. 2002. Riverine landscapes: taking landscape ecology 

into the water. Freshwater Biology 47:501-515.

Wiley, R. W. 2003. Planting trout in Wyoming high-elevation wilderness waters. Fisheries (Bethesda) 28(1):22-27.

Willson, M. F., and K. C. Halupka. 1995. Anadromous fish as 

keystone species in vertebrate communities. Conservation Biology 9:489-497.

Wiltzius, W. J. 1985. Fish culture and stocking in Colorado, 1872-

1978. Colorado Division of Wildlife, Report 12, Denver.

Winston, M. R., C. M. Taylor, and J. Pigg. 1991. Upstream extirpation of four minnow species due to damming of a prairie stream. 

Transactions of the American Fisheries Society 120:98-105.

With, K. A. 2002. The landscape ecology of invasive spread. Conservation Biology 16:1192-1203.

Wofford, J. E., R. E. Gresswell, and M. A. Banks. 2005. Influence 

of barriersto movement on within-watershed genetic variation of

coastal cutthroat trout. Ecological Applications 15:628-637.

Wohl, E. E. 2001. Virtual rivers. Yale University Press, New Haven, 

Connecticut.

Yamamoto, S., K. Morita, and A. Goto. 1999 Geographic variations 

in life-history characteristics of white-spotted charr (Salvelinus 

leucomaenis). Canadian Journal of Zoology 77:871-878.

Yamamoto, S., K. Morita, I. Koizumi, and K. Maekawa. 2004. 

Genetic differentiation of white-spotted charr (Salvelinus leucomaenis) populations after habitat fragmentation: spatial-temporal 

changes in gene frequencies. Conservation Genetics 5:529-538.

Young, M., and A. Harig. 2001. A critique of the recovery of greenback cutthroat trout. Conservation Biology 15:1575-1584.

Young, M. K. 1994. Mobility of brown trout in south-central Wyoming streams. Canadian Journal of Zoology 72:2078-2083.

Young, M. K., editor. 1995. Conservation assessment for inland 

cutthroat trout. USDA Forest Service, Rocky Mountain Station, General Technical Report RM-GTR-256, Fort Collins, 

Colorado.

Young, M. K. 1996. Summer movements and habitat use by Colorado 

River cutthroat trout (Oncorhynchus clarki pleuriticus) in small, 

montane streams. Canadian Journal of Fisheries and Aquatic 

Sciences 53:1403-1408.

Young, M. K., and P. M. Guenther-Gloss. 2004. Population characteristics of greenback cutthroat trout in streams: their relation to 

model predictions and recovery criteria. North American Journal 

of Fisheries Management 24:184-197.

Young, M. K., P. M. Guenther-Gloss, and A. D. Ficke. 2005. 

Predicting cutthroat trout (Oncorhynchus clarki) abundance 

in high-elevation streams: revisiting a model of translocation 

success. Canadian Journal of Fisheries and Aquatic Sciences 

62:2399-2408.

Young, M. K., A. L. Harig, B. D. Rosenlund, and C. Kennedy. 

2002. Recovery history of greenback cutthroat trout: population 

characteristics, hatchery involvement, and bibliography. USDA 

Forest Service, Rocky Mountain Station, General Technical 

Report: http://www.fs.fed.us/rm/pubs/rmrs_gtr88.

Young, M. K., R. B. Rader, and T. A. Belish. 1997. Influence of 

macroinvertebrate drift and light on the activity and movement 

of Colorado River cutthroat trout. Transactions of the American 

Fisheries Society 126:428-437.

Young, M. K., R. N. Schmal, T. W. Kohley, and V. G. Leonard. 

1996. Conservation status of Colorado River cutthroat trout. 

USDA Forest Service, Rocky Mountain Forest and Range Experiment Station, General Technical Report RM-GTR-282, Fort 

Collins, Colorado.

Zimmerman, C. E., and G. H. Reeves. 2000. Population structure 

of sympatric anadromous and nonanadromous Oncorhynchus 

mykiss: evidence from spawning surveys and otolith microchemistry. Canadian Journal of Fisheries and Aquatic Sciences 

57:2152-2162.
The U.S. Department of Agriculture (USDA) prohibits discrimination in all its 

programs and activities on the basis of race, color, national origin, age, disability, 

and where applicable, sex, marital status, familial status, parental status, religion, 

sexual orientation, genetic information, political beliefs, reprisal, or because all 

or part of an individual’s income is derived from any public assistance program. 

(Not all prohibited bases apply to all programs.) Persons with disabilities who 

require alternative means for communication of program information (Braille, 

large print, audiotape, etc.) should contact USDA’s TARGET Center at (202) 

720-2600 (voice and TDD).

To file a complaint of discrimination, write to USDA, Director, Office of Civil 

Rights, 1400 Independence Avenue, S.W., Washington, DC 20250-9410, or call 

(800) 795-3272 (voice) or (202) 720-6382 (TDD). USDA is an equal opportunity 

provider and employer.

The Rocky Mountain Research Station develops scientific information and technology to improve management, protection, and use of 

the forests and rangelands. Research is designed to meet the needs 

of National Forest managers, Federal and State agencies, public and 

private organizations, academic institutions, industry, and individuals.

Studies accelerate solutions to problems involving ecosystems, 

range, forests, water, recreation, fire, resource inventory, land reclamation, community sustainability, forest engineering technology, multiple 

use economics, wildlife and fish habitat, and forest insects and diseases. 

Studies are conducted cooperatively, and applications may be found 

worldwide.

Research Locations

Flagstaff, Arizona Reno, Nevada

Fort Collins, Colorado* Albuquerque, New Mexico

Boise, Idaho Rapid City, South Dakota

Moscow, Idaho Logan, Utah 

Bozeman, Montana Ogden, Utah 

Missoula, Montana Provo, Utah 

*Station Headquarters, Natural Resources Research Center,

2150 Centre Avenue, Building A, Fort Collins, CO 80526

RMRS

ROCKY MOUNTAIN RESEARCH STATION