United States

Department

of Agriculture

Forest Service

Rocky Mountain

Research Station

General Technical 

Report RMRS-GTR-290

September 2012

Climate Change, Forests, 

Fire, Water, and Fish:

Building resilient landscapes, streams, 

and managers

Charles Luce, Penny Morgan, Kathleen Dwire, Daniel Isaak, 

Zachary Holden, and Bruce Rieman
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Luce, Charles; Morgan, Penny; Dwire, Kathleen; Isaak, Daniel; Holden, Zachary; 

Rieman, Bruce. 2012. Climate change, forests, fire, water, and fish: Building 

resilient landscapes, streams, and managers. Gen. Tech. Rep. RMRS-GTR-290. 

Fort Collins, CO: U.S. Department of Agriculture, Forest Service, Rocky 

Mountain Research Station. 207 p.

Abstract

Fire will play an important role in shaping forest and stream ecosystems as the 

climate changes. Historic observations show increased dryness accompanying 

more widespread fire and forest die-off. These events punctuate gradual changes 

to ecosystems and sometimes generate stepwise changes in ecosystems. Climate 

vulnerability assessments need to account for fire in their calculus. The biophysical 

template of forest and stream ecosystems determines much of their response to fire. 

This report describes the framework of how fire and climate change work together 

to affect forest and fish communities. Learning how to adapt will come from testing, 

probing, and pushing that framework and then proposing new ideas. The western U.S. 

defies generalizations, and much learning must necessarily be local in implication. 

This report serves as a scaffold for that learning. It comprises three primary chapters on 

physical processes, biological interactions, and management decisions, accompanied 

by a special section with separately authored papers addressing interactions of fish 

populations with wildfire. Any one of these documents could stand on its own. 

Taken together, they serve as a useful reference with varying levels of detail for land 

managers and resource specialists. Readers looking for an executive summary are 

directed to the sections titled “Introduction” and “Next Steps.”

Acknowledgments

This synthesis was funded by the Joint Fire Sciences Program. The authors wish to 

thank a large number of people who have participated in discussions on synthesizing 

these ideas. We would particularly like to thank three colleagues who joined us in a 

workshop in February 2009 but did not have the time to join in the preparation of this 

GTR: Matt Dare, Claire McGrath, and John Buffington. We would also like to thank 

the authors of the Advanced Topics papers: Robert Gresswell, Jason Dunham, Helen 

Neville, Amanda Rosenberger, Mike Young, and John Rinne.
Climate Change, Forests, Fire, Water, 

and Fish:

Building resilient landscapes, streams, 

and managers

Charles Luce, Penny Morgan, Kathleen Dwire, Daniel Isaak,

 Zachary Holden, and Bruce Rieman

with

Advanced Topics on Fish Populations and Fire 

by various authors

United States Department of Agriculture

Forest Service

Rocky Mountain Research Station

General Technical Report RMRS-GTR-290

September 2012
Contents

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

Key Debates....................................................................................................................... 3

Framing Solutions............................................................................................................... 4

Part I: The Physical System .......................................................................................8

A. Climate ................................................................................................................................... 8

Climate Change.................................................................................................................. 8

Fire................................................................................................................................... 14

B. Hydrology ............................................................................................................................. 15

Climate Change................................................................................................................ 15

Fire................................................................................................................................... 21

Fire and Climate Change: Feedback and Cumulative Effects............................................. 26

Implications for Aquatic Biota .......................................................................................... 27

C. Stream Temperature .............................................................................................................. 28

Climate Change................................................................................................................ 28

Fire................................................................................................................................... 30

Implications for Aquatic Biota .......................................................................................... 31

D. Geomorphology.................................................................................................................... 32

Climate Change................................................................................................................ 32

Fire................................................................................................................................... 33

Part II: Biological Systems......................................................................................39

A. Forests, Climate Change, and Fire ......................................................................................... 39

We Live in a Fire Environment.......................................................................................... 39

Changing Fire Regimes..................................................................................................... 40

Burn Severity.................................................................................................................... 42

Burn Severity Indicators Inferred From Satellite and Field Data......................................... 45

Will Future Fires be Larger and More Severe, and If So, So What? .................................... 47

Fuel Treatment Effectiveness............................................................................................. 47

Bark Beetles and Burn Severity ......................................................................................... 48

Key Uncertainties............................................................................................................. 49

Fire Management Strategies.............................................................................................. 50

Summary.......................................................................................................................... 51

B. Riparian Forests, Climate Change, Fire.................................................................................. 51

Riparian Vegetation, Values, and Connection to Streams................................................... 51

Disturbances in Riparian Areas......................................................................................... 59

Flooding........................................................................................................................... 60

Fire................................................................................................................................... 61

Climate Change, Fire, and Riparian Values and Functions ................................................ 72

C. Fish, Fire, Forest Management, and Climate Change............................................................. 74

The Responseo of Fish Populations to Fire ........................................................................ 76

Interaction With Land Management.................................................................................. 79

Interactions With a Changing Climate .............................................................................. 81

Connections to Conservation............................................................................................ 85
Part III: Management Actions and Decisions..........................................................87

A. Actions.................................................................................................................................. 87

Action Choices................................................................................................................. 87

Effectiveness of Actions .................................................................................................... 88

B. Framing Issues and Decisions................................................................................................ 90

The Broad Scope of Debate.............................................................................................. 90

Simplifying the Frame....................................................................................................... 92

Applying the Principles .................................................................................................... 94

Evaluating Risks................................................................................................................ 98

Addressing Uncertainty .................................................................................................. 100

C. A Changing Climate ............................................................................................................ 102

A Thought Experiment .................................................................................................... 104

Integration with Water Resources Management .............................................................. 105

Decision Making for an Uncertain and Dynamic Future................................................. 107

Next Steps............................................................................................................111

References ...........................................................................................................120

Advanced Topics on Fish Populations and Fire .....................................................157

Fire and Fish: a Synthesis of Observation and Experience ............................159

Bruce Rieman, Robert Gresswell, and John Rinne

Genetic Variation Reveals Influence of Landscape Connectivity on 

Population Dynamics and Resiliency of Western Trout in 

Disturbance-Prone Habitats.........................................................................177

Helen M. Neville, R.E. Gresswell, and J.B. Dunham

Fish Life Histories, Wildfire, and Resilience—A Case Study of 

Rainbow Trout in the Boise River, Idaho.......................................................187

Amanda E. Rosenberger, Jason B. Dunham, and Helen Neville

Aquatic Species Invasions in the Context of Fire and Climate Change .........195

Michael K. Young
USDA Forest Service RMRS-GTR-290. 2012. 1

Introduction

As the climate changes in the western United States, streams are warming, low 

flows in summer are declining, and winter floods are occurring more often in places where snowmelt is the main source of water (Stewart and others 2005; Hamlet 

and Lettenmaier 2007; Luce and Holden 2009; Isaak and others 2010). Some of the 

changes have been subtle, others more noticeable, and they are expected to shift distributions of fishes (Rieman and others 2007; Wenger and others 2011a; Wenger and 

others 2011b). At the same time, the terrestrial ecosystems surrounding the mountain 

streams of the West are changing in response to the same climatic signals. Drier years 

and drier summers have often led to more large fires, many of which are more severe (Dillon and others 2011). Further, fire regimes are shifting, with fires becoming 

more frequent in some places and less frequent in others, and potential conversion of 

forests to shrubs in some places (Pierce and others 2004; Breshears and others 2005; 

Westerling and others 2006; Morgan and others 2008; Westerling and others 2011).

Fires have long been prevalent in Western mountain landscapes. Many, but not all, 

ecosystems benefit from the biomass consumption, cycling of nutrients, rejuvenation 

of vegetation, and changing vegetation composition and structure after fires (Agee 

1993). Indeed many species and ecological communities in the western United 

States depend on fire in some form. Some benefit from frequent fires that consume 

small amounts of fuel, while others, seemingly paradoxically, thrive as a result of 

infrequent but severe fires that consume most of the available fuel in their path. Thus, 

fire itself has been long recognized as a considerable influence, apart from any consideration of a changing climate, on forest and stream ecosystems (e.g., Bisson and 

others 2003; Shakesby and Doerr 2006; Hessburg and others 2007).

The number of large fires has increased in recent decades, and future annual area 

burned is likely to increase further with concurrent concerns over costs of fire management and threats to safety of people and property (NWCG [National Wildfire 

Coordinating Group] 2009; Spracklen and others 2009; Littell and others 2010). 

Although not all environments are equally prone to fire, and humans have been 

very effective at detecting and suppressing the majority of fires when they are small 

(Stephens and Ruth 2005), forest fires will continue to occur. Global, national and 

regional trends of increasing number of large fires in recent decades are likely to 

continue with implications for both terrestrial and aquatic systems.

Fire and related disturbances will be an agent of climate change in shifting forest 

ecosystems (Dale and others 2001; Jentsch and others 2007; Turner 2010). Tree mortality can be caused directly by climate, or it may be induced by fire that is in turn 

responding to climate. Sometimes, the loss of the current forest canopy can pave the 

way for new species and even life forms (e.g., shrubs and grasses). Thus, climate 

change and climate variability have both direct and indirect implications for fish, 

streams, and aquatic ecosystems. Fire may become a critical point in the progression 

of individual forest stands or streams, where ecosystems may either gradually shift 

in response to climate change punctuated by fire and recovery (Figure 1a), much 

like they always have, or where ecosystems are relatively non-responsive to climate 

between events that provide the catalyst to adjust to new climate conditions (Figure 

1b). This new transitioning role of fire as “coup de grace” will pose new challenges 

for land managers who are well versed in the cyclic dynamics of forests. Of course, 

this simple model must be thought of in different terms in the context of changing 

disturbance frequency and severity as well. Providing sustained ecosystem services 

through seemingly unpredictable change-points may represent a primary challenge 

for managers.
2 USDA Forest Service RMRS-GTR-290. 2012.

While natural systems have evolved adaptations to the kinds of disturbances provided by fire, plant and animal populations may not be resilient to fires when fire 

regimes change, or when the landscape context of those fires changes. Large trees 

that survived many surface fires in the past may die in high severity fires, and regeneration of new trees may fail if fire recurs before young trees grow to fire-resistant 

size. Where serotinous cones have aided rapid post-fire regeneration, such regeneration will be less successful if fires recur before trees are old enough to produce 

abundant cones. For species relying on recolonization through dispersal from unburned refugia, very large, severe fires may present too great a barrier. Trees may not 

regenerate successfully following high severity fires at lower timberline if the postfire environment is less conducive than in the past or if invasive plants pose severe 

competition. An awareness of how the chain of consequences from climate change 

interacts with natural adaptations will be critical to forming solutions that maintain 

valued ecosystem components and processes into the future.

Within the pantheon of adaptation and mitigation concepts and approaches, two 

terms, resistance and resilience, stand out as critical ideas (e.g., Holling 1973; Waide 

1988; Millar and others 2007 and see Resilience text box). Resistance is the ability 

of an ecosystem to experience stressors but not change. For example, old ponderosa pine and Douglas-fir trees with thick bark are very resistant to surface fires. 

Engineers describe resilience as the ability to return to a given state despite sometimes formidable changes. Ecologists, sociologists, and psychologists share a more 

generalized definition of resilience as the capacity to absorb and weather change in 

a way that both combines and transcends the engineering concepts. Nonetheless, 

the engineering-oriented metaphor highlights the point that if climate changes, there 

seems to be a difficulty in applying either concept, as resistance must eventually be 

overcome, and it’s difficult to “bounce back” if the driving pressures are maintained 

if not growing. The concept of facilitated change fills in this difficult space. For example, while thinning might be seen as a resistance step for fire in one context, it can 

also serve to help forests cope with a changed water balance in a more predictable 

manner than without treatment.

Environmental Change

Ecological Departur

e

a

b

Figure 1: Conceptual roles for disturbance 

in a changing climate. Disturbance could 

continue to operate much as it always 

has (a), with unique disturbance/recovery 

patterns, or it could become the catalyst 

that forces ecosystems to shift rapidly and 

via alternate and uncertain pathways (b) in 

response to climate.
USDA Forest Service RMRS-GTR-290. 2012. 3

Preparation in many forms contributing to both resilience and resistance will be important, as will appropriate responses during and after major disturbances. No longer 

will simple protective responses to events suffice, nor even simple protective preparations. A set of strategic measures encompassing whole landscape perspectives 

using combinations of protective, monitoring, and corrective approaches will be 

necessary to manage a dynamic system suffused with uncertainty from both chance 

events and incomplete understanding. There will be tradeoffs between current and 

future risks. Management actions taken in the present will generally pose some risk, 

especially in the short term. The question is whether the imposed risks outweigh potential future risks. Even if they do, there are questions about scaling imposed risks, 

like how much at once and how much do we leave to chance in the short to medium 

term. While none of these questions have universal answers, there are contexts that 

support one approach versus others, and attentive managers teamed with researchers 

can learn how to describe the tradeoffs rationally.

Key Debates

With respect to forests, critical issues revolve around fire and fuels management 

including mechanical fuel reduction, intentional fire treatments, and natural fire 

treatments. Each comes with attendant risks, such as fires with unintentionally high 

severity or size, long duration and severity of smoke exposure from fires, potential 

for increase in invasive species, and impacts of roads where they are needed to facilitate management. There are costs and benefits with these actions, just as there are 

costs and benefits to no action. Decisions about where to prioritize work are a critical 

piece of the decision-making process. These decisions are made most frequently in 

the contexts of human habitation and threats to forests from fire, insects, and disease.

These decisions are sometimes difficult (and constrained) without considering the 

riparian and aquatic components of the ecosystem. Within riparian zones, most 

treatment options, including no action, have consequences for unique plant communities and adjacent streams. Considering aquatic communities brings in a range 

of other issues for water and aquatic management, some of which compete, or seem 

to compete, with decision space for forest management. Roads, which provide important access for silvicultural treatments and fire response, now form a threat not 

just to native vegetation, but also to stream communities, intensifying the tradeoffs. 

Reframing the decision goals to optimize both aquatic and terrestrial conditions can 

reveal opportunities in place of tradeoffs, particularly in previously managed areas 

with an existing road system (Rieman and others 2000; Rieman and others 2010).

Earlier syntheses of fire effects on fish and streams (Bisson and others 2003; Rieman 

and others 2003a and papers therein) provided new ideas and science that helped 

bridge the complexity of balancing the multiple resources. A principal idea presented 

in those papers was a greater reliability on natural dynamics to create resilient forest 

and stream ecosystems. The idea was appealing both from an economic perspective, 

due to reducing fuel and fire management expenses, and the perspective of persistence of key aquatic resources. The notion that while local fish populations might 

be severely reduced, they had life history adaptations that allowed them to persist 

in the long term stepped away from a static view for healthy aquatic ecosystems. 

The emerging view is that fire has played not just an important role in Western forests, but the streams running through them as well. This conceptualization ties well 

to a broader understanding of forest rejuvenation after fire and offers alternatives 

to artificial divisions between forest and stream ecosystems in any given place in 
4 USDA Forest Service RMRS-GTR-290. 2012.

favor of viewing it as one jointly cycling ecosystem. Scientists and managers have 

warned, however, that where connectivity and fragmentation of forests and streams 

was changed through historical land management, the resilience that is critical to 

natural cycling may be compromised.

The complexities added by climate variability change the discussion about forests, 

fish, and fire. Where the understanding that a goal of improved resilience to fire 

could commonly solve both aquatic and forest issues (Rieman and others 2010), 

potentially with relatively low expense and public opposition, we are now more 

commonly faced with choices between some kind of active intervention versus prospective loss of species locally. Once, we could identify a fairly simple dichotomy 

between (1) places where wildfire could operate freely without dramatically changing the natural dynamics of an environment and (2) places where some restoration 

of forests or streams would be necessary before wildfire would be unlikely to subsantially alter the natural dynamics. A concept called “historical range of variability” 

(e.g., Keane and others 2009) was used to describe natural dynamics in forest or 

stream conditions. There are likely still a very few places where the dynamics will 

be comparatively unchanged, but most places in the western United States will no 

longer have the same temporal and spatial scalings of the dynamic processes, limiting our ability to mimic cycles of the past.

We are seemingly back to similar issues that defined the debate about forests versus 

fish a decade ago, only with more intensity. Forests are now more urgently in need 

of treatment, and that need may be more geographically widespread. Aquatic ecosystems are becoming even more sensitive to either management or uncharacteristic 

fire. The limitation of resources for treating or responding to changes is similar, pointing again to prioritization as an important first step in reducing apparent conflicts. 

Human disruption of forest landscapes through timber harvest and construction of 

new roads has, however, decreased. Solutions are likely to be challenging, and most 

proposals for active intervention are likely to be controversial in the public arena 

(Spies and others 2010). While there is a recognition that dynamics are critical, the 

fact that they are no longer the same dynamics leaves questions in the minds of scientists, land and water managers, and the public as to how to proceed.

Framing Solutions

A key, though difficult, step will be articulating goals for the future (Rieman and 

others 2010). Communication of the goals will need to address the cross-disciplinary 

nature of the problem and be explicit in definitions and values that frame the goals. 

Climate change shifts the decision space in significant ways. It alters what the 

ultimate goals look like, from one where we consider the (comparatively) simple 

harmonic (cyclic) dynamic of forests burning and regrowing in patches to one of a 

moving target, sometimes gradual, sometimes rapid.

Envisioning future solutions will be facilitated through exploration of forest and 

stream ecosystem dynamics across landscapes containing multiple populations or 

patches and over long time periods containing many events and ecosystem response 

trajectories (White and Jentsch 2001; Jentsch and others 2003; Jentsch 2007). The 

nature of interventions, if any, should draw from a deep understanding of which 

components of ecosystem-stream dynamics are most important to their structure and 

function and how the spatial and temporal scales of disturbances interact through the 

biological system to produce the observed biodiversity. Nonlinearities in physical 
USDA Forest Service RMRS-GTR-290. 2012. 5

and biological processes, including either threshold-like or buffering behaviors, will 

identify both heightened and dampened vulnerability. As shifts occur in the physical 

climate and novel patterns begin to emerge, then, managers may develop new coping 

ideas taking advantage of this understanding of ecosystem dynamics.

Climate change is shifting what is and is not possible in some areas, and goals that 

once included protection of some species in a given location may no longer be tenable. If the critical dichotomy 10 years ago was whether wildfire or fuel treatments 

were worse for aquatic ecosystems, today it could be glibly paraphrased as whether 

we are more interested in maintaining select species populations or maintaining ecosystem function (Rieman and others 2010). The question revolves around values, 

what is possible within constraints of changing climate, and how much intervention 

we are willing to accept, or pay for, in forests. To be effective, interventions may 

have to be at scales grand enough to address the issues, with consequent needs to 

manage the potential for unintended outcomes of our decisions.

The scientific contributions to these policy issues lie in exploring the constraints 

fundamentally imposed by climatic changes, the biophysical context, and what capacity we have as land and water managers to alter how ecosystems will respond. 

Constraint can be partially viewed as parallel to vulnerability analysis, where shifts 

in habitat suitability over time are explored for species or communities (e.g., Parson 

and others 2003; Turner and others 2003; Füssel and Klein 2006; Millar and others 2007). Habitat suitability is defined in part by disturbance regimes as well as 

basic climatic factors, and this particular aspect of vulnerability is key to working 

through issues related to fire. Our ability to alleviate ecosystem stresses and stretch 

possibilities around these constraints is found in adaptation. Although the concept 

of resilience resonates for applications related to fire, resistance and facilitation will 

likely also be important (Millar and others 2007), and the distinction may be blurry 

in application.

Although most vulnerability analyses have been done without the full knowledge 

of the likely interactions and indirect effects of climate change, a logical extension would include changes in expected fire regimes associated with climate shifts 

and the interactions of those fire regimes with species and biotic communities. The 

expression of disturbance regimes can in fact be a stronger determinant of species 

ranges than the optimal temperature and precipitation requirements of the species 

(Pickett and White 1985). Because fire severity and size have a significant influence 

on aquatic habitats, this more complex scope for vulnerability analysis to include 

upslope and riparian vegetation conditions and risks is a key step in building holistic 

plans for adaptation of forest landscapes and watersheds.

The added challenges posed by climate change are revising not only how we assess 

risks to aquatic resources but also how much we can rely on resilience as a primary 

adaptation strategy. Many fish populations, particularly in more natural settings and 

without pressures from invasive species, seem to be well poised for resilience to disturbances from wildfire. However, it is generally unclear which populations would 

continue to persist under the combined effects of stream warming and increased 

fire size, severity, and frequency, and for how long. There are also questions about 

how much time restoration practices might buy, if, for example, habitat gained with 

reconnection of fragmented habitats through culvert removal were eventually lost to 

thermal barriers and shrinkage of local patches.

Even given the recognition that rapid climate change is altering ecosystems beyond 

a natural capacity to adjust, there is substantial contention about the use of active 
6 USDA Forest Service RMRS-GTR-290. 2012.

management in adaptation (Spies and others 2010). Philosophical discussions of the 

goals of management intervention without the context of a specific physical landscape can be interesting, but a long literature search would suggest that they might 

be unproductive absent a rooting in a real landscape (e.g., Cissel and others 1999; 

DellaSala and Frost 2001; Rhodes and Baker 2008). Discussion of future management will necessarily be filled with details, and with some uncertainty. Some details 

will be about tradeoffs between goals ranging from maintaining specific genetic 

resources (e.g., subspecies), to maintaining ecosystem functions, to maintaining 

general biotic assemblages (e.g., trout or forests) (Rieman and others 2010). Some 

details will be about how particular goals are achieved, probably reflecting a general 

bias toward those that are least intrusive while still somewhat effective. Other details 

will be about the relative risks to different resources and about acceptable levels of 

risk. In the translation of goals to specific objectives in the landscape, it is useful 

to frame alternatives as where we can maintain or restore process, and where it is 

necessary to impose more control (Rieman and others 2010). Recognizing and understanding uncertainty should not be a barrier to action, but rather a basis to inform, 

refine, and revise action through monitoring or adaptive management.

Ultimately, solutions that satisfy this high-dimensional and uncertain decision space 

require information and creativity. At any given location, it will be necessary to 

understand what climate has been doing and where it might be going, and the uncertainties inherent in the forecast. It will be important to understand physical and 

biological sensitivities to the changes, some layered through indirect effects and 

feedbacks. Learning from careful monitoring of climate shifts and their effects may 

be the only source of information for some processes. It will also be necessary to 

have ideas about potential beneficial actions and the relative value of actions from 

place to place.

In this monograph, we describe occurring and predicted changes to components of 

the climate and ecosystem to increase the understanding of how ecosystem responses 

to disturbance may be changing. Where we can, we delve into the intersection of fire, 

forests, streams, and changing climate to discuss current questions and debates and 

the relative contexts and conditions that might shape decision frames. Throughout, 

we try to provide some background to assist readers in understanding the physics and 

biology sufficiently to follow the rapidly evolving science and to spur creativity in 

local problem solving. We emphasize recent science.

We begin with a brief review of how the physical system of terrestrial and aquatic habitat, including climate, hydrology, and geomorphology, is being affected by 

climate change and fire, and then we proceed to discuss ecological changes to the 

upland and riparian vegetation and aquatic systems. Throughout, there is the opportunity to contrast the relative and combined effects of changes driven directly by 

the climate and those related to fire. The changes to the physical environment and 

upslope/upstream terrestrial ecology set the stage for a discussion of how aquatic 

communities will feel pressure through multiple pathways. We close by framing 

important components of the discussion about analyzing vulnerability and building 

future management options.
USDA Forest Service RMRS-GTR-290. 2012. 7

Resilience

Resilience is one of the most commonly discussed 

ecological concepts with respect to fire and climate 

change. Based on such broad usage, one might 

suppose that it is a 

concretely defined 

term. There are, 

however, nuances that 

are sometimes unclear 

to new readers, and 

different connotations 

may appear in a single 

document. Though 

subtle, interesting 

concepts, underlie the 

distinctions, and they 

are worth exploring.

The etymology of 

resilience is pretty simple from the Latin, 

re- “back” and “salire” “to jump”. Engineers 

tend to focus on this aspect of the word in 

application to material properties. Ecologists 

have a broader usage as “the 

capacity to absorb” (Walker 

and Salt 2006; Gunderson and 

others 2010). This usage is fairly 

similar to psychological and 

sociological usage. Ecologists 

sometimes contrast resilience 

with resistance to emphasize 

potential for recovery, and in 

common usage, rigidity is taken 

as an antonym for resilience. Yet 

resistant attributes can be an important component 

of resilience. The general upshot is that resilience is a 

quality describing the ability to withstand the “slings 

and arrows of outrageous fortune.”

Mathematics is a terse and generally unequivocal 

language; so it seems like a potentially fruitful place 

to turn for tightening the definition. Indeed, ball 

and cup analogies illustrating phase portraits and 

flow diagrams of partial differential equations are 

favorites for mathematically inclined ecologists. 

In mathematical terms, the various meanings of 

resilience relate to the concept of “stability.” In the 

literature of mathematical dynamics, there are, by 

one account, 57 different definitions of stability 

(Glendinning, 1994), providing on the one hand 

a vast potential insight about various themes on 

resilience, and, on the other, too many shades of 

nuance to be practical for general discussion.

Analogies for resilience are fairly common, and 

perhaps more helpful. Engineers like to talk about 

springs, while foresters seem to prefer bristlecone 

pines. These capture the “rebounding” and the 

“tough” aspects of the concept very well. Other 

apt analogies might be a swarm of gnats or a rock 

thrown into a pond. In these cases, there is no 

outward appearance of bouncing or resisting, but 

one would be hard pressed to say whether anything 

happened a few minutes after the disturbance.

All of these analogies relate in some form to a 

variation on stability, and in terms of biological 

process, they illustrate the capacity of individuals 

or populations to heal after being harmed, or to 

avoid being harmed in the first place. This kind of 

resilience reflects the most common perception 

with respect to trees and fish after fire. This is 

also the conceptualization of resilience applied in 

the development of forest and 

fishery resource management 

models that have been used to 

set “sustainable” harvest levels.

There is a need also to discuss 

the resilience of ecosystems 

more generally, e.g, how well 

ecosystems retain their resilience 

over time—a sort of resilience of 

resilience. This could also be framed 

as resilience of ecosystems to multiple 

interacting stressors. As the climate 

changes and as human developments 

proceed, some of the processes that 

provide an ability to avoid, absorb, 

resist, or recover from disturbance 

are changing. Some examples:

• Longer growing seasons at high elevations allow 

more fuels to grow and diseases and insects to 

be more effective, increasing risks for trees that 

historically survived through isolation;

• Fragmentations by roads has impinged on the 

reestablishment of fish populations;

• More frequent fires do not allow sufficient time 

between events for resistant species to attain 

sufficient size and species drawing on postfire 

reproduction to sufficiently mature;

• Forests or fisheries managed to maximize yields 

have demographics that are less able to survive 

major disruptions.

Although concrete definitions, synonyms, or even 

analogies seem elusive, the antonyms seem clear. 

A resilient system is not vulnerable or sensitive. 

Understanding the complex pathways of resilience 

in forest and stream ecosystems will help analyze 

vulnerability to future changes.
8 USDA Forest Service RMRS-GTR-290. 2012.

Part I: The Physical System

A. Climate

Patterns of air temperature and precipitation, the minimums and maximums, 

the seasonal patterns, and the correlation in timing between the two are critical 

elements of climate. Many biota have evolved some degree of specialization to 

particular temperature ranges or particular amounts of available water. Some biota 

trade specialization in extreme environments against open capacity for growth, 

while others take full advantage of mild and low variability climates. As the climate changes, the adaptations of various species and life forms will be tested. The 

increased number of large fires in recent decades across the western United States 

is explained, in part, by climate (Westerling and others 2006; Holden and others 

2012). The National Wildfire Coordinating Group (2009) has predicted that the 

annual area burned across the United States will increase to 10-12 million acres/

year (up considerably from the 10 year averages of 3.8 million acres in 1990s and 

7.1 million acres 2000-2008) due to a combination of factors, including climate 

change. As changes in climate result in increased length of fire season (Running 

2006; Westerling and others 2006) and increased tree mortality from bark beetles 

and drought (Breshears and others 2005), we will experience important positive 

feedbacks among fire, climate, and other disturbances with important implications 

for aquatic ecosystems.

Climate Change

Increasing concentrations of greenhouse gasses are causing the atmosphere to 

become warmer (see textbox on climate change mechanics). The changing heat 

balance of the earth is also changing atmospheric flow patterns and redistributing 

the wind streams that carry water vapor from oceans to land (Solomon and others 

2007; Archer and Caldeira 2008; Fu and others 2010b), changing the precipitation. 

Temperature increases to date are already substantial compared to historical and 

paleoclimatic records and clearly tied to anthropogenic greenhouse gas emissions 

(Solomon and others 2007). Estimates of the future rate of change depends on the 

rate that greenhouse gases are added to the atmosphere, producing a number of 

estimates that depend on the economic/regulatory scenario (Figure 2, Table 1).

The changes are complex, and not all places will warm equally, nor will precipitation change in the same way or to the same degree everywhere. In the northern 

hemisphere, it is expected that the warming will be more pronounced in arctic and 

Antarctic regions (Figure 3), where more precipitation is also expected. The belt 

of deserts in the subtropics (25-35 degrees N latitude) will likely spread northward 

with expansion of the Hadley cells, a primary component of the earth’s circulation. 

Beyond these generalities by latitude, the actual changes to any region depend 

on the relationship of the landmass to ocean currents. A combination of Global 

Circulation Models (GCMs) and historical analyses allow us to estimate what the 

future may bring to any given location. The implications for complex mountain 

terrain are poorly understood (Solomon and others 2007, Chapter 11).

Western North America is predicted to warm at rates comparable to global averages (Solomon and others 2007). For the A1B emission scenario, this is on the 
USDA Forest Service RMRS-GTR-290. 2012. 9

Figure 2: Global temperature 

trends showing the 20th 

century rise in temperature and 

projections for continued global 

average temperature increases 

depending on alternative 

carbon dioxide emission 

scenarios for the future 

(Solomon and others 2007).

Table 1: Descriptions of the carbon emission scenarios used in IPCC reports from Special Report on Emission Scenarios

(SRES: Nakićenović and Swart 2000).

Scenario Storyline (from SRES) Description

Cumulative Carbon 

Emission

(Gt 1990-2100)

A1B

Rapid and successful economic development, in which 

regional average income per capita converge - current 

distinctions between “poor” and “rich” countries eventually 

dissolve. The scenario reflects a strong commitment to 

market-based solutions, high savings and commitment to 

education at the household level, high rates of investment 

and innovation in education, technology, and institutions 

at the national and international levels, and international 

mobility of people, ideas, and technology. 

Initially fastest carbon 

emission growth rate 

with declining emissions 

starting by mid-21st 

century

1499

A2

Characterized by lower trade flows than A1B, relatively 

slow capital stock turnover, and slower technological 

change. Less emphasis on economic, social, and cultural 

interactions between regions are characteristic for this 

future, and economic growth is uneven and the income 

gap between now-industrialized and developing parts of 

the world does not narrow.

Accelerating carbon 

emission over 21st 

century

1862

B1

High level of environmental and social consciousness 

combined with a globally coherent approach to a more 

sustainable development.

Slowest CO2 emission 

growth, emissions 

declining by mid century 

and emissions below 

2000 levels by 2100.

983
10 USDA Forest Service RMRS-GTR-290. 2012.

Climate Change Mechanics

The energy balance of the earth is pretty simple. The sun shines on the earth, warming it, and the earth 

“shines” back into space. Because there is no other way to move heat into space, these two energy fluxes 

are nearly equal.

Most of the light from the sun is in the shortwave portion of the light spectrum, that is light we can see. 

Most of the light emitted by the earth is in the longwave portion of the spectrum, which is not visible to 

humans. Clouds and particles in the atmosphere do little to interrupt or reflect incoming solar radiation, 

but longwave radiation can be captured by greenhouse gases.

Most of the radiation emitted by the earth’s surface passes directly through the atmosphere into space. 

Fortunately for us, some is briefly captured by greenhouse gases in the atmosphere, which re-radiate it: 

half back down toward earth, and half continuing on to space. So our atmosphere acts a bit like a blanket 

with respect to radiant energy.

Humans have a substantial influence on the amount of CO2 in the atmosphere, and it has been increasing 

over time. Although levels of CO2 have naturally varied in the past, in part with solar cycles, levels now are 

well beyond any measured over the last few hundred thousand years, covering several solar cycles (Figure 

T2-1).

0

50

100

150

200

250

300

350

400

450

-800,000 -700,000 -600,000 -500,000 -400,000 -300,000 -200,000 -100,000 0

Carbon Dioxide Concentration (ppm)

Years before 2010

EPICA Dome C and Vostok, 800 ky bp to 1813

Law Dome, 1010 to 1975

Mauna Loa, 1959-2010

Figure T2-1: Composite of CO2

data from 800 ky before 

present to 2010. Green: Mauna 

Loa data (Tans and Keeling 

2011), Rust: Law Dome ice 

core data (Etheridge and others 

1996), Blue: composited 

Vostok and Dome C ice core 

data (Petit and others 1999; 

Monnin and others 2001; 

Siegenthaler and others 2005; 

Lüthi and others 2008)

As greenhouse gas concentrations in the atmosphere increase, more of the longwave photons emitted 

by the earth’s surface are caught by greenhouse gas molecules, and about half of them are returned to 

the earth’s surface. The radiative forcing increase from CO2 as of 2005 was about 1.5 W/m2 (Solomon 

and others 2007), or a little less than 1 miniature tree light on every square meter. As a result, the earth is 

warming well beyond variations seen in proxy records we have (Mann and Jones 2003; Mann 2008; Figure 

T2-2).
USDA Forest Service RMRS-GTR-290. 2012. 11

Figure T2-2: Overlap of the 

published 1300-year northern 

hemisphere temperature 

reconstructions included in 

IPCC 4th Assessment with 

instrumental record shown as 

a black line for the last 150 

yrs. Overlap is determined 

by the degree of agreement 

among the uncertainty 

ranges (±1 standard error 

of estimate) of the 10 

reconstructions (from Figure 

6.10 of Solomon and others 

2007).

There are differences in warming caused by increased greenhouse gases compared to increased solar radiation. 

The radiation blanket analogy might make it easier to explain. When you sleep in a cold room at night, it is 

usually your feet that get cold first, and adding a warmer blanket will warm your feet up. Contrast this to 

standing around a camp fire with cold feet; you usually need to actually expose your feet to the heat from the 

fire to warm them up. Increased solar radiation would be expected to warm places with lots of sunlight, like 

the tropics, more than places without much sunlight, near the poles. A CO2 ‘blanket,’ though, warms the polar 

regions more than the tropics. What we have observed so far is that the poles have been warming more than 

the tropics, reducing the meridional temperature gradient (Gitelman and others 1997; Braganza and others 

2004).

Figure T2-3: The meridional (equator to 

pole) temperature gradient over time 

(raw and smoothed) plotted with 

CO2 concentrations. CO2 data from 

Law Dome (Etheridge et al., 1996), 

MTG data from (Karamperidou et 

al., 2010). Note that the MTG axis is 

negative, and that the increasing trend 

represents a warming arctic compared 

to tropics.

1880 1900 1920 1940 1960 1980 2000

200 250 300 350 400

Year

CO2 (ppm)

-0.84

-0.82

-0.80

-0.78

Meridional Temperature Gradient (C/lat)

CO2

MTG

MTG-loess

Similarly, although you probably don’t pay much attention to it, the top of a blanket in a cold room is actually 

still warmer than the room, because you are warming it from underneath. If you put on a thicker blanket, you, 

on the bottom side of the blanket, will feel warmer, while the top of the blanket gets closer to the cold room 

temperature. If instead of putting on a thicker blanket, you had the opportunity to let more sun shine on the 

blanket, both the top and the bottom of the blanket would warm. What we have observed so far is that the 

mean temperature of the earth’s surface is warming while the stratosphere is cooling (Oort and Liu 1993; 

Golitsyn and others 1996; Guo and others 2008).

1880 1900 1920 1940 1960 1980 2000

200 250 300 350 400

Year

CO2 (ppm)

−0.84

−0.82

−0.80

−0.78 Meridional Temperature Gradient (C/lat)

CO2

MTG

MTG−loess
12 USDA Forest Service RMRS-GTR-290. 2012.

order of 2-6°C by 2100 (Figure 4). In contrast, the temperature has warmed about 

0.7°C relative to a natural atmosphere over the 20th century (Solomon and others 2007). This seemingly modest amount of warming has been linked to many 

changes in the western United States, including plant phenology (Cayan and others 

2001), snowpack reduction (Mote and others 2005; Pierce and others 2008), and 

earlier streamflows (Stewart and others 2005).

Projected changes in precipitation suggest strong declines in the Southwestern 

United States but seem less certain in the Northwestern United States (Figure 5). 

There is, in general, much less agreement between GCMs on precipitation than 

on temperature and pressure (Figure 6). Thus, even in areas where the sign of the 

change is not in question, there is still substantial disagreement about how large the 

change will be, with major implications regarding the magnitude of consequences 

(e.g., Barnett and Pierce 2008, 2009; Rajagopalan and others 2009).

Other predictions about future precipitation relate to variability and extreme 

values. One prediction is that precipitation events will be more intense when they 

occur (Trenberth 1993). This is derived from the slope of the saturation vapor 

pressure increasing with temperature, thus for a given change in temperature (e.g., 

from lifting over mountains), more water would be extracted from a given change 

in temperature. Another general prediction is increased variability in precipitation resulting from more variable storm tracks (Easterling and others 2000). Both 

results imply an increased flood risk from precipitation events (Easterling and 

Figure 4: Temperature increases 

relative to 1901-1950 average 

temperature in western North 

America (from Figure 11.11 in 

Solomon and others 2007). Bars 

on right show ranges for B1 

blue, A1B orange, and A2 red.

Figure 3: Multi-model mean surface 

temperature warming relative to 

1980-1999 mean temperature 

(from Figure 10.8 in Solomon and 

others 2007)
USDA Forest Service RMRS-GTR-290. 2012. 13

others 2000; Hamlet and Lettenmaier 2007). However, there is also recognition 

that increased dryness or length of dry spells could also have significant ecological 

consequences (Easterling and others 2000; Dale and others 2001; Westerling and 

others 2006; Holden and others 2012).

Variations in precipitation are influenced by sea surface temperature anomalies, such as ENSO, PDO, PNA, NAO, and AMO (e.g., Dettinger and others 

1998; Cayan and others 1999; Clark and others 2001; McCabe and others 2004; 

Abatzoglou 2011) that influence patterns of global air pressure and therefore circulation of air masses. These phenomena operate at frequencies of one cycle every 

few years (ENSO range) to decadal or multi-decadal cycles (NAO and PDO ranges). Reconstructions of long-term streamflow (which relate to long-term patterns 

in precipitation) show significantly more variability over deep time compared to 

current variations (Figure 7) than do the marked shifts in temperature in recent 

decades (Figure T2-2). As a consequence of the strong natural variability, it is not 

as easy to discern the effects of anthropogenic climate change on precipitation and 

drought as it is to discern (and attribute) the effects of climate change on temperature (Easterling and others 2000).

Figure 5: Precipitation anomalies (multi-model mean, A1B, 2080-2099 relative to 1980-1999). Stippling indicates areas where 

at least 80% of the models agree on the sign of the change.
14 USDA Forest Service RMRS-GTR-290. 2012.

Figure 7: Streamflow reconstruction of annual flow volume of the Yellowstone River based on Tree Ring widths. Annual 

flows are in a solid black line bounded by dashed lines at 2-standard errors and with a 10-yr moving average smooth in a 

heavier black line. Reproduced from Graumlich et al. (2003) using data at ftp://ftp.ncdc.noaa.gov/pub/data/paleo/treering/

reconstructions/montana/yellowstone_flow.txt.

0

20

40

60

80

100

Barometric Pressure Skill

Air Temperature

Shortwave Radiation

Humidity

Wind Speed

Longwave Radiation

Precipitation

Figure 6: Relative skill of general 

circulation models in predicting 

different climate variables over 

Australia (Johnson and Sharma 

2009). Skill was indexed based 

on degree of agreement among 

models.

Fire

Fire effects on climate tend to be either very local or global (note: the effects 

of climate on fire are discussed in sections II-B and III-A). The most important 

local change post-fire in forests is the loss of above-ground canopy cover, and 

this is seldom 100%. Vegetation cover may increase relatively quickly following 

less severe fires, but until it does, there may be less shading leading to increased 
USDA Forest Service RMRS-GTR-290. 2012. 15

daytime temperature near the ground and in streams. The change in evapotranspiration rates and canopy interception of precipitation can alter soil moisture and 

therefore stream recharge (see Hydrology section below). While common wisdom 

would suggest that the forest canopy would buffer against radiative cooling of the 

ground in winter, the canopy still experiences radiative cooling, and the resulting 

cold air flows down below the canopy, where it may keep conditions cooler than 

the general atmospheric temperatures. Precipitation intensity increases have been 

noted over deforested areas in the Amazon (Chagnon and Bras 2005) related to 

the decreased albedo; however, circulation patterns in the tropics differ substantially from those over North America, with lower horizontal wind speeds aloft, 

and similar processes have not been examined in temperate latitudes. Globally, 

the particulates from smoke from forest fires can influence atmospheric processes 

(Fromm and Servranckx 2003).

B. Hydrology

The response of stream and forest ecosystems to shifts in climate will be mediated through the changes in hydrology. An overarching issue in much of the 

western United States is simply the availability of water. The Interior West is a 

dry place, and even the wetter portions of the western United States are dry in the 

summer. Minor changes in the water balance or timing can have more exaggerated 

effects on biota because of the competition for this valuable resource. The influence of water on disturbance regimes, such as insects, disease, fire, or flood, is 

another important linkage.

Wildfire is, itself, often an outcome of reduced water availability to forests, and 

it provides an important feedback to the hydrologic system, with the potential both 

to ameliorate and to exacerbate changes already occurring in the climate system. 

An important question is the relative contribution of wildfire to hydrologic changes locally and at the basin scale. A key aspect of the discussion on climate change 

is how much hydrologic change results from climatic change versus how much 

change results from land use and land cover shifts. Because there is a large legacy 

of research on the hydrologic effects of land use and land cover changes, this 

provides some leverage for understanding the potential effects of climate-induced 

effects. There is also the issue of cumulative effects through multiple pathways, 

such as the combined effects of fire and climate change together on water yield or 

flooding.

Climate Change

Hydrologic changes in the western United States in recent decades include both 

changes to timing of streamflow and the water balance. There are linkages between the two, in that changes in precipitation can cause changes to the timing of 

streamflow (Luce and Holden 2009). The principle changes attributable to anthropogenic warming are changes to snowpacks (Pierce and others 2008), and include 

reduced precipitation as snow compared to rain (SWE/P) (Knowles and others 

2006), reduced snowpack on April 1st (Mote and others 2005; Regonda and others 

2005), and earlier runoff timing (Stewart and others 2005). Changes to precipitation related to climate change are expected in the Southwestern United States 

in the coming century caused by spreading Hadley cells (Seager and al. 2007; 

Johanson and Fu 2009), but changes in the Northwestern United States are uncertain, leaving attribution difficult (Easterling and others 2000). Historical changes 
16 USDA Forest Service RMRS-GTR-290. 2012.

show increases in Southwestern streamflows (Regonda and others 2005) and declines in the Northwestern United States (Luce and Holden 2009); such changes 

are partially consistent with general expectations for climate cycles (Dettinger and 

others 1998). Dry-year streamflow is better correlated with time, however, than 

with indices of low frequency variability (e.g., PDO), leaving questions about the 

relative contribution of cycles versus monotonic climate change (Luce and Holden 

2009). The land use contribution to changes in streamflow are likewise a source 

of uncertainty as are the changes caused by increased potential evapotranspiration 

(Hoerling and Eischeid 2007).

Precipitation is the largest term in the terrestrial water balance, and any incoming precipitation can be partitioned into either evapotranspiration or runoff 

(through surface, near surface, or deeper groundwater). Most precipitation in the 

western United States falls in fall and winter, leaving a dry summer. Most of the 

precipitation in the western United States also falls in mountains. It is not surprising then that about 75% of runoff in the western United States is currently derived 

from precipitation that falls as snow (Service, 2004), and equally unsurprising that 

concerns about snowpack changes are among the most important in the western 

United States (Barnett and others 2005).

Snowpacks in many parts of the western United States are sensitive to variations 

in temperature (Mote and others 2005; Regonda and others 2005), and therefore 

to anthropogenic increases in greenhouse gases (Pierce and others 2008). Declines 

in snow water equivalent and earlier melt dates over the last half century also 

have a relationship to trends in precipitation and runoff, for which the connection 

to greenhouse gas concentrations is more uncertain within GCMs. Regardless of 

cause, the primary trend over the last 60 years has been for less snow in the mountains of the western United States (Barnett and others 2008). As a result, the spring 

freshet has become both shorter and smaller (less volume). Because these trends 

are partially related to temperature, which is projected to continue increasing, the 

expectation is that they will continue. Most of the western United States has dry 

summers, and the earlier and smaller spring runoff predicted for the future will 

hold important implications for both biota and farmers.

Warming temperatures cause less precipitation to fall as snow and more to fall 

as rain (Knowles and others 2006). Conceptually we expect to see higher snowline 

elevations for individual storms (Casola and others 2009). This means that some 

fall and winter storms that historically produced more snow will now produce runoff, shifting some mountain streams from snowmelt-dominated hydrographs, with 

peak runoff in the spring, to rain-dominated or transitional hydrographs, where the 

timing of flows is more related to the timing of precipitation (Stewart and others 

2005). In the western United States, that means more streamflow in fall and winter 

and, consequently, less in the spring and summer (Figure 8). Such changes will 

happen soonest at mid-elevation sites, above already rain-dominated streams but 

below places where winter temperatures will remain cold enough for snow for 

some time (Regonda and others 2005; Pierce and others 2008; Nayak and others 

2010).

Temperature-related shifts in timing also have implications for flood and flood 

effects on biota. Since fall and winter are the main precipitation seasons, a shift 

from snow to rain means that the likelihood of floods in late fall and winter could 

increase, with consequences for fall-spawning fish (Wenger and others 2011b). 

Floods are likely to increase in magnitude in many basins as well, both because 

of the increased occurrence of rain-on-snow events (Lettenmaier and Gan 1990; 

Hamlet and Lettenmaier 2007) in currently spring-snowmelt-dominated basins and 
USDA Forest Service RMRS-GTR-290. 2012. 17

because of increasing precipitation intensity in rain-dominated basins (Easterling 

and others 2000).

An important discussion is evolving in the Pacific Northwest about the roles 

and causes of changing precipitation in the regional hydroclimate. Earlier work 

suggested a lack of trend in flows in the western United States (e.g., Mote and 

others 2005; Regonda and others 2005) in part due to underestimating the statistical importance of increased variance over time. More recent work has identified 

regional trends of declining streamflows over the last half century (Moore and others 2007; Luce and Holden 2009; Clark 2010; Fu and others 2010a), particularly 

in the Northwestern United States and with a more pronounced decline in runoff 

in drier years (Figure 9). This is an example of the principle that both means and 

variances are shifting, increasing the likelihood of some rare events (Jentsch and 

others 2007).

An important question is whether the changes are precipitation or transpiration related. Some hypothesize that trends in mountain streamflow are related to 

precipitation (Luce and Holden 2009; Clark 2010; Fu and others 2010a), others 

hypothesize warmer temperatures are increasing evapotranspiration (Hoerling and 

Eischeid 2007), and still others suggest changes to land use and land cover (Wang 

and Hejazi 2011). This raises speculation, for example, that decreases in streamflow 

have been caused by increased forest cover due to fire suppression. Examination 

of the Historical Climatology Network of weather stations suggests no trend in 

precipitation (Mote and others 2005), supporting a stronger focus on temperature 

related changes to hydroclimatology, which are more easily tied to anthropogenic 

greenhouse gas increases. There is, however, some question as to whether the precipitation gage network represents precipitation trends in the mountains because 

the gages in the network are primarily at lower elevations (Mote and others 2005). 

0

50

100

150

200

250

Oct Nov Dec Jan Feb Mar Apr May Jun Jul Aug Sep

Monthly Water Yield (millions 

m3)

1990s 2080s

Figure 8: Average annual hydrographs for the Middle Fork Boise R. simulated using the VIC 

model based on historical (1990s) climate compared to the projected climate of the 2080s 

under an A1B scenario. Data derived from VIC runs done by University of Washington 

Climate Impacts Group and U.S. Forest Service Rocky Mountain Research Station (Elsner 

and others 2010; Wenger and others 2010).
18 USDA Forest Service RMRS-GTR-290. 2012.

An important concept differentiating among alternative causes is that increased 

demand for water from either increased forest cover or warmer temperatures could 

not be satisfied if there is insufficient water. Essentially, this describes the distinction between potential and actual evaporation. Several have noted that changes 

in forest cover alter water yield primarily in wet years (Troendle and King 1987; 

Zhang and others 2001; Ford and others 2011). The findings of decreasing trends, 

particularly in the driest years and not in the wettest years, would not support the 

hypothesis that observed streamflow changes are caused by increased evaporative 

demand, leaving precipitation as the most likely driver.

This leaves the question of whether such precipitation changes are simply part 

of the regional climate cycles or can be attributed to increases in anthropogenic 

greenhouse gases. Causal linkages between decreasing streamflow and precipitation in some regions and anthropogenic climate change are not as easily identified 

as they are for temperature because observed variations in precipitation and 

streamflow are within bounds of historical and estimated paleoclimatic variation. 

Unfortunately, GCMs are notoriously poor at predicting precipitation (Johnson and 

Sharma 2009), so they are difficult to apply in formal attribution studies to discern 

the relative contributions of natural and anthropogenic changes on precipitation. 

Empirical statistical analysis supports primarily an anthropogenic contribution for 

the very low frequency (sometimes called secular) component of the trend with 

climate cycles playing an important role in variations over the course of a few 

years (Luce and Holden 2009). There is also theoretical support that the changes 

Figure 9: Trends in streamflow means and quantiles 1948-2006 for Johnson Creek at Yellow Pine, 

Idaho (USGS gage sta. 13313000). The dashed red line is the trend in the mean annual flow (24% 

decline, P=0.049), solid black line is the trend in the median (31% decline, P=0.025), the lower 

dashed black line is the 25th percentile flow (1 in 4-year low annual flow; 47% decline, P=0.01), 

and the top dashed black line is the 75th percentile flow (1 in 4-year high annual flow; 5% decline, 

P=0.82). This is a fairly common pattern in the Northwest, with dry years increasingly dry and wet 

years about as wet as they have been.

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1950 1960 1970 1980 1990 2000

200 400 600 800 1000

JOHNSON CR AT YELLOW PINE ID

Annual Flow (mm)
USDA Forest Service RMRS-GTR-290. 2012. 19

are tied to anthropogenic changes; the decreasing meridional temperature gradient 

and ocean-land temperature contrast discussed in the earlier textbox (Gitelman 

and others 1997; Braganza and others 2004) would both predict decreases in precipitation for this region. Decreasing meridional temperature gradients reduce the 

baroclinicity, or storminess, at midlatitudes, which is reflected in the storm record 

(McCabe and others 2001). The faster warming of the land compared to the ocean 

means that a water vapor content that is in balance with the ocean temperature 

would result in reduced relative humidity over the more rapidly warming land surface, reducing orographic precipitation in mountains (Simmons and others 2010).

Even if precipitation-related changes in snowpack have not resulted from anthropogenic climate change and have a connection to natural climate cycles, they 

are still important influences on overall snowpack patterns observed in the last halfcentury and have played a dominant role in higher elevation snowpacks (Regonda 

and others 2005; Moore and others 2007; Luce and Holden 2009). The date that 

snow completely melts off of a site is earlier for shallow snowpacks than deep 

snowpacks, all else being equal. It is less well understood that the timing of melt 

for shallow snowpacks is more sensitive to the amount of accumulation than for 

deep snowpacks, resulting in a non-linear relationship between timing and snowpack accumulation, in turn creating a non-linear relationship between total annual 

streamflow and the timing of streamflow from high elevation basins (Figure 10). 

Concave downward relationships can yield statistically significant changes in flow 

timing related only to changes in total flow, without any effects from temperature 

(Luce and Holden 2009). While some shifts in timing are occurring because there 

100 150 200 250

Annual Streamflow (mm)

16 Mar

16 Apr

16 May

01 Apr

01 May

01 Jun

Center of Timing

Figure 10: Relationship of snowmelt timing to total annual streamflow. Because sun angles 

are higher later in the spring, the last inch of snow can take weeks to melt if exposed 

by March but can melt in a few days if not exposed until June. The concave downward 

relationship has implications for causes of observed trends in streamflow timing (Luce 

and Holden 2009).
20 USDA Forest Service RMRS-GTR-290. 2012.

is more rain and less snow, others are occurring simply because there is less snow 

(Figure 11).

Summer streamflow provides habitat for fish rearing, carries food downstream 

to fish, and helps maintain cool stream temperatures. It is also related to soil moisture during the summer growing season for forests. Summer flows are mostly 

correlated to annual flows, and many streams are showing declines in summer 

flows (e.g., Luce and Holden 2009; Leppi and others 2011). Besides the evapotranspiration hypotheses already discussed, an additional mechanism suggested 

for declining summer flows and soil moisture is the earlier melting of the snowpack caused by warmer temperatures discussed above (e.g., Mote and others 2005; 

Stewart and others 2005; Westerling and others 2006; Barnett and others 2008). 

Because both temperature and precipitation are changing, both are contributing to 

the effect with different contributions in different places. Higher elevation basins, 

for instance, may be primarily responding to precipitation variability, while lower 

elevation basins in the Cascades may be responding more to temperature changes 

(Mote and others 2005).

With increasing variability in streamflow between years and lower low 

streamflows, the geologic context of streams may increase in importance. 

Groundwater-dominated systems such as provided by karst or recent volcanic geologies buffer short-term variations in streamflow driven by climatic variations, 

though at some cost through increased sensitivity to dry spells lasting several 

years (Lall and Mann 1995; Shun and Duffy 1999). Tague and Grant (2009) noted 

C

B

A

Figure 11: The flow timing seesaw. If the blue boxes represent buckets of runoff from a 

watershed, the center of timing for streamflow (blue triangle) is the balance point for when 

those buckets runoff over the year (A). If some of the flow starts coming off sooner because 

of earlier melt or falling as rain, the center of timing shifts earlier (B). If some of the flow in 

the summer is lost altogether, such as if precipitation is declining in a snowmelt dominated 

system, the center of timing also shifts earlier (C). Experiments that only measure the center 

of timing cannot distinguish between cause (B) and cause (C). Information is also needed 

about trends in flow or precipitation.
USDA Forest Service RMRS-GTR-290. 2012. 21

an ironic exception where timing shifts in snowmelt may yield greater absolute 

changes in low summer streamflow in deep groundwater-fed systems primarily 

due to the fact that shallow groundwater systems are already nearly dry in late 

summer.

Fire

Hydrologic changes induced by fire are generally seen as somewhat more 

“spectacular” than the changes driven more directly by climate change. Extensive 

rilling, gullies, and debris flows related to post-fire runoff from water-repellent 

soils, for example, are sometimes dramatic after wildfire (Klock and Helvey 1976; 

DeBano 1981; Doerr and others 2000; Cannon and others 2001; Istanbulluoglu and 

others 2002; Neary and others 2003; Shakesby and Doerr 2006; Moody and Martin 

2009). There are other more subtle changes, however, including changes to snowmelt, water yield, and low flows. As a result, peak flows in streams may be 200 to 

450 times higher post-fire than pre-fire, though it is more frequently reported that 

post-fire peak flow is less than 10 times that of peak flow pre-fire (Shakesby and 

Doerr 2006). In some smaller basins, peak flows bulked with debris have destroyed 

gage sites and not been recorded (e.g., Woodsmith and others 2004)

Runoff changes after wildfire have primarily been attributed to changes in soil 

properties. Many have studied the formation of water repellency after fire (see, 

for example, Shakesby and Doerr 2006 for review). Surface sealing has been suggested as another mechanism for increased runoff after fire (Rowe 1948; Swanson 

1981; Benavides-Solorio and MacDonald 2001; Meyer and Pierce 2003).

Post-fire water repellency typically occurs in a shallowly buried layer of soil 

and prevents infiltration of water through that layer where it occurs (DeBano 1981) 

(Figures 12 and 13). The layer is hypothesized to be formed when waxy substances 

in accumulated leaf and needle litter are volatilized by fire and recondense on 

cooler soils particles deeper in the soil (DeBano 1981; Doerr and others 2007). 

Because of the dependence on substances found in vegetative litter, it does not 

occur everywhere, but seems to be most commonly associated with particular vegetation communities, including (but by no means limited to) chaparral, eucalyptus, 

and subalpine fir, and are more likely where fires burn severely. It is also most 

Figure 12: Schematic of 

water repellency effects 

on infiltration and runoff 

generation after fire. The 

water repellent layer 

impedes infiltration leading 

to saturation of the wettable 

layer and runoff over the 

surface (after DeBano 

1969).
22 USDA Forest Service RMRS-GTR-290. 2012.

common on coarse-textured soils. Although water repellent chemicals (mostly fatty acids associated with plants) are present on soil particles prior to fire, and fires 

consume much of that water repellent material, water repellent areas may have as 

little as 1-4% of the original infiltration capacity after severe fires (DeBano 1981; 

Doerr and others 2000). Potentially the volatilization and recondensation of the 

chemicals has an annealing effect.

Water repellency is sensitive to the soil moisture state, and soils with water 

repellent substances generally repel water when dry (Doerr and others 2007). 

Water repellent soils are actually wettable, but only very slowly through vapor 

diffusion processes. As a consequence, soils rarely display water repellency in the 

wetter parts of the winter and spring, because soil moistures are maintained by frequent precipitation and melt. Dry and hot summers associated with Mediterranean 

climates of the western United States are ideal for bringing out water repellent 

behaviors.

Besides the annual disappearance and reappearance of water repellency with wetting and drying, there are longer-term patterns. The most frequently cited is a study 

by Dyrness (1976), showing some repellency remaining 6 years after a fire. Based 

on the sampling done in that study, the remaining repellency represents less than 

1/8th of the area, however, which is an important consideration for broader-scale 

effects (Shakesby and Doerr 2006). Sampling done after fires near Boise, Idaho, 

and Bozeman, Montana, showed a relatively rapid decline from nearly 90% water 

repellency in severely burned areas to less than 50% within 3 years (Figure 14). 

Consideration of the hydraulic conductivity (Megahan and Clayton 1986) and precipitation intensity characteristics for the area would suggest that in excess of a 

20-year precipitation event would be necessary to generate runoff after three years. 

This relates to the general observation that significant runoff and erosion events 

typically occur within 1-2 years of fire (Robichaud and Brown 1999; Shakesby and 

Doerr 2006).

Soil sealing is caused by the disaggregation of soil particles by raindrops, due 

primarily to high-energy raindrops falling on friable soils that are unprotected by 

vegetation or organic matter layers. Finer particles at the soil surface impede infiltration, and in extreme cases may form a crust. Much of the work on soil surface sealing 

has been done with agricultural and other severely disturbed soils (Mohammed and 

Kohl 1987; Bosch and Onstad 1988; Luce 1997), but similar behaviors have also 

been seen on burned soils under intense rainfall (Larsen and others 2009). Because 

soil particle aggregate stability is increased by organic matter (Kemper and Koch 

1966), losses of organic matter through heating and increased post-fire decomposition rates may be an important contributor to the vulnerability of soils after fire. 

Figure 13: Photograph of wetted layer over a water repellent layer after rainfall simulation with blue dye showing 

substantial similarity to the schematic by Debano (1969). The scale bar is 1-m between the legs.
USDA Forest Service RMRS-GTR-290. 2012. 23

Decreases in hydraulic conductivity related to surface sealing seem to be less severe 

than those associated with water repellency. For example, Larsen and others (2009) 

measured about a 50% decline in infiltration rate by applying intense rainfall to recently burned soils in Colorado. Nonetheless, in locations where water repellency is 

not as prevalent, and precipitation is intense, even minor reductions in soil infiltration can have severe consequences for runoff generation.

Peak runoff rates after fire are generally tied to intense precipitation events, such 

as convective storms (thunderstorms). Water repellency is less prevalent in winter 

and spring when soils are wet, thus the timing of the most severe repellency coincides with the timing of convective storms. Soil sealing requires high raindrop 

energy to disperse soil aggregates, and infiltration rates after sealing are still much 

greater than snowmelt rates, thus soil sealing is also more important under summer 

storms. The consequence of the tie to convective storms is that runoff effects from 

post-fire events tend to be localized. In a database of 600 severe post-fire flood and 

related events in the western United States, the largest basin with a reported event 

was 122 km2, and 99% of the basins were less than 25 km2 (Gartner and others 

2005). After the Tillamook burn in the Oregon Coast Range in 1933, Anderson and 

others (1976) estimated a 45% increase in the peak flow of two basins close to 400 

km2 in size the first year after the fire. In the Boise River, no increase in peak flow 

was noted at the 2,000-km2 scale despite measurable changes to water yield and 

some dramatic events in basins up to 20 km2 in extent (see textboxes on water yield 

increases after fire and debris flow scale in the Boise River Watershed).

Snowmelt changes after wildfire are important as well. Changes in snowmelt rate 

relate to the increased exposure of the snowpack to solar radiation and wind where 

vegetation cover post-fire is reduced. Increases in solar radiation post-fire have been 

linked to advances in the timing of snowmelt by 1 to 2 weeks, but not to increases 

in snowmelt-related peak flows in high elevation areas (Megahan and others 1995; 

Troendle and others 2010; also see textbox on Boise River). However at a lower 

elevation site, changes in the soil water balance and increases in accumulated snow 

combined with rapid melt during rain-on-snow increased peak flows and caused 

debris torrents in burned and salvage-logged basins (Klock and Helvey 1976). 

Turbulent transfer of heat from warm air can dramatically increase snowmelt rates 

and can be increased by forest harvest (Harr 1986). While the protective influence of 

0

10

20

30

40

50

60

70

80

90

100

2000 2001 2002 2003

Year

% Water Repellent

LFL

SM

UFL

ZZ

Figure 14: Declining 

fractional water 

repellent area over 

time following fires 

near Boise, Idaho, and 

Bozeman, Montana. 

The four lines are four 

separate burned areas 

(unpublished data).
24 USDA Forest Service RMRS-GTR-290. 2012.

Water Yield Increases After Fire in 

the Boise River Watershed

It can generally be said that when trees are 

removed from the landscape, runoff increases 

(Zhang and others 2001; Andréassian 2004; Brown 

and others 2005). There are, however, questions 

about whether water yield increases realized in 

small experimental basins (typically less than 

10 km2) translate into increases from large basins 

on the order of a few thousand km2 (Troendle 

1983; Troendle and others 2010). There are also 

questions about whether the larger fires that have 

been occurring in recent decades will translate to 

greater risks of flooding in large basins post-fire.

A principle obstacle to researching the question has 

been the lack of ability to manipulate vegetation 

experimentally over large fractions of a major river 

basin. A series of fire events in the Boise River 

basin, between 1992 and 2003, however provide 

an opportunity to examine streamflow changes 

from a large basin (Figure T3-1). The middle fork 

of the Boise River has been gauged since 1912, 

while the South Fork was gaged in 1946. The two 

adjacent basins have similar size and elevations, 

making them likely candidates for paired watershed 

analysis (Table T3-1). Several fires occurred in the 

middle fork basin between 1992 and 2003, the 

largest in 1994, burning about 45% of the basin 

area (Table T3-2), while the South Fork saw little 

disturbance. Both basins had some historical 

logging and other disturbances, but nothing on a 

scale to match the coverage of the fires.

The Middle Fork experienced a post-fire increase 

in water yield of about 5%, which translates to 

about 50,000 acre-feet of water annually (Figure 

T3-2). Despite severe changes in peak flows in 

some severely burned small basins (Figure T3-3) 

there was no increase in peak flows at the basin 

scale (Figure T3-4). The seasonal distribution of the 

flow increases was primarily in winter and spring, 

although late summer low flows also increased a 

little (Figure T3-5). There was a decrease in early 

summer flows related to advancing the hydrograph 

because of faster melt caused by increased solar 

radiation. Between 1948 and 2006, mean annual 

Middle Fork runoff decreased 18% (Luce and 

Holden 2009), and the increases associated with 

the wildfire were small in comparison.

Table T3-1: Area and elevations of the two gages.

Area

(km2)

Gauge

Elev. (m)

Mean

Elev. (m)

South F. 1645 1286 2085

Middle F. 2150 993 1936

Table T3-2: Fire events in the Middle 

Fork Boise River basin, 1992-2003

Year Acres in MF % Basin

1992 30,000 6%

1994 152,000 29%

2000 30,000 6%

2003 23,000 4%

Figure T3-1: Middle and South Fork Boise watersheds with 

fire perimeters as of 2006.

0.0

0.2

0.4

0.6

0.8

1.0

1.2

1.4

1.6

1.8

2.0

0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4

Annual Runoff Middle Fork Boise (Billion M3)

Pre-Fire (1946-1994)

Post-Fire (1995-2005)

R2 = 0.95

Annual Runoff South Fork Boise (Billion M3)

Figure T3-2: Paired watershed analysis regression for 

annual water yield.
USDA Forest Service RMRS-GTR-290. 2012. 25

Figure T3-3: Steel Creek debris flood following the 

Hot Creek Fire.

0

50

100

150

200

250

300

350

400

450

0 50 100 150 200 250

Annual Peak Flow South Fork Boise (CMS)

Annual Peak Flow Middle Fork Boise (CMS)

Pre-Fire (1946-1994)

Post-Fire (1995-2005)

1964 Storm

R2 = 0.83

Figure T3-4: Paired watershed analysis regression for annual 

peak flows.

-10

-5

0

5

10

15

20

OCTNOVDECJANFEB

MAR

APR

MAY

JUN

JUL

AUGSEP

Month

Change in Yield (millions 

m3

)

Figure T3-5: Monthly distribution of the water yield changes. 

Increases are shown through most of the year. The 

decreases in June and July are related to faster snowmelt 

in burned areas and relate to the much higher increases in 

April and May.

trees is typically attributed to the canopy, research on wind turbulence suggests that 

the stems could be important as well, particularly considering many branches still 

remain post-fire (Poggi and others 2004a, b).

The loss of forest canopy also reduces the loss of water through evaporative processes. Less precipitation is intercepted and subsequently evaporated and less water 

is transpired by trees, though this depends on level of tree mortality and response of 

other vegetation post-fire (Adams and others 2011; Guardiola-Claramonte and others 

2011). Annual water yields may increase post-fire (Shakesby and Doerr 2006), as they 

have in many forest harvest experiments (Stednick 1996; Andréassian 2004; Brown 

and others 2005). In general, water yield increases would be greater in wet locations 

and in wet years than in drier locations and drier years. Lower evapotranspiration has 
26 USDA Forest Service RMRS-GTR-290. 2012.

also been observed to result in higher soil moisture contents later in the dry season 

(Klock and Helvey 1976). Thus, we expect that some of the increased yield would 

benefit late season flows (see textbox on Boise River streamflow).

Fire and Climate Change: Feedback and Cumulative Effects

Fire extent is projected to increase in response to increased drought and lower 

precipitation combined with warmer temperatures (Littell and others 2009). Likely 

the effects of climate change on fire, vegetation, and streams will be synergistic. 

In the western United States, more large fires (Westerling and others 2006) and 

more widespread fires are more likely when early, warm springs are followed by 

warm, dry summers in the forests of the U.S. Northern Rocky Mountains and elsewhere (Westerling and others 2006; Morgan and others 2008; Littell and others 

2009). Westerling and others (2006) found that fire seasons were 78 days longer 

1986-2003 than in the previous 16 years (1970-1985) across the western United 

States, thanks to a combination of climate and fuel conditions. Based on their 

data, Running (2006) highlighted the 6-fold increase in area burned and the 4-fold 

increase in number of large fires in the same time period. Fire extent is projected 

to increase under projected climate changes. For instance, Spracklen and others 

(2009) predict that area burned will increase by a factor of 2.75 by 2050 in the 

Rocky Mountains and by a factor of 1.54 by 2050 across the western United States. 

Littel et al. (2010) predict increased area burned in many different regions across 

the western United States in response to changes in temperature, precipitation, and 

soil moisture. These projections are based upon correlations between fire extent 

and climate in historical records for recent decades (Westerling and others 2006), 

multiple decades (McKenzie and others 2004; Morgan and others 2008; Littell and 

others 2010), and multiple centuries as inferred from crossdated fire scars on trees 

(e.g., Kitzberger and others 2007; Heyerdahl and others 2008) and from charcoal 

in lakes (Whitlock and others 2003) and debris flows (Pierce and others 2004).

Although fuel accumulation (though more specifically fuel architecture) has 

been implicated in the increase in wildfire frequency and extent in recent decades 

(Kilgore 1973; Parsons and DeBenedetti 1979; Agee 1993; Graham and others 2004), it is part of a complex interaction of multiple variables that influence 

the vegetation and fire patterns we experience on landscapes today (Figure 15). 

Topography strongly influences patterns of burn severity in the Pacific Northwest 

and Southwestern United States (Holden and others 2009; Dillon and others 2011), 

as north-facing slopes are more likely to burn severely than north-facing slopes 

at the same elevation and high elevation forests often burn more severely than 

lower elevation forests (Holden and others 2012). North-facing slopes are often 

relatively moister than south-facing slopes, with soils with higher organic matter 

Fire

Ve Topography getation

Land Use Climate

Figure 15. The interactions among 

factors influencing fire patterns are 

complex. Adapted from Canadian 

Forest Service (2001). Forest fire: 

context for the Canadian Forest 

Service’s science program. <http://

www.nrcan-rncan.gc.ca/cfs-scf/

science/context_fire/index_e.html>
USDA Forest Service RMRS-GTR-290. 2012. 27

and with higher biomass, which once dry can burn severely (Dillon and others 

2011). Further, fires occur less often there than on adjacent south-facing slopes 

(Heyerdahl 2001) and biomass productivity is often higher so that when they burn, 

north-facing slopes are likely to burn more severely. Heyerdahl and others (2001) 

found that climate acts as a “top-down” factor strongly influencing fire extent, and 

that local factors such as topography, fuels, and vegetation influence fire “bottomup,” resulting in local differences on contrasting aspects. Dillon and others (2011) 

found that topography had a greater influence on burn severity than did climate for 

1,521 fires in Pacific Northwest and Southwest regions of the United States that 

burned 5.7 million ha 1984-2006. Likely the relative importance of fuels, weather, 

topography, vegetation, and climate vary greatly from place to place, and surely 

will be affected by land use, including fire exclusion resulting in changing fuel 

conditions. In years of widespread fires, fires are often large, suggesting that local 

fuels and microclimate have less influence on fire spread when it is especially hot, 

dry, and windy.

The relative role of climate/weather is likely different between forests and 

rangelands and among different rangeland and forest ecosystems (Collins and others 2006; Littell and others 2009) Where fine fuels are important to carrying fire, 

the factors affecting their abundance, including precipitation during their growing 

season, grazing, and wind, affect fire intensity and extent.

It is highly likely that fires and other disturbances will be an agent of climate 

change in altering vegetation and that effects may be cumulative. The term cumulative effects includes additive, compensating, and synergistic effects (Reid 1993). 

If we compare the effects of wildfire and climate change from the paragraphs 

above, there are several ways in which direct effects can combine with positive 

feedback. For instance, climate change is advancing snowmelt timing through 

reduced accumulation (Knowles and others 2006) while fire increases snowmelt 

rates and further advances snowmelt timing (Figure T3-5). There may also be 

compensatory feedbacks. For example, as streamflows decline in many parts of 

the western United States through reduced precipitation (Service 2004; Luce and 

Holden 2009), wildfire could result in a portion of that reduced water input reaching streams.

More complex interactions could increase some of the more severe consequences 

of climate change and wildfire. For instance, if warmer winter temperatures cause 

increased rain-on-snow flood risks in winter over more areas (Hamlet and Lettenmaier 

2007), losses of forest canopy associated with fire could increase those risks. Salvage 

logging would likely exacerbate risks of turbulent transfer of melt to the snowpack. 

Increasing precipitation intensity interacting with water repellency and soil sealing 

processes could magnify post-fire runoff events.

Implications for Aquatic Biota

The network of streams and rivers comprises the habitats of fish and other 

aquatic organisms. Too little flow can pose a reduction in habitat amount, quality, and connectivity, and too much can scour or sweep organisms downstream. 

Timing can be important too. The decreases in low flows, particularly in the driest years, has the direct impact of reducing the volume of pools and habitat, but 

also reduces velocities and water surface area and therefore the delivery of food 

from upstream sources (Harvey and others 2006). Decreases in low flows could 

also cause some sections of stream to become so dry as to become impassible to 

migrating fish (Rieman and McIntyre 1996), which would compound the effects 

of water withdrawals in some situations, including groundwater withdrawal.
28 USDA Forest Service RMRS-GTR-290. 2012.

Higher flood flows and debris flow-related flood events have complex effects as well, depending on timing and frequency. High streamflows scour redds 

(Montgomery and others 1996; Tonina and others 2008) or sweep fry downstream (Fausch and others 2001) when they occur at the right time of year. 

Fall-spawning fish, such as bull trout (Salvelinus confluentus), are expected to be 

more vulnerable as peak flows shift from spring to winter months in historically 

snowmelt dominated basins, because their eggs may still be in the gravel or their 

fry inadequately prepared for high flows when they occur (Wenger and others 

2011a). Debris flows have a much more limited footprint in the streamscape, 

but they typically remove all aquatic organisms from a given reach of stream, 

requiring recolonization. The speed with which affected reaches are recolonized 

will depend on the proximity of unaffected populations and the presence and 

abundance of migratory individuals.

C. Stream Temperature

For aquatic ecosystems, particularly for those containing rare salmonids, stream 

temperature is a critical variable structuring species distributions, patterns of 

abundance, and life history characteristics (Brannon and others 2004; Pörtner and 

Farrell 2008; Wenger and others 2011a). Both climate change and fire have strong 

influences on the energy balance of streams, primarily increasing temperatures, 

meaning that shifts to stream temperature regimes are among the principal processes driving changes to fish populations (Dunham and others 2003; Rieman and 

others 2007; Isaak and others 2010).

Climate Change

The direct effect of climate change on stream temperatures is increased incoming longwave radiation (see textbox for a description of the energy balance). 

Warmer air masses with higher emissivity will generate greater incoming radiation 

both day and night. Because water acts as a black body toward longwave radiation, 

the additional incoming radiation increases the temperature of the water. Warmer 

air masses will also increase the temperature of the forest canopy, again increasing downwelling longwave radiation. Direct warming from sensible heat transfer 

will likely be comparatively small (Leach and Moore 2010) and could easily be 

offset from increased evaporation from reduced relative humidity. Strong correlations between stream temperature and air temperature have made air temperature a 

proxy in estimating future stream temperature (Mohseni and others 2003; Rieman 

and others 2007; Wenger and others 2011b).

Indirect effects from climate change relate to changes in water availability either through streamflow or forest cover changes (see below for the fire effects). 

Declines in summer flows driven by declines in annual flows and earlier snowmelt 

(e.g., Cayan and others 2001; Luce and Holden 2009) mean that there is less water to heat in the months when the water is hottest. While the wetted width (area 

exposed to heat exchange) of streams will also decrease, it will not decrease as 

much as the depth and velocity (Dunne and Leopold 1978), yielding a net warming. Historical analyses of stream temperature also show a significant sensitivity to 

streamflow (Kiffney and others 2002; Isaak and others 2010; Kelleher and others 

2011).
USDA Forest Service RMRS-GTR-290. 2012. 29

The Stream Energy Balance

The relationship of streamflow to fire and climate change is governed by the energy balance. Figure T4-1 

shows the primary components of the energy balance of a stream reach. Because stream temperatures 

can change fairly rapidly, e.g., from one hour to the next, the diagram applies conceptually at hourly or 

shorter time scales; however, these fluxes are commonly summarized over longer periods (e.g., Webb and 

Zhang 1997). Solar radiation goes from being zero, at night, to being more than an order of magnitude 

greater than any other stream surface energy flux in the middle of the day (Sinokrot and Stefan 1993). 

Averaging over several days in one study, net radiation (short and longwave) was on the order of 70% of 

the incoming heat, with friction and sensible heat making up most of the rest, while radiation was 37% of 

the outgoing heat, with evaporation, bed conduction, and sensible heat playing significant roles (Webb 

and others 2008). Forest cover is a significant control on solar radiation so variations in forest cover play 

an important role in variation of stream temperature in forested environments (Johnson 2004; Moore 

and others 2005a; Dunham and others 2007). Turbulent fluxes (latent and sensible) generally oppose one 

another at hourly time scales because air warmed over the course of a day is typically also drier increasing 

evaporation; so if taken together they can be a minor component of the energy budget. Forests and 

streambanks can serve to protect smaller streams from the wind as well (Moore and others 2005b). Bed 

fluxes serve mostly to dampen and lag stream temperature responses to surface temperature forcings, 

and they respond primarily to the magnitude of daily stream temperature oscillations. They are more a 

part of the internal dynamics of a complex stream/bed/aquifer system than an external driver.

Qin, Tin

Qout, Tout

Qg, Tg

Rsn Rln Fs

Fb

Ff

Fe

Figure T4-1: Components of the stream energy balance. Q’s are water fluxes, and T’s are temperatures 

of inflowing (in), ouflowing (out), and groundwater (g) respectively. Together these comprise the 

advective heat fluxes to and from a stream reach. Radiant fluxes, denoted by R, are separated as 

net shortwave (sn) from the sun, and net longwave (ln). The sensible (Fs) and latent (Fe) heat fluxes 

are together called turbulent heat fluxes and represent the energy carried by wind in terms of cool 

or warm air convection on the stream (sensible) or evaporation or condensation on the stream 

(latent). The friction flux (Ff) is a function of the volume of water and slope of the reach. Bed fluxes 

include both conduction into solids of the bed and shallow groundwater exchanges (hyporheic) 

that are not just incoming groundwater adding to the streamflow.
30 USDA Forest Service RMRS-GTR-290. 2012.

Historical trends in stream temperature show increases in many places in recent 

decades even without land cover changes (Langan and others 2001; Petersen and 

Kitchell 2001; Morrison and others 2002; Bartholow 2005; Hari and others 2006; 

Isaak and others 2010). While stream temperatures have been rising in concert 

with air temperatures, rates of warming are generally less than air temperature 

rates, and not all places are warming equally (van Vliet and others 2010). For example, some streams in mountains, particularly with glaciers or snowfields, show 

a distinct buffering due to increased snowmelt inputs from increased summer melt 

rates (Hari and others 2006), and greater groundwater inputs can buffer warming 

as well (Kelleher and others 2011).

Fire

Stream temperatures post-fire increase where vegetation shading the stream 

is reduced. Short wave radiation is one of the largest inputs to stream temperature, and fire can substantially open the canopy, particularly over smaller streams 

and with associated debris flows. Estimated increases in stream temperature due 

to fire range from 0.5°C – 4°C for mean temperatures and 2.5°C – 10°C for 

maximum temperatures (Helvey 1972; Amaranthus and others 1989; Hitt 2003; 

Dunham and others 2007; Isaak and others 2010). Increases depend on stream 

size and canopy removal, and the effects of the combination of fire and debris 

flow can be much greater than fire alone. In a study of small streams in the Boise 

River basin (less than 1000 ha) burned streams were on average 3.4°C warmer 

(maximum daily) than unburned streams, though with substantial variability in 

response, and streams that had experienced both fire and passage of a major 

debris flow were on average 7.9°C warmer (Dunham and others 2007)(Figure 

16). Relative to biological criteria, these changes translated to about a 20% increase in probability of exceeding 20°C in burned streams, while those with a 

Unburned

Burned

Burned & DF

9.4 11.4 13.4 15.4 17.4

Temperature (°C)

9.4 11.4 13.4 15.4 17.4

Figure 16: Maximum summer 

temperatures in streams 

after fire. Data from 10 data 

loggers placed along 

9 streams (Dunham and 

others 2007).
USDA Forest Service RMRS-GTR-290. 2012. 31

debris flow as well showed about an 80% increase for streams between 1400 and 

1600 m in elevation.

Recovery of stream temperatures over time after fires and debris flows is 

important to the dynamics of aquatic populations. Unfortunately there are few 

measurements of long-term recovery following fire. Dunham and others (2007) 

showed only minor recovery from about a 3°C increase in mean and maximum 

temperatures measured annually for more than a decade post-fire on a stream 

where only fire occurred (Figure 17). High solar angles during summer mean 

that trees and shrubs must be tall or very close to a stream to cast much shadow 

during periods of highest heat loading, so recovery may well take a few decades, 

depending on growth rates of adjacent vegetation post-fire and the size of stream.

Implications for Aquatic Biota

Aquatic biota interact with stream temperature in many ways. Poikilotherms 

(cold-blooded animals, like fish and many other aquatic organisms) have metabolisms that are regulated by the ambient temperature (Pörtner and Farrell 2008). 

This means that under warmer temperatures their metabolism runs faster with the 

consequences that they will need more energy (food) to survive, less of the food 

they consume will go to growth, and they may sexually mature earlier (Dunham 

and others 2007). If winter and spring temperatures increase earlier in the year, 

eggs will incubate more rapidly and young fish will emerge from the gravel 

earlier in the year. Changes in emergence timing and in growth may affect the 

development (or non-development) of migratory individuals from a given rearing population. Different fishes have different physiological adaptation to specific 

thermal regimes, and different species have tolerances for different temperature 

ranges (Reist and others 2006; McCullough and others 2009), which shows in the 

spatial and elevation distributions of fishes (e.g., Wenger and others 2011a).

Depending on the context, therefore, the ecological consequences of these 

physiological responses may be the outright loss of habitat suitability in stream 

reaches that become too warm or increased susceptibility to displacement of coldadapted fish by relatively warm-adapted fish in stream locations where overlap 

occurs. For example, bull trout are generally displaced by brook trout from reaches Temperature difference (°C)

Year

12.0

9.0

6.0

3.0

0.0

1992 1994 1996 1998 2000 2002 2004 2006

Figure 17: Temperature 

differences between 

Cottonwood Creek, 

which burned in 1994, 

and Roaring River in the 

Boise River basin. Open 

symbols are summer 

mean temperatures 

and filled symbols are 

the summer maximum 

(Dunham and others 

2007).
32 USDA Forest Service RMRS-GTR-290. 2012.

where the two species overlap and cutthroat trout are often displaced from entire streams by encroaching brook trout, rainbow trout, and brown trout (Wenger 

and others 2011a). Stream warming, through a variety of means, therefore, is predicted to shrink the extent of habitat patches for cold-water fish of conservation 

concern and increases the isolation of populations by pushing them farther into 

headwater streams (Rieman and others 2007; Isaak and others 2010; Wenger and 

others 2011b). At the same time, decreases in low flows and increased debris flood 

responses in steep tributaries may shrink habitats from above, further restricting 

populations and increasing the potential for debris flow disturbances. If decreases 

in low flows and temperature-related growth and productivity changes also decrease the number of migratory fish from these areas, the populations may increase 

in their vulnerability to individual fire or flood events.

D. Geomorphology

Climate Change

Over geologic time, variation in climate has left profound marks on the landforms of the western United States. Glacially carved valleys are the most well 

recognized remnants of shifting climates, but we also see regionally extensive 

pluvial lakebeds. The connection of more contemporary climate variations to the 

incision and aggradation of arroyos and streams in arid and semi-arid regions has 

been a rich subject of research (e.g., Bull 1991), as well as subject of debate about 

the relative effects of climate and land use.

Some of the more direct relationships between climate and geomorphology of 

forested fluvial systems relate to the transport capacity of streams. Because of 

the strongly non-linear shape of sediment transport relationships (e.g., Parker and 

Klingeman 1982; Buffington and Montgomery 1997), flood flows are more important than total annual water yields. Thus, shifts in annual water yield may be 

less important than the potential of increased floods due to higher precipitation 

Vegetation density (areal coverage and root mass)

Sediment yield

arid semi-arid temperate humid

Climate

Driving force (Precipitation)

Resisting force (Vegetation)

Figure 18: Conceptual 

relationship between 

sediment yield and 

climate. The sediment 

yield curve (solid) is 

based on the empirical 

relationship of Langbein 

and Schumm (1958). The 

dashed line represents 

the resistance to erosion 

by vegetation while the 

dotted line represents 

the relative driving force 

by precipitation, with 

the greatest difference 

between the two is in 

semi-arid climates (Goode 

and others 2011).
USDA Forest Service RMRS-GTR-290. 2012. 33

intensity (e.g., Easterling and others 2000; Hamlet and Lettenmaier 2007) or increased probability of occurrence of rain-on-snow floods (Lettenmaier and Gan 

1990; Nayak and others 2010).

A long literature on climate and landscape evolution notes that the direct effects 

of some climatic changes, e.g., precipitation, may be dramatically overshadowed 

by the vegetational response (e.g., Langbein and Schumm 1958; Bull 1991; Kirkby 

and Cox 1995; Tucker and Bras 1997; Istanbulluoglu and Bras 2006; Collins and 

Bras 2008). For example, Figure 18 (Goode and others 2011) shows a conceptual 

relationship between sediment yields from river basins and the mean annual precipitation. On the one hand, increasing precipitation should increase the volume 

of sediment removed; on the other, it supports more vegetation, which modulates 

the effects of precipitation. At intermediate precipitation levels, vegetation growth 

is interrupted by frequent disturbances, yielding a peak in sediment yield, while at 

lower precipitation levels, the scarcity of water dominates.

Climate is an external driver to the complex interchanges between vegetation growth, fuel accumulation, weather events, and fire frequency and severity. 

Paleoclimatic research links periods of drought, severe fire, and severe erosion 

events using tree rings, fire scars, pollen from lakebed sediments, and charcoal 

in alluvial fans (Meyer and others 1992; Swetnam and Betancourt 1998; Briffa 

2000; Meyer and Pierce 2003; Whitlock and others 2003; Pierce and others 2004; 

Marlon and others 2006). At shorter time scales, it is clear that years of widespread 

fire are linked to severely dry and warm years (e.g., McKenzie and others 2004; 

Heyerdahl and others 2008; Morgan and others 2008; Littell and others 2009). As 

we shift toward a drier and warmer climate in the western United States, there is 

an expectation of greater areas burned annually (e.g., Running 2006; Littell and 

others 2009; Spracklen and others 2009) and the geomorphic consequences of fire.

Fire

The geomorphic consequences of fire are widely recognized; they are sometimes dramatic (Luce 2005; Shakesby and Doerr 2006; Moody and Martin 2009) 

and sometimes subtle (e.g., Ryan and others 2011). Hillslope and steep channel 

processes, such as surface erosion and mass wasting, receive the greatest attention 

(e.g., Benavides-Solorio and MacDonald 2001; Cannon and others 2001; Miller 

and others 2003; Pierce and others 2004; Moody and Martin 2009; Robichaud 

and others 2009b), while the disposition of channels with aquatic habitat is comparatively poorly discussed, despite more direct connections to aquatic ecology 

(Benda and others 2003a; Scheidt 2006; Lisle 2008). Most post-fire erosion studies 

focus on relatively short-term and small-scale processes, because they are acute 

and intense, with relevance to human life and property as well as aquatic ecology. 

There is, however, a growing recognition of decadal- to century-scale geomorphic 

dynamics distributed across stream networks and their role in evolving aquatic 

ecosystems (Reeves and others 1995; Benda and Dunne 1997a, b; Rieman and 

Clayton 1997; May and Gresswell 2003; Miller and others 2003; Scheidt 2006).

Loss of vegetative protection after fire along with alteration of soil properties 

increases the potential for surface erosion and mass wasting. The loss of trees reduces interception of raindrops by tree crowns and reduces root strength in the soil. 

Similarly, losses of trees, shrubs grass and surface organic layers expose the soil 

surface allowing it to be splashed and washed away more readily. Increased water 

repellency and surface sealing increase the runoff, as discussed earlier, and the loss 

of soil organic matter at the surface increases the disaggregation of soil particles 

allowing easier transport.
34 USDA Forest Service RMRS-GTR-290. 2012.

Many measurements of surface erosion from plots in many different environments demonstrate dependence on fire severity, slope, precipitation intensity, time 

since fire, soil characteristics, pre-fire vegetation, and aspect among other gradients 

of measurement. Unfortunately, the large number of methods used in estimating 

erosion and changes from pre-fire conditions preclude simple synthesis. Because 

studies are generally ad hoc after fires, systematic assessments of erosion processes over a range of fire severities, soils, and climate are lacking, so we cannot 

quantify how likely severe erosion events are or will be. Instead we direct readers 

to earlier reviews (e.g., Shakesby and Doerr 2006; Moody and Martin 2009). Rates 

for surface and rill erosion reported therein range from tens to a few hundred Mg/

ha in the first few years following fire.

Mass wasting events, such as debris flows, directly disrupt aquatic habitat, 

potentially extirpating local populations and simplifying habitats in the streams 

where they pass. Paradoxically, these large events also provide large amounts of 

coarse material such as gravel, cobbles, and logs that ultimately add to the habitat 

complexity and quality of streams where they deposit (Reeves and others 1995; 

Benda and others 2003a). It is the relationship of populations to these reorganizing 

events, their occurrence and extent, and the recovery over time that we argue is 

most critical to aquatic ecology (Dunham and others 2003). This conceptualization is not premised solely on the fact that a great deal more sediment is produced 

from mass wasting events in small channels than from upslope areas (e.g., Santi 

and others 2008; Moody and Martin 2009), but it also recognizes a fundamentally 

different interaction between mass wasting events and aquatic populations and 

habitats in comparison to sediments detached and transported by water alone.

Post-fire debris flows result from two primary causes: initiation by landslides 

and initiation by bulking during gully excavation (Cannon and Reneau 2000; 

Cannon and others 2001; Istanbulluoglu and others 2002; Istanbulluoglu and others 2003; Miller and others 2003; Santi and others 2008). Climatic influences may 

favor the frequency of one type of initiation compared to another (Wondzell and 

King 2003), but both initiation mechanisms are active across steep landscapes in 

the West. Erosion rates from these kinds of events are typically in the range of 

several hundred Mg/ha (Istanbulluoglu and others 2003; Meyer and Pierce 2003; 

Istanbulluoglu and others 2004; Moody and Martin 2009; Cannon and others 

2010). A primary difference between the two mechanisms is that the bulking debris 

flow events are most common during the first significant rainfall event (Cannon 

and others 2001), potentially tied to the durability of water repellency, whereas 

the window of landslide susceptibility may be on the order of a decade or more 

post-fire (e.g., Sidle and Ochiai 2006). Events as common as the 2-3 year (return 

interval) precipitation event have been noted as triggers for some of the larger 

debris flow events (e.g., Breidenbach and others 2004). Debris flow passage and 

deposition occur in distinct areas of a stream network and have greatly differing 

effects on habitat and biota.

The passage of a debris flow scours sediment, wood, and other biota from the 

stream, which become entrained in the flow. Because of the severe effects of debris flow passage to aquatic populations and habitats, a primary concern is the 

relative scale of debris flow tracks compared to the total habitat in a patch (see 

textbox on debris flow scale). If the amount of stream channel simultaneously 

affected by debris flows covers most of the suitable habitat in a patch, or if there 

is such poor connectivity to patches from which fish can recolonize, it would be 

difficult for the remaining fish to persist. However, if debris flows only affect a 

small fraction of a habitat patch at any given time, it is more likely that the patch 

will persist in time. Although there may be periodic dips in the population numbers 
USDA Forest Service RMRS-GTR-290. 2012. 35

Debris Flow Scale in the 

Boise River Basin

Understanding how disturbance affects biota 

requires some understanding the nature of 

disturbances, particularly the most severe ones. For 

fishes in mountain streams, one of the most severe 

disturbances is debris flows. An important question 

is how much continuous habitat is affected during 

any given event.

We mapped debris flows across the Boise R. Basin 

from aerial photographs taken in 1969, 1979, 

1988, and 1996 (Figure T5-1). The 1969 photos still 

showed the effects of a severe storm in winter of 

1964, while the 1996 photos showed the outcome 

of a 1994 fire and 1995 thunderstorm (Figure T5-2).

Mapped stream segments were analyzed by 

calculating the probability that any two segments 

separated by a given distance along a stream 

were both simultaneously disturbed (Figure T5-3). 

These figures show that disturbed segments tend 

to cluster below 20 to 25 km network distance, 

and that at longer distances, they are randomly 

distributed.

Dunham and Rieman (1999) show the probability 

that thermally suitable bull trout habitat is actually 

occupied depends on patch (watershed) size. 

The relationship levels off above 100 km2, which 

corresponds to a stream network length of 20 km. 

Figure T5-1: Example aerial photo with disturbed 

and undisturbed streams.

Historical Fires

1920

1939

1949

1953

1989

1990

1992

1994

Streams

Disturbed Segments 0 1 5 0 20

Kilometers

1996

Figure T5-2: Map of debris flow affected segments from the 

1996 photos.

0

0.1

0.2

0.3

0.4

0.5

0.6

0 10 20 30 40 50 60

Network Distance (km)

Probability that both segments are disturbed given

that one is 

1969

Figure T5-3: Probability that two segments separated by a 

given distance are simultaneously disturbed. Thin black 

lines show expected probability if disturbed segments 

were randomly distributed.

Figure T5-4: Map of all debris flows with bull trout 

patches from Dunham and Rieman (1999). 

Unoccupied patches tend to be small, on the order 

of debris flow affected basins.
36 USDA Forest Service RMRS-GTR-290. 2012.

for such patches, the capacity to reestablish within affected tributaries constitutes 

a key factor in resilience (Dunham and Rieman 1999). This view of size of debris 

flow is different from the typical volume orientation but represents an ecologically 

relevant perspective (Miller and others 2003). Debris flow mapping in the Boise 

River suggests that there are few continuous debris flow tracks with greater than 

about 20 km of stream length (see textbox on debris flow scale). A related way to 

look at the problem is by the size of drainage basins affected. From the database of 

Gartner and others (2005) looking at the western United States, the 95th percentile 

basin size was about 15 km2, while the median basin size was about 1.2 km2. By 

either analysis, debris flows most commonly impact basins smaller and steeper 

than are typically stable and productive fish habitats, but they occasionally affect 

tributary streams that are large enough to contain fish. Mapping tools considering 

debris flow movement and constraints can be helpful in evaluating the risk and 

appropriate scaling of aquatic habitats for persistence (Benda and Cundy 1990; 

Cannon and others 2010; Rieman and others 2010).

Debris flows deposit material when channel slopes decrease or the valley floor 

widens. In contrast to debris flow passage, deposition of sediment by major events 

appears to be important in the maintenance of diverse and high quality aquatic habitat (Reeves and others 1995; Benda and others 1998; Benda and others 

2003a). Gravels, large boulders, woody debris, and soil entrained in debris flows 

bring in large cover, habitat structure features, spawning substrate, and nutrients. 

Furthermore, the large deposits form reach-scale heterogeneity in stream slope, 

contributing to habitat diversity. The disposition of debris flow deposits depends 

on the circumstances of the event and the configuration of the receiving channel. Debris flows may deposit large fans that are stable over decades to centuries 

(Benda and others 2003a), deposit fans that are rapidly reworked and transported 

(Lisle and others 2001; Cui and Parker 2005), or be lost in the flood event that 

initiated them (e.g., Meyer and Pierce 2003). The first type of event may have the 

strongest influence on aquatic habitat complexity.

Resilience is the propensity of an ecosystem to recover from an acute event 

(Holling 1973, 1986; Walker and Salt 2006). Part of the resilience of aquatic ecosystems is based in the life-history strategies of affected species (Rieman and 

McIntyre 1995, 1996; Dunham and Rieman 1999), but another aspect is the recovery of physical habitat in time (e.g., Minshall and others 1989; Reeves and others 

1995; Benda and Dunne 1997a, b; Gresswell 1999; Scheidt 2006). The temporal 

dynamic of recovery also relates to the spatial scaling in terms of defining synchrony of disturbance (Poff and Ward 1990). If recovery takes one year, then events 

separated by a few years may effectively be independent, whereas if recovery takes 

decades, they may be effectively synchronous. Based on sediment transport theory, 

we suggest that basic channel form and sediment characteristics could settle fairly 

rapidly (1-5 years) post-flood because channel forming floods are fairly frequent 

(Wolman and Miller 1960). This is supported by observations of recovery of channel form within a decade after major floods (Wolman and Gerson 1978) and rapid 

recovery of basic channel characteristics after fire events (Potyondy and Hardy 

1994). Studies examining the long-term changes in aquatic habitat following fire 

suggest a more complex picture relating to the supply and fate of wood proximal 

to the stream and tributary debris flow paths (Reeves and others 1995; May and 

Gresswell 2003; Scheidt 2006; also see textbox on instream large wood dynamics 

in riparian section). Therefore, while the basic components of a habitat are available within a few years after disturbance, recovery of optimal habitat conditions, 

which depend on a host of other parameters, may take several decades post-fire 

(Gresswell 1999).
USDA Forest Service RMRS-GTR-290. 2012. 37

The contribution of these episodic events to basin-scale sediment yields is an 

important consideration with respect to the impacts on fish, vis-a-vis the idea of 

“pulse” versus “press” disturbance (Yount and Niemi 1990). Kirchner and others (2001) compared long-term (~10,000 years) sediment yields measured using 

cosmogenic 10Be isotopes to sediment yields measured over 30 years using sediment traps and suspended sediment sampling (Figure 19). They found that the 

long-term average rate was an order of magnitude higher than the contemporary 

rate, implying that sediment production was episodic in nature for these basins. 

Istanbulluoglu and others (2004) followed up with comparison to post-fire erosion 

rates to establish a magnitude and scale of the episodicity required to generate the 

relationship. Their conclusion was that the long-term sediment yields from forests 

can be explained by events as severe as the ones measured post-fire occurring 

on the order of a few hundred years apart, implying long relaxation periods after 

events.

How important are roads in this context? Some of the watersheds studied by 

Kirchener and others (2001) had roads; some did not. While the roaded watersheds produced more sediment, and one study of the first 4 years after construction 

Figure 19: From Goode et al., (2011). Sediment yields for individual post-fire erosional events, 

long-term basin averages, short-term basin averages (~30 yrs.), and road-surface erosion. 

Individual post-fire erosional events include measurements of debris flow and gully tracks in 

the Boise and Payette River Basins (Meyer and others 2001; Istanbulluoglu and others 2003) 

and deposit estimates (Breidenbach and others 2004) and model predictions (Cannon and 

others 2010). Long-term basin averages are from analysis of cosmogenic nuclides in fluvial 

sediments (Kirchner and others 2001). Short-term averages for small basins (< 20 km2) are 

from catchbasin dams (1950s-1980s, Kirchner and others 2001) and are subdivided by the 

presence or absence of roads in the basin. Short-term averages for larger basins are predicted 

from sediment rating curves and daily stream flows (1920-2000, Kirchner and others 2001), 

supplemented with data from King et al. (2004) using the same methods and period of record 

as that of Kirchner et al. (2001). Basin-average road-surface erosion is predicted from the 

GRAIP model (Black and others 2010), with values updated from Prasad (2007) based on 

measurements of road-surface erosion from the Middle Fork Payette watershed (Black, unpub. 

data). Event-based road-surface erosion values are from observed, post-construction erosion 

(4-year average yield, Ketcheson and others 1999).
38 USDA Forest Service RMRS-GTR-290. 2012.

showed substantial sediment additions (Ketcheson and others 1999), the magnitude 

of road erosion is extremely small in comparison to post-fire sediment inputs in the 

same time frame (Figure 19). One key difference is in the “pulse” versus “press” 

disturbance of fires compared to roads. Observations of post-fire deposition show 

a rapid recovery, followed by long periods with few additions of fines (a “pulse” 

process). In contrast, road sediments are produced in less abundance but every 

year (a “press” process). In addition, fire sediments usually require a significant 

storm to be generated, whereas roads produce runoff and sediment in almost every 

precipitation event. The cumulative effect from frequent “press” disturbances on 

aquatic biota may far exceed the direct effect of even major “pulse” disturbances.

A great deal of effort after fires goes into the control of potential erosion using 

post-fire stabilization techniques. Some methods used include contour felled logs, 

straw wattles placed on contour, surface application of straw or engineered wood, 

and aggressive grass seeding (Robichaud and others 2000). The general focus of the 

techniques is on control of surface erosion processes, which is reflected in largely 

plot-scale evaluation methods (e.g., Wagenbrenner and others 2006; Robichaud 

and others 2008). The importance of mass wasting processes in small steep channels, however, makes it difficult to extrapolate from such studies to speculate on 

the broad-scale efficacy of these treatments. Because most techniques relate to the 

control of sediment movement as opposed to controlling water, we may ultimately 

expect limited performance for preventing post-fire debris flows. Two caveats to 

this statement are that 1) contoured stabilization methods and surface mulches 

both extract some water, either by ponding or intercepting and 2) hillslope-derived 

sediment may contribute to the bulking of debris flows. It has been noted that the 

effectiveness of treatments declines with return interval of the precipitation event 

(e.g., Wagenbrenner and others 2006).
USDA Forest Service RMRS-GTR-290. 2012. 39

Part II: Biological Systems

A. Forests, Climate Change, and Fire

This brief synthesis of recent fire science is focused on forest vegetation dynamics and burn severity that have implications for fish and streams. We build upon 

recent reviews of related topics and recently published research. We suggest reading the syntheses on related topics in other sections in this document.

We Live In a Fire Environment

Fires will continue to occur, and they will sometimes be large and burn 

intensely—we need to plan accordingly. Biomass accretion exceeds decomposition 

in most forest and rangeland ecosystems. Fire is a global herbivore (Bond and 

Keeley 2005), consuming accumulated biomass when fires ignite and weather 

conditions are conducive. As a result, every place has a fire history, though it 

differs from place to place (Agee 1993). Recurring fires have shaped ecosystems 

and species adaptations. Despite very intensive efforts at fire suppression, we have 

experienced extensive fires in many years with many large fires in recent decades.

Climate change and other aspects of global change means that area burned by 

wildfires is expected to increase to as much as 10-12 million acres per year over 

the next five years nationwide (NWCG [National Wildfire Coordinating Group] 

2009) and to double, triple, or more in some regions of the country (Littell and others 2009; Spracklen and others 2009; Littell and others 2010), but not to the same 

degree everywhere. Most of the area burned in any given region results from just 

a few years of widespread fire, and it is in these years that climate is an important 

driver of fire extent (McKenzie and others 2004; Morgan and others 2008; Littell 

and others 2009). When many large fires burn synchronously, threats to people and 

property are high, our ability to suppress fires can be overwhelmed, and fires have 

important cumulative effects on smoke production, carbon, water, and nutrient 

budgets as well as habitats for many species of conservation concern (McKenzie 

and others 2004; Morgan and others 2008; Spracklen and others 2009).

Managers must balance the costs of fire suppression, ecological benefits and 

impacts of fires, fire fighter safety, protecting people and property, and the ecological realities of increasing wildland urban interface (Theobald and Romme 2007), 

invasive species (Brooks and others 2004), changing climate (Solomon and others 

2007), and changing perceptions of risk. Doing so will require strategic fire management that integrates fuels management, fire prevention, fire use, multiple fire 

suppression strategies, restoration, and other management in support of effective 

landscape-scale fire management across lands of intermingled jurisdiction (NWCG 

[National Wildfire Coordinating Group] 2009). With 10,000 homes burned in wildfires 2002-2006 (Gude and others 2008) and much focus on fuels management, 

protecting people and homes from fires continues to be a major, and expensive, 

fire management goal. Yet allowing fire to play a more natural role in some locations is a goal for many federal land management agencies (NWCG [National 

Wildfire Coordinating Group] 2009). Managing the rising costs of fire suppression 

and threats to people and property is a goal of all fire managers (NWCG [National 

Wildfire Coordinating Group] 2009).
40 USDA Forest Service RMRS-GTR-290. 2012.

Changing Fire Regimes

Fire is one of many disturbances that have shaped landscape dynamics for millennia. Fire is part of the resulting natural variability to which many species are 

adapted and is an essential component of most terrestrial ecosystems. Yet not all 

fires are alike, and fire regimes vary from place to place. The degree of change in 

fire regime from past to present (and therefore the future) varies greatly (Figure 20). 

The different patterns of recurring fires by frequency, severity and other characteristics are classed into fire regimes (Table 2). Since the early 1900s, humans have 

significantly altered historical fire regimes in many parts of the world. People use 

fire, suppress fires, and otherwise change when and where and how fires burn 

with their direct (fuels management, fire suppression) and indirect (roads, logging, 

grazing, limiting vegetation management, etc.) actions. Major trends are evident. 

Where fires currently occur less often than they did historically, we generally see 

an increase in woody biomass in many ecosystems—some with native species, 

some with nonnative species. Many argue that this is a result of climate change, 

while others attribute this to very effective fire suppression and other land use; 

likely both sets of factors contribute (Dombeck and others 2004; McKenzie and 

others 2004; Morgan and others 2008; Littell and others 2009; NWCG [National 

Wildfire Coordinating Group] 2009). In many ecosystems worldwide, introduced 

annual grasses have fueled much more frequent fires than occurred in the past. It is 

referred to as the grass-fire cycle: the more grass, the more fire, and the more fire 

the more grass (Vitousek and others 1996; Brooks and others 2004). As a result, 

there are many ecosystems worldwide experiencing fires much more frequently 

than in the past (Vitousek and others 1996; Brooks and others 2004; Shlisky and 

others 2007).

Figure 20: Fire Regime Current Condition (FRCC) class map version 2000 (from Schmidt and others 2002). Red 

areas show the greatest departure from historical fire regimes and green show the least. Gray areas are nonforested.
USDA Forest Service RMRS-GTR-290. 2012. 41

The Fire Regime Condition Class is a recent national mapping effort aimed 

at identifying locations where departures of fire and vegetation conditions is low 

(class I), moderate (class II), or great (Class III) (www.frcc.gov, Figure 20). Severe, 

stand-replacing fires are the norm in some ecosystems and therefore represent little 

departure, yet they represent a great departure (depending on size and other variables) in other ecosystems, including many dry forests.

Humans have altered vegetation and associated fire regimes. Humans alter the 

seasonality, frequency, extent, and severity of fire. The greater the degree of departure in fire frequency and severity and in vegetation, the greater the changes in 

biodiversity and other ecological values, and the more uncharacteristic the effects 

of fires will be when they occur (Shlisky and others 2007). Species may not be 

well adapted to the uncharacteristic fire regimes that develop when fire frequency 

and severity and vegetation composition are very different from historical range of 

variability (Keane and others 2008). More than 80% of the ecoregions examined 

worldwide have degraded or very degraded fire regimes (Shlisky and others 2007). 

Forest fire regimes have changed most where fires were historically frequent, as 

was the case in many grasslands and in dry forests (Agee 1993; McKenzie and others 2004). In mixed-conifer forests at moderate to high elevation, historically most 

fires were small and a few accounted for most of the area burned and mixed and 

stand-replacing fires were the norm (Schoennagel and others 2004). Departures of 

current from historical fire regimes in fire frequency and severity can be characterized better in some ecosystems than others, for we know relatively less about 

historical fire regimes in grasslands, shrublands, woodlands, and wet forests than 

we do in dry forests.

Some patterns emerge from an analysis of fire regimes relative to climate, topography, and vegetation types. First, the wettest and coldest forest sites do not 

burn often, but when they do, they can burn severely (Morgan and others 2008; 

Dillon and others 2011). A shift to warmer springs and warmer, dryer summers 

could result in more years of widespread fires, and such shifts could be contributing to the extensive area burned in many large fires in recent decades in the forests 

of the U.S. Northern Rockies (Running 2006; Morgan and others 2008; Littell and 

others 2009) and elsewhere. The degree to which current and future fire regimes 

Table 2: Fire regimes can be grouped by how often fires recur and the degree to which ecosystems change when they 

burn (from Barrett and others 2010), http://www.fire.org/niftt/released/FRCC_Guidebook_2010_final.pdf.

Group Frequency Severity Severity description

I 0–35 years Low / mixed

Generally low-severity fires 

replacing less than 25% of the 

dominant overstory vegetation; can 

include mixed-severity fires that 

replace up to 75% of the overstory

II 0–35 years Replacement

High-severity fires replacing greater 

than 75% of the dominant overstory 

vegetation

III 35–200 years Mixed / low Generally mixed-severity; can also 

include low severity fires

IV 35–200 years Replacement High-severity fires

V 200+ years Replacement / any 

severity

Generally replacement severity; 

can include any severity type in this 

frequency range
42 USDA Forest Service RMRS-GTR-290. 2012.

are different from historical regimes depends, in part, on the relative importance 

of climate and fuels in influencing fire occurrence, extent, and severity. While the 

relative importance of climate and fuel varies among forest types (Schoennagel 

and others 2004), we do not yet fully understand the implications of interactions 

among climate, vegetation, land use, topography, other disturbances, and fires. The 

complex interactions and potential for feedback among these factors lends uncertainty to predictions, particularly in complex topography.

It is one of the many paradoxes of fire that as fires have become less frequent, 

future fires are increasingly likely to burn more intensely and severely. People can 

choose to live with fire, allowing more choice about where and when fires burn, or 

to continue suppressing fires and suffer the effects of large fires, burning under extreme conditions that we cannot control. Integrated Fire Management approaches 

balance sustainable levels of effort with consequences for people and the environment (Figure 21).

Fire regimes, the pattern of recurring fires through time, reflect the interaction 

between vegetation, topography, climate, and land use. Precipitation, temperature, and soils influence where plants can grow, while disturbance, succession, 

and competition affect where they occur and the ongoing landscape dynamics. 

These factors and the amount of fuel available to burn are, in turn, major factors in determining the rate of spread, intensity, and severity of fires (Rothermel 

1972). Climate and weather interact with topography and vegetation to influence 

fire behavior and effects. People influence vegetation structure and composition, 

and they suppress and ignite fires.

Burn Severity

Burn severity is an attribute of fire regimes used to express the degree of ecological change that results from a fire (Lentile and others 2006; Lentile and others 

2007; Keeley 2009) (Figure 22). We now understand that fire is beneficial and 

often essential to most ecosystems, and ecologists and managers are often focused 

Multi-national

National

Local

Fire Use Strategies

Beneficial Fires

i.e. Benefits outweigh risks

Prevention/Suppression

Strategies

Detrimental Fires

i.e. Risks outweigh benefits

Adaptive

Management

Feedback Sustainable Ecosystems & Livelihoods

Analysis: Potential environmental, social, economic

Benefits & Damages

Integrated

Fire Solutions

Increased benefits, Decreased damages

Figure 21. Integrated Fire Management encompass fire use, prevention, and suppression with an understanding of sustainable 

ecosystems and livelihoods (from Myers 2006).
USDA Forest Service RMRS-GTR-290. 2012. 43

on the post-fire effects when an area burns. Thus, fire severity or burn severity 

(these terms are often used interchangeably) is perhaps the most significant but 

least understood attribute of fire regimes. We focus on describing recent research 

on burn severity here.

There are a number of fire effects of ecological consequence, including vegetation loss, tree mortality, soil heating, and soil effects. The term “burn severity” is 

imprecise and carries inherently negative connotations (Lentile and others 2006; 

Keeley 2009). For instance, in many vegetation types, “severe” stand replacing 

fires are characteristic and within the historical range of variability, yet calling 

these severe implies that they are also catastrophic and by definition undesirable. 

While a fire that results in mortality of the aboveground biomass is indeed one that 

results in great change (and is therefore often assigned high burn severity), many 

plant and animal species thrive after such events; some are even dependent upon 

severe fires. Thus, a severe fire is not necessarily catastrophic or “bad,” and all fires 

will have some desirable and some undesirable effects, even for streams and fish. 

In some cases, the ecosystem changes resulting from an absence of fire may result 

in less desirable ecosystem charactertistics and perhaps more severe fire effects 

once the area burns in subsequent wildfires.

Burn severity is a continuous measure of multiple fire effects, including but 

not limited to overstory vegetation mortality, soil heating, and fuel consumption 

(Lentile and others 2006; Lentile and others 2007; Keane and others 2008; Keeley 

Figure 22: Burn severity is used to describe the degree of change due to fire and is often based upon fire 

effects, including but not limited to overstory vegetation mortality, soil heating, and fuel consumption 

(Keeley 2009, Lentile et al. 2006, Lentile et al., 2007, Keane et al. 2008). Burn severity can be 

classified for forests (as shown here), woodlands, shrublands, and grasslands in the field and from 

satellite or airborne remote sensing immediately or one year post-fire (from Lentile et al., 2006).
44 USDA Forest Service RMRS-GTR-290. 2012.

2009).For purposes of mapping and communication, burn severity is commonly 

classified into four or more classes (Figures 22 and 23), including unburned, low, 

moderate, and high. Other classifications are possible. Such classes are often inferred from satellite imagery and field observations. In reality, burn severity is a 

continuous measure on a variety of variables.

When fire frequency changes, burn severity also changes. Where fires become 

less frequent, fuels often accumulate sufficiently between fires such that subsequent fires burn more intensely and severely. This is the case with many dry forests 

that historically burned frequently with many surviving large trees. Currently, 

Figure 23: Soil burn severity is commonly described in classes based upon fire effects on overstory and soils.

a) Low burn severity (typically in places with

light fuels) may be patchy and may not

even show complete combustion of

herbaceous species.

b) Moderate burn severity may show black

ash or blackened duff at the surface.

c) High burn severity presents mostly bare

soil with some white and black ash at the

surface.
USDA Forest Service RMRS-GTR-290. 2012. 45

where dry forests are densely populated with small trees and ladder fuels, even 

large-diameter trees of otherwise fire-resistant trees, like ponderosa pine, are more 

likely to die in subsequent fires.

If a fire burns very intensely, the high rate of energy release may mean that foliage in the crowns of many overstory trees or shrubs are consumed, and most are 

top-killed. Some may resprout. A fire does not have to burn intensely to be severe, 

however (Keeley 2009). Fires smoldering for many months in the organic soil of a 

peat bog, or creeping around burning the accumulated duff and logs in a subalpine 

fir forests while not burning any tops of the shrubs and trees, can nonetheless result 

in vegetation mortality and soil effects that would be judged severe. In such fires, 

the flame lengths may be very low (suggesting that fire intensity is low), but the 

heating of the soil and consumption of organic matter is sufficient to kill many of 

the roots of trees, shrubs, grasses, and forbs; thus subalpine fir trees can be readily 

killed by cambial damage to roots and bole. Typically, fires that burn intensely are 

also severe as occurs when both tree crowns and dense mats of feather moss burn 

in black spruce forests in Alaska, or when crown fires occur and burn large areas 

in subalpine forests. Fires in many grasslands burn very intensely (with rapidly 

moving fire fronts and high flame lengths), but where the grasses are well adapted 

to resprout vigorously post-fire, there may be little difference between burned and 

adjacent unburned grasslands within just a few months or years, and the degree 

of ecological change as a result of the fire is small and the fire is not severe. Such 

rapid burning cannot endure long with sparse fuels; so subsoil heating is slight.

For soils, for streams, and for aquatic organisms, severity of fire matters. How 

fires burn is often more important than if an area burned. Severity is often related 

to fire frequency—less frequent fires are often more severe, simply because there 

is more fuel to burn.

Burn Severity Indicators Inferred From Satellite and Field Data

Ideally, indicators of burn severity will be ecologically meaningful, measureable in the field and remotely from air or satellite, and readily interpretable. They 

must be useful in describing ecosystem recovery and condition, including vegetation, carbon, water, and nutrients (Lentile and others 2009). Often burn severity 

is mapped from satellite imagery, such as Landsat, in order to get rapid, consistent evaluation across large areas. The Normalized Burn Ratio (NBR), differenced 

Normalized Burn Ratio (dNBR: Key and Benson 2006), and RdNBR (Relative differenced Normalized Burn Ratio) are widely used (Miller and Thode 2007) indices 

for creating the soil burn severity maps used in post-fire rehabilitation assessments 

(Robichaud and others 2007; Parsons and others 2010) and in ecological assessments based on one-year differences due to fire. These ratios effectively measure 

the relative degree of vegetation and soil char between pre- and post-fire conditions and they have been related to one-year post-fire vegetation cover (Holden and 

others 2005; Smith and others 2005; Lentile and others 2006; Hudak and others 

2007). These continuous measures are often broken into classes, often based on 

field assessments made with the Composite Burn Index (CBI) (van Wagtendonk 

and others 2004; Brewer and others 2005; Cocke and others 2005). Unfortunately, 

there are few studies evaluating dNBR or CBI against quantitative biological or 

ecological measures of post-fire effects (Hudak and others 2007; Robichaud and 

others 2007; Santis and Chuvieco 2007; Smith and others 2007), especially more 

than one year post-fire; such studies would help users to know how effectively 

burn severity can be inferred from satellite imagery and how it influences subsequent vegetation trajectories. These indicators are widely applied in mapping burn 
46 USDA Forest Service RMRS-GTR-290. 2012.

severity for individual fires and across the United States in the national Monitoring 

Trends in Burn Severity (MTBS, www.mtbs.gov) program (Cocke and others 

2005; Epting and others 2005; Miller and Thode 2007). There are limitations to 

these approaches (Roy and others 2006; Smith and others 2007; Lentile and others 

2009). Briefly, limitations include the following: subjective and qualitative nature 

of breaks between classes and field assessments using CBI exists; it is difficult to 

scale from the points where CBI or other on-the-ground assessments are made to 

the landscapes and watersheds over which assessments are needed; the spectral 

bands used are not ideal; RdNBR was developed to address the less-than-optimal 

performance of dNBR in woodlands; shrublands and grasslands have such low 

biomass that even if most is consumed the absolute change is not high; and the 

measures are mostly affected by vegetation consumption when we also want to 

know soil heating and fuel consumption (Roy and others 1987; Hudak and others 

2007). Further, though NBR from a single post-fire image may suffice, differencing helps to address site conditions that can affect inferences from imagery, thus 

helping to ensure that difference is due to fire, not other site conditions. Scaling 

is an issue because the post-fire effects measured in the field typically reflect finescale processes, but also impact coarse spatial (watershed to regional) and temporal 

(decadal) scales. Char fraction has been suggested by Lentile and others (2009) as 

a potentially versatile measure of postfire ecological impact that also influences the 

terrestrial carbon and water cycles. However, no single indicator of burn severity 

will be ideal for evaluating burn severity across all ecosystems affected by fires.

Burn severity includes overstory mortality, consumption of biomass, and soil 

heating (Keane and others 2008). Many people assess burn severity using broad 

categories of overstory tree or shrub mortality. Burn severity varies with fuel and 

the environmental conditions before and during combustion (Ryan and Noste 

1985). Assessments of burn severity based on satellite imagery are more likely to 

be accurate for high burn severity (Hudak and others 2007). In part, this is because 

it can be difficult to see soil effects through overstory vegetation canopy that is left 

where fires burn with low or moderate severity, and both low and moderate severity are spatially heterogeneous at fine scales. Further, we are likely more accurate 

in locating large patches of high burn severity.

All fires are patchy. Even when large fires burn through thousands of hectares 

in a few hours, burn severity is seldom uniformly severe (Schoennagel and others 

2004; Keane and others 2008). Pattern is very important to biodiversity and vegetation dynamics. In terms of post-fire soil erosion, we are often most concerned 

with large patches of stand-replacing fire, especially if these are on steep ground 

and on erosive soils. Fire effects on the soil are typically very heterogeneous; even 

within a large patch of severely burned overstory, there are often unburned vegetation patches. Because soil effects are more uniform where overstory vegetation 

was largely killed, and because removal of that vegetation means that we “see” the 

soil from remotely sensed imagery, mapping high burn severity is more accurate 

than mapping of moderate and low severity burns (Hudak and others 2007).

Burned Area Emergency Rehabilitation (BAER) teams concerned about postfire erosion focus on large patches that are burned severely, especially when those 

are on steep slopes with erodible soils (Robichaud and others 2009a; Parsons and 

others 2010). Starting with Burned Area Recovery Characteristics maps that are 

based on dNBR from satellite imagery (usually Landsat with its 30-m resolution), BAER teams evaluate the local conditions. Because they wish to identify 

those locations where fires have greatly affected soils, in the field they may look 

for bare, reddened soil with white ash. Soils with cover are less likely to erode. 

Where litter, duff, vegetation or other organic layers remain post-fire, and where 
USDA Forest Service RMRS-GTR-290. 2012. 47

vegetation survives or rapidly recovers, soils are less likely to be displaced even in 

relatively high-intensity rain storm events (Robichaud and others 2007 and references therein).

Will Future Fires Be Larger and More Severe, and If So, So What?

People worry that fires will be larger and more severe as climate changes (see 

earlier section for a summary of recent work on fire and climate). No doubt fires 

will happen. Spracklen and others (2009), Littel and others (2010), and Westerling 

and others (2011), among others, have predicted that fires will become larger and 

more severe as fires respond to changing temperature and precipitation and lightning patterns. The effects will not be the same everywhere. Some regions will 

have more fires in some years, and in some years, fires will be widespread across 

multiple regions (NWCG [National Wildfire Coordinating Group] 2009).

Miller and others (2009) found that recent fires in the Sierras and Cascades are 

more severe than historically. They evaluated satellite images before and after fires 

for a 22-year period. Dillon and others (2011) found that annual proportion burned 

severely increased in just one of the three U.S. Southwest ecoregions and two of 

three ecoregions in the U.S. Northwest 1984-2007. They attributed the change 

to a combination of fuels and climate. More analyses like these based upon the 

MTBS data (www.mtbs.gov) are likely soon. Dillon and others (2011) found that 

topography exerts significant “bottom-up” controls on patterns of burn severity 

with north-facing slopes often burning severely, probably due to higher productivity and more biomass, and a low likelihood of burning except under relatively 

extreme weather conditions. Burn severity in the Gila Wilderness is influenced by 

time since and degree of burn severity of previous fires (Holden and others 2010).

Fuel Treatment Effectiveness

Managers implementing fuel treatments use and manage disturbance, taking advantage of the strong interplay between fire and vegetation. Nearly 30 million acres 

have been treated to reduce fuels and fire hazard on federal lands with additional treatments on private and state lands (NWCG [National Wildfire Coordinating 

Group] 2009; Schoennagel and others 2009), most of which have been implemented 

since adoption of the National Fire Plan in 2000. More are planned especially as the 

wildland urban interface has grown (Theobald and Romme 2007). Nonetheless, the 

area with fuel treatments and other active vegetation management is far outpaced by 

wildfire and by insects and disease, which argues for strategic planning of fuels and 

fire management (NWCG [National Wildfire Coordinating Group] 2009). The goals 

of fuel treatments commonly include reducing wildfire risks to communities and the 

environment, and improving ecosystem resiliency to wildfire effects (USDA and others 2006). Fuel treatments are designed to reduce fire hazard with the goal of altering 

fire behavior, thus easing fire suppression efforts (Graham and others 1999; Graham 

and others 2004) and the escalating costs of fire suppression and threats to people and 

property (NWCG [National Wildfire Coordinating Group] 2009). In forests, thinning 

from below to remove small trees can reduce crown fire hazard, but if wildfires occur 

before the residual surface fuels are treated (e.g., before hand piled wood is burned), 

tree mortality can be high. Treated areas are less likely to stop a fire, but can be useful during fire suppression. Treatments to reduce fire hazard often focus on thinning 

from below to reduce vertical (ladder fuels) and horizontal continuity of fuels, as 

well as treatments to reduce the amount of fuel available on the ground. Grazing is 

a common treatment in grasslands and shrublands. Mechanical treatments, such as 
48 USDA Forest Service RMRS-GTR-290. 2012.

mastication, chipping, piling by hand or machine, and compaction, as well as burning treatments including piling and burning and broadcast burning, are all designed 

to reduce the amount of fuel available to burn in subsequent wildfires.

In recent reviews, Graham and others (1999; 2004; 2009) found abundant evidence that forest fuel treatments can reduce fire intensity and fire severity, and that 

treatment effectiveness varies with time, location, and time since treatment. In 2007, 

many fuel treatments were subjected to wildfires and multiple case study assessments 

judged them effective. That fuel treatments work to alter fire behavior, make fire 

suppression easier, and make fires less severe is supported by simulation modeling 

(Pollet and Omi 2002) and recent case studies using remote sensing and field assessments where 2007 fires burned into treated areas (Fites and others 2007; Harbert and 

others 2007; Murphy and others 2007; Martinson and others 2008; Hudak and others 

2011). Fuel treatments challenged by 2007 fires, many of which burned under relatively extreme conditions, were generally judged successful unless treatments were 

not complete or homes were readily ignited by burning embers. Wimberly and others 

(2009) found similar results. Continued effectiveness will depend on maintenance 

and retreatment (Graham and others 2009). Hartsough and others (2008) review the 

costs of alternative fuel treatments for dry forests.

While there is general agreement that removing and reducing fuels will reduce fire 

intensity, not all agree that fuel treatments work. They are likely more successful in 

dry forests and adjacent to buildings (Graham and others 1999; Graham and others 

2004; Graham and others 2009), but it is unclear how this will change with climate. 

Until the recent case studies, many assessments have been qualitative or based on 

simulation models with little empirical data. Rhodes and Baker (2008) argued that 

fuel treatments were unlikely to be burned. Using extensive fire records for Western 

U.S. Forest Service lands, they estimated that a given fuel treatment had a 2-8% 

probability of being burned in a moderate- or high-severity fire within 20 years of 

implementation. Thus, it is important to ensure that fuel treatments are ecologically 

appropriate, socially acceptable, and feasible as vegetation management treatments 

(Schoennagel and others 2004; Graham and others 2009). Further, fuel treatments 

seldom stop fires, though fire fighters can effectively use them in fire suppression efforts. There is general agreement that fuel treatments immediately adjacent to homes 

are more effective (e.g., in the home ignition zone, Cohen 2000), and that fuel treatment effectiveness will vary among forest ecosystems and with the fire behavior and 

weather (Schoennagel and others 2004). Further, what is ecologically appropriate, 

sustainable, and socially acceptable will vary from place to place (Graham and others 2009).

Bark Beetles and Burn Severity

Insects, including bark beetles and defoliators, are major disturbances that along 

with fire, wind, and human action have shaped forest composition, pattern, and structure. Landscapes are dynamic, and most places are in some state of recovery from 

disturbance. Interactions between insects and fire as agents of forest disturbance have 

many implications for landscape dynamics, carbon, sustainability, and resilience, but 

the interactions are poorly understood, especially at landscape scales. Although we 

generally lack good historical data on the extent of tree mortality from diseases and 

insects, the area affected is increasing and will likely continue to increase. Between 

1997 and 2001, the five-year trend ranged between 2 to 3 million acres affected per 

year in U.S. forests. From 2002 to 2007, extensive tree mortality occurred on approximately 5 to 12 million acres per year (NWCG [National Wildfire Coordinating 

Group] 2009). Similar trends are apparent around the globe (Allen and others 2010).
USDA Forest Service RMRS-GTR-290. 2012. 49

There is strong agreement from observational and modeling studies that 

extensive tree mortality due to bark beetles, defoliators, and other agents can affect 

the available fuels and crown fire behavior (Schoennagel and others 2004; Jenkins 

and others 2008; Hoffman and others 2011; Simard and others 2011). However, 

there are conflicting reports in the peer-reviewed literature on the influence of 

insect-induced changes to stand structure on resulting fire behavior (Jolly and 

others 2012b) and outcomes are less certain for burn severity. Many people think 

trees killed by insect outbreaks are more likely to burn and burn severely (Geiszler 

and others 1980; Knight 1987) because foliage is drier and more flammable (Jolly 

and others 2012a) and there are more fine fuels on the ground and eventually more 

large wood (Page and Jenkins 2007; Jenkins and others 2008). Others argue the 

opposite (Bigler and others 2005; Simard and others 2011) because they feel that 

the species composition and forest structure, especially less continuous fine crown 

fuels, following insect outbreaks are less conducive to fire occurrence and spread 

(Veblen and others 1994; Bebi and others 2003; Kulakowski and others 2003). 

Likely, crown fire hazard is high where the proportion of trees killed by bark beetles 

is high and the red needles are still in the trees, but then the hazard decreases as 

red needles fall (sometimes called the gray stage) (Hoffman and others 2011). Tree 

mortality could also increase fire hazard by increasing amount of solar radiation 

in the subcanopy, which is not only effective in drying surface fuels (Hoffman 

and others 2011) but also in causing earlier snowmelt (Pugh and Small 2011). The 

effects of tree mortality vary rapidly with time since outbreak and likely vary with 

site (Page and Jenkins 2007). Fire-induced tree injury could favor insect attack of 

stressed trees (McCullough and others 1998), and this was evident in some studies 

(Bradley and Tueller 2001; McHugh and others 2003; Wallin and others 2003; 

Cunningham and others 2005), but not others (Elkin and Reid 2004), or no relation 

was found (Sanchez-Martinez and Wagner 2002).

Lynch and others (2006) found that the extent of the 1988 fires in Yellowstone 

National Park were related to bark beetle outbreak 13-16 years prior (but not to bark 

beetle outbreak 5-8 years prior), drought, and aspect. Prior mountain pine beetleinduced tree mortality increased the odds of burning by 11%. They concluded that 

for fires following mountain pine beetle outbreaks, the effect of the changed stand 

structure and composition (increased understory vegetation) resulting from canopy 

mortality were more important than the increase in fuels. Bigler and others (2005) 

found that prior stand structure that resulted from multiple disturbances, including 

bark beetles, affected burn severity in spruce forests in Colorado. Kulakowski and 

others (2003) found that areas affected by a 1940s spruce beetle outbreak burned 

less often by a 1950 fire than would be expected at random. It is very likely that 

this effect varies with drought severity, insect species and associated tree species, 

and the extent and timing of mortality relative to time of the burn. It is likely that 

these relationships vary with insects, time since outbreak, severity of the outbreak, 

and climatic conditions (Jenkins and others 2008). Hoffman and others (2011) 

found that fire intensity and crown consumption increased with level of mortality 

in mixed conifer stands.

Key Uncertainties

The many interactions among fire, vegetation, topography, land use, and climate, 

and between fire and other disturbances, will likely lead to non-linear, synergistic, 

and unexpected effects as climate changes. The effectiveness of post-fire management, including salvage and rehabilitation, is poorly studied, especially in 

streamside areas. There are many unanswered questions. How will projected future 
50 USDA Forest Service RMRS-GTR-290. 2012.

changes in vegetation composition and structure associated with climate warming 

and disturbance influence future fire extent and ecological effects? At what point 

do landscapes become fire-limited? In other words, when do we have enough past 

wildfires and fuel treatments such that future fires become self-limiting? How will 

severity and spatial patterns of fire change as fires become more extensive with 

climate warming? How will ecological effects of shifts in seasonality of burning 

(e.g,. earlier fires) influence ecological effects? Despite these uncertainties, fires 

are occurring and decisions about management strategies must be made before and 

during fires.

Fire Management Strategies

Fire management, including suppression and management of intentional fires 

and lightning-ignited fires, fuel treatments, and other vegetation management 

treatments, are widely applied, often with the goal of altering the size and severity 

of subsequent wildfires. These actions can be taken at local to landscape scales. 

Fires respond to and interact with the vegetation dynamics that are often a legacy 

of past disturbance.

Fire managers and scientists now widely acknowledge the critical and natural 

role fire plays in most terrestrial ecosystems. The last two decades have seen a 

paradigm shift from a suppression orientation to an acknowledgement of the need 

to use wildland fire as a whole-landscape management tool. This shift in thinking acknowledges that in the western United States, where the scale of land use 

change due to fire exclusion has been large relative to the size of stand treatments, 

use of larger, naturally occurring fire is essential. However, fire management is 

complicated by many of the factors described above, including climate change and 

growing western U.S. populations in fire-prone areas.

Many efforts are focused on homes to reduce the likelihood that they will ignite in a rain of embers when surrounding wildlands burn (www.firewise.org, 

Cohen 2000). If homes were less likely to ignite, more different fire management 

strategies would be possible. However, designing “fire-smart” landscapes that are 

resilient to the effects of fire on both ecological and social systems is challenging, 

both from a technical and social perspective. Increasing public awareness of the 

importance of fire as a natural process, and the individual homeowners’ role in 

facilitating fire management, will be key to the successful implementation of fire 

use strategies in the future.

Effective fire management at the landscape scale will require thoughtful assessment and means to take advantage of past fires, prescribed fire treatments, and 

local topography and other conditions to understand and manage at the landscape 

scale if we are to address the implications of fires for streams and fish, and for the 

many other landscape values. Likely, thinning or other fuels treatments alone will 

not be enough to alter the size, severity, and occurrence of fires. Increased use of 

naturally occurring fires is helping managers meet the need for fuel reductions in 

some places, and there is a need to resolve risks and challenges in others to open 

opportunities to take advantage of natural fires in others.

The fire management challenges are many. Fire organizations are under intense 

pressure to reduce costs and ensure the safety of fire personnel while protecting 

people and property, addressing smoke impacts on human health and visibility, 

and realigning public perceptions about fire and fire impacts. They do so through 

concerted efforts before, during, and after fires. Initial attack is largely successful, 

so the few large fires burn under very hot, dry, and windy conditions. This reinforces perceptions that fires are always large, intense, and threatening to people 
USDA Forest Service RMRS-GTR-290. 2012. 51

and their property or other valued resources. Many of these large fires are managed 

for months, and when there are some large fires there are typically many large fires, 

severely challenging fire suppression resources and budgets to pay for them. Since 

2000, fire managers have sought to provide strategic, comprehensive plans. These 

will be even more important in the near future given implications of projected 

climate change, drought and fuel conditions, demographic shifts in human society, 

public expectations in the wildland urban interface, budget limitations, and demand for fire suppression resources to respond to other natural disasters (NWCG 

[National Wildfire Coordinating Group] 2009).

Summary

It is quite likely that fires will mediate the effects of climate change on forests 

and associated aquatic ecosystems. How forests will respond is uncertain given 

that the effects of climate are both direct and indirect, but we can expect to see 

changes in where, why, and when fires burn. Ecological effects of those fires will 

vary depending on where and when fires occur. Whether and how forest species 

will adapt depends on how climate variability affects them and the extent, frequency, and severity of fires. Practical solutions depend on framing constructive 

approaches that facilitate future ecological and social resilience to those fires.

B. Riparian Forests, Climate Change, Fire

This brief synthesis focuses on characteristics that differentiate riparian areas 

from uplands in considerations of fire, forests, and climate change. The valued 

habitat functions provided by riparian vegetation are discussed, as well as how 

these may change with shifting climate and management actions. The role of natural and human disturbance in shaping riparian communities is described, with 

emphasis on the role of fire.

Riparian Vegetation, Values, and Connection to Streams

Uniqueness and Natural Variability of Riparian Vegetation

Riparian plant communities are frequently the most floristically and structurally diverse vegetation in a given region (Naiman and others 1993; Naiman and 

others 1998; Pollock and others 1998; Tabacchi and others 1998; Naiman and others 2005). Stream-riparian corridors are characterized by multidimensional spatial 

gradients that change within a watershed in response to elevation, aspect, lithology, stream size, and local and regional geomorphology and hydrology (Naiman 

and others 2005; Wohl and others 2007). Streamside vegetation reflects these local physical features (Baker 1989; Friedman and others 2006). Because of their 

transitional location at the land water ecotone, riparian vegetation may include 

upland, riparian, and wetland species, and a range of life forms and functional 

groups (Pollock and others 1998). High levels of biodiversity in riparian areas are 

maintained by spatial habitat heterogeneity (Pabst and Spies 1999; Sarr and others 

2005).

The diversity of riparian areas is also attributed to the temporal variability in 

natural disturbances, such as floods, debris flows, landslides, and wildfire (Gecy 

and Wilson 1990; Naiman and others 2005). Hydrogeomorphic disturbances, including seasonal variability of flow and sediment erosion, transport and deposition 
52 USDA Forest Service RMRS-GTR-290. 2012.

contribute to the shifting mosaic of physical landform patches and associated biotic communities along stream-riparian corridors (Poff and others 1997; Corenbilt 

and others 2009; Merritt and others 2009). Successional patterns of riparian plant 

community development are driven by responses to natural and anthropogenic disturbances, physical variables, and plant species attributes (Baker 1989; Merritt 

and others 2009). There are also feedbacks between riparian plant species and 

the physical environment. These involve plant features that influence sediment 

deposition and accumulation and lead to biostabilization of streambanks and floodplains. Riparian plant characteristics include mechanical resistance and flexibility, 

root anchorage ability, and post-disturbance regeneration via sprouts and seedlings 

that influence sediment deposition and accumulation (Petitt and Naiman 2007; 

Corenbilt and others 2009). Thus, the diverse composition and structure of riparian vegetation are a result of the interdependence of physical and biotic processes 

over time (Bennett and Simon 2004)

The natural variability of riparian plant communities can pose management 

challenges and is apparent in the many classifications that have been developed 

for national forests and states in the western United States (e.g., Hansen and others 

1995; Manning and Padgett 1995; Crowe and Clausnitzer 1997; Carsey and others 

2003). Most classifications are based on plot-level vegetation sampling but indicate 

the dependence of streamside plant distributions on elevation, hydrogeomorphic 

features, landscape position, and location within watersheds. These classifications 

have served as management tools, and may be useful in determining the vulnerability of some riparian community types to climate change. Current challenges 

for riparian management include: (1) the integration of existing riparian classifications with developments in landscape ecology that highlight the role of landscape 

position and location within watersheds; (2) prediction of changes to riparian vegetation in response to climate-related shifts in temperature and precipitation given 

local and regional characteristics, watershed condition, and disturbance regimes; 

and (3) maintenance of valued riparian functions.

Valued Functions of Riparian Plant Communities

Riparian areas cover a relatively small area in any given watershed, yet they 

provide critical ecological functions (Brinson and others 2002; Naiman and others 

2005). They are disproportionately important for maintenance of water quality and 

quantity (water storage and aquifer recharge), habitat for aquatic and terrestrial 

biota, sediment retention, stream bank building and maintenance, and provision 

of services of economic and social value (Gregory and others 1991; Prichard and 

others 1993; revised 1995, 1998; Naiman and Decamps 1997; Brinson and others 

2002; Naiman and others 2005). The focus here is on the contribution of riparian 

vegetation to the maintenance of aquatic habitat for native fishes, specifically: (1) 

provision of shade for thermal modification of stream temperature; (2) inputs of 

large wood for instream habitat complexity; (3) allochthonous organic matter inputs to aquatic food webs; and (4) provision of streamside habitat and stabilization 

of streambanks. Each of these functions could be altered at the reach scale with 

changes in riparian vegetation, including short-term responses to fire and longerterm responses to changing climate.

Stream temperature: Along many stream segments, riparian vegetation attenuates the input of solar radiation. Direct sunlight warms streams, particularly 

during periods of low flow. During winter, lack of cover can affect stream temperature by permitting radiant cooling to the sky, potentially resulting in the formation 

of anchor ice (Ashton 1989). Riparian and topographic shading moderates these 
USDA Forest Service RMRS-GTR-290. 2012. 53

thermal fluctuations. Stream temperature has tremendous ecological importance 

for aquatic biota and for ecosystem processes such as productivity and nutrient 

cycling (Sweeney 1992; Allan and Castillo 2007; McCullough and others 2009). 

Water temperature strongly influences growth, development, and behavioral patterns of aquatic biota directly and because of its influence on dissolved oxygen 

concentrations (Sweeney 1993; McCullough and others 2009). Stream temperature 

is an important factor determining the distribution of fish in freshwater streams, 

and most species of concern have limited temperature tolerances (Torgersen and 

others 1999; Dunham and others 2007; Isaak and others 2010).

Stream water temperature varies markedly within and among stream systems 

(Poole and Berman 2001; Caissie 2006). Natural influences on water temperature 

include topographic shade, upland and riparian vegetation, ambient air temperature and relative humidity, altitude, latitude, discharge, water source, and solar 

angle and radiation (Poole and Berman 2001; Ebersole and others 2003). Various 

approaches to modeling stream temperature have been developed; in general, these 

either examine components of an energy budget with deterministic models, or develop regression or stochastic models based on relationships between air and water 

temperatures (Caissie 2006). Whatever approach is used, riparian vegetation is implicitly included in the radiation terms, since riparian shade protects streams from 

excessive heating or radiation. For the upper Boise River basin, western Idaho, 

Isaak and others (2010) developed a series of multiple regression models to determine the relative importance of input variables on summer stream temperature 

(means and maxima). Consistent with other studies, they found that three critical 

input variables were air temperature, stream flow (which together describe the time 

variation in temperature), and radiation (the most significant geographically varying quantity). They also evaluated the role of fire on stream temperature and found 

that stream temperature increases over the study period averaged 2-3 times greater 

within the burned portions of watersheds than basin averages, and that increases 

in radiation accounted for 50% of the warming. These results highlight the role of 

both upland and riparian vegetation in moderating incoming radiation and reducing stream temperatures, particularly following fire.

Effectiveness of vegetation in providing stream shade varies with topography, 

channel size and orientation, extent of canopy cover above the channel, and vegetation structure. Streams in different regions and stream segments in different 

parts of a basin vary in response and sensitivity to disturbance and human activities that alter vegetative shading (Poole and Berman 2001). However, stream 

shading by riparian and upland vegetation is one of the few factors that can be 

actively managed to achieve stream temperature targets, as reflected by riparian 

Best Management Practices and designation of riparian buffer widths (Beschta and 

others 1987; Belt and others 1992). With predictions of rising stream temperatures 

in response to changing climate and increased incidence of fire, more focus will 

be directed toward manipulation and restoration of riparian vegetation to increase 

shade (Davies 2010; Furniss and others 2010). An important aspect of prioritizing 

future restoration efforts will be to identify stream reaches where increasing or 

maintaining riparian shade could protect or extend the longitudinal influence of 

cold groundwater influxes (Isaak and others 2010). Potentially, extension of fuel 

reduction treatments into riparian areas may also reduce the fire risk or decrease 

the severity of wildfires along stream-riparian corridors. Although reduction of 

riparian fuels may reduce effective shade in the short-term, i.e., for several posttreatment years, vegetative recovery following treatment (or wildfire) may proceed 

more quickly and vigorously and prolong shade benefits over decades.
54 USDA Forest Service RMRS-GTR-290. 2012.

Inputs of large wood for instream habitat complexity: Over the last three 

decades, extensive literature has documented the hydrological, ecological, and 

geomorphic effects of instream large wood and reported on the role that large wood 

plays in linking aquatic, riparian, and upland portions of watersheds (Lienkaemper 

and Swanson 1987; Bilby and Bisson 1998; Gregory and others 2003a). Large 

wood strongly influences channel form in small streams, creating pools and waterfalls and affecting channel width and depth (Montgomery and others 2003). Many 

aquatic species use pools formed by large wood as habitat and in-stream wood for 

cover (Bilby and Bisson 1998; Wondzell and Bisson 2003). The presence of large 

wood in streams affects erosion, transport, and deposition of sediment, as well as 

the creation and growth of gravel bars and channel and floodplain sedimentation 

(Montgomery and others 2003). Dams formed by accumulations of large wood 

increase channel complexity and facilitate deposition of organic matter, thus providing a food source for numerous invertebrate species and contributing to nutrient 

cycling and retention (Bilby and Bisson 1998; Wondzell and Bisson 2003). The influence of wood in affecting stream morphology depends on the size of the stream 

and the size of the wood pieces (Bilby and Ward 1989; Marcus and others 2002; 

Wohl and Jaeger 2009). The function of LW in forming fish habitat, especially 

plunge and dammed pools, is strongly influenced by the location of the stream or 

reach within a given watershed (Richmond and Fausch 1995).

Less well documented are the different processes of wood recruitment, retention, transport and turnover, and longitudinal distribution of wood pieces and jams 

within stream networks (but see May and Gresswell 2003; Wohl and Goode 2008; 

Wohl and Cadol 2011). These are important considerations for estimation of instream large wood targets and the long-term management of streamside forests 

and in-channel habitat. Chronic inputs of large wood to stream channels occur as 

a result of bank erosion, windthrow, and mortality of individual trees from adjacent hillslopes and riparian areas (McDade and others 1990; Bragg 2000; Benda 

and others 2003b; Reeves and others 2003). Large pulses of wood may originate 

from near channel sources following fire (Figure 24), windthrow, or insect infestations, or pulses may be transported from other portions of a watershed by debris 

torrents, avalanches, or landslides (Bilby and Bisson 1998; Bragg 2000; Benda 

and others 2003b). The relative importance of chronic LW inputs vs. episodic, 

disturbance-related inputs varies in time and space (Benda and others 2003b) and 

is reflected in wood distribution at multiple scales. In one of the few empirical 

studies to quantify the longitudinal distribution of instream LW, Wohl and Jaeger 

(2009) surveyed wood pieces in 50 contiguous stream segments, each segment 

25 m in length (total surveyed length = 1250 m per stream) along 12 streams in 

the Colorado Front Range. Their results suggested that local valley and channel 

geometry, i.e., valley-bottom width, gradient, and sequence of channel changes, 

exerted a stronger influence on patterns of longitudinal wood distribution than either time since last forest disturbance or progressive downstream trends associated 

with larger drainage area. They also found that the combination of forest stand age, 

longitudinal sequences of wood recruitment sources (hillslope and riparian), and 

channel geometry significantly influenced reach-scale wood loads and aggregation 

patterns. These findings represent one point in time; the temporal variation in LW 

loads, which includes disturbance-related inputs and wood movement, creates additional complexity.

Retention and transport of instream LW depends on wood piece dimensions, notably diameter and piece length relative to channel width, stream flow regime, and 

channel characteristics. Reported values for wood residence time in streams vary 

from weeks to centuries (Wohl and Goode 2008), although residence time of log 
USDA Forest Service RMRS-GTR-290. 2012. 55

Figure 24a: Post-fire inputs of large wood 

(source=hillslope) to Boulder Creek, 

Bridger-Teton National Forest, Wyoming. 

Photo taken in 2007, seven years following 

the Boulder Fire (2000). At the time of the 

photo, approximately 75% of the hillslope 

and riparian ‘recruitable wood’ had entered 

the stream along this reach.

Figure 24b: Post-fire inputs of large wood (source = riparian) to Boulder Creek, Bridger-Teton National Forest, 

Wyoming. Photos were taken in 2007 (left) and 2011 (right), 7 and 11 years following the Boulder Fire 

(2000). In 2011, approximately 90% of the riparian ‘recruitable wood’ had either entered the stream channel 

or fallen on the floodplain along this reach.
56 USDA Forest Service RMRS-GTR-290. 2012.

jams is longest in small headwaters and tends to decrease with increased drainage 

area (Martin and Benda 2001).

In second and third order streams, however, most researchers have reported fairly 

rapid turnover (<10 years). In five Colorado mountain streams, Wohl and Goode 

(2008) found that reach-scale wood loads and logjam locations remained relatively 

constant during an 11-year monitoring study. Although results from other regions 

vary, instream LW has been shown to be mobile and dynamic, and the physical factors influencing in-channel wood distribution and loads are similar (Lienkaemper 

and Swanson 1987). Most published studies have presented data on chronic inputs; 

because tracking individual LW pieces is time-consuming and labor-intensive (Wohl 

and others 2010), few studies have monitored individual reaches beyond 10 years, 

particularly the fate of LW pieces following fire and other disturbances. Field surveys 

of short durations have assisted in defining natural variability in wood recruitment 

and storage for a few forest types, but questions about long-term dynamics, watershed patterns, and integration of disturbance processes—difficult to address based on 

sparse empirical data alone—have led to modeling efforts.

Instream LW dynamics have been simulated using deterministic and stochastic models that incorporate a range of recruitment, transport, and decay processes 

(Bragg 2000; Bragg and others 2000; Benda and others 2003b; Gregory and others 

2003a; Gregory and others 2003b; Meleason and others 2003). Simulation models 

have been run at reach and watershed scales, using empirical or derived data on 

upland and riparian vegetation and terrain. Some have specifically included disturbances, notably fire, landslides and mass failure, forest harvest and insect outbreaks. 

Most models have been developed in the Pacific Northwest, reflecting the history 

of LW research and existence of empirical data for this region (Beechie and others 

2000; Gregory and others 2003b). Model objectives have focused on recruitment 

dynamics; input variables include streamside forest attributes, rates of wood delivery to the stream, and depletion from decay, transport, and breakage. By necessity, 

existing models are very simplistic representations of riparian forests, and most do 

not address the role of channel characteristics on the distribution of wood. To date, 

model assumptions have not been well supported by empirical data. Despite these 

limitations, the development of quantitative wood supply models has highlighted the 

importance of riparian forest processes and improved understanding of the role of 

disturbance in LW recruitment to streams. Future model development and application 

will be necessary to predict and manage for instream LW over varying time periods, 

across stream networks, and with different scenarios that incorporate climate-related 

disturbances, including changes in streamflow regimes and fire frequency.

Conceptual models of LW distribution and dynamics are generally based on a 

simplified landscape view of stream networks or watersheds, classified into three 

dominant morphologies: high gradient, small headwaters; intermediate, 3rd and 4th 

order stream segments; large, low-gradient, meandering streams and rivers (Marcus 

and others 2002; Swanson 2003; Wohl and Jaeger 2009 and see text box on large 

wood dynamics). As noted above, vegetation, physical constraints, and natural hydrologic, sediment, and disturbance regimes differ markedly in these portions of 

river and riparian landscapes and strongly influence LW distribution and dynamics. 

Instream LW loads are generally highest in the headwater portions, where trees are 

large and small channel size and stream power limit mobility (transport-limited). In 

intermediate stream reaches, correlations have been documented between wood load 

and drainage area, elevation, channel width, bed gradients, and total stream power. 

Although few data have been collected over the required time periods, intermediate channels appear to display a dynamic equilibrium, where LW pieces are moved 

out at approximately the same rate that they enter the channel (Marcus and others 
USDA Forest Service RMRS-GTR-290. 2012. 57

2002; Wohl and Goode 2008). In large, low-gradient streams and rivers, the spatial 

distribution of LW varies widely but is supply-limited due to reduced areal contact 

with riparian edges. This broad framework serves as a starting point for estimating 

reasonable LW targets and rates of chronic recruitment, due to bank erosion and 

mortality. Although more challenging, it may also prove useful in assessing the role 

of shifting climate-related disturbance regimes in the delivery and movement of instream wood.

Contributions to aquatic foodwebs: Organic matter in streams can either be 

produced by aquatic organisms (autochthonous) or enter the stream from other 

sources (allochthonous) (Allan and Castillo 2007). Autochthonous organic matter 

is generated through photosynthetic production by autotrophic organisms of the 

aquatic community (vascular plants, bryophytes, algae, bacteria, and protists) and is 

driven by the amount of light reaching the stream surface. In contrast, allochthonous 

organic matter originates directly from riparian or upland vegetation in the form of 

leaves, twigs, and other fine litter and indirectly as terrestrial invertebrates (Bisson 

and Bilby 1998). The input, use, retention, and transport of allochthonous organic 

matter in streams frequently drive carbon and nutrient dynamics and affect aquatic 

biota (Webster and Meyer 1997). For many low order streams in forested watersheds, the energy for aquatic food webs is largely derived from allochthonous inputs 

(Vannote and others 1980; Newbold and others 1982). Allochthonous plant sources 

vary widely in nutritional quality and require different degrees of in-stream processing and conditioning by microbes and invertebrates (Webster and Benfield 1986; 

Allan and Castillo 2007). In some areas, seasonal inputs of terrestrial insects from 

riparian areas are an important food source for drift feeding fish species (Young and 

others 1997); such inputs are highest from closed canopy riparian areas dominated 

by deciduous plant species (Edwards and Huryn 1996; Nakano and others 1999; 

Baxter and others 2004; Baxter and others 2005). For floodplain forests, it has been 

suggested that the effectiveness of riparian vegetation in providing allochthonous 

inputs to streams declines at distances greater than approximately one-half a tree 

height away from the channel (FEMAT [Forest Ecosystem Management Assessment 

Team] 1993). Removal of riparian vegetation by fire reduces the amount and quality 

of allochthonous inputs and promotes autotrophic production by increasing available light (Bisson and Bilby 1998; Malison and Baxter 2010), causing shifts in the 

feeding guild composition of stream macroinvertebrate communities and changes 

in aquatic trophic pathways that affect fish productivity (Edwards and Huryn 1996; 

Bisson and Bilby 1998; Bisson and others 2003).

Streambank stability: Riparian vegetation can directly affect stream channel 

characteristics, particularly streambank habitat and stability (Gregory and Gurnell 

1988; Davies-Colley 1997; Simon and Collison 2002; Pollen and others 2004). 

Root systems protect stream banks through armoring (Stokes and Mattheck 1996; 

Abernathy and Rutherford 2001) and bind bank sediment, thus contributing to bank 

stabilization, reduction of sediment inputs to streams (Dunaway and others 1994), 

and development and maintenance of undercut banks (Sedell and Beschta 1991). 

There are marked differences among riparian species and vegetation types in root 

characteristics and their influence on bank stability (Lyons and others 2000; Simon 

and Collison 2002; Wynn and others 2004). Management activities, such as logging 

and grazing, and natural disturbances, such as fire and debris flows, can directly affect 

stream bank stability through alteration of riparian vegetation. Removal of woody 

riparian vegetation with beneficial rooting characteristics can result in erosion of 

alluvial streambanks. Removal of herbaceous vegetation can decrease retention and 

accumulation of sediment, possibly influencing floodplain soil development (Thorne 
58 USDA Forest Service RMRS-GTR-290. 2012.

Instream Large Wood (LW) 

Dynamics: Influence of 

Disturbance and Changing Climate

The numerous roles and benefits of instream large 

wood (LW) have been extensively documented 

in the recent fluvial geomorphology, forestry, 

fisheries, and stream ecology literature. In smaller 

streams, large wood pieces strongly influence 

channel form, creating pools, log steps, waterfalls, 

and affect both stream width and depth. In second 

order and larger streams, channel features formed 

by LW, such as pools and eddies, provide critical 

habitat for native fish species. Aquatic species 

also use in-stream logs as cover, feeding and 

colonization substrates. Instream LW contributes 

to the growth of sandbars and gravel bars, affects 

the erosion, transport and deposition of sediment, 

and increases channel complexity. Due to these 

benefits and the consequent need to manage for 

instream LW, the amounts, characteristics and 

physical controls on instream wood loads have also 

been studied. Less well documented, however, are 

the timing and processes of wood recruitment, 

transport and turnover, and the ever-shifting 

longitudinal distribution of wood pieces and jams 

within stream networks. Although some aspects 

of LW dynamics have been modeled, models 

predicting instream wood supply and transport are 

not generally available for most regions. Challenges 

for managers are the assessment of current wood 

loads for their streams of concern, setting targets 

for loads and distribution over stream segments 

(meters to kilometers) and different time scales 

(years to centuries), and the incorporation of 

a disturbance perspective in plans for future 

management.

Conceptual models of LW distribution and 

dynamics are generally based on a simplified 

landscape view, which classify stream networks 

or watersheds into three dominant portions: high 

gradient, small headwater streams at the highest 

elevations; intermediate, 3rd and 4th order stream 

segments; and large, low-gradient, meandering 

streams and rivers (Fig. T6-1). Vegetation, physical 

constraints, and natural hydrologic, sediment, 

and disturbance regimes differ markedly in these 

portions of stream-riparian corridors and exert 

strong controls on LW distribution and transport. 

Instream LW loads are generally highest in the 

headwater portions, where trees are large and 

small channel size and stream power limits mobility 

(transport-limited). In intermediate stream reaches, 

correlations have been documented between 

wood load and drainage area, elevation, channel 

width, bed gradient and total stream power. 

Wood load decreases downstream, although the 

frequency of wood jams, or accumulations, is 

highest along these intermediate stream portions. 

In large, low-gradient streams and rivers, the spatial 

distribution of LW varies widely, but is frequently 

supply-limited due to reduced areal contact with 

forested riparian edges. Disturbance processes are 

integral to the understanding and management 

of instream LW, and the relative importance of 

different disturbance types also changes with 

landscape position. In headwaters and many 

intermediate reaches, fire, windthrow, landslides, 

and debris flows are responsible for delivering large 

pulses of wood to streams. In intermediate and low 

gradient portions of channels, flooding, waterlogging and lateral channel migration influence LW 

inputs, redistribution and accumulation patterns. 

Anthropogenic disturbances that can affect delivery 

of LW to streams, such as mining, timber harvest, 

and road construction, also differ with elevation 

and location within a watershed.

At the reach or stream-segment scale, simple 

wood budgets can be estimated through 

calculation of inputs and outputs, which forces 

consideration of the relative contribution of 

different disturbances for recruitment, breakage, 

and movement of LW pieces. The conceptual 

framework of watershed-level LW distribution can 

be combined with estimated wood budgets for 

longitudinal segments and thus serve as a starting 

point for setting LW loading targets. Management 

and restoration of instream LW involves linking 

riparian forest processes with hydrology and 

fluvial geomorphology; planning for impacts of 

climate change requires consideration of altered 

disturbance regimes and the responses of riparian 

vegetation. Indirect outcomes of climatic warming 

will be (1) changes in LW recruitment to streams, 

including greater near-term inputs following fire or 

blow-down, and possible decreased inputs in the 

long-term; (2) for some stream segments, the size 

and species of trees entering streams will change, 

with fewer large, old-growth trees; (3) changes 

in transport rates and seasonal redistribution of 

instream LW, in response to shifting flooding and 

drought cycles; (4) local changes, such as increased 

frequency of debris flows, depending on lithology 

and precipitation patterns. 
USDA Forest Service RMRS-GTR-290. 2012. 59

Figure T6-1. Conceptual model of large wood loads and spatial distribution along mountain streams. Although the 

model was generated from data collected in streams of the Colorado Front Range, the basic framework applies to 

most streams throughout the forested, mountainous West (Wohl and Jaeger 2009). 

1990). Impacts of local alterations to riparian vegetation that affect bank stability and 

other geomorphic processes may have effects that extend downstream.

Disturbances in Riparian Areas

The development and maintenance of riparian environments are largely regulated by physical processes and natural disturbance regimes (Naiman and others 

2005). Stream and river systems are naturally dynamic, changing at multiple spatial and temporal scales, frequently in response to episodic disturbance events. 

Geomorphic and hydrologic processes, including disturbances such as flooding and 

debris flows, have largely shaped streamside environments. Riparian plant species 

exhibit a range of adaptations that contribute to rapid recovery of streamside habitat after disturbance (Dwire and Kauffman 2003; Merritt and others 2009). In this 

section, the following natural disturbances, their interactions, and their influence 
60 USDA Forest Service RMRS-GTR-290. 2012.

on riparian areas are discussed: flooding, fire, debris flows, insects, and beaver. A 

brief overview of the impacts and legacy of land use and management, i.e., human 

disturbance, is also presented.

Flooding

Riparian environments are intrinsically linked to the dynamics of stream hydrographs, including flooding. Stream hydrographs show the seasonal and interannual 

variability in flows, and they display characteristic forms depending on the local 

climate, particularly precipitation patterns, and the size and shape of the watershed. Hydrograph peaks correspond to flood events that inundate floodplains, 

scour streambanks, and transport sediment and large wood onto bars and floodplains. Many mountain streams are strongly influenced by spring snowmelt and 

display distinct peaks during spring runoff (Stewart 2009). Low- and mid-order 

streams, and the riparian environments bordering them, are sensitive to individual 

precipitation events, resulting in dynamic hydrographs characterized by multiple 

peaks (floods) over a year. Larger rivers and their riparian environments are less 

sensitive to individual precipitation events because the scale of the basin usually 

surpasses the size of the storm. Also, flow in larger rivers integrates the flow of 

upstream tributaries, some of which may not be flooding. Some arid-land streams 

are intermittent or ephemeral, without surface flow for extended periods; their 

hydrographs reveal seasonal floods, such as those associated with monsoonal rainfall (Stromberg and others 1993). Floods in headwaters initiate flood waves that 

propagate as they travel and accumulate in downstream sections. Thus, the same 

flood event will affect riparian environments in distinct ways depending on location within the watershed, and flood impacts will differ in high-energy portions of 

stream networks relative to low-gradient, meandering portions (Bendix and Hupp 

2000).

Four flood characteristics are important to riparian and floodplain ecosystems: 

magnitude, frequency, timing, and duration. Magnitude refers to the maximal discharge associated with an individual flood and reflects the intensity and severity of 

the event; variations in flood magnitude within a given watershed are expressed as 

recurrence intervals (Gordon and others 2005). The range of flood magnitudes for 

a given stream segment depends mainly on climate and the upstream catchment 

area. Frequency is the temporal pattern of flood recurrence, either over seasons or 

multiple years. Timing of floods is linked directly to precipitation or snowmelt runoff patterns. Flood duration is the amount of time that the riparian area (floodplain) 

is flooded, either seasonally or during individual flood events. Flood duration varies as a function of topography; low-lying areas close to channels flood first and 

are last to drain and thus experience longer flooding duration than other portions of 

the floodplain. Some aspects of these four flood characteristics are changing with 

shifting climate and are discussed in more detail in the earlier hydrology section. 

Many regions are already experiencing changes in magnitude, frequency, and timing of flood events relative to the period of record.

Flooding is an integral, essential disturbance for riparian ecosystems that has 

both geomorphological and hydrological hydraulic impacts (Hupp and Osterkamp 

1996; Bendix and Hupp 2000). Hydraulic impacts include mechanical damage, 

saturation, and transport of sediment, organic material, large wood, and plant propagules. Geomorphological impacts include the shaping of fluvial environments. 

The structure, composition, and distribution of riparian vegetation are strongly 

related to fluvial geomorphological processes and forms. In many cases, species 

occurrences can be linked directly to specific fluvial landforms created by known 
USDA Forest Service RMRS-GTR-290. 2012. 61

flood events (Rood and others 1998). Floods can erode streambanks and undercut, 

topple, and remove standing riparian vegetation. Entrained wood and debris can 

batter riparian trees (Johnson and others 2000), and vegetation can be buried by 

sediment deposited by floodwaters. The mosaic of riparian vegetation can reflect 

the role of floods in the differential destruction of previous vegetation, distribution of substrates and geomorphic surfaces, and in the transport of propagules. 

Depending on post-flood conditions and the climatic context, major floods can 

foster the establishment of vegetation stands or reset successional processes in 

riparian plant communities (Rood and others 1998). There are also feedbacks; 

streamside vegetation physically constrains flood flows, traps sediment and floating debris, and contributes to the erosional resistance of streambanks.

Streamflow regimes have exerted selective pressures on riparian plant species, 

resulting in morphological, physiological, and reproductive adaptations to flow 

attributes (Poff and others 1997; Naiman and others 2005; Poff and others 2007; 

Merritt and others 2009). Many riparian plants are specifically adapted to flooding, 

as well as sediment deposition, physical abrasion, and stem breakage associated with flooding (Karrenberg and others 2002; Naiman and others 2005; Merritt 

and others 2009). For example, the reproductive phenology of common riparian 

woody species, including cottonwoods and many willows, is synchronized to coincide with the seasonal hydrology and rainfall of specific regions (Mahoney and 

Rood 1998; Rood and others 1998). Cottonwood seed dispersal coincides with the 

seasonal retreat of floodwaters when moist seedbeds are available for successful 

germination and colonization. In addition to sexual reproduction by seeds, many 

riparian plant species reproduce by clonal growth (i.e., vegetative or asexual reproduction); multiple sprouts can result from burial during floods and abrasion during 

floods can stimulate stump sprouts (Karrenberg and others 2002).

The disruption of natural flow regimes through diversions, damming, withdrawals, and levees has focused attention on the dependence of riparian species on 

streamflow attributes and different portions of regional hydrographs. In the Rio 

Grande Valley (New Mexico), water withdrawal and flow regulation, including 

the cessation of spring floods, has simplified the valley, which transitioned from 

a mosaic of multiple channels, marshes, wet meadows, and forests to a system 

constrained by levees bordered by a narrow width of riparian forest (Molles and 

others 1998). Similar examples are common throughout the western United States 

and elsewhere worldwide. As the U.S. population continues to grow, increasing 

demands are being placed on water originating or flowing through Forest Service 

administered lands. Managing the limited water supply to meet multiple and 

sometimes competing uses is an ongoing and complex responsibility. Efforts to 

provide water for multiple uses include defining baseline environmental instream 

flow prescriptions that sustain and regenerate riparian habitats and communities. 

Characterizing environmental flows includes a flood component and also addresses 

flow requirements for channel maintenance, in-channel habitat, and maintenance 

of water quality (Richter and Richter 2000; Rathburn and others 2009).

Fire

Wildfire has played a critical role in shaping ecological heterogeneity across landscapes of the West (Agee 1993). Fire has also influenced the species composition, 

structure, and environmental conditions of the riparian and aquatic communities 

associated with stream networks that drain these landscapes (Gom and Rood 1999; 

Gresswell 1999; Everett and others 2003; Skinner 2003; Reeves and others 2006; 

Petitt and Naiman 2007; Stromberg and Rychener 2010). Research on riparian 
62 USDA Forest Service RMRS-GTR-290. 2012.

fire frequency and severity has primarily been conducted in forests of the Pacific 

Northwest (see text box on fire histories in riparian areas). However, results are 

consistent with observations elsewhere and indicate that most riparian areas burn 

either similar to adjacent uplands or less frequently and more moderately than 

uplands. Reviews have summarized research on the role of fire as a natural disturbance in stream-riparian ecosystems, especially in mountainous environments 

(Bisson and others 2003; Dwire and Kauffman 2003; Petitt and Naiman 2007); 

recent work has advanced understanding of post-fire recovery in different settings 

(Mellon and others 2008; Jackson and Sullivan 2009; Malison and Baxter 2010).

Different scenarios of generalized fire behavior and effects in riparian areas have 

been proposed. Pettit and Naiman (2007) described four cases of fire effects, post 

disturbance impacts, and riparian recovery based on their observations of wildfire 

in Kruger National Park, South Africa. The four cases were categorized by stream 

gradient (high or low) and amount of rainfall (high or low). Halofsky and Hibbs 

(2008) developed a sequence of hypotheses to test the relative effect of riparian 

vegetation, valley bottom topography, and upland fire variables on riparian fire 

severity. The relative role of these driving factors varies locally and regionally but 

can be used to predict how wildfire may burn along specific stream segments. Key 

considerations address the connection to the larger landscape and include: location within the watershed relative to precipitation regime (snow vs. rain influence, 

Wohl and others 2007); topography, such as aspect and shifts in stream gradient 

and slope relative to uplands; geomorphology, such as changing width of the channel and valley floor; and riparian vs. upland vegetation and fuel characteristics.

We present four generalized scenarios of fire behavior and effects in riparian areas and speculate about potential responses to climate change (Table 3). Variations 

of these four scenarios occur and different combinations may be observed in the 

same watershed or during the same wildfire. The relative likelihood of occurrence 

for any scenario is largely driven by vegetation and fuel indicators, basic topographic variables, and characteristics of the fire and fire weather.

Riparian areas burn like adjacent uplands: This scenario is most likely to 

occur along stream reaches where the riparian vegetation, terrain, and general 

topography are similar to uplands. Stream reaches that drain shrub-dominated 

portions of drainage networks, such as shrub-steppe ecosystems throughout the 

portions of the Great Basin, or stream segments that drain the lower parts of stream 

networks in shallowly dissected terrain with low local relief, are likely to burn as 

frequently and severely as adjacent uplands. Other examples occur in the upper 

portions of drainages at high to moderate elevations in fairly steep terrain with 

steep stream valleys. This scenario could also occur under conditions of severe fire 

weather, i.e., when a large fire carries across the entire landscape and could overwhelm both the influence of local topography and vegetation differences between 

riparian and upland areas.

Riparian areas burn less frequently and/or less severely than adjacent uplands: In contrast to the above, this scenario is most likely to occur where riparian 

conditions are distinctly wetter or more mesic than upland vegetation. It is the 

most commonly documented scenario in the literature, especially for forests of the 

Pacific Northwest (please see textbox on riparian fire histories). In forested riparian reaches, particularly those located in deeply dissected terrain with north–facing 

aspects that foster cold-air drainage and cool riparian microclimates, fires tend to 

burn less “hot” and less frequently than nearby uplands. However, even within 

similar vegetation associations and in lower portions of drainage networks, the 

relative frequency of fire scars has been found to increase linearly with distance 

from the stream (Everett and others 2003; Skinner 2003).
USDA Forest Service RMRS-GTR-290. 2012. 63

Table 3. Four generalized scenarios of fire behavior in riparian areas. Variations on these four scenarios occur and different 

combinations may be observed in the same watershed or during the same wildfire. Ecological outcomes are given, as well 

as speculation regarding potential responses to shifts in temperature and precipitation regimes. Please see text for additional 

explanation.

Fire Behavior in 

Riparian Areas

Generalized Scenarios When and Where? Ecological Outcome 

Potential Responses to 

Changing Climate 

Riparian areas burn 

like adjacent uplands, 

i.e., wildfires burn 

with similar frequency 

and severity.

When large fires burning under severe 

fire weather exceed the influence of 

local topography and riparian / upland 

vegetation differences;

Where riparian vegetation, terrain and 

topography are similar to uplands.

Impact depends on fire severity, 

season, extent, but generally 

moderate to high.

Slow to moderate recovery, 

via seedlings and resprouting. 

Fire adapted species will likely 

survive.

Likely to increase, i.e., 

more riparian area will 

burn similar to uplands; 

differences in riparian /

upland burn patterns will 

decrease.

Riparian areas burn 

less frequently and/

or less severely than 

adjacent uplands.

Where terrain fosters cold-air 

drainage, higher humidity, cooler 

microclimate relative to uplands;

Where riparian vegetation is 

distinctly different from uplands 

(more hardwoods, higher herbaceous 

component, higher fuel moisture);

Where saturated soil conditions, 

presence of riparian wetlands, or 

hydrologic inputs from hillslopes 

influence fire behavior.

When fires burn with low intensity.

Low to moderate impact.

Moderate to rapid recovery, via 

seedlings and resprouting. Fire 

adapted species (esp. conifers) 

will persist.

Likely to decrease; with 

rising temperatures and 

shifts in precipitation 

patterns, more riparian 

areas will burn similar to 

uplands.

Riparian areas burn 

more frequently and/

or more severely than 

adjacent uplands.

Where fuel abundance/accumulation 

is higher in riparian areas than 

uplands (due to riparian management 

or natural conditions);

When riparian areas serve as 

chimneys or corridors for fire spread, 

e.g., where steep terrain and narrow 

stream valleys influence fire behavior.

Impact depends on fire severity, 

season, extent; can be high 

impact, with destruction of most 

of the riparian community.

Slow to moderate recovery, 

via seedlings and resprouting. 

Post-fire invasive species are a 

potential concern.

May increase or decrease;

Locally dependent on fuel 

characteristics, and/or 

physical context; dependent 

on severity and behavior of 

fire event. 

Riparian areas serve as 

fire breaks.

Where large perennial stream and 

river valleys create significant breaks 

in fuel characteristics and continuity;

Where saturated soil conditions, 

presence of riparian wetlands, or 

hydrologic inputs from hillslopes 

influence fire behavior;

When fires burn with low intensity.

Low impact; rapid recovery. May increase or decrease;

locally dependent on 

physical context; dependent 

on severity and behavior of 

fire event. 
64 USDA Forest Service RMRS-GTR-290. 2012.

Fire History, Severity, Behavior and 

Burn Patterns In Riparian Areas 

Relative to Uplands

Ecologically diverse riparian corridors are 

maintained by active natural disturbance regimes, 

and fire has likely played an important role in 

shaping the heterogeneity of riparian vegetation 

communities. In many forested watersheds, 

riparian vegetation is comprised of a mosaic 

of different successional stages similar to the 

surrounding upslope landscape, suggesting that 

fire and other disturbances historically played a 

similar role in shaping both riparian and upland 

ecosystems.

The current understanding of fire return intervals 

in riparian areas, however, is limited to a few 

studies conducted in the Sierra Nevada Mountains 

(California) and Pacific Northwest region of the 

USA (Table T7-1). Results are varied and depend 

on the physical characteristics (slope, aspect) and 

disturbance history of the study watersheds, prefire moisture conditions, and degree of difference 

between riparian and upland vegetation. In 

general, fires are thought to occur less frequently 

within riparian forests than adjacent upslope 

forests. This is the case for mesic forests in Klamath 

Mountains, California, where fire return intervals in 

riparian areas were approximately twice as long as 

intervals in adjacent upslope forests (Skinner 2003). 

Fire return intervals are usually less frequent in 

streamside areas with notably different vegetation 

relative to uplands, especially higher densities of 

understory shrub and herbaceous species. In mixed 

conifer in the northern Sierra Nevada Mountains 

(CA), Van de Water and North (2010) found 

that fires burned fairly continuously across the 

landscape and were historically common in both 

riparian and upslope forests. At high elevations 

along low order streams, where the vegetation 

composition of riparian areas is similar to that of 

adjacent uplands, streamside areas are also likely to 

burn as frequently as the surrounding uplands.

Patterns observed for fire behavior, intensity, 

and severity in riparian areas are similar to those 

reported for fire frequency. Fires in riparian areas 

may be less severe, as severe, or more severe 

than in adjacent uplands, depending on the local 

topography, vegetation characteristics (especially 

fuel moisture and loading), and fire weather. 

Several recent studies have specifically addressed 

fire behavior in riparian areas, contributing to 

current understanding of riparian burn patterns. 

In different conifer-dominated forests of Oregon, 

Halofsky and Hibbs (2008) compared fire severity in 

riparian vs. upland plots following the Biscuit Fire 

(Klamath-Siskiyou region) and the B&B Complex 

Fire (east Cascades). They found that the strongest 

predictors for riparian overstory fire severity were 

upland fire severity, riparian vegetation indicators, 

and local topography, including stream width and 

gradient, and adjacent hillslope steepness. Their 

study sites had diverse riparian understories, with 

varying levels of the deciduous tree and shrub 

component. Riparian understories generally burned 

less severely than upland understories, and the 

authors attributed this burn pattern to high fuel 

moisture in riparian vegetation, and cooler, moister 

streamside microclimates.

These plot-based results are consistent with a 

geospatial analysis that utilized remotely sensed 

Burned Area Reflectance Classifications (BARC) to 

compare upland vs. riparian burn intensity for four 

large fires (Fisk and others 2004): the Hayman and 

Missionary Ridge Fires in Colorado; Rodeo-Chediski 

Fire in Arizona; and the Stanford Fire in Utah. 

Authors found that riparian areas burned ‘less 

hot’ than upslope areas. However, riparian burn 

values related positively to upslope burn values; 

i.e., the hotter the watershed burned, the hotter 

(on average) the riparian areas burned. Results also 

indicated that smaller, lower order streams burned 

more like uplands, while riparian areas along larger, 

higher-order streams burned less like surrounding 

uplands. These studies are place-based; more 

information is needed on regional variation and 

for a range of riparian plant associations. However, 

the results largely support previous anecdotal 

observations regarding riparian burn patterns and 

relations between upland and riparian fire severity.

Looking forward, climate change predictions 

include rising air temperatures and shifts in 

precipitation regimes. In the western United 

States, these changes will result in longer summer 

droughts, potentially leading to smaller differences 

in riparian vs. upland fuel moisture in the shortterm and reduction in the areal extent of riparian 

zones over time. More intense wildfires are also 

predicted, so differences in fire frequency, severity, 

and intensity in riparian areas vs. uplands will likely 

diminish. 
USDA Forest Service RMRS-GTR-290. 2012. 65

Location Forest Type

Riparian Fire 

Return Interval 

(years)

Sideslope Fire 

Return Interval 

(years) Citation

Blue 

Mountains, OR 

Dry, Douglas-fir 

and Grand Fir 

series 

13-36 10-20 (Olson 2000)

Elkhorn 

Mountains, OR 

Dry, Ponderosa 

Pine, Douglas-fir 

series 

13-14 9-32 (Olson 2000)

Salmon River 

Mountains, ID 

Dry, Ponderosa 

Pine and 

Douglas-fir series 

11-19 9-29 (Barrett 2000)

Cascade Range, 

WA 

Dry, Ponderosa 

Pine and 

Douglas-fir series

15-26 11-19 (Everett and 

others, 2003) 

Northern 

Sierra Nevada 

Mountains, CA

Dry, Ponderosa/

Jeffrey Pine series

10-871 (8-422) 10-561 (6-582) (Van de Water 

and North 

2010) 

Dry Forest Type Average 12-36 10-31

Cascade Range, 

OR 

Mesic, Douglasfir series 35-39 27-36 (Olson and 

Agee 2005)

Klamath 

Mountains, CA 

Mesic, Douglasfir series 16-42 7-13 (Skinner 2003) 

Mesic Forest Type Average 26-41 17-25 

1including only fire events recorded on two or more specimens at a given site

2 Includes every fire event recorded on every specimen

Table T7-1: Fire return intervals for riparian versus upslope forests.

*Table modified from Table 1 in Stone and others (2010)

Riparian areas burn more frequently and/or more severely than adjacent 

uplands: This scenario has been reported by the fire control/fire management 

community (Barrows 1951; Countryman 1971). It has been observed where steep 

terrain and narrow stream valleys create more heat and serve as chimneys or chutes 

that promote updrafts and convective heating of the fire, causing it to carry upslope 

and upstream at a rapid rate of spread with high intensity (Skinner 2003). This 

fire behavior is most likely to occur in the middle or upper portions of drainage 

networks with south-facing aspects, along small perennial or intermittent stream 

channels. Although we are not aware of research that has quantified the vegetative 

conditions that influence this fire behavior, we suspect that riparian vegetation is 

either (1) similar to upland vegetation in stand and understory composition and 

fuel characteristics; or (2) contains higher levels or denser fuel loads, particular 

ladder fuels, than adjacent uplands (Agee 1993). If fire suppression, “hands-off” 

riparian management, or natural processes have contributed to higher accumulations of fuel loads in streamside areas relative to uplands, and if pre-fire moisture 

levels are low due to drought or season, riparian fire severity may be greater than 

adjacent uplands. High riparian fuel loads, especially if uplands have been harvested or actively managed for fuel reduction, can influence fire spread by serving as 
66 USDA Forest Service RMRS-GTR-290. 2012.

“wicks.” This fire behavior was observed during the Angora Fire, Tahoe National 

Forest, California, in late June 2007 (Murphy and others 2007). Prior to ignition, 

the Angora Creek Stream Environment Zone (SEZ, or riparian area) contained 

heavy dead woody fuel loadings. A retrospective evaluation of the Angora Fire 

behavior noted that “dense stands of trees in the Angora SEZ likely contributed to 

the rapid spread upslope to Angora Ridge and across the slope to the base of Tahoe 

Mountain” (Murphy and others 2007; Figure 25). This fire burned over 250 structures on private property, cost approximately $160,000,000 in property loss and 

suppression costs, and has drawn attention to the role of riparian corridors and fuel 

conditions on fire behavior (Murphy and others 2007; Safford and others 2009).

This scenario is locally dependent on fuel characteristics, physical context, and 

the characteristics of a given fire event. However, the contributing riparian conditions may become more common with shifts in temperature and precipitation 

regimes. Although not well documented, riparian areas may also burn more severely in arid landscapes where frequent, low-intensity fires limit fuel abundance 

in uplands, while fuel accumulates in streamside areas. During periods of drought, 

differences in the riparian-vs.-upland microclimate and fuel moisture may be high 

enough to promote plant growth, stand development, and fuel accumulation in 

riparian areas, but not high enough to protect riparian forests from fire. This scenario is of particular concern for resource managers and fuels specialists in some 

locations in the Great Basin and Southwestern United States, where woody encroachment into riparian areas has increased streamside fuel loads.

Riparian areas serve as fuel breaks: This scenario is most commonly observed where large perennial stream and river valleys create significant breaks in 

fuel characteristics and continuity. Wide stream channels, alluvial terraces with 

extensive gravel bars, and large, sparsely vegetated areas with wet soils may function as fuel breaks. Other examples include wet meadows, stream segments with 

a high herbaceous component, and willow-dominated reaches or riparian areas 

Figure 25: Stream Environment Zone (SEZ) along Angora Creek following the Angora Fire, Tahoe NF, California 

(2007). Dense, continuous stands of trees contributed to rapid spread rates (to the NNE) down this stream 

corridor. Arrow points in direction of wind and fastest fire spread (NNE). Note greater density of trees within 

the SEZ (roughly outlined in red). Moister portions of the SEZ (outlined in yellow) burned less severely than 

surrounding areas. (Photo originally published in USDA, R5-TO-025, August 2007.)
USDA Forest Service RMRS-GTR-290. 2012. 67

with a notable hardwood tree and shrub component. These meadow segments are 

frequently located in wider, lower gradient portions of stream networks that may 

receive significant hydrologic inputs (surface and subsurface) from surrounding 

hillslopes, resulting in saturated soil conditions and the presence of riparian or 

slope wetlands. They may be sites of past and current beaver activity that has 

modified the channel and flooded portions of the valley bottom. Saturated soils 

combined with high fuel moisture can stop the advance of fire or cause a fire to 

“jump” from hillslope to hillslope and not burn in the streamside area. In some 

cases, fire characteristics and upland conditions can influence the extent to which 

riparian areas function as fire breaks. If a fire is burning with low-intensity, riparian areas along low gradient, perennial streams may serve as effective barriers to 

fire spread.

Seasonality also plays a role in fire behavior and fire severity and may influence 

each of these scenarios. In mixed conifer stands of the Sierra Nevada Mountains 

(California), Van de Water and North (2010) found that depending on forest type, 

the majority of fire scars in both riparian and upland areas occurred during the 

late summer and fall. Later in the season, as trees become dormant, foliar moisture decreases, increasing the probability of a crown fire (Agee and others 2002). 

However, the ratio between the current year’s growth and older foliage influences 

moisture content as the seasons change. The ratio between old and new foliage 

depends on species and environmental factors such as elevation, site fertility, and 

light (Agee and others 2002). Drought cycles can also be credited for lower foliar and fuel moistures and have been correlated with increased fire occurrence. 

Although this correlation is stronger in uplands, riparian areas also experience 

more fires during times of drought (Van de Water and North 2010).

Debris Flows

In many headwaters and other steep, erosive landscapes, landslides, mass failures, 

and resulting debris flows are common natural disturbances. The occurrence of 

debris flows depends on topography, underlying geology, and soil and vegetation 

characteristics and is frequently associated with fire, past management activities 

(roads and forest harvest), and storm events. In mountainous areas, debris flows 

can play a major role in routing sediment and wood stored on hillslopes and 

in low-order channels and delivering it to higher-order channels (May 2002; 

Istanbulluoglu and others 2003). Because low-order streams lack capacity for 

fluvial transport of large wood, they can accumulate and store large volumes of 

sediment and wood (Swanson and Lienkaemper 1978; May and Gresswell 2003). 

Debris flows episodically transport and redistribute this material to downstream 

portions of the stream network. In mid-order streams in the Oregon Coast range, 

the contribution of wood from debris flows ranged from 11 to 57% of the total 

volume of wood in the channel (May 2002). In the Boise Basin of western Idaho, 

sediment delivered by debris flows has been shown to be beneficial for fish 

spawning habitat (Benda and others 2003a). Although considered a major hazard 

in mountain regions worldwide (Coe and others 2008), debris flows are natural 

processes that contribute to the shifting mosaic of stream and riparian habitat 

patches along stream networks.

The impacts of debris flows on riparian areas are not well documented (but see 

Johnson and others 2000; May and Gresswell 2003; Wohl 2006), although they 

have been anecdotally noted in the geomorphology literature. In steep stream segments, where the channel becomes the runout path, debris flows can scour riparian 

areas, removing soil and vegetation, including large streamside trees. During a 
68 USDA Forest Service RMRS-GTR-290. 2012.

large flood (~ 100 year recurrence interval) in the Cascade Range of Oregon, riparian trees were uprooted and removed for nearly 1.5 km downstream of the debris 

flow tributary channel (Johnson and others 2000). This debris flow contained large 

accumulations of congested large wood, which contributed to the toppling and 

removal of riparian trees. At tributary junctions or along larger channels, debris 

flows can also deposit large volumes of sediment, burying portions of existing 

riparian habitat while creating new geomorphic surfaces for potential vegetation 

colonization and establishment (Gecy and Wilson 1990). Despite limited research, 

debris flows have exerted considerable localized influence on forested riparian areas in mountainous regions.

The occurrence of debris flows in relation to wildfires is of great concern 

throughout the western United States, particularly in steep terrain. Numerous studies have documented increased frequency of debris flows following large-scale, 

severe fires (Swanson 1981; Meyer and others 1992; Cannon and others 2001; 

Istanbulluoglu and others 2002; Istanbulluoglu and others 2003; Pierce and others 2004; Gabet and Bookter 2008; Santi and others 2008). In the Oregon Coast 

range, May and Gresswell (2003) found that a pulse of debris flow activity occurred following the last stand-replacement fire on mid- and upper-slope positions. 

In their study basins, the most recent fire in the upper slopes did not directly impact 

the lower elevation channels or valley bottoms, but the influence of the fire was 

propagated through the stream network by debris flows in the tributaries. In central 

Idaho and northeast Wyoming, Meyer and Pierce (Meyer and Pierce 2003) used 

14C-dated geologic records to examine evidence of past debris flows and fire frequency in relation to long-term climatic reconstruction (last 10,000 years). They 

concluded that drought and a warming climate have contributed to severe wildfires 

and postfire sedimentation, both past and present, and that the incidence of fire 

may increase with future warming. Much remains to be learned about the frequency, magnitude, and spatial extent of debris flows in different regions, as well 

as the rate and direction of temporal recovery for stream, riparian, and hillslope 

ecosystems. This is an active area of research, particularly for physical scientists, 

but is becoming increasingly multidisciplinary as the impacts of debris flows on 

aquatic ecosystems are being investigated at different spatial and temporal scales.

Insect Outbreaks

Insect outbreaks are a recurring natural disturbance in forested ecosystems, but 

current beetle outbreaks (mountain pine beetle, spruce beetle, ips, and others) are 

among the largest and most severe in recorded history (Bentz and others 2010). 

Mechanisms contributing to the widespread outbreaks are complex and influenced 

by multiscale factors, but most insect populations are highly sensitive to changes 

in temperature and moisture. As noted above, air temperature is projected to 

increase across North America, particularly at high latitudes and elevations; 

associated changes in precipitation patterns will result in earlier and longer dry 

seasons across the western United States, with a greater frequency and duration of 

droughts (Seager and al. 2007; Solomon and others 2007). These climatic changes 

will affect the condition, distribution and productivity of forest trees species, as 

well as associated insect populations. An emerging literature addresses climate 

change influence on native bark beetle populations, which have evolved with 

forest tree hosts as natural disturbance agents (Jenkins and others 2008; Bentz and 

others 2010). Here, we briefly discuss the potential impacts of climatically caused 

shifts in the extent and frequency of forest insect outbreaks on uplands and streamriparian corridors. Although most research in the western United States has focused 
USDA Forest Service RMRS-GTR-290. 2012. 69

on native beetles in coniferous forests, it should be noted that similar trends are 

likely occurring with insect species that utilize and parasitize riparian hardwood 

species, particularly cottonwoods and willows (Kendall and others 1996).

Warming temperatures are predicted to dramatically affect insect outbreaks in 

forested areas (Bale and others 2002) by increasing water stress on the host trees 

while conferring physiological advantages to the insects. The cumulative effect of 

forest harvest patterns, fire suppression, and climate change, especially drought 

and mild winters, has already resulted in large, contiguous landscapes susceptible 

to bark beetle outbreaks (Jenkins and others 2008; Bentz and others 2010). Some 

forest types are dominated by fairly even-aged stands within the preferred size 

class range for native beetles; others contain a high percentage of old, large diameter, and low vigor host trees. Flexibility in the life-history strategies of some insect 

populations appears greater than previously anticipated and rapid genetic adaptation of insects to seasonal changes in temperature has already been documented. 

Warmer temperatures could disrupt climate controls on winter mortality, generation duration, and developmental and emergence timing of insects, thus increasing 

survival and the probability of population success. As temperatures rise, the area 

suitable for both adaptive seasonality and winter survival for insects is predicted 

to grow, thus expanding the potential range of some species as they move into new 

niches. Bark beetle outbreaks will vary regionally because of differences in feedbacks driving beetle populations and physiological differences among host tree 

species. Although a high degree of uncertainty and complexity exists, bark beetle 

outbreaks driven by climate change may shift some forest ecosystems beyond their 

natural boundaries of resilience.

Elevated temperatures are also associated with drought conditions that exacerbate tree stress. An important consequence of climate change is higher frequency 

and severity of droughts (Seager and others 2007), which will influence distribution of forest tree species and increase susceptibility to bark beetle attack. Using 

existing data for 130 North American tree species and associated climate information, McKenney and others (2007) predicted that the average range for a given tree 

species will decrease in size by 12% and will shift northward by 700 km during 

this century. Relative to current distributions, by 2060 the range of Engelmann 

spruce, a common riparian species and principal host for spruce beetle, is projected to decrease by 47% within the contiguous United States. Beetle outbreaks 

increase tree mortality rates and can result in subsequent replacement by other tree 

species and plant associations (Veblen and others 1991). Bark beetles are linked 

to their host trees and will undoubtedly influence the formation of new western 

North American forests, including riparian forests. Broad-scale tree migrations are 

predicted to occur this century. Riparian areas provide mesic refugia for some 

conifer species at the margins of their current distributions. As these distributional 

boundaries retreat and expand for Western conifer species, bark beetles may play 

a significant role in colonizing and killing stressed individuals at the margins. 

Characterization of thresholds regulating species distributions (insects and trees) 

may be an important component of forest management in a changing climate, both 

in uplands and along stream-riparian corridors.

Complex feedbacks relate to increased incidence and consequences of bark 

beetle outbreaks. Fire, an important forest disturbance that is directly influenced 

by climate change (Westerling and others 2006), can reduce the resistance of surviving trees to insect attack. Insect-caused canopy mortality alters the amount, 

composition, and arrangement of living and dead biomass in various fuel complexes. Currently, this is a major concern throughout portions of the western United 

States impacted by the recent mountain pine beetle epidemic. Relationships and 
70 USDA Forest Service RMRS-GTR-290. 2012.

consequences of the interactions between fire and beetle outbreaks are poorly 

understood, complex, and spatially and temporally dynamic (Jenkins and others 

2008). However, as fire and insect-caused mortality are transforming Western forests, addressing their interactions is necessary in the development and application 

of forest management strategies.

Beaver

Beaver (Castor canadensis) profoundly influence the short- and long-term 

composition, structure, and function of riparian environments throughout stream 

networks in mountainous regions of the western United States. As agents of natural disturbance, beaver both use riparian areas as habitat and alter the hydrology, 

geomorphology, biogeochemistry, and biota of the stream segments they occupy 

(Naiman and others 1998). The beaver is a considered a keystone riparian species 

due to its extensive influence on fluvial corridors (Pollock and others 1995). Prior 

to their large scale removal in the late 1800s, beaver occupied nearly all stream 

habitat types from the arctic to northern Mexico (Naiman and others 1998); their 

removal is considered a major disturbance in itself (Wohl 2001, 2006). In catchments where beaver are abundant, there may be 2 to 16 dams/km of stream length 

and each dam may retain between 2000 and 6500 m3 of sediment (Naiman and 

others 1998).

Beaver cut and utilize riparian woody species to build dams in first- to fourthorder streams and in side-channel and floodplains of larger rivers (Johnston and 

Naiman 1990). Dams are generally built on low gradient stream segments; however, where beaver population densities are high, dams may be built in steeper 

gradient portions (Collen and Gibson 2001). Dams retain water and sediments, 

forming ponds that inundate and frequently flood surrounding trees, altering upstream and downstream riparian environments, and creating wetland habitat. The 

cyclic pattern of pond creation and abandonment has produced a shifting mosaic 

of habitat patches and left a legacy on riparian plant community composition and 

distribution in many stream networks (Pollock and others 1995; Naiman and others 1998). Some abandoned ponds are rapidly recolonized by riparian plants and 

return to pre-ponded conditions in a few years to decades. Depending on topography, soil characteristics and other factors, other ponds may develop distinct and 

stable wetland or meadow features that persist for decades or centuries, enhancing 

species and habitat diversity.

The hydrologic effects of beaver dams and dam-building activities can extend 

well beyond the boundaries of the pond, both upstream and downstream within the 

fluvial corridor. Beaver dams alter the patterns of stream discharge by decreasing 

current velocity and enhancing the depth, extent, and duration of inundation associated with floods. They also elevate the water table during both high and low 

flows for stream segments upstream of dams. On the upper Colorado River, beaver 

dams caused water to move around them as surface runoff and subsurface seepage 

during both high- and low-flow periods, and dams attenuated water table decline 

in the drier summer months (Westbrook and others 2006). Beaver can influence 

hydrologic processes during both peak flow and low flow periods, thus creating 

and maintaining hydrologic regimes suitable for the formation and persistence of 

wetlands (Westbrook and others 2006). Geomorphic effects include the retention 

and redistribution of sediment and organic matter, flooding and erosion of streambanks, and expansion of the extent of flooded soils (Pollock and others 1995; 

Naiman and others 1998). Beaver also affect plant community composition and 

the spatial-temporal dynamics of the vegetation through selective herbivory and 
USDA Forest Service RMRS-GTR-290. 2012. 71

foraging practices (Pollock and others 1995). Beaver cut trees and shrubs to feed 

on bark, preferring trees with soft, brittle bark, including common riparian species 

such as aspen, willows, alder, maple and ash. Felled trees may be used in dam 

construction or they may be left in place, adding structural complexity to riparian zones (Barnes and Dibble 1988). Foraging by beaver significantly impacts the 

composition and successional dynamics of riparian vegetation, particularly where 

beaver densities are high (Barnes and Dibble 1988; Johnston and Naiman 1990).

Beaver activities and dams can have both positive and negative effects on fish 

community composition and habitat (Collen and Gibson 2001). Dams stabilize and 

warm stream temperature, increasing localized productivity but potentially having 

a negative effect on coldwater species. Spawning sites may be flooded and silted, 

upstream migration may be impeded, and habitat may be created for predators, 

with detrimental effects on desired fish species. However, the increased habitat 

complexity may provide refugia and cover. Because stream discharge is stabilized, 

channel scouring and bank erosion are decreased, organic matter and nutrients are 

retained, and invertebrate and fish production may increase (Collen and Gibson 

2001). Reintroduction of beaver has been suggested as a possible “adaptation action” to climate change that may improve watershed resilience (Furniss and others 

2010). However, the practicality and benefits of introducing or restoring beaver 

populations will vary according to location and should be considered in conjunction with a management plan to control their densities.

Human Disturbance of Riparian Areas: Land Use and Management

Natural disturbances and processes have influenced the development and current 

condition of riparian and stream habitats as briefly described above (McAllister 

2008). Along many stream and river segments, however, the effects of past and 

present human disturbance may be more pervasive than natural processes. Human 

effects can be broadly considered with respect to five categories: flow regulation/

alteration, water pollution, channel alteration, decreased biotic integrity, and land 

use (Wohl 2006). Direct human impacts on stream-riparian corridors result from 

activities conducted within the stream channel itself that alter channel geometry, 

the dynamics of water and sediment movement, or aquatic and riparian communities. Examples include construction of dams or diversions, channelization, removal 

of beavers, and placer mining (Wohl 2006). Less direct human impacts result from 

activities within the watershed that alter the movement of water, sediment, large 

wood and nutrients, or introduce contaminants into the channel. Examples include 

road-building, forest harvest, urbanization, agricultural cropping, and grazing. 

Human impacts frequently interact or lead to changes in the timing, frequency, or 

magnitude of natural disturbances. For example, activities such as forest harvest 

and road building can accelerate the frequency and volume of debris slides and 

hillslope sediment loss; and grazing can increase erosion due to changes in bank 

stability. Several extensive reviews have described the impacts of human disturbance and land use on streams, rivers, and riparian areas (Patten 1998; Wohl 2001; 

Brinson and others 2002; Naiman and others 2005; Wohl 2006).

In the context of climate change, increased alteration of streamflow is a critical 

human disturbance affecting the ecological integrity of many aquatic and riparian ecosystems (Furniss and others 2010). Alteration in stream flow, including the 

amount, timing, and duration of flow, all contribute to changes in the geomorphology, physical processes, ecological condition, and biological characteristics of 

the stream channel and associated riparian habitat (Poff and others 1997; Merritt 

and others 2009). Land use practices, such as agriculture and urbanization, have 
72 USDA Forest Service RMRS-GTR-290. 2012.

added to the disruption of natural hydrologic regimes within stream networks (Poff 

and others 1997). As noted previously, reconciling the increasing human demand 

for water with the dependency of stream and riparian biota on natural flow attributes remains one of the most difficult challenges in the face of climate change 

predictions.

Land use and management has changed considerably with time and past practices 

have been discontinued or modified to mitigate environmental impacts. However, 

the legacy of human disturbance and land use continues and must be considered in 

current management strategies. For example, historical practices, such as removal 

of wood from rivers for navigation and fish passage, splash damming, tie drives, 

and clearing of riparian trees has resulted in simplification of stream channels and 

streambanks, reduction in the areal extent of riparian areas, and local decreases in 

amounts of instream large wood (Sedell and Froggart 1984; Young 1994). A legacy 

consequence of timber harvesting is the marked long term reduction in recruitment of large wood to streams in logged basins. Livestock grazing in the past has 

resulted in significant impacts to riparian vegetation and soils. Fire suppression in 

uplands and riparian areas has resulted in an increase of fuel loads within areas that 

typically experienced low-severity historical fire regimes (Ellis 2001; Dwire and 

Kauffman 2003). The legacy effects of past human disturbance influence current 

and future condition and potential of streamside areas; their continuing impacts 

must be considered when defining riparian management targets, planning restoration projects, and strategizing on climate change adaptation and mitigation actions.

Climate Change, Fire, and Riparian Values and Functions

Riparian areas are dynamic environments, influenced by strong disturbance 

regimes, and characterized by considerable habitat heterogeneity and multidimensional gradients. The range of riparian ecological processes, values, and functions 

depend on physical characteristics associated with location within the basin and 

stream network. The influence of climate change on riparian areas, with consequent shifts in precipitation, stream flow characteristics, and fire severity and 

frequency, also depends on the physical context of a given reach or stream-riparian 

segment. Because of their spatial position in watersheds, riparian areas integrate 

interactions between aquatic and terrestrial environments and can be sensitive to 

disturbance and management both upslope and upstream. In addition, the interactions and feedbacks among natural and human disturbances depend on location 

within a watershed, physical context, and land use legacy (Nakamura and others 

2000; Rood and others 2007; Rieman and others 2010).

The impacts of climate change will influence different stream-riparian ecosystems in different ways. In high elevation headwaters, stream segments in alpine 

and treeline environments will be affected by variability in annual snowpacks and 

higher temperatures through the growing season. In subalpine and montane forested riparian areas, riparian vegetation may be most affected by shifts in streamside 

microclimates. Riparian tree species composition is commonly similar to surrounding uplands, but with higher frequency of more mesic species, like Engelmann 

spruce, and greater understory diversity and productivity. Although influenced 

by streamflow and shallow subsurface drainage that may emerge near streams, 

many conifer-dominated riparian areas could be characterized as micro-climate 
USDA Forest Service RMRS-GTR-290. 2012. 73

Non-Native Invasive Plants in 

Riparian Areas of the Western 

United States

Riparian corridors are particularly susceptible to 

invasion by non-native plant species. The same 

factors and processes that foster high native 

species diversity along streams also facilitate 

dispersal, establishment, and persistence of 

undesirable species. These factors include: active 

natural disturbance regimes, resulting in frequent 

turnover and a range of habitats; high degree of 

connectivity that facilitates the movement and 

transport of plant propagules from upslope and 

upstream; and resource availability, particularly 

moisture and nutrients. Common management 

activities in riparian areas—including grazing, 

road construction, mining and recreation—

can accelerate the spread and persistence of 

invasive species. Many stream valleys are also 

transportation corridors, which frequently serve as 

conduits for the movement of nonnative species. 

The alteration of natural stream flow regimes 

through damming, diversion, and groundwater 

extraction has directly and indirectly contributed 

to the increase of non-native plant abundance 

along streams and rivers. This has been extensively 

reported for low gradient, low elevation streams 

and rivers in arid and semi-arid regions, where 

native cottonwood and willow species have 

been displaced by saltcedar, especially Tamarix 

ramosissima and T. chinensis, Russian olive and 

Siberian elm. Extensive streamside stands of 

saltcedar and other nonnative woody species can 

result in high streamside fuel loads, potentially 

influencing fire behavior and frequency. Shifts in 

precipitation, as regionally predicted by climate 

change models, will affect stream flow and hillslope 

hydrologic inputs to streams, likely reducing the 

areal extent of riparian areas over time, and 

conferring additional advantage to invasive woody 

species, particularly in low gradient floodplains.

Climate change predictions also include rising air 

temperatures, which influence the distribution of 

both native and invasive species and thus affect 

species assemblages. The distribution of saltcedar 

is thought to be limited by sensitivity to frost, and 

the frequency of its occurrence in the western 

United States has a strong positive relation with the 

mean annual minimum temperature (Figure T8-1). 

In contrast, Russian olive occurs most frequently at 

sites with mean annual minimum temperatures of 

-8.7 ̊C or lower, suggesting a chilling requirement 

for bud break or seed germination. Shifts in the 

ranges of both species—tamarix moving further 

north, Russian olive moving to higher elevations—

will continue to affect native riparian plant 

communities throughout arid portions of the west. 

Non-native herbaceous species have also invaded 

riparian areas; in many rangelands and forests, 

these are the same noxious species that occur in 

uplands, and include leafy spurge, nonnative brome 

grasses and thistles, knapweeds, toadflaxes, and 

knotweeds. Existing distributions of other riparian 

native and non-native species are also adjusting, 

depending on biotic and abiotic constraints, 

and will further complicate efforts to maintain 

functional, native riparian plant communities.

The basic principles of managing invasives apply in 

riparian areas as in uplands: minimize the potential 

for their dispersal and establishment; focus on 

prevention of spread; maintain resistant native 

ecosystems; and conduct incremental treatments 

within an adaptive management framework. 

However, control or eradication of invasive species 

is more difficult in riparian areas because use of 

mechanical and chemical treatments is restricted 

or not allowed along many streams. Prescribed 

fire is increasingly used to control invasives in 

riparian areas; use is likely to increase as more is 

learned about treatment effectiveness in different 

ecoregions. Diligent post-treatment and postfire 

management is critical, however, since certain 

invasive species (e.g. cheat grass and reed canary 

grass), have been observed to increase in riparian 

areas for several years following fire. Follow-up 

restoration activities may be required to reestablish 

native riparian plant communities in some 

locations. Interactions and feedbacks between fire, 

climate change variables, and response of invasive 

plant species are likely to increase in complexity, 

requiring expanded coordination for invasive 

species control at watershed and regional scales. Probability of Occurrence

Mean Annual Minimum Temperature (°C)

0.5

0

-40 -20 0

E. angustifolia

T. ramosissima

Figure T8-1. Gaussian logistic regression of occurrence 

of salt-cedar, Tamarix ramosissima, and Russian 

olive, Elaeagnus angustifolia, as a function of mean 

annual minimum temperature. Vegetation data were 

collected at 475 randomly selected stream gaging 

stations in 17 western states; temperatures were 

derived from weather station data recorded from 

1961-1990 (Friedman and others 2005).
74 USDA Forest Service RMRS-GTR-290. 2012.

dependent. They are frequently cooler and moister due to spatial position in steep 

watersheds, cold air drainage, and topographic shading. With increasing air temperatures, riparian microclimates may warm and coniferous streamside vegetation 

may become more similar to upland vegetation. During wildfires, these riparian 

areas may increasingly burn like surrounding uplands (Table 3).

Throughout the western United States, willow-dominated riparian areas occur 

in broad valley bottoms, including unconfined and glaciated valleys with low 

slopes (<3%) in montane and subalpine settings (Patten 1998; Rocchio 2006). 

While floods and streamflow are important regulators of willow ecosystems, 

other major drivers are beaver and shallow subsurface drainage that contributes 

to maintenance of high water tables (Gage and Cooper 2004). Groundwater recharge can originate from deep glacial till, hillslopes with highly fractured rock, 

and long, slow-draining hillslopes. Typically, at higher elevations, the magnitude 

of hillslope discharge is higher as a consequence of snowmelt runoff and precipitation events. The relative importance of streamflow and hillslope discharge for 

maintenance of willow ecosystems depends on elevation, geology, season, and 

other factors (Westbrook and others 2006; Wolf and others 2007; Westbrook and 

others 2011). Climate change will affect both streamflow and patterns of groundwater discharge and may result in the spatial contraction of willow ecosystems 

and local loss of species near limits of their distributional ranges. The dry-down 

of willow ecosystems may limit their ability to serve as fuel breaks during wildfires (Table 3).

Cottonwoods are keystone riparian species, dependent on flooding for recruitment and stand replacement, and dependent on streamflow for stand maintenance. 

Streamflow-ecology relationships have been described for different cottonwood 

species, geomorphic settings, and regions in western North America, mostly in 

response to dams and other flow alteration (Rood and Mahoney 1990; Stromberg 

and Patten 1991; Braatne and others 2007; Wilding and Poff 2008; Poff and 

others 2009; Merritt and Poff 2010). These relationships will also prove useful 

in predicting potential shifts in distribution and condition of cottonwood stands 

in response to altered streamflow due to climate change. As noted above, a key 

challenge in securing sustainability of cottonwood and other riparian ecosystems 

is developing a framework that incorporates predicted changes in streamflow 

characteristics and guides the development of environmental flow standards for 

regional planning (Poff and others 2009).

C. Fish, Fire, Forest Management, and Climate Change

Forest streams provide some of the coldest and cleanest waters, and very high 

ecological, recreational, and intrinsic values are placed on the trout and salmon 

species that require the high quality water. Many of these fishes are now listed as 

sensitive, threatened, or endangered (Rieman and others 2003b), and sometimes 

their presence is cited as a protected value to support fire suppression or fuel abatement projects. Fire does indeed represent a challenge, as do land management and 

climate change, but fire has also played an important role in the development of 

the fish communities in the western United States. Understanding the ecological 

dynamics in relationships between fish and fire is an essential step in successfully 

managing forests and streams in a changing climate. This chapter provides a basic 

overview of the topics necessary to understand the dynamics and provide the logic 

supporting the continuing synthesis. The reader is directed to the compendium of 

“Advanced Topics on Fish Populations and Fire” at the back of this publication 
USDA Forest Service RMRS-GTR-290. 2012. 75

for a richer and more thorough review of the general topic and several important 

aspects. Following is a list of the articles:

Advanced Topics on Fish Populations and Fire

• Fire and Fish: A Synthesis of Observation and Experience

By Bruce Rieman, Robert Gresswell, and John Rinne

• Genetic variation reveals influence of landscape connectivity on population 

dynamics and resiliency of western trout in disturbance-prone habitats

By Helen M. Neville, R.E. Gresswell, and J.B. Dunham

• Fish life histories, wildfire, and resilience—a case study of rainbow trout in 

the Boise River, Idaho

By Amanda E. Rosenberger, Jason B. Dunham, and Helen Neville

• Aquatic species invasions in the context of fire and climate change

By Michael K. Young

The relationship of fish to fire is complex, like it is for other ecological systems. 

Aquatic communities in the West evolved with fire, along with the forests. Similar 

to effects on forests, there may be severe short-term negative consequences of fire 

for the individuals or local populations, but those may be coupled to long-term 

benefits for habitat complexity, quality, and productivity. Disturbance and recovery are key processes in many ecosystems (Pickett and White 1985), providing 

analogs for learning. A chief lesson is that in many circumstances, the full range of 

dynamics from mild to severe performs important functions in ecosystem renewal 

and cycling. Severe fire, and severe mass wasting erosional consequences following fire, often play important, positive ecological roles in aquatic ecosystems.

Fisheries and land managers often have natural protective instincts relative to 

headwater streams that have our coldest and cleanest water and our rarest fishes. 

Decades of work have demonstrated how forest management, which is conceptually less intrusive than severe wildfire, has contributed to declines in habitat quality 

and populations of salmonids across the West (Salo and Cundy 1987). Even within 

the context of recent harvest practices where the focus is on ecosystem integrity 

more than timber volume, there are potential consequences of both interfering with 

recovery dynamics and attempting to soften the role of fire.

Reconciling these different views of aquatic systems and fishes, fragile yet 

also tough, can be enlightened by considering the ecological processes protecting 

them. Millions of years of coevolution between forests and fish have developed 

some measure of ecological stability through resilience, a capacity to recover 

from not just population reductions, but major habitat altering events. Salmonid 

populations have substantial resilience to disturbance from fire or flood through 

expression of diverse life histories (Dunham and others 2003) that may include 

variation in patterns or extent of migration and timing of critical events in the life 

cycle. Metapopulation dynamics, wherein populations at one locale are supported 

by individuals dispersing from others (Levins 1969), provide an additional degree 

of buffering for aquatic populations as they do for many other wildlife species 

(Rieman and Dunham 2000). Seeking to engineer stability by resisting or controlling disturbance has inadvertently undermined some of the resilience, particularly 

through activities that fragment or isolate habitats from one another (Fausch and 
76 USDA Forest Service RMRS-GTR-290. 2012.

others 2009). A key difference in outcomes is that there are many opportunities 

over time for resilience processes to succeed, and a single failure in time does not 

consign a given population to permanent oblivion, whereas resistance dependent 

measures may have greater sensitivity to individual failures.

Even as we come to terms with this interesting dichotomy and begin to articulate 

more robust strategies for dynamic management, we are faced with the implications of a changing climate on aquatic systems. The nonstationary behavior of 

streams under future climates (Barnett and others 2008) with trends in streamflow 

timing (Stewart and others 2005), increasing variability in streamflows (Pagano 

and Garen 2005; Luce and Holden 2009), and warming streams (Isaak and others 

2010) challenges most notions of stability, dynamic or otherwise. The ecological 

response of fish populations to the stresses imposed by fire, land management, and 

climate change can help us see how they may interact in the future to affect fishes. 

This understanding can form a foundation for management response to climate 

change.

The Response of Fish Populations to Fire

The immediate and short term effects of fire are commonly harmful to individual fish and even local populations, but the intensity of the effect varies. Direct 

heating of water by fire and dissolution of ammonia and other chemicals from 

smoke has resulted in fish kills (Minshall and others 1989; Earl and Blinn 2003; 

Spencer and others 2003), but fish also appear to simply avoid affected areas if 

refugia are available (Rieman and Clayton 1997). Introductions of toxic material 

and ash flows shortly after fire have resulted in local extirpations (Rinne and Neary 

1996; Rinne 2003; Rinne and Carter 2008). Anecdotal observations of lower concentration introductions of ash and sediment have shown little immediate change 

in other circumstances (Sawtooth National Forest 2007). Major debris flow events 

in steep channels (e.g., Cannon and Reneau 2000; Miller and others 2003) almost 

certainly remove the fish that are present at the time.

The response of fish populations to these impacts are varied (see Rieman and 

others, this volume for additional detail). Some extirpations are permanent (e.g., 

Rinne 2003), while some locations see reestablishment of fish populations within 

a relatively short time (e.g., Jakober 2002; Howell 2006). Sublethal temperature 

increases after canopy removal have been observed to alter the growth and maturation of fish (Dunham and others 2007). Although fine sediment increases are 

documented to interfere with life stages that use gravel interstices (Everest and 

others 1987; Chapman 1988; Thurow and King 1991), the brief period of vulnerability to surface erosion post fire (Shakesby and Doerr 2006) and the large 

transport capacity of rivers to rapidly remove fine sediment (Lisle and others 2001; 

Burton 2005) seem to make fine sediments less important than other factors for fish 

post-fire status.

The long-term benefits of fire effects have been noted as well. The renewal of 

spawning gravels is cited (e.g., Reeves and others 1995; Benda and others 1998) 

(Figure 26). Inputs of nutrients released by fires may also provide at least a temporary boost in productivity (Spencer and others 2003; Malison and Baxter 2010). 

Fires are one of many disturbances that regulate sunlight coming to streams; so 

they contribute to maintaining a diversity of invertebrates that use both algae growing in streams as well as detritus falling from riparian forests (Minshall 2003). The 

legacy left behind by fire, including both the renewed material availability and the 

presence of fish to use those materials, is important in the net benefit of a fire.
USDA Forest Service RMRS-GTR-290. 2012. 77

Life history diversity, which considers the variation in life cycle stages, timing, 

and patterns of behaviors, is an important source of resilience in fish populations 

(Hilborn and others 2003; Moore and others 2010; Schindler and others 2010). 

Commonly identified life history categories include the range and extent of migration (from purely resident to anadromous) along with tremendous variation in the 

timing of life history events. In general, migratory fishes that move to large rivers, 

lakes, or the ocean grow to larger sizes than fish that do not migrate, commonly 

returning to natal streams more fecund than resident fish. Besides being able to 

contribute to increased population growth rates, these fish also spend less time in 

the smaller steep tributaries and are less likely to be directly impacted by violent 

post-fire impacts (Rieman and Clayton 1997). Fish that migrate away from natal 

streams have their own set of hazards to navigate, but in general, populations with 

Life History Diversity and Resilience

Life history variation or diversity, is a key concept in the biology of native fishes. Considerable interest has 

focused on notions of “resilience” (Dunham and others 2003; Bisson and others 2009; Bottom and others 

2011) and “biocomplexity” (Hilborn and others 2003; Schindler and others 2010) that follow from diversity 

as foundations for conservation management and habitat restoration (ISAB 2011). There are three key 

points linking these ideas. First, variation in life histories within and among populations contributes 

to increased productivity, persistence in variable environments, and the capacity of populations to 

recover from major disruptions (Dunham and others 2003; ISAB 2011). Resilience is the maintenance 

of these functions in the face of disturbance (Bottom and others 2011). Second, the expression of any 

potential variation depends on the diversity, connectivity among, and access to, the full range of habitats 

and environmental gradients available throughout the life cycle (e.g., Healey and Prince 1995). And 

third, it also depends on genetic diversity underlying the potential range of life history expression and 

its continuing evolution. In salmonids, life history variation commonly has been classified by distinct 

migratory forms. Interior salmonid life histories are often characterized, for example, as individuals that 

remain in natal waters their entire life (resident forms), migrate as juveniles from natal streams to larger 

rivers (fluvial forms) or to lakes or reservoirs (adfluvial forms), to mature before returning to spawn. 

Rieman and Clayton (1997) argued that the existence of these distinct life histories was key to the recovery 

and persistence of bull trout populations that experienced catastrophic disturbance associated with 

large wildfires. But, life history variation can be far more than several distinct migratory or non-migratory 

patterns. Salmonid populations, in particular, may show strong variation in egg size, incubation time, and 

size and timing of alevin emergence (McGrath and others 2008), subsequent patterns of foraging and 

growth (Nielsen 1992), size and age at juvenile migration (Downs and others 2006), migratory route and 

time spent in transitional habitats (Burke 2004), size and age of maturity, seasonal timing and distribution 

of adult migration and spawning (Doctor and others 2010; Crozier and others 2011), and for species that 

do not die after spawning, the frequency of migration and spawning events. The result can be a very 

large number of possible combinations contributing to the potential diversity of life histories within and 

among populations and even within family groups. The diversity that is actually expressed results through 

the interactions of individuals and populations with each other and the available habitats and streams, 

through phenotypic plasticity and evolution. These environments and interactions are embedded within 

larger watersheds and landscapes that are themselves shifting, reorganizing, and evolving through time 

and space. Together these represent complex systems or ‘biocomplexity’ that is arguably key to resilience 

in the larger perspective (Hilborn and others 2003; Schindler and others 2010). In this view, life history 

variation is both a product of, and a necessary condition for, persistence in a dynamic and changing world.
78 USDA Forest Service RMRS-GTR-290. 2012.

a diversity of strategies will have greater resilience with respect to a range of disturbance events.

Dispersal of fish from populations in nearby unburned streams or reaches 

is another important mechanism for refounding and supporting populations in 

burned streams that contributes to resilience (e.g., Rieman and Clayton 1997; 

Howell 2006). The number of proximate or interconnected habitat “patches” can 

be a useful indicator of this form of resilience (MacArthur and Wilson 1963; 

Dunham and Rieman 1999; Hilderbrand and Kershner 2000). Interconnected 

habitats and large patches are more resilient because they are less likely to experience synchronous disturbance from debris flows or other events and may 

contain some larger, more productive streams that can support less productive 

headwater streams. Debris flows transit through smaller streams, where their 

passage can be destructive to biota and habitats, but they usually deposit upon 

entering large streams, where the new material contributes to habitat complexity. 

High complexity of the stream network, e.g., having multiple branching tributaries as compared to a single threaded configuration, can also add to the robustness 

of the patch (Gresswell and others 2006). Patch size, complexity, connectivity, 

and the presence of multiple life histories also combine to produce populations 

that are less likely to suffer from small population size effects on genetics (see 

Neville and others in Advanced Topics of this volume).

The amount of habitat needed to ensure population persistence (in light of 

disturbance and environmental variation) is not precisely known, although available lines of evidence suggest something in the range of 20-40 km of suitable 

stream length for bull trout and less for other species (Rieman and McIntyre 

1995; Dunham and Rieman 1999; Dunham and others 2003; Peterson and others 

2008; Fausch and others 2009; Cook and others 2010). Observations of the size 

of occupied versus unoccupied patches that were suitable for bull trout (based 

on temperature) showed an increasing probability of presence leveling off above 

100 km2 (Dunham and Rieman 1999), which equates to roughly 40 km of suitable habitat (Isaak and others 2010). A graph comparing results of the bull trout 

presence/absence data and the post-fire debris flow data (1997 photos) gives a 

sense of the relationship between the two (Figure 27).

Plotting genetic difference as a function of distance between populations often 

results in a positive relationship. An analysis of genetic relationships in the Boise 

Figure 26: Salmon redds on recent 

debris flow deposit in the Middle 

Fork Salmon River in a location 

where spawning did not occur 

previously because of a lack of 

suitable substrate (photo courtesy 

of Russ Thurow).
USDA Forest Service RMRS-GTR-290. 2012. 79

River supported the notion that gene flow was much stronger at shorter distances 

especially under 20 km (Whiteley and others 2006). Debris flow mapping in 

the Boise River basin also shows an increasing synchrony of severely scoured 

reaches at scales less than 20 km (see textbox on debris flow scaling) in spite of 

much larger fire extents. These observations support speculation that the current 

structure and resilience of populations may emerge through the patterns of disturbance and recovery of the past.

Interaction With Land Management

If we draw from the general concept that the structure and resilience of current 

populations of fish reflect millennia of natural disturbances, including wildfires, 

but note that most local extirpations and declines have happened since the introduction of land management, including wildfire suppression, the question that 

arises is, “What is the difference?” This framing may place too much at the foot 

of land management, when issues like invasive species, introduced diseases, water 

diversion and management, and climate change also contribute to specific incidents. Nonetheless, this framing opens up a discussion of the contrasts between 

what superficially would appear to be impacts of a similar nature.

Land management comprises multiple activities that interact with streams in 

ways that are unique from fire. While forestry practices on individual stands, or 

even watersheds and landscapes, are conceptually less severe with a focus on reduced impacts to soils and dispersed impacts in space, roads are unprecedented 

components of managed landscapes, which have consequences disproportionate 

to their area (Luce and Wemple 2001). Early forest management on both private 

and public lands were designed to preserve soil resources both to protect site productivity and water resources (Hays 1969; Pyne 2002). Guidelines now seek to 

disperse canopy removal impacts over broad areas and avoid practices that result 

in soil degradation, even at site scales. While the positive effects of periodic mass 

wasting inputs were noted earlier, it is not clear that the advent of cautious harvest 

practices nor the reduction in burned acreages from fire suppression have substantially altered the long-term stochastic properties of these events. Rather it would 

appear that historical declines in aquatic species status might be more tied to the 

expansion of the road network (Lee and others 1997; Baxter and others 1999; 

Trombulak and Frissell 2000). Probability

0 5 10 15 20 25 30 35 40

0.0

0.2

0.4

0.6

0.8

1.0

Stream length (km)

0

0.0

0.2

0.4

0.6

0.8

1.0 Bull trout

occurrence 

Channel

reorganization

Figure 27: Probability of bull trout presence within a 

patch based on total stream length in the habitat 

(based on Dunham and Rieman 1999) versus 

the probability that channel segments within a 

given distance of stream length might experience 

debris flows at the same time. Neither study 

directly measured these quantities, but this is an 

interpretation of how they might be compared.
80 USDA Forest Service RMRS-GTR-290. 2012.

Roads have numerous physical effects on fish habitats including: habitat fragmentation, chronic fine sediment introduction, more frequent sediment from mass 

wasting, and channel constraint. Of these, fragmentation may be particularly 

important with respect to resilience to fire disturbances. Most obviously, fragmentation prevents migratory fish from re-founding or supporting a severely depressed 

population. If fragmentation prevents the expression of migratory life histories, removing culvert barriers post fire may be substantially less effective. Fragmentation 

by roads may lead to reduced genetic diversity, leaving populations less well prepared for shifts in conditions that could occur post fire even without the more 

catastrophic population resets associated with debris flows (see Neville and others in Advanced Topics of this volume). Barriers constructed to protect against 

non-native fish invasions can have similar consequences to road fragmentation 

(Peterson and others 2008; Fausch and others 2009).

Chronic fine sediment from roads reduces habitat productivity and survival of 

embryos and juveniles (Chapman 1988). Overall, this effect can restrict population growth rates, reducing resilience to individual events. Individual mass wasting 

events from roads are similar in nature to other mass wasting events. The risk of 

mass wasting from roads is highest in the initial decade after construction and declines over time, unless road maintenance stops, which can dramatically increase 

the risk. In basins where harvest was done carefully and incrementally, the serial 

construction of new roads may have generated essentially a chronic mass wasting 

scenario (see, e.g., Colombaroli and Gavin 2010)

Some contrasts of land management to fire have focused on sediment yield from 

harvested areas (e.g., Istanbulluoglu and others 2004; O’Laughlin 2005; Roloff 

and others 2005). In low gradient areas, a series of careful harvests with soil protection can produce less sediment than a single severe fire (O’Laughlin 2005). In 

steeper areas, long term sediment yields are similar, but the event sizes tend to 

be different (Istanbulluoglu and others 2004). The understanding provided above 

regarding aquatic ecology resilience to disturbance, however, suggests that differences in long term sediment totals may not be a useful decision variable (Luce and 

others 2005; Luce and Rieman 2010). Episodicity clearly has a direct influence 

on the consequences for fish as does the spatial distribution of synchronous major 

disturbances. The sediment yield studies have been faulted for failing to include 

road erosion impacts (Rhodes 2005). The addition of road erosion would likely 

contribute little to long term sediment yields (Goode and others 2011), however, 

considering the chronic additions would provide interesting and biologically relevant contrast.

Land management also includes fire prevention and suppression. Fire suppression practices certainly have the capacity to increase sediment loading, but they 

are likely minor additions compared to inputs from large severe fires. Some fire 

retardants are toxic to fish, and others impose a chemical oxygen demand on the 

water; as a consequence, fire retardant application near streams can be hazardous 

to aquatic systems (Little and Calfee 2002; Pilliod and others 2003; Giménez and 

others 2004). The introduction of diseases and non-native aquatic species from untreated pumping equipment also poses a hazard but can be managed with vigilance.

Changes in fuel loads caused by fire suppression over much of the 20th century 

are commonly discussed as an emergent risk for aquatic systems (Bisson and others 2003; Hessburg and Agee 2003; Rieman and others 2010). It is not obvious 

that such changes have led directly to extirpations historically, but the increased 

continuity of fuels and flammability in some forest types pose an increased risk of 

larger and more continuous fire in those locations, which directly relate to strategies fish have adapted to cope with fire. In explorations of the potential for forest 
USDA Forest Service RMRS-GTR-290. 2012. 81

restoration to reduce risks in the South Fork Boise River, some intermediate sized 

basins showed increased persistence probabilities from reducing fuel continuity 

(Dare and others 2009). Many of the places with strongly altered fuels or fire regimes were directly affected by forest harvest and attendant road construction, thus 

they seldom coincide with current habitats of sensitive, threatened, or endangered 

species (Rieman and others 2000). Some of the habitats are suitable, however, so 

these places may represent opportunities for joint restoration of forest and aquatic 

habitats to more natural fire regimes (Rieman and others 2010).

Discussions of fuel change issues and fish have focused largely on forests (e.g., 

Bisson and others 2003), perhaps because of debate over forest management and 

restoration policy (e.g., DellaSala and Frost 2001). Fuel changes (and consequent 

fire regime changes) caused by shifts in range species, particular the replacement 

of sagebrush (Artemesia tridentata) communities by cheatgrass (Bromus tectorum) 

and other non-native brome grasses has the potential to affect many aquatic communities in the West as well. Invasive riparian species have major implications 

for streamside fuel structures too (see textbox on invasive species in the riparian section). Land management, particularly road management, fire suppression 

management, and post-fire restoration practices, have strong influences on the introduction and spread of non-native plants, which may play out as a long-term risk 

issue for aquatic systems because of their close coupling to the terrestrial ecology.

Interactions With a Changing Climate

Changes in climate described earlier will influence fish most directly through 

stream temperature increases and changes in flow regimes (Rieman and Isaak 

2010; Wenger and others 2011b). Most trout and salmon are adapted to relatively 

cold water and typically use some of the higher elevation waters in basins where 

they are present. This means that there may be limited ability for some populations 

to shift to higher elevation streams. While temperature increases will place additional stress on populations in stream reaches where temperatures are warmer than 

optimal, there are some exceptionally cold streams at high elevations where productivity may increase with warming (Harig and others 2000; Coleman and Fausch 

2007). Where warming air temperatures increase the probability of flood events in 

winter, fall spawning species like bull trout, brook trout, or the five Pacific salmon 

species with eggs incubating during the winter may be at greater risk. Climate 

change may cause additional indirect effects to populations through changes in 

wildfire size, frequency, and severity and alterations to riparian ecosystems (please 

refer to the sections on wildfire ecology and riparian ecology preceding this section 

for more background). Increased wildfire presence in the landscape could contribute to keeping riparian canopy less dense and stream temperatures warmer.

The combination of temperature and streamflow changes will reduce the size of 

headwater patches of the species adapted to the coldest temperatures (Figure 28). 

Patches may also effectively shrink from above in locations where streamflows are 

declining and streams become too small, and in places where increased rain-onsnow inputs are driving more frequent mass wasting in steep headwater channels. 

Similar changes may also reduce connectivity within and among habitat patches, 

with barriers being imposed by reductions in low streamflows (Luce and Holden 

2009; Leppi and others 2011) or high temperatures.

Different species have different sensitivity to changes in temperature and flow 

regime, and habitat suitability models can inform expectations for shifts in fish distributions related to climate change (Rieman and others 2007; Wenger and others 

2011a; Wenger and others 2011b) (Figures 29, 30). Contrasting and complementary 
82 USDA Forest Service RMRS-GTR-290. 2012.

Figure 28: Decline in patch 

size and consequent 

probability of being 

occupied starting from 

(a) current conditions 

with (b) decreases due 

to warming temperatures 

[red areas] and (c) low 

flow decreases [dashed 

lines] and debris flow risks 

[yellow lines].

Figure 29: Comparative sensitivity 

of four trout species to stream 

temperature A and the frequency 

of winter high flows B. Green = 

cuthroat trout; blue = brook trout; 

red = rainbow trout; brown = 

brown trout. (from Wenger and 

others 2011b)

0.0

A

10 15 20 25

Occurrence probability

0.2 0.4 0.6 0.8

Air temperature (C)

0.0

B

0 1 2 3

0.2 0.4 0.6 0.8

Winter high flow frequency

4
USDA Forest Service RMRS-GTR-290. 2012. 83

Figure 30: Modeled trout distributions under present and future (A1B) climate change scenarios. Gray streams are relatively 

unsuitable and black streams are relatively suitable (from Wenger and others 2011b)
84 USDA Forest Service RMRS-GTR-290. 2012.

Figure 30: Continued.
USDA Forest Service RMRS-GTR-290. 2012. 85

effects of different processes on different species creates a complex set of potential 

responses. For example, changes in winter flood frequency may be less important 

for bull trout if temperature excludes them from habitats where flood frequency is 

increasing; and spring-spawning cutthroat that would otherwise be negatively influenced by temperature changes may actually benefit from increased winter floods 

that reduce competition with fall-spawning brook trout.

Hydrologic changes will likely precipitate changes in water management 

(Barnett and others 2008), which in turn will have consequences for many aquatic 

systems. In the Southwest, where water withdrawals are a common challenge for 

fish (Rieman and others 2003b), projections for decreased flows and increased demands will likely exacerbate current constraints. Summer flow declines projected 

for other locations may have similar consequences where water withdrawals are a 

substantial portion of summer streamflow now. Shifts in snow accumulation and 

melt have been cited as reasons for proposing new reservoirs, expansion of existing 

reservoirs, and altering management of others (Goode and others 2011). Changes 

in reservoir management have potential to affect migratory life histories using the 

reservoirs as well as those populations using cold tailwaters below the reservoirs. 

Because migratory life histories using reservoirs can have such a dramatic influence on post-event recovery for populations affected by fires and debris flows, the 

implications of climate change for reservoir management may be important for 

fish in many locations in the western United States.

Connections to Conservation

The nexus of the four primary stressors for fish—fire, land management, nonnative invaders, and climate change—poses a difficult challenge. Land and aquatic 

managers have critical questions about what to do, and where to do it, and a primary concern is often the conservation of native fish populations. The next section of 

this report discusses the complexity of integrating the joint conservation concerns 

of terrestrial and aquatic systems together, but it is worthwhile here to reiterate 

important concepts related to conserving trout populations. Because the resilience 

of fish populations to fires is strongly influenced by diversity in life histories, the 

extent of habitat networks necessary to support persistent populations serves to 

focus the discussion.

Activities that increase connectivity among suitable habitat patches and existing 

populations and building or rebuilding local populations in and around large patches 

are likely to increase robustness to disturbances and species persistence probabilities. In some contexts, however, concerns may exist about increasing connectivity 

for invasive species as well, and such decisions may need to consider multiple local conditions (Peterson and others 2008; Fausch and others 2009). Under certain 

circumstances, climate change may reduce threats of invasion by some species 

(Wenger and others 2011a; Wenger and others 2011b). The most common specific 

activity for increasing habitat size and connectivity is replacement and removal 

of road culverts that act as barriers to fish movement. Restoration of local habitat 

quality to mitigate stresses that constrain population productivity can also encourage more migratory behaviors as well, because productivity of natal habitats is 

important to migratory life histories (Lucas and Baras 2001).

Management options for ameliorating warmer stream temperatures include 

maintaining or restoring instream flows or reservoir design and operation, especially on larger reservoirs that stratify during the summer (Neumann and others 

2006; Olden and Naiman 2009). Temperature management through canopy manipulation is generally not a reasonable approach for affecting significant portions of 
86 USDA Forest Service RMRS-GTR-290. 2012.

a stream network, particularly larger streams, and is ultimately vulnerable to wildfire. Exceptions might occur in short sections of streams flowing through meadows 

if these areas have been significantly degraded by livestock grazing and riparian 

vegetation and bank structure are substantially altered.

Priority areas for activities that improve resilience to events may not be the same 

places as priorities for conservation of rare aquatic species. For example, natural 

areas, like designated Wilderness areas, commonly serve to house “stronghold” or 

“core” populations of trout, and maintaining representative core populations is a 

key part of a conservation strategy. Improvements to long-term species outlooks 

will also be achieved through restoration of areas that can support or be supported 

by the core populations; so most opportunities for improvement in persistence will 

likely be in areas partially degraded by historical land management. Priority areas for investment of restoration funding would be places that are most robust to 

climatic changes (e.g., those that could be made large for the future as well as the 

present). Places that are large might have a lower priority than places of intermediate size that can be expanded, because greater gains in net persistence would come 

from making a questionable patch into a core area. While there is an inclination to 

grow large patches to greater size, there are also benefits to having multiple large 

patches with somewhat greater geographical separation. A key issue, therefore, 

is developing more precise definitions of patch sizes needed for persistence by 

different species and understanding how alteration of disturbance regimes from 

interactions of fire and climate may affect these patch sizes in the future.

Over time there will also be discussion about the values that we want to conserve (e.g., Rieman and others 2010). Values range from just having some fish 

in a creek to representation of rare genotypes or phenotypes that might represent 

important evolutionary legacies. Between, there are values associated with economics (having enough of the right kinds of fish to attract fishing) and ecological 

functions. To some extent the range in value connects to the potential to substitute 

other fishes or other processes for maintaining fish populations (e.g., hatcheries). 

Ultimately, however, retaining genetic and phenotypic or life history diversity will 

be a particularly important conservation goal related to the changing climate because it offers the primary base from which evolutionary adaptation can take place 

(Haak and others 2010; Williams and others 2011), and see Rosenberger and others 

in Advanced Topics of this volume.
USDA Forest Service RMRS-GTR-290. 2012. 87

Part III: Management Actions and Decisions

Managers of forest, wildlife, and water resources have heard the call for increasing resilience of ecosystems (Walker and Salt 2006; Millar and others 2007; Baron 

and others 2009; Heller and Zavaleta 2009; Joyce and others 2009; Keane and 

others 2009; Palmer and others 2009; West and others 2009; Turner 2010). The primary question managers have is what to do about it. The question has three distinct 

aspects: 1) understanding the actions that could be taken, 2) how to decide which 

actions to apply, and 3) when and where to apply them. The general set of actions 

available around fire effects on aquatic systems has a fairly limited scope, although 

there are many variations on the key themes. While there is limited information 

on the effectiveness of many actions, there is an understanding of the mechanisms 

by which they affect risks. Making decisions about solutions has been difficult, 

but advances in understanding of aquatic ecosystem response to fire supports new 

ideas in framing key decisions (Bisson and others 2003; Rieman and others 2010).

A. Actions

Action Choices

The choices available to managers to reduce risks associated with wildfire are 

somewhat limited in their general nature, although there may be many variations 

in details available for each to increase local suitability. The general classes are:

1. Fuel treatment

2. Aquatic and/or riparian habitat restoration

3. Hillslope restoration

4. Fire suppression

5. Post-fire stabilization

Fuel treatments include a large range of potential activities ranging from carefully managed mechanical removal of specific fuels, to various levels of directly 

applied fire, to wildfire itself (see the section on Forests, Climate Change, and 

Fire earlier for more discussion on fuel treatments). The most common goal of 

fuel treatment programs is to reduce fuels, particularly near human infrastructure, 

to alter fire behavior and intensity, aid fire suppression, and reduce burn severity 

(Graham and others 2004). A more general ecological goal may be described as 

keeping the fuel levels consistent with the type of ecosystem: for example maintaining few ground and ladder fuels in a ponderosa pine forest. One objective of 

managing fuels is to reduce burn severity of future fires and consequent effects on 

vegetation and soils, with potential reductions in risk to aquatic ecosystems from 

thermal and sediment impacts (Hessburg and Agee 2003). Fuel treatments are also 

used for terrestrial habitat restoration and control of invasive plant species, particularly in riparian areas (Stone and others 2010).

Aquatic habitat restoration is directed at increasing diversity and complexity 

of aquatic habitats, which generally supports more productive and resilient populations of fishes. Examples of activities include adding large woody material 

to form pools and provide cover, reconstructing meanders to deepen pools and 

provide more hydraulic variability, riparian planting to improve shade, or road 
88 USDA Forest Service RMRS-GTR-290. 2012.

culvert replacement to allow migratory fish passage. The specific activities are 

quite diverse but share the general approach of reducing risks to aquatic systems 

by improving some aspect of the in-stream and near-stream habitat. As might be 

discerned from the section on aquatic ecosystems, restoration of connectivity may 

be critical for many populations, although the risk from invasive species must be 

considered.

Hillslope restoration consists of a substantially more restricted tool set. In most 

forest ecosystems, it refers to road repair, upgrade, or decommissioning with the 

goal of reducing erosion and mass wasting. In a broader sense there are related activities applied to mined, overgrazed, or logged sites. We distinguish these pre-fire 

activities from post-fire stabilization (see below), which is focused on preventing 

(resisting) losses to populations. Reducing impacts from sediment over the longer 

term before fire occurs helps to build populations and communities that are more 

diverse and productive and, thereby, more resilient to fire effects.

In the absence of strategic planning and implementation of other restoration and 

enhancement actions, fire suppression becomes the default activity for reducing 

risks to both forest and aquatic resources. Although there are conceptually shortterm reductions in risks associated with putting out the fire, the continued presence 

of fuels could lead to a longer-term hazard. Thus the choice not to suppress a fire, 

within the context of a broader plan, can be seen as a fuel management strategy. 

Fire suppression has been very successful, but the few fires that escape initial attack can burn intensely and severely. Because conditions will eventually occur 

wherein fire suppression won’t be successful, suppression should be only one tool 

in the “tool box” that is a broader plan for fire management for long-term ecosystem resiliency.

Emergency post-fire stabilization practices are done after fires to “suppress” 

post fire erosion events. Depending on agencies involved, these measures may be 

referred to as Emergency Stabilization and Burned Area Rehabilitation (ESR) or 

Burned Area Emergency Response (BAER). Both programs separate the shortterm stabilization activities from the long-term rehabilitation activities; the latter 

generally have a goal of preserving ecological integrity of hillslope vegetation. The 

nature of stabilization activities is diverse, but the most commonly applied measures focus on restoring strength to the soil to keep soil particles in place. Runoff 

control strategies are mostly experimental. In general, emergency stabilization is 

authorized for protection of human life or property, although it can be applied for 

protection of special resources as well, including threatened or endangered aquatic 

fauna. As with fire suppression, it is done in a preventative fashion; so it does 

little to build resilience in the ecosystem, and aquatic ecologists note that periodic 

influxes of gravels and nutrients from erosion and mass wasting are important to 

aquatic habitats in the longer term (Reeves and others 1995). Most stabilization 

practices offer protection only for relatively common storms.

Effectiveness of Actions

It is worth discussing the understanding and uncertainties of the different choices to clarify expectations for different types of actions. The above listed actions 

have been evaluated to some degree, but generally for only the most proximal goal 

of the action—for example, changes in fuel loading for fuel treatments, improvement in number of pools for aquatic habitat restoration, decreased fragmentation 

for culvert replacements, and reductions in sediment from road decommissioning 

or post-fire treatments. The more distant, long-term goals of protecting threatened aquatic populations or protecting forest ecological processes are much more 
USDA Forest Service RMRS-GTR-290. 2012. 89

difficult to evaluate because of the large mix of influences on those outcomes. The 

evidence for these actions improving resilience is derived primarily from mechanistic logic. The complexity of interactions from multiple influences, however, has 

left questions about even something as seemingly direct as habitat improvement 

(Bash and Ryan 2002; Palmer and others 2005).

Many forest and stream ecosystems in the western United States contain species that have been present for at least a few million years and that have occupied 

more or less the same areas as they do now for the last several millennia despite 

historical fluctuations (Dunham and others 2003; Keane and others 2008). Such 

persistence in the face of natural dynamics supports an expectation that reliance on 

natural adaptations is a reliable conservation approach. Unfortunately the reality 

is that conditions now and expected in the near future depart substantially from 

those of the last few hundred thousand years. As noted earlier in this document, 

air temperatures are increasing in ways that are unprecedented, and temperature 

is important to both forest ecosystems and aquatic species. Although hydrologic 

cycles are changing with more uncertainty, they are changing nonetheless, and for 

the most part changes have been in a direction less favorable to present ecosystems. In addition to climate changes, there are many changes in the last century 

from invasive terrestrial and aquatic species, changing fire regimes, and displacing 

native species to less favorable habitats. On top of these pervasive changes, affecting wilderness and developed areas alike, the dams and other water management 

infrastructure have imposed significant, and usually irreversible, constraints on the 

naturally adapted system.

One simple classification of actions related to fire is between those that are done 

prior to fire to build resilience or resistance, and those that are done in response to a 

fire to prevent or reduce harm (Dunham and others 2003). The effectiveness of prefire solutions in building resilience depends in part on the potential for managed 

systems to operate within the limits of the adaptations. Changes in forest structure 

and climate that alter the spread and intensity of fires, changes in connectivity of 

habitats caused by infrastructure or invasive species, and changes in habitat quality affecting productivity all impose limits on a general strategy relying on natural 

resilience, meaning that solutions depending on this strategy need to address, or at 

least evaluate, multiple potentially limiting factors.

Actions focusing on responses to emergencies can be considered less reliable. 

Large wildfires, for example, are usually escaped suppression efforts. Similarly, 

most post-fire strategies include a combination of protection of human infrastructure and temporary evacuation of people to increase reliability of protection of 

human life and property. In some rare instances, fish have been evacuated for 

protection of a small population. If a fish population’s persistence depends on successfully suppressing fire or its effects in the short-term, it is necessarily at greater 

risk than a population that has the capacity to weather a fire event and then rebound 

post-fire.

Some assessment of pre-fire treatments has been done. Fuel management, for 

example, does not result in fully controlled or completely “tame” wildfires, nor is 

it 100% effective. It is important to recall that some severe fire effects are desirable for gravel, nutrient, and energy inputs to streams for long-term maintenance 

of aquatic systems. The implementation of aquatic habitat restoration is usually 

more directly controllable, e.g., the number of additional logs or constructed pool 

features is specified in a construction contract. What is less well understood about 

aquatic habitat restoration is whether productivity is actually increased, or whether 

fish move from poorer quality habitats nearby, with no net gain in production from 

a stream (Bash and Ryan 2002; Palmer and others 2005). Road decommissioning 
90 USDA Forest Service RMRS-GTR-290. 2012.

assessments generally show substantial improvement, but remediation of steep 

slopes is problematic, and it is difficult to entirely erase road impacts (Luce 1997; 

Madej 2001; Switalski and others 2004).

The effectiveness of post-fire stabilization and rehabilitation has received more 

scrutiny. Although the intention of the treatments is to protect human life and 

property by reducing the probability of severe erosion events, the reality is that 

only the smallest events are prevented or reduced, and larger events overwhelm 

treatments (Wagenbrenner and others 2006). While smaller erosion events can represent a threat to some isolated fish populations in small streams (Rinne 2003), 

many populations can rebound from events that do not completely displace them 

(Rieman and others 2003a). As a consequence, many post-fire stabilization treatments probably benefit homeowners more than fish. Post-fire removal of migration 

barriers is less beneficial than doing it beforehand, because pre-fire removal allows 

for development and dispersal of migratory life histories from the stream itself, 

which is a more reliable source of recolonization (Dunham and others 2003). Postdisturbance removal would at least make the stream available for reestablishment 

from dispersing fishes.

Hopefully, the reader takes away the message that no single approach will be 

adequate to guarantee fish persistence in any single location, much less across the 

diversity of situations in the West. A key concept is reliability of the approach, and 

reliability analysis can be a useful approach in thoughtfully and efficiently deciding which steps to take where.

B. Framing Issues and Decisions

The Broad Scope of Debate

A primary issue facing land, wildlife, and water managers is to understand what 

can be done to improve the prospects of fishes, particularly threatened, endangered, 

or sensitive species and stocks, in the face of wildfire and a changing climate. 

Sometimes this goal has been framed in a way that is competitive with the health 

of the terrestrial landscape, usually by way of recognizing the threats that forest 

management poses for aquatic ecosystems (Rieman and Allendorf 2001; Bisson 

and others 2003). The challenge that all resource managers face is in developing 

resilient and resistant landscapes encompassing both streams and forests.

Managers from different disciplinary backgrounds, often in agencies or departments with differing missions, have developed tools and approaches to supporting 

those ecosystem components with which they are most familiar or have the most 

control. Thus, foresters have focused on forests, while civil engineers develop 

strategies to maintain water supplies, and fisheries biologists find ways to maintain 

aquatic ecosystems. Some solutions that are optimal for one resource may be less 

optimal for another, or even harmful to it. As a consequence, from disciplinary 

perspectives, concerns of other disciplines are sometimes viewed as constraints.

Forest managers have focused on fuel reduction through manual thinning or 

application of both intentional and unintentional fire. There is, however, a great 

deal of concern about effects of management intervention (e.g., DellaSala and 

Frost 2001; Rhodes and Baker 2008). Direct vegetation management represents 

a continued and in some cases additional threat to aquatic systems through 

management-related disturbances, including roads. However, even natural fuel 

treatments through fire use can be controversial in some circumstances (Holden 

and others 2010). Solutions are commonly suggested with somewhat universal 
USDA Forest Service RMRS-GTR-290. 2012. 91

framing, e.g., we need to thin the forests to prevent catastrophic or uncharacteristic 

wildfire or thin forested riparian areas to prevent damage in the event of a fire 

(O’Laughlin 2005; Roloff and others 2005). A primary framing for foresters is in 

suggesting that the short-term impact produces long-term benefits.

Aquatic managers have focused on erosion prevention, usually road related, as 

well as some aquatic habitat restoration. Road decommissioning or related restoration techniques can be helpful in reducing chronic sediment loading (Switalski and 

others 2004). Removal or replacement of road culverts to improve fish passage 

has been viewed as an important activity to increase the connectivity and size of 

habitat patches (Clarkin and others 2005). Aquatic restoration also includes actual 

habitat manipulation to add wood, for example, even though the benefits of such 

activities are less well understood (Palmer and others 2005). Introduced species 

are a critical issue in many areas, and consideration of the relative benefits of connectivity or habitat restoration to invasive species may be important in some cases 

(Peterson and others 2008).

Forest and aquatic managers, alike, have drawn heavily on strategies depending 

on fire suppression and suppression of the effects of fire through post-fire stabilization. Although fire suppression as a general approach to ecosystem management 

is not widely supported, it is accepted in situations where there is a threat perceived to a valuable resource, e.g., people and their property or endangered fishes. 

Post-fire stabilization is generally not perceived as negative for terrestrial systems, 

except for practices that introduce invasive plants (e.g., Monsen and Shaw 2001; 

Shaw and others 2005). These approaches are primarily applied for high value or 

irreplaceable resources, but cost and effectiveness are critical issues. Treatments 

are not completely reliable, and there is some irony that the cost and effort is great 

enough that one might be led to expect they provide comprehensive protection as 

opposed to a last ditch effort.

Managers of municipal watersheds have often had to work closely with forest 

managers to protect city water supplies from fine sediment due to timber harvest 

within forested municipal watersheds. The recent increase in frequency of large 

fires and the potentially severe impacts of wildfire on water quality have made 

some water managers proponents of fuel treatments and aggressive fire suppression 

and post-fire treatments within watersheds (Graham 2003). Although an additional 

perceived benefit of the fuel treatments is increased streamflow, research does not 

support the hypothesis, particularly for more fire-prone forests where fuel treatments are ecologically recommended (Troendle and others 2010).

Increased withdrawal of water may be an impending issue with respect to fire 

and fishes in a changing climate. There is concern about the potential need for 

increased irrigation in a changing climate to satisfy higher evaporative demands 

and longer growing seasons. Stream segments dewatered for irrigation may pose 

critical barriers to migration, particularly in more arid parts of the West (Rieman 

and others 2003b). In some cases, technological fixes may be available to shift 

withdrawal locations, but more commonly there may need to be discussion of 

water rights for instream uses. In locations where climate change is driving deeper droughts or lower summer streamflows (e.g., Luce and Holden 2009; Leppi 

and others 2011), water diversion issues may become more severe and urgent. 

Similarly, the construction and operation of dams for water storage to offset timing 

shifts in streamflow could impair migration.

Despite a desire for a blanket answer covering a range of climates and landscapes, solutions depend on a complex set of contexts. None of the tool sets is 

without controversy or consequence to other resources. Forest management may 

threaten aquatic systems, particularly through roads, but road deconstruction could 
92 USDA Forest Service RMRS-GTR-290. 2012.

limit future forest management options and may reduce fire suppression success. 

While we protect the forests for water quality and supply, withdrawal or storage 

of that water for use may impact fisheries, and affect riparian conditions. Many 

of the solutions may have high and unpredictable costs as well, adding issues of 

economic efficiency to an already complex ecological problem.

The number and dimension of issues impinging on decisions about fire is high, 

and the decisions can be difficult and, sometimes, overwhelming. Multiple competing interests and issues can create an impasse that could in itself yield an outcome 

that is optimally detrimental. While there are no magic bullets for cutting through 

all of the different considerations, there are ways of looking at the problem that can 

simplify some aspects.

What is important is setting general principles that help to: 1) build frameworks 

and logic for broad decisions, 2) simplify the issues for managers and interested 

publics who have diverse, often non-technical backgrounds, and 3) suggest process 

and perspective to help solve problems and puzzles where the “knots” are hardest.

Simplifying the Frame

There are a few critical ideas that are helpful for simplifying the complexity. In 

part they help build a hierarchy by noting overarching priorities, and they point to 

interesting features of the problem itself that reduce conflict. We summarize them 

as five general principles to apply to aquatic-terrestrial planning for fire:

1) Holistic approaches are required,

2) Spatial arrangement has relevance,

3) The system is dynamic,

4) Sustainable solutions are needed, and

5) Timing may be critical.

Perhaps these are more reminders than principles to people well versed in natural resources management; nonetheless, they provide guidance to sort through the 

myriad choices presented to us.

The need for a holistic approach has already been stated and would seem to 

contribute to the complexity described above. It is repeated here as the first principle because a clear expression that there is only one ecosystem to manage helps 

immediately deemphasize solutions that harm one component of the ecosystem to 

preserve another. Although common usage of “ecosystem” (including our own) 

treats different locations within a watershed along lines of scientific discipline 

(e.g., riparian, aquatic, forest, rangeland, and terrestrial ecosystems), the interconnectedness of these parts is an important feature of the fish-fire-forest problem. 

Solutions treating just one aspect of the ecosystem may be considered under particular circumstances; however, if such circumstances are limited in space (see next 

principle), there may be alternatives that are more broadly beneficial or solutions 

to one problem that are benign to other ecosystem components.

The spatial arrangement of forests and aquatic habitats at risk has a profound 

influence on reducing apparent conflicts. Flows of energy and material through the 

landscape control the degree of interaction between land and water, and thoughtful mapping and zoning may be applied with these concepts in mind to reduce 

conflict compared to more generalized application of solutions (e.g., Cissel and 

others 1999; Dellasalla and others 2004; Rieman and others 2010). For example, 

forest management or fire, either one, would do little to impact fishes upstream, 

unless, for instance, a culvert blocking upstream fish passage were placed. By 
USDA Forest Service RMRS-GTR-290. 2012. 93

systematically mapping where restoration may be needed to help either forests or 

aquatic habitats, there are opportunities to highlight large areas where no work is 

required, places where only aquatic work might be required, and places where forest work would not affect sensitive aquatic habitats. That remaining portion of the 

landscape where forest work could degrade aquatic habitats would then become 

the focus. The joint spatial alteration of fish and forest habitat through historical forest harvests represents some further opportunity to improve conditions in 

the same places for both aquatic and terrestrial ecosystem components (Rieman 

and others 2000; Dellasalla and others 2004; Rieman and others 2010). Although 

work required to restore these forests might be directly affecting already degraded 

aquatic habitat, identification and remediation of the causes of that degradation in 

concert with the forest treatments could produce net benefit for both ecosystem 

components.

The fact that the system is dynamic favors solutions seeking to build resilience 

over those trying to protect against dynamism. If the motivation behind a forest 

treatment is to make fire more manageable (e.g., more easily suppressed) than it 

might be under natural variations in fuel load, then there may be negative consequences for ecosystems. Both the forests and the aquatic habitats are adapted 

to fire and have co-existed successfully with fire for a few millennia. Although 

climate change has already altered fire regimes in some locations substantially 

compared to the 20th century, past variations in fire synchrony and associated mass 

wasting have matched current levels within these millennia (Kirchner and others 

2001; Meyer and Pierce 2003; Whitlock and others 2003; Pierce and others 2004), 

implying that adequate biological mechanisms exist to survive widespread and 

severe fire and its consequences. Where historical anthropogenic effects have impaired the resilience of aquatic or forest systems by altering the spatial structure or 

connectivity of habitats, risks are higher.

Sustainability relates to the level of external effort, as energy and materials, 

required to maintain system processes. Solutions requiring persistent large outlays to maintain a particular condition through a combination of fire suppression 

and thinning, for instance, would be expensive and probably impractical for longterm application except where very high values, like homes or other infrastructure, 

are protected. If we look across the broader landscape, there are insufficient resources available to public land management agencies to correct current issues 

immediately. For example, there is a $4.5 billion backlog in road maintenance, 

some related to water quality impairment (U.S. Forest Service 2011 Budget 

Justification). Sustainable restoration practices require prioritization of the most 

important issues, e.g., determination and targeting of the most critical places and 

the treatments with the most effect on desired outcomes. As work is done and the 

dynamics of disturbance play out across the landscape, periodic reevaluation may 

be beneficial.

Timing is critical because disturbance is imminent in a dynamic landscape. 

Given both the high technical complexity of designing landscape-scale solutions 

to persistence in a dynamic environment and the huge challenges in convincing a 

diverse public that it is all in their best interest (i.e., in the interest of their pocket 

books, their houses, their safety, and the environment), it is tempting to put off 

concrete decisions and actions until a wildfire provides a seemingly unquestionable mandate. There is a certain degree of hubris in waiting for a fire to occur 

before acting, however, and it is increasingly recognized that both the forest and 

streams could suffer in the aftermath of such an event without some preparation. 

It is also expensive and dramatically limits the scope of choices available to managers. One need only go as far as one’s own dentist or doctor to hear the benefits 
94 USDA Forest Service RMRS-GTR-290. 2012.

of preventative care. It applies for forests and fish as well. For example, the existence of migratory stocks provides one of the stronger guarantees that a population 

will persist despite a short term setback; however, migratory stocks must be available prior to the occurrence of a major disturbance (Dunham and others 2003). 

Likewise, fuel treatments maintaining ecologically appropriate fuel levels, vertical structure, and spatial patterns create greater opportunities for managers to 

use natural ignitions to continue to maintain the situation. Perhaps some would 

note this principle looks redundant since a strategy relying on fire suppression and 

post-fire stabilization alone would allow fuel buildups that reduce forest resilience, 

imposes a static conceptualization for the forest and aquatic habitats, and requires 

substantial resources. In other words, it conceptually violates the first, third, and 

fourth principles. We specify the timing as a unique principal for emphasis and 

clarity. We accept that in the short-term, there may be no reasonable alternatives 

to suppression of fires and post-fire erosion in some locations. Prioritization may 

place restoration work in other areas first, for example, or physical isolation may 

preclude major improvements in fish migration.

The point of these five principles is that even though there are seemingly conflicts between management actions for different disciplines, there are also parallels 

and complements in process. Taking advantage of complementary processes in 

planning requires understanding the fundamental behaviors shared across resources and acknowledging realistic constraints on managers. Individually, the five 

principles look like truisms. Taken together, however, these five principles allow 

for a first order evaluation of most proposed actions or strategies.

Applying the Principles

Most individual proposed activities, e.g., a road decommissioning, a culvert 

replacement, or a fuel-reduction project, would fail screening by the five principles if designed outside of a more comprehensive plan that describes the spatial 

arrangement and sequencing of projects to reestablish dynamic ecosystems of multiple resources in a financially sustainable way. For example, an individual NEPA 

analysis stating that a particular pile and burn project would reduce the risk of 

wildfire and therefore sedimentation in the stream, cannot really address whether 

the sediment input from that site has any relevance to fish populations (negative 

or positive), or whether a different project would better achieve goals of sustainability or restoration of dynamic processes. Thus, there are two reasons why a 

stand-alone proposal of this fuel treatment project would be inadequate: 1) there is 

no context of spatial or temporal prioritization (e.g., watershed analysis or cumulative effects analysis), and 2) it suggests an inappropriate scoping (e.g., a belief that 

sediment load is the primary issue to address for streams). Although fuel treatment 

projects are usually proposed with benefits to vegetation in mind, when threatened 

or endangered fishes are potentially affected, a benefit (or at least a lack of risk) to 

them must be shown. Simply scheduling treatments in areas without threatened or 

endangered fish to avoid the regulatory problems, however, equivalently misperceives the value of an integrated plan.

The first point about lacking a larger contextual relationship to other projects 

is well recognized by land managers and has resulted in technical planning initiatives that are spatially comprehensive, such as watershed analysis (FEMAT 1993), 

fire management plans (NWCG 1995), and transportation planning (e.g., Forest 

Service Handbook 7709.55). Although all of these are carried out using interdisciplinary teams, the very fact that there are different kinds of plans (along disciplinary 

lines) reveals a lack of interdisciplinarity in their development or inception. A 
USDA Forest Service RMRS-GTR-290. 2012. 95

brief reading of such plans or the manuals for their development shows a series 

of descriptions of issues with one particular resource, the proposed solutions, and 

comments on the impacts to other resources. Such plans are commonly tiered to 

more comprehensive area planning documents (such as Forest Plans in the Forest 

Service or Resource Management Plans in the BLM), but even those tend to have 

a series of chapters with resource-specific guidelines. Considering that there is 

probably more effort involved in developing four or more plans for any given land 

area (and trying to roughly tie them together each time) than a single complex 

intertwined one, it is not much of an assumption to believe that the issue is not so 

much a lack of will as a lack of a well stated frame to build them on. Although the 

hypothesis that building consensus among multiple individuals from diverse disciplinary backgrounds is difficult, it cannot be ignored.

The five principles can serve as a frame for interdisciplinary planning across the 

issues of fire, fuels, roads, and aquatic habitat. Above, we noted how each resource 

could be viewed as conflicting with another; however, when viewed through the 

frame provided by the five principles, the importance of the complex interrelationships among them becomes more apparent. In short, although there would seem to 

be many potentially competing needs, in reality, only a few would address problems holistically and sustainably, and these would acknowledge disturbance and 

recovery in patches over the landscape.

Viewing the problem in this way leads to steps that can shed light on the complex relationships and help diverse teams decide which actions to take:

1) Identify resource specific needs, limitations, or vulnerabilities 

(multidisciplinary step)

2) Identify where they are in conflict (and conversely not)

3) Identify where sequencing (order of multiple activities) could ameliorate 

conflicts

4) Prioritize and schedule non-conflicting tasks

5) Creatively solve remaining issues—describe and quantify risks and means 

to obtain feedback to guide future management

The first step is familiar, identifying the needs of the ecosystem in recognition of 

a dynamic system and the need for sustainable solutions. Some may frame this step 

in terms of ecosystem restoration, sometimes as restoration of process. It can also 

be framed as a vulnerability analysis, or alternatively as a reliability or persistence 

analysis. One approach is to identify what makes the forests or aquatic resources 

vulnerable to fire and suggest what steps could provide resilience or resistance to 

those vulnerabilities. It is a multidisciplinary process, with specialists in each discipline using their understanding to clearly articulate the specific aspects of current 

conditions that could result in an anomalous outcome from wildfire. This could be 

seen essentially as a diagnostic examination: for forests determining where they 

are at risk for rapid spread and or homogeneous severity, and for aquatics determining what would maximize survivability following fire.

Although this step is primarily multi-disciplinary, there should be clear recognition that a holistic solution is being sought. Overstated “needs” for individual 

ecosystem components can lead to unnecessary conflicts being identified in the 

second step. Each team member must recognize that there is only one ecosystem, 

and that while there may seem to be tradeoffs between individual resources in an 

ecosystem, the objective of the interdisciplinary team is to find the balance that 

best allows the ecosystem to thrive without substantial ongoing external investments. There is potential at the stage of identifying needs for interpersonal skills 
96 USDA Forest Service RMRS-GTR-290. 2012.

to override objective assessments. There is a need for a combination of personal 

determination, restraint, and leadership by the team members to separate interpersonal conflict from real resource conflicts.

Identification of conflicts or agreement and complementary needs is the key 

interdisciplinary step. As outlined in the general framing that is commonly experienced, there may be some general expectations of conflict a priori about managing 

forests in places with sensitive fish species. Unfortunately, this expectation is uninformative for decisions, and it is important to note where activities for the benefit 

of one component do not impair another, for example, identifying where impacts 

from forest management practices would affect areas downstream of sensitive life 

stages of fish, or where a road removal would not reduce access for fuel treatments. Rieman and others (2000) provide an example where impacted forest and 

stream ecosystems tend to be in the same places and superficially would appear in 

conflict, e.g., impaired fish populations being particularly sensitive to additional 

management. Recognition, however, that such sites may represent opportunities 

for more comprehensive treatment could reveal more opportunity on the landscape 

than conflict. Such recognition could only occur in an interdisciplinary process. 

Misidentified and overgeneralized identification of conflict can also impede finding solutions, and management of personal and interpersonal factors is important 

in this step.

Roads are a primary source of conflict in management of public lands so warrant 

additional discussion. Roads are not a resource; they are a tool to manage, access, 

or benefit other resources. Roads only have value for the resources they access. In 

this framing, roads do not have “needs”; however, if they serve to more sustainably 

manage a resource or obtain the benefits of a resource, the roads could be viewed 

as a “need” for that resource. More objectively, they represent a value with respect 

to a particular resource, e.g., a recreational resource becomes more accessible, or 

a mineral resource becomes economically viable. The environmental and financial 

costs of the road could then be objectively evaluated against such a value. A key 

point to consider is what could be substituted for the road when determining its 

value. A well designed and maintained road could ultimately be more sustainable 

than other forms of access, depending on frequency of access and hauling requirements and the topography to be crossed. Roads are a shared value when they are 

used to access multiple locations, perhaps with different ownership and different 

land uses, so coordination of road systems is often needed.

Some conflicts could be reduced through sequencing of treatments. Most fuel 

treatments require roads to be practical. Fortunately, fuel treatments are likely 

most needed in places that have roads now (Rieman and others 2000). Although 

those roads may be the most practical means for a fuel treatment now, if the fuel 

treatment can be maintained in the future through means not requiring roads (e.g., 

wildfire use), those roads can be decommissioned. If they might be needed in the 

future, and values warrant ongoing costs, they can be upgraded to be resilient 

to storms. This kind of coordination allows both improvements to terrestrial and 

aquatic components. Note, however, in any arrangement whereby future risks are 

reduced following a short period of increased risk could be self-defeating. For 

example, in the case of a species extirpation related to a temporary risk increase, 

no amount of future risk reduction would be of any benefit. If the increase in risk 

is more than nominal, and there are no additional measures that can be taken to 

reduce the risk, the proposed activity would need to be identified as potentially 

conflicting and set aside for creative thinking about how to quantify and manage 

through the uncertainty.
USDA Forest Service RMRS-GTR-290. 2012. 97

Prioritization among the non-conflicting actions requires identifying which 

treatments provide the greatest return in terms of the objectives for the ecosystem, again a thoroughly interdisciplinary process. Various criteria are available for 

prioritization and, ultimately, monitoring progress. Two of the biologically critical ones are persistence of sensitive aquatic stocks and vertical/spatial structuring 

of forest stands. Elements overlap, but there are also many independent aspects. 

The key question for managers may be to determine which action sets (including 

those with sequencing) most affect persistence and improve forest structure. It has 

been suggested that places with some intermediate levels of historical alteration/

disturbance may be the most productive for restoration results (e.g., Dellasalla 

and others 2004) (Figure 31). These places are not necessarily so degraded that 

connectivity can only be gained in increments of several meters, nor are they in 

places where improvements represent icing on the cake. Locations where persistence probabilities for fish species are already high would be lower priority than 

places where fragmentation has put species at significant risk. Places where the 

original genetic diversity values are lost are lower priority than preserving those 

currently at risk, but still extant, although eventually reestablishment of extirpated 

populations could be a sustainable approach to conserving the remaining diversity.

In places where conditions prohibit improvement in the condition of some ecosystem components without harm to other components, more creativity will be 

needed. A combination of historical practices, changing climates, and introduction 

of species has resulted in many locations where conflicts are real. In such places, 

risks for both forests and fish are likely high, with potential for severe effects from 

wildfire or loss of a local fish population. In these circumstances, choices must be 

made, and even deferral is a decision with consequences. These places may show 

conflict because the sensitivity of the ecosystem is highest and risks are greatest, 

and these are the places in greatest need of decisions and actions.

We also note the utility of the first four steps in helping with the tough “knots” 

that are left over. Within the map of non-conflicting treatments developed from the 

first four steps is a list of practices or activities that will alleviate stresses on the 

most stressed parts of the ecosystem. For example, restoring structure of forests 

in one area ultimately benefits places nearby by altering fire spread and fire risks Value of Restoration

Level of Distruption

Figure 31: Most benefits 

from restoration are likely 

to be found from efforts 

in places only mildly or 

moderately degraded.

Disruption
98 USDA Forest Service RMRS-GTR-290. 2012.

from off site. Similarly, restoration of fish migratory patterns in larger river systems could be a critical step in ultimately providing stronger persistence for what 

is now an isolated habitat.

Even the most creative solutions will require accepting risk and impacts, at least 

in the short term, to reduce longer term risks. As noted earlier, the balance between 

short- and long-term risk is important to consider. If a short-term high-impact action can be mitigated through short-term high costs to effect a change that makes 

a situation more sustainable in the long term, high costs may be justified. A small, 

isolated fish population, with unresolvable downstream invasive species issues, for 

example, poses a difficult problem. One can reduce fire risks upstream with forest 

management, but risks from the management and roading could pose more certain 

risk to this isolated population. Potentially, however, helicopter access and hand 

work on finer fuels could dramatically reduce management related impacts. The 

cost would be high in the short term, but if it allowed wildfire use in the future to 

maintain fuel levels, it may be more sustainable in the long term. If such steps produced an anomalous fuel condition for the forest (e.g., thinning a mixed fir forest), 

but the condition could ultimately be maintained with less expensive broadcast 

burning, it would still represent a more sustainable and likely more reliable way to 

conserve an isolated population than relying on suppression alone.

Decisions about values include more than forests and fish. Recreation access 

uses roads, and there will come a question as to which is more important: that 

particular access or the aquatic resources kept at risk by the road. In streams facing the potential of invasion from native species, the choice to place a barrier or 

not involves tradeoffs in risk between the invasion and potential for loss of genetic 

diversity in some places versus the risks associated with isolating a population, 

and people will need to decide between the different values the fish provide in that 

stream (Rieman and others 2010). Dams for water supplies are sensitive to sediment from management, thus limit options for land management. They are also 

sensitive to the major water quality changes from fire, such as metals, nitrates, and 

sediment. Fish are also sensitive to these and are made even more sensitive by the 

fragmentation associated with the dams. Two choices are to “engineer” the ecosystem tightly and dampen disturbances, which is expensive to the land managers, or 

to design a more reliable water supply system (e.g., with other intakes or groundwater reserves for temporary use) that could operate within the dynamic landscape. 

Both choices are expensive, but the costs are borne by different parties.

The difficult conflicts will require complex and creative thinking. Ideally we 

want to go back to the five principles to develop a holistic solution that can be sustained in time recognizing landscape dynamics and using space and time to build 

strategies. A variety of constraints will impinge on some of these principles, and 

sometimes the best solution does not follow all of the principles in the short term. 

There are too many stories to analyze here, but the sample should demonstrate the 

reality and complexity of the situations, but also their potential rareness in the general landscape. If we can work in the places that are not difficult first (e.g., where 

the fish are resilient to management effects or where the management can help fish 

habitat), perhaps we can lessen the stresses on the difficult places too.

Evaluating Risks

Balancing tradeoffs or optimizing for priorities are inherently quantitative in 

nature; even in a world of dynamic and stochastic processes and events, there are 

gaming theories that could be applied to rationally decide the most astute course 

of action (Bishop 1978). Such rational decisions, however, require objective 
USDA Forest Service RMRS-GTR-290. 2012. 99

valuations of alternative outcomes, which are rarely agreed upon with respect to 

natural resources, and in any event are not technically allowed as a basis for decisions on endangered species. Where the decisions bear on traditionally valued 

resources like trees or homes, or even where they bear on replaceable populations 

of fishes, there is the potential to apply rigorous tools. When the mix of values 

involved includes officially protected species, the decision frame is dramatically 

reduced to evaluate whether an action increases or decreases risks to the species 

in question.

As a consequence, the risk to these species becomes an overriding factor in the 

decision making, and quantification means evaluating risks of alternative actions. 

Technically, risk is defined as the probability of loss multiplied by the value. If we 

are only evaluating increases or decreases in risk for a given resource, however, we 

can assume the value remains fixed, if immeasurable, and focus on probabilities 

of loss or persistence. Within a stochastic framework, probability is commonly 

expressed in terms of time, e.g., a probability of 0.01 yr-1 would be a 100 year 

event. Since we know that fire events are a limiting condition for fish, and because 

fire severity/size characteristics are rarely expressed in this temporal probability 

framework, one could further limit analysis to whether a population persists in 

the eventuality of a fire. For a management action with a short-term impact, the 

equivalent question is whether the population would survive the action and its 

consequences. The question is whether the reduction in future risks from the management action outweigh any temporary increases in risk with the management 

action.

With respect to fishes and fire, the key concepts that apply to long term persistence of a local population are metapopulations and life history diversity. 

Metapopulation theory addresses the spatial interactions between patches of habitat 

and patches of disturbance, examining how local populations that are lost to disturbances can be refounded from other nearby occupied habitat patches (Rieman 

and McIntyre 1995; Rieman and Dunham 2000). Life history diversity relates to 

whether there may be migratory individuals from a particular stream that would 

be elsewhere (downstream in a reservoir or the ocean) when the fire-related effects 

occur (e.g., Hilborn and others 2003). Both means of repopulating after disturbance depend on either a large local population or a well-connected population. 

Essentially, these are populations where disturbances cannot simultaneously affect 

all potential refounding sources, including fragments of a given habitat patch, other 

nearby populations, or downstream migrants. The primary metrics that emerge as 

critical control on persistence are the 1) sizes of habitat “patches” and the 2) connectivity of those patches. Productivity within patches can be important too, in that 

it interacts with their size and connectivity. If habitat quality is low, reproduction 

in undisturbed habitat will not be effective to reinforce losses elsewhere. Also if 

productivity is low, migration may not occur, or could be less successful because 

small migrating fish are more susceptible to predation.

Sometimes, the view that sediment is an impact to aquatic biota is overgeneralized, and the comparative “risk analysis” is boiled down to a comparison of 

total sediment mass from fire versus that from management (Istanbulluoglu and 

others 2004; O’Laughlin 2005; Roloff and others 2005). Unfortunately this oversimplification of the ecological response of streams to sediment inputs leads to 1) 

stalemate in some circumstances (in areas with steep slopes), 2) negative decisions 

(when chronic road sediments are less than the total long-term pulsed sediment), 

or 3) low efficiency decisions (e.g., to rely on post-fire stabilization). A contrast of 

assessments done for relatively low gradient slopes (O’Laughlin 2005) to one on 

steeper slopes (Istanbulluoglu and others 2004), for instance, shows that while the 
100 USDA Forest Service RMRS-GTR-290. 2012.

total sediment over the long term is higher from fire than forest management on 

low gradient slopes (considering surface erosion only), there is little difference in 

estimated sediment mass between fires and forest harvest for steep slopes where 

mass wasting might occur (Luce and others 2005). Such an analysis would show 

no improvements for environments with steep slopes, providing little guidance. 

More problematically, this kind of decision criterion would preferentially select 

frequent fine road sediments inputs over infrequent but more severe inputs from 

fires, which could have negative implications for aquatic habitat in the long term 

(e.g., Reeves and others 1995). Modeling the aquatic habitat response based on 

total sediment loads does not recognize the dynamism of the coupled forest-stream 

ecosystem but instead drives management toward a low sediment, low variability 

state, that cannot persist without large energy and financial inputs, such as fire suppression and post-fire stabilization.

An alternative model is to consider the length of aquatic habitat simultaneously 

affected by debris flows, temperature, and sediment effects within a given habitat 

patch, and its connectivity to migration corridors. Decisions under such a model 

are likely to emphasize resilience options. Within this kind of persistence model, 

one could describe the effects of temporary sediment loads on population productivity versus the effects of wholesale habitat loss. Such a model is also more likely 

to reflect the benefits of careful timing and location of various treatments (e.g., 

forest thinning, road decommissioning, or culvert replacement) to avoid synchronous risks over a patch or stream segments. More importantly, however, it would 

clarify which measures do the most to increase connectivity, productivity, and size 

of population units.

Persistence analysis should also address solutions requiring sustained external 

inputs. Sustainability relates to how reliant a strategy is on future financial and 

energy inputs. Reliability depends both on whether a particular effort-intensive 

option is effective and whether or not it gets done. There is a certain irony to managing systems that can survive a catastrophic wildfire but not a recession. It can be 

difficult to show performance and accountability for “disasters averted” (e.g., extinctions), particularly when those disasters are rare in time. This makes it difficult 

to maintain funding for approaches that require continuing or repeated financial 

outlays. This concept is relevant to strategies incorporating fuel treatments, fire 

suppression, and post-fire stabilization as elements. If the fuel treatments promote 

resilience to natural fire events and allow for more passive management of the 

system in the event of fire, they are sustainable. If, however, fuel treatments are 

done to reinforce suppression efforts, e.g., an increase in resistance to fire, then it 

is likely that such treatments would require sustained efforts to maintain. The uncertainty in future funding (and therefore implementation) for a proposed program 

could be directly incorporated into a formal risk/reliability analysis estimating persistence probabilities.

Addressing Uncertainty

One of the values of a more formal and quantitative assessment of risks is in 

clarifying and defining uncertainties. Disagreements about the perceived risks associated with either fire or forest management can be strong sources of debate 

about endangered species. Discussion about the acceptability and relative tradeoffs 

between particular risks can be long, expensive, debilitating, and difficult to resolve 

(Rieman and others 2010). By placing decisions within a risk analysis framework, 

two different aspects of uncertainty are separated: those related to chance or luck 

(e.g., future weather) and those related to what we know about the system (e.g., the 
USDA Forest Service RMRS-GTR-290. 2012. 101

probability of different future weather events). The more clearly and specifically 

the risks are defined, the easier it is to determine the degree to which uncertainty 

about outcomes affects the balance of risk. For example, although there may be 

wide error bars on the amount of fine sediment reduction that may be achieved 

from post-fire stabilization, we also know that fish populations are more sensitive 

to rare mass wasting events. Because the post-fire stabilization has little influence 

on debris flow events, the uncertainty about performance of treatments on fine 

sediment loads has little leverage on the outcome for fish.

Costs of uncertainty are manifested in a variety of ways. The most common is 

unresolved disagreements, which are ultimately reflected in agency budgets and 

morale. Incorrectly applied action or inaction (sometimes related to unresolved 

disputes) can also result in local extinctions of threatened and endangered species, 

which is potentially the worst outcome. Alternatively we may see only a loss in 

other values, such as recreation, grazing, or timber, derived from forest and stream 

ecosystems. Wasted restoration efforts may seem like an economic issue, but they 

may also represent continuing risks in other places. None of the outcomes from 

poor or incomplete information are positive, but some are more severe than others. 

Discerning the places and circumstances where ignorance is relatively benign is an 

important step forward in generating potential options for decisions. Those topics 

where not knowing something can generate severe consequences represent a clear 

priority for generating information.

This synthesis illustrates that we have a general understanding of the relationship of some fishes to wildfire in semi-quantitative terms (Bisson and others 

2003). One key uncertainty is the scale of habitat patches required for persistence 

in time (Rieman and others 2010). While observations of currently occupied and 

unoccupied habitats provides some guidance, actual testing of the fish population 

response to fire is lacking. Although uncertainty has been expressed over the degree to which various fuel treatments affect the severity and spread of fire (e.g., 

Rhodes and Baker 2008; Stone and others 2010), there is agreement that disturbance patch sizes of natural forests tend to be smaller than in regulated forests 

or spatially homogeneous forests (Miller and Urban 2000; Hessburg and others 

2007; van Wagtendonk and Lutz 2007). Disturbance patch size may be important 

to some fish populations, but less so to others. For instance, populations in very 

small patches can be affected by small disturbances, so they are relatively insensitive to the distribution of fires beyond a nominal size. A better understanding of 

which fish populations benefit more depends on the relationship of patch scale to 

long term persistence. The degree to which stream heating occurs after riparian 

thinning, its recovery rate, and the degree to which thinning reduces losses in fire 

remain unanswered questions that affect decisions for specific projects (Stone and 

others 2010). Perhaps the more important question is how stream heating would 

affect patch size or geometry, which depends on the distribution of fishes, and the 

degree to which they are constrained by temperature. More broadly, and we suggest more problematically, the lack of information about the current status and 

condition of fish populations and diagnostic information about their limiting conditions at local population scales most strongly limits risk assessments.

Risk reduction is not just about performing some action in the landscape, it is 

also about making fewer mistakes, which will require better information. Inventory, 

monitoring, traditional research, and adaptive management are all important in 

gathering the information that will be most important to decisions. In applying 

the five principles, we outline a procedure that identifies areas where activities are 

not in conflict, offering ideal opportunities to learn from earlier actions in local 

areas. Analysis of fish populations and habitats in the wake of large fires would 
102 USDA Forest Service RMRS-GTR-290. 2012.

also address some of the key uncertainties, particularly related to scale of sustainable habitats. Although members of interdisciplinary teams who are charged with 

assessing risks under various alternatives recognize the value of inventory, there 

seems to be less understanding of the value of such information to those providing 

funding. There is, in general, much more support to perform actions than to design 

them, which suggests a need to better estimate the value of inventory and monitoring efforts for informing actions.

Finally, we should note that what the climate will do in the future may be one 

of the largest uncertainties before us. Although it is clear that temperatures will 

increase, much less is known about precipitation trends. Fluid dynamic theory 

consistently predicts increases in precipitation at high latitudes (>50 degrees) and 

broadening of the arid sub-tropical (~30 degrees) zone. Between, there is large 

uncertainty (Solomon 1986). Within the western United States, for example, there 

is substantial uncertainty with respect to changes in orographic precipitation from 

mountains (Dettinger and others 2004; Kirshbaum and Smith 2008). One of the 

clear needs that emerges from evaluating risks is the need to pay attention to what 

changes actually occur and what the responses are to those changes. These sensitivities provide important context to which habitats will be most resilient in the 

face of fire in the future.

C. A Changing Climate

In contrast to the comparatively immediate and substantial effects of fire, changes in weather statistics associated with climate change seem subtle or gradual, 

although persistent. An important question is how to think about the consequences 

of climate change (the more gradual but steady effect) on how forests and aquatic 

ecosystems respond to and could be managed with respect to fire. The climate 

change adaptation literature has already noted the importance of increasing resilience to temporary disturbance in ecosystems (Millar and others 2007), which begs 

the question of what more can be done than what has been suggested for increasing 

the resilience of forests and stream ecosystems to fire (Bisson and others 2003).

Two primary issues are important to discuss: 1) the altered contexts of forest 

and stream ecological dynamics, and 2) an increased urgency, particularly for restoration of historically disrupted habitats. In earlier sections we saw how climate 

change impacts are constraining habitat patch sizes and reducing habitat connectivity for fishes. Given that size and connectivity are critical parameters for the 

resilience of fish populations to fire, climate trends are causing changes that oppose what managers would like to create. Climate, for example, has contributed to 

the increased number of large fires in the western United States (e.g., Westerling 

and others 2006; Morgan and others 2008; Dillon and others 2011; Holden and 

others 2012). The rapidity of the changes is challenging ecosystem responses, and 

places where resilience has been impaired by management, such as through isolation of streams (Dunham and others 2003) or homogenization of forests (Hessburg 

and others 2007), now have compounded risk factors.

Climate changes are also reducing the effectiveness of resilience adaptations 

under natural conditions, suggesting that strategies for managing fire that assume 

stationarity in climate could be risky. Millar and others (2007) have suggested 

measures for adaptation to climate change, including building resistance to the 

changes and taking measures to help transitions occur more manageably, as opposed to catastrophically. They offered suggestions to essentially embrace the 

coming changes, some of which seem controversial, e.g., facilitating range shifts. 
USDA Forest Service RMRS-GTR-290. 2012. 103

Some suggestions are derived from basic principles in conservation biology and 

can be revisited in terms of resistance, resilience, and facilitation for the future. 

Their key points are [with editorial license for emphasis]:

1. Experiment with [networks of] refugia

2. Realign disrupted conditions

3. Increase redundancy

4. Expand genetic diversity guidelines [introduce individuals from other parts 

of range]

5. Facilitate migration [but applied within current ranges]

6. Promote connectivity [but reduce contagion of fire, disease, and pests]

7. Manage for [and take advantage of] asynchrony

In Millar and others (2007) some of the ideas presented were offered in the 

context of helping make transitions to new habitats successful, but all of these 

steps could be equally well applied in resisting effects of climatic changes on loss 

of local representation of species and promoting resilience. If such steps delay the 

consequences of climate change, they allow more time for decisions about species, 

adaptation of species, or successfully reducing carbon loading in the atmosphere. 

How these ideas are emphasized relative to one another and how they are functionally implemented may look different in aquatic and forest ecosystems.

Climatic changes can be viewed simplistically as shifting suitable habitat ranges 

higher in elevation and farther north. Elevation-wise distribution shifts can be accomplished more easily than latitudinal shifts, owing to much shorter distances 

for a given temperature change and the lack of any need to cross potentially large 

unsuitable terrain. This latter issue can be particularly troublesome for fishes, who 

are constrained to move within waterways, many of which are blocked by dams or 

geologic features. In general, remaining aquatic habitats for cold-water fishes will 

be higher in watersheds with steeper channels that are more prone to post-fire debris flow disturbances. They will also be smaller streams with greater vulnerability 

to drier dry years and lower summer flows (e.g., Barnett and others 2008; Luce 

and Holden 2009; Leppi and others 2011). Even species that are less temperature 

sensitive may have reduced net productivity and fewer migratory individuals if 

temperatures warm (Dunham and others 2007). One of the primary effects of climate change is to pinch populations into increasingly small, isolated, lower quality, 

and dangerous habitats. One of the key weaknesses of refuge based conservation 

programs is that a refuge can be lost to an individual event (Williams and others 

2011). Recognition of this condition leads to a generally pessimistic view of the 

fate of species and connected ecosystems and could cause one to question the utility of suggestions about expending effort on building resilience. We would like to 

counter that pessimism.

The changes occurring to ecosystems are diverse in nature, and not all habitats are changing at the same rate. Some will change faster than ecohydroclimatic 

models would suggest, while others are changing more slowly. There are several 

examples describing mechanisms of fine grain heterogeneity providing sustained 

water flows or cooler temperatures, at least temporarily (Luce and others 1998; 

Baxter and Hauer 2000; Hari and others 2006; Lundquist and Cayan 2007; Tague 

and Grant 2009; Millar and Westfall 2010; Holden and others 2011). These areas 

are sometimes referred to as climatic refugia or microrefugia (e.g., Noss 2001; 

Dobrowski 2011). While sometimes these are cooler or moister habitats now, the 

important point is that habitats in future microrefugia should be relatively insensitive to regional climate changes. For fish, we might expect spotty distributions in 
104 USDA Forest Service RMRS-GTR-290. 2012.

the future related to local phenomena of cold air drainage, snow drifting, alluvial 

valley fill aquifers, aquifers in volcanic terranes, glaciers, or related phenomena. 

Some of these would be less enduring than others. The remaining patches created 

by these kinds of processes will tend to be small and isolated, thus at extreme risk 

to fire or other disturbances. They may also be at higher risk to small populations 

effects related to genetic diversity (e.g., Ellstrand and Elam 1993) (and see Neville 

and others in Advanced Topics of this volume). Forests may see similar local variations in topography and microclimate that help protect against fire, drought, and 

pests. Some of these variations in geology and topography provide the complex 

ecological landscapes we now see. Fires during extremely dry conditions promote 

homogenization of landscapes, with most of the area prone to severe fire regardless 

of local conditions (Dillon and others 2011). Fires during less extreme conditions 

promote heterogeneity because they are more responsive to local microclimate and 

fuel conditions. By consuming fuels and promoting heterogeneity in fuels, they 

also reduce the likelihood of later homogenizing events. Thus using wildfire during conditions that might normally allow suppression could be an important part 

of an adaptation strategy.

A Thought Experiment

If we consider the situation of heterogeneous, fragmented habitats in an idealized sort of way, we can examine how the principles outlined above might work 

together. Climate change can be viewed as creating disconnected networks of 

small refugia. If we identify the anticipated network of refugia and potential refugia, we can take steps early in the process to reinforce them with appropriate 

realignment (a term meaning restoration but recognizing that some aspects cannot be restored) of conditions disrupted through historical management and land 

use. We could also reintroduce individuals to some of the small suitable but currently unoccupied habitats, presumably that have been used in the past but may be 

currently unoccupied as a result of recent disturbance and migration obstacles or 

barriers. Recolonizing these habitats now would increase the redundancy available 

in the refugia network. Because these sites are not currently occupied, we could 

introduce individuals from various populations farther south or at lower elevation 

that may already have some adaptation toward future conditions in the unoccupied 

habitat. Manually maintaining (and controlling) gene flow among the habitats and 

providing restocking of extirpated patches as they occur is a related extension, 

wherein connectivity between habitats would be provided through management 

intervention where the option of simply opening or maintaining passage has been 

preempted. While more expensive, providing artificial connectivity between populations also offers control over spread of some introduced species and contagions.

This idealized approach draws on the idea of asynchrony in disturbance. 

Asynchronous disturbances mean that some places are disturbed while others are 

not, thus forming a foundation for resilience. If one views a snapshot of a dynamic asynchronous landscape, it will look heterogeneous. A snapshot of synchrony, 

on the other hand, looks like uniformity. Some tree species are well adapted to 

large conflagrations having seeds that are resilient to fire (e.g., serotinous cones 

on lodgepole pine or jack pine), producing large even-aged stands that themselves 

encouraged spread. Unfortunately, such an adaptation is ill-suited for a climate 

that more frequently yields severe fire weather conditions. Creative fragmentation 

of forest landscapes can help maintain current stands of such species as islands 

within patches of more frequently burning uneven aged stands. The trick is finding 
USDA Forest Service RMRS-GTR-290. 2012. 105

the places that burn less frequently and encouraging reproduction of fire-sensitive 

species in those locations.

Futures with dotted refugia of climatically misfit forests and fishes are unlikely to evolve naturally, and hopefully the reader recalls that this is a thought 

experiment within an idealized example. Such networks require planning and intervention, and some cost to maintain. The level of effort will depend on how much 

area will be manipulated and the degree and nature of intervention involved. We 

lack the capacity to keep whole landscapes from transforming, however, by taking 

advantage of natural sources of heterogeneity, there is some capacity to build small 

reserves for specific genetic resources. Tradeoffs between values and costs are an 

important consideration.

Unless substantial changes are made in greenhouse gas production, all of these 

steps may ultimately fail in any given location. One should recognize that gradual 

step wise changes could help bring (and build) genetic variations that are better 

adapted to warmer climates to more suitable ranges in a slow and orderly fashion, 

without placing species currently in those locations at increased risk from longer 

distance introductions through large scale facilitated migration. The gradual nature 

of the process we describe is also more adaptable to take advantage of (or respond 

more quickly to) incorrect climatic projections or temporary reprieves created by 

low-frequency oceanic temperature cycles (e.g., such as described by Mann and 

others 1995a; Mann and others 1995b; Jain and Lall 2001). There are characteristics in this process that may also allow taking advantage of differential effects of 

climate change on native versus non-native species, if we have information about 

the relative pressures (Bradley and others 2009; Wenger and others 2011b).

This was an exercise in idealization, but the important idea is that the situation is not without hope; there may be substantial capacity to slow or delay the 

most severe consequences of climate change, e.g., species extirpation or extinction. It is also useful to remember the generalized principles: 1) taking advantage 

of natural heterogeneity, 2) restoring historical degradation, 3) connecting refugia, 

4) building redundancy and representation of genetic variability, and 5) developing 

flexibility around disturbance dynamics.

Integration With Water Resources Management

Climate change is driving changes to human demands on wildland watersheds. 

Water is increasingly more valuable across most of the West, and there are increasing demands to store water on public lands and draw water from rivers. Some of 

the uses will be the traditional irrigation uses, but an increasing amount of water 

use may go to producing energy to offset greenhouse gas related sources. Again 

there will be the tug of war between positive and negative effects for aquatic biota.

Some of the more simple principles may apply. For example, many requests 

for infrastructure additions may not coincide with the most sensitive fish habitats 

(parallel to Rieman and others 2000). Raising the height of an existing dam a small 

amount can provide additional storage for comparatively little expense and alters 

little additional habitat. Another example is restoration of historical impacts to 

meadows to increase shallow aquifer storage and release during low flow periods 

(Loheide and Gorelick 2007). Keeping ideas about the distribution of impacts over 

the landscape in mind can be helpful in reducing additional impacts from water 

resource development.

An important issue for aquatic, riparian, and water managers is how climate 

change may alter precipitation amounts. A principle reason for increased requests 

for additional impoundment in the West is the changing snowpacks. Warmer 
106 USDA Forest Service RMRS-GTR-290. 2012.

temperatures melt snowpacks sooner, and the storage that the snow provides to 

hold water until it is needed in the spring and summer will be less available (e.g., 

Barnett and others 2005). Unfortunately, storage in reservoirs or meadows also 

increases evapotranspiration. In places where precipitation is rising or remaining 

constant, this may pose little difficulty. However, where precipitation is declining, 

additional storage could exacerbate the problem. Most irrigation related reservoir 

systems in the West do not have multiple year storage, and in extreme cases, the 

reservoirs could be dry or very low during years with the greatest need. Even 

where multiple year carryover is available, water allocated to its limit could render storage facilities ineffective if not carefully managed (e.g., Barnett and Pierce 

2008, 2009; Rajagopalan and others 2009).

Reservoir operation will become both more important and more difficult, 

and improved information and forecasts about incoming flows will become increasingly necessary. Reservoir operations have potential impact (or utility) for 

downstream fishes and riparian vegetation and for upstream migratory fishes that 

use the reservoirs as part of their life cycle. Reservoir management can also reduce 

the need for storage increases. Some dams release water from deep in the reservoir, 

and thus can provide cool water for tail stream fisheries. If these releases are too 

cold or too strong, however, the quality and utility of the habitat as refuge can be 

limited. Upstream migrating fishes can be affected by channel processes and abandoned channels in the drawdown areas in the upper reaches of reservoirs. They 

can also be affected by food and temperature relationships within the reservoir. 

Because of the multiple complex, interlaced, and sometimes competing demands 

for reservoir operation, improved information about inflows, particularly mid- to 

short-term forecasts, can be useful for optimizing storage near the end of the melt 

season, which could reduce the need for additional storage facilities. Unfortunately, 

if contemporary trends of increasing interannual variability in flow continue, designing operations to benefit aquatic and riparian biota may be challenging. It is 

likely that improved data on snowpack, soil moisture, and precipitation data could 

be used to substantially improve forecasts relevant to dam operations, however, 

and the increasing value of that information should be assessed. What may have 

seemed too expensive in the past may be more reasonable with increasing value 

of information.

Municipal water supplies may be more vulnerable in the future to wildfire. One 

challenge will be deciding between fuel management activities, with expected 

erosion and pollution from forest roads, versus risks from wildfires that are less 

easily controlled. Either choice carries risks and solutions with different costs. 

Alternative water sources can be useful because impacts of post fire erosion events 

to water quality can be relatively brief. Developing a multi-tiered water sourcing 

scheme could improve reliability of water supplies while allowing more flexibility 

in fire management. For example, the San Francisco water system, with a range 

of back-ups from large cisterns within the city, to reservoirs close to the city, to 

groundwater reserves further south and finally a major reservoir proximal to the 

city is an example provided for engineering reliability analysis where failure of the 

system post earthquake can be dire (Scawthorn and others 2005). Relying on forest 

managers, fire fighters, and federal emergency stabilization engineers to prevent 

water quality impairment to municipal water resources effectively transfers costs 

of providing reliability away from the water supply agencies (and their rate payers). Consequently, accounting of costs to different entities may be important in 

clarifying the relative benefits and costs of different strategies.

Cutting down forests to release more water has been a subject of interest for 

water and forest managers for as long as forests have been scientifically managed 
USDA Forest Service RMRS-GTR-290. 2012. 107

(e.g., Bates and Henry 1928; Andréassian 2004). The opportunity to both reduce 

fire risk and improve water yield through forest thinning makes the idea all the 

more tempting. The actual performance of thinning in increasing water yields, 

however, is not well established (e.g., Troendle and King 1987; Troendle and others 2010). In high elevation forest environments, increases in snowpack due to 

reduced canopy have been documented (Wilm 1944; Wilm and Dunford 1948; 

Troendle 1983). However, success has not been unequivocally determined for thinning (e.g., Wilm and Dunford 1948; Troendle and King 1987; Ffolliott and others 

1989), because it is thought that the residual stand can develop rapidly to utilize the 

available water. An additional concern is that drier forests and drier years produce 

less additional water from forest harvest (Troendle and King 1987; Brown and 

others 2005; Ford and others 2011). As a consequence, increased water yields are 

easier to make in places and years they are least useful. Also, almost all of the work 

has been done on experimental watersheds smaller than 10 km2, where treatments 

are done on complete watersheds. The example from the Boise River fires may 

be germane to many circumstances (see textbox on Boise River hydrology in the 

hydrology section). With roughly 45% of the basin in moderate to severe fire, there 

was a statistically significant increase in water yield of 5%. The additional 50,000 

acre-feet is substantial; however, the 200,000 acres of removed canopy represent 

more area than most forest managers can realistically hope to treat and maintain in 

a condition with low canopy within current budgets.

There are numerous complexities when water resource management is considered in decisions about fire, forests, and aquatic ecosystems. The intention here is 

not to provide a full summary, but rather to build awareness about it because water 

is a critical resource to forests, fish, AND people. The values to people will likely 

have heavy weight where there are conflicts with other values. There are some 

activities, particularly investment in information infrastructure (like weather and 

streamflow gaging stations) that can help better balance between the needs of these 

resources.

Decision Making for an Uncertain and Dynamic Future

The uncertainty of elements of climate change raises questions about what climate change means for decisions regarding future management of risks to forests 

and fish. Is it all too uncertain to make any plans? Does knowing how the climate 

is changing make a difference to decisions about fire and fish?

Optimizing forest management has long been discussed with respect to timber 

harvests as framed in the maximum sustainable yield and normal forest concepts 

(e.g., Hawley 1921). Formalized mathematical optimization under constraints of 

multiple uses and other resource values has been used in operational forest planning for several decades now (Johnson and others 1986). Within such optimization 

techniques, however, are assumptions about certainty of outcomes. Some have 

promoted robust decision making (Regan and others 2005) as a better model for 

planning under climate change because of the focus on reliability and prevention 

of failure by considering worst case analyses. However, even implementations of 

robust decision making approaches still require more information than land managers might have.

The classic minimum path length (or cost) problem provides an illustration. 

Figure 32a provides the common framing with a set time (cost) for each path. 

Optimization algorithms can be applied to this and significantly more complex 

networks to find the shortest path, which the reader can find for this simple figure. 

Figure 32b frames the problem in a way consistent with robust decision making, 
108 USDA Forest Service RMRS-GTR-290. 2012.

which essentially considers the “worst” case, i.e., finding the path with least cost 

if everything goes wrong. In this case, the range of potential times is bracketed. 

In this sense, robust decision making takes uncertainty into account in a way that 

focuses on preventing failure, and a less uncertain path becomes optimal, even 

though the original path could sometimes be faster. This is very similar in some 

ways to the first problem. The third Figure, 32c, shows something that might be a 

more common situation for decision makers, uncertainty that is poorly characterized or so uncertain as to be uninformative. A temptation might be to select the 

path with the most information (least uncertainty), but this is no more rational than 

selecting any other path and is essentially built from beliefs about what is behind 

the question marks. Neither classic optimization algorithms nor robust algorithms 

can solve this problem a priori, and decisions must be made as events unfold, a task 

described as dynamic decision making (Brehmer 1990).

Why might there be question marks on some paths? The cost may depend on 

previous paths taken, on the outcome from a previous uncertain path, or environmental events that took place in the time to cross one path. That is, answers to 

some of these questions may be unknowable until the first step is taken. A classic 

example of dynamic decision making provided by Brehmer (1990) is of forest fire 

management describing three criteria for dynamic decision making: 1) a series 

of decisions is needed, 2) the decisions are interdependent, and 3) the context or 

environment for the decisions changes autonomously as well (Edwards 1962). In 

short, they are the generally common situations associated with complex natural 

resource management decisions.

In the framing of dynamic decision making, information resources and the flow 

and processing of information become paramount concerns. The principal processes associated with improving performance are

1) Feedback (collecting information on environment and outcomes of 

decisions)

2) Feedforward (predictive modeling)

3) Cognition (collating information, assessing conditions, and making 

decisions)

Studies on dynamic decision making examine strategic combinations and 

characteristics of these three processes to improve learning and performance 

(Gonzalez 2005). Several emergent ideas seem generalizable across a number of 

fields of endeavor. Lag in feedback dramatically decreases learning and ultimate 

Figure 32: Conceptual path length diagrams for (a) classic optimization,( b) robust optimization, and (c) dynamic 

decision making. In (a) path lengths are given with uncertainty, in (b) and (c) path lengths are given across a 

range; however, some of the path characteristics are unknown in (c). 

(8)

(2)

(4)

(6)

(3)

(1)

(4)

(2)

(1)

(8,9)

(2,7)

(4,5)

(6,9)

(3,4)

(1,2)

(4,10)

(2,3)

(1,4)

(8,?)

(2,?)

(4,5)

(6,9)

(?,?)

(1,2)

(?,?)

(2,?)

(1,4)
USDA Forest Service RMRS-GTR-290. 2012. 109

performance (Brehmer 1990). Feedback-control alone does not work well, but is 

an important component in feedforward and cognitive support tools. Although it 

is well recognized that retrospective examination of outcomes as a stand-alone 

feedback strategy is inefficient and produces suboptimal results (Gonzalez 2005), 

it can become a preferred mode of control because it requires less cognitive effort 

and simpler task models. Decision makers may be unlikely to recognize that they 

could be doing better if they attain normative success with feedback-only controls, in part because a task model developed from a feedback-control strategy has 

difficulty describing improved performance from alternative strategies (Brehmer 

1990). People are the decision makers and how they take in information, even the 

format of the information, can affect how rapidly they learn and improve and their 

ultimate performance (Atkins and others 2002). There are strong parallels between 

DDM principles and the principles involved in high performing organizations and 

increasing safety in complex and demanding tasks such as wildland fire fighting 

(Weick and Sutcliffe 2001; Black and others 2008).

Dynamic decision making theory describes how people learn to make reasoned 

decisions on complex problems with only partial data within constrained time periods. At the heart of it, it is about improving learning about complex situations. Why 

is learning important, and how is it related to climate change, forests, fish, and fire? 

Because adaptation is about learning. Evolution, a quintessential example of adaptation, is strictly about learning and encoding the information into genetic material 

(e.g., Williams 1966). Research helps us understand the kinds of tools people can 

best use to learn efficiently and ultimately make better decisions. Application of 

the ideas to climate change has captured improvement in agricultural performance 

at different time scales from different sources of information (Risbey and others 

1999). Any visit to a modern wildfire incident command also shows an example of 

bringing in multiple sources of information to increase the reliability, efficiency, 

and degree of control offered by fire management operations. By expanding on 

concepts already applied in fire management and including information and relationships relevant to aquatic ecosystems, better decisions can be made for fires as 

well as in pre-fire planning (Dunham and others 2003; Rieman and others 2010).

Application to problems with fish, forests, fire, and climate change would suggest several key strategic components. The recent history of research on dynamics 

in natural systems has already taught us the value of diversification as a long-term 

structural/strategic approach in developing resiliency (e.g., Dunham and others 

2003 for fishes). For shorter time scales where human intervention may be necessary, several information resources require some development. Better information 

about the conditions, distribution, and utilization of habitats for aquatic species is a 

key piece, including information about riparian canopies and stream characteristics. 

In essence, a better inventory of the aquatic/riparian habitat and species is needed. 

In comparison, we currently have substantial information about upland forest condition. More information about the temporal variability, at both interseasonal and 

interannual time scales, for precipitation, temperature, snowpacks, stream flow, 

and stream temperature will be critical in forecasting habitat changes. The lack of 

precipitation information at higher elevations (e.g., Mote and others 2005), where 

most of the projected habitat for the most sensitive species lies, makes it difficult 

to identify relationships that would help predict the most resilient habitats.

Anticipating changes over both long (decades) and short (weeks to months) time 

scales will be helpful. While the use of GCM projections in estimating future climates is an obvious, if fuzzy, tool, there may also be utility in data that can provide 

season-ahead forecasting to support fire management. Since we know that fire will 

be a critical mediator of climate associated impacts to forest and aquatic systems, 
110 USDA Forest Service RMRS-GTR-290. 2012.

improving management and direction of fires, both intentional and unplanned, may 

ultimately be one of the most important tools for mitigating the potential impacts 

of climate change. Spring time information about fine spatial scale snowpack and 

soil moisture information can also help make planned fuel treatments more successful (Holden and Jolly 2011). Avoiding or repairing short term acute impacts 

may be valuable in managing a network of small refugia. This means that there is 

value to climate data for forest managers, in a way that may be parallel to farmers 

and others whose livelihood depends on weather (Risbey and others 1999; Hansen 

and others 2011).

Although there is a need to improve information resources about habitats and 

climate to improve control, much of the general approach already outlined for 

activities is aligned with general expectations of climate change. Current management interventions should continue to be the common sense application of habitat 

realignment to mend disruptions from historical management that was focused 

on generating wood products. Probably the key step to take at this time, however, 

is gathering information on habitats, stands, streams, and aquatic stocks to understand variable sensitivity to the effects of climate. Managers, specialists, and 

scientists should continue to anticipate new issues and new constraints (maybe 

new opportunities) that may require action to conserve special habitats. Perhaps 

the greatest management focus will be determining priorities. The rare components 

with greatest value will likely be afforded management attention and resources. 

If we use current management to encourage heterogeneity in landscapes, we can 

allow vegetation and other habitat changes to occur in some places while focusing management efforts on those locales and circumstances that are least sensitive 

to climate change. Within this context, however, we should avoid current refugia 

that are at limits of distributions and invest in the places that will serve as future 

refugia.

Finally there is a need for courageous leadership. Courage is needed because 

challenging goals must be pursued, with conviction. Learning is only accomplished in the context of a challenge. As Brehmer (1990) noted, if we feel we are 

doing acceptably well, there is no incentive to undertake the additional effort to 

collect and analyze data for improved decisions. If we only go so far as to envision 

an “acceptable” future, we may be fortunate enough to realize just that. Only if 

we envision a future where we have diverse species representation in many of the 

same general areas it is now, are we likely to obtain an outcome resembling that 

future. To accomplish such a task is a substantial challenge; to accomplish it within 

reasonable financial constraints will be an even greater challenge.
USDA Forest Service RMRS-GTR-290. 2012. 111

Next Steps

Fire is an important ecosystem process for forests, riparian areas, and streams. 

It is an agent of renewal and redistribution, and the biota that live with it have 

adapted a fine balance between the strain on individuals and local populations and 

the benefits that flow from renewal. Despite the use of the word “disturbance” 

to describe fire as an event, we need neither classify fire as “good” or “bad”; it 

need only be acknowledged as part of the diversity in nature. There are biota and 

ecosystems that depend on fire, and they do not often occur in places without fire. 

While the austere aesthetics of a recently burned forest are usually considered to be 

an acquired taste, there is a well noted appreciation for the beauty and simplicity of 

the landscapes, species, and ecosystems shaped by fire.

Biota have learned about fire through evolution. The signature of fire is encoded in their DNA and some of the resulting species, like lodgepole pine and 

ponderosa pine, have unique morphologic features that are classroom examples of 

fire adaptation. Many species have also learned to cope with the inconsistency and 

unpredictability of fire through development of a diversity of phenotypes adapting in slightly different ways. Some characteristics that are adaptive to fire may or 

may not have evolved in response to fire, but their representation in populations is 

reinforced by fire. For example, migratory life behaviors in fish have other major 

evolutionary benefits, like productivity and fecundity, but those benefits come with 

the cost of higher risk of predation or other incidents during migration. In so far as 

fire provides disbenefit to species that do not migrate, it selects for those species 

that do.

Much of that learning is now being put to the test. The way that fire operates in 

the landscape, as defined by relationships between frequency, size, and severity, is 

changing. Those relationships are fairly direct outcomes of weather, and trends of 

warming and drying in the spring and summer lead generally to an expectation for 

greater frequency, size, and severity of fires. Although global circulation models 

cannot model the effect well, other theoretical support and historical observations 

remind us that interannual variability is changing, and we are likely to see more 

extremes. The warmest driest summers are becoming warmer and drier even more 

so than average summers. The big and the severe fires will still be an outcome of 

extremes in weather, just as they are now. Shifts in the driest summers and the 

hottest days will be the most informative to predicting changes. These kinds of 

conditions conspire not just to produce the most flammable fuels, but also to make 

the atmosphere the most unstable and most prone to strong local winds.

Ecosystems are being affected directly by changes to climate as well, and some 

of the changes may reduce the effectiveness of natural adaptations to fire. Stream 

temperature changes, for example, may push thermally sensitive fishes into smaller, more isolated habitats that are then more vulnerable to post-fire debris flows or 

droughts. Trees and other plants more stressed by heat and water deficits are less 

able to fend off disease, pests, and invasive species, increasing their vulnerability 

to fire as well.

Fire will be one of several agents through which climate will change ecosystems, and losses in individuals and changes to microclimate associated with fire 

could be the final step in some local extirpations. For populations and ecosystems 

of conservation concern, the outcomes of fire will become more and more important. Many of the historical and recent challenges for land managers around 

fire have been prompted by the interaction of land management with wildfire and 
112 USDA Forest Service RMRS-GTR-290. 2012.

its effects (Miller and Urban 2000; Bisson and others 2003; Hessburg and Agee 

2003; Rieman and others 2010). The challenge for managers now is to blend their 

understanding of fire under natural variation (highlighting natural resistance and 

resilience) with the consequences of land management activities on responses, and 

an awareness of how climate change will further alter both the dynamic fire events 

and the response for forest and stream ecosystems.

Uncertainty about future conditions and events may seem like the greatest impediment to developing adaptation approaches. Although many consequences of a 

changing climate are fairly certain, like temperature increases and the subsequent 

effects on snowpack declines, earlier springs, and drier summers, precipitation related consequences are less certain, with uncertain sign (wetter versus drier) in 

some places. Precipitation is the largest term in the water balance, and uncertainties 

in precipitation have major implications with respect to potential adaptation, such 

as reservoir operation or expansion (Barnett and Pierce 2008, 2009; Rajagopalan 

and others 2009). These uncertainties also have substantial impact on future fire 

occurrence and size (Holden and others 2012). Substantial uncertainty surrounds 

how changes in extremes will transpire, as they are not well represented in GCM 

simulations. If we see more extreme events, that may affect ecosystems profoundly 

even if the average conditions change little. Besides these general uncertainties 

about future climates, there are uncertainties related to individual disturbance 

events. Although there is a growing acceptance that many kinds of forest disturbance are more likely, there remains the substantial uncertainty for a particular 

location about if and when, and then how big and how severe.

Preparing for climate change has two primary dimensions: preparing the landscape and preparing the managers. An important tactic for promoting landscape 

stability in the form of improved resilience and resistance is restoration of management-derived impacts. There are also tactics oriented toward anticipation of 

future conditions, such as thinning to adjust to future water balances. Even mild 

anticipatory actions yield some controversy, and stronger forms of facilitation (like 

introducing species in further north areas) are not widely endorsed in the literature.

Far, far less has been said about preparing managers or agencies for climate 

change. Ironically, people may be the most adaptable part of the system—at least 

in so far as managers recognize themselves as important and effective agents within ecosystems. Predictability is a key issue for adaptation. In so far as we can really 

see what climate change is going to do with some accuracy, we can start to prepare 

the landscape itself. If we are wrong, though, we are really just meddling or tinkering. For example, preparing for reduced water budgets does not prepare for greater 

variability, where one to a few wet years can undo efforts to reduce vegetation only 

to have thickly vegetated stands exposed to extreme dry conditions.

Given the large uncertainties in long-term climate projections at regional to 

sub-regional scales, an important strategic concept is responding intelligently and 

in a timely manner. This does not mean sitting around and waiting for something 

to happen! Effective responses result from accurate anticipation. Even if we cannot anticipate the details of climate and disturbance processes 40 years from now, 

there are some envelopes we can draw. As the distance to the time window for 

projection decreases, the envelope of possibilities grows narrower in some dimensions. Although uncertainty for particular events and locations will exist up to the 

moment lightning strikes or a debris flow happens, we have a general recognition 

that they will ultimately happen in many landscapes.

Natural resource management agencies have dealt with this nature of uncertainty for as long as most have existed. Fire suppression organizations provide several 

idealizations and lessons that can be applied to the context of land management 
USDA Forest Service RMRS-GTR-290. 2012. 113

in a changing climate. Over the decades of professional wildland fire suppression 

development, there have been some general indicators of performance that are 

somewhat universally accepted: not losing important places to fire and not losing 

people to fire or accidents. To carry out their tasks better and more safely, they 

have developed weather and fire behavior forecasting skill at both seasonal and 

incident time scales (e.g., the National Fire Decision Support Center). They have 

also identified the critical points in their landscapes, analyzed how to defend them, 

and, where necessary or beneficial, actually prepared the places that put them or 

their charges at highest risk. While there is an impression that fire managers mostly 

do their work during fires, it is only when they are most visible. There is also a 

great deal of time preparing gear and equipment, preparing themselves, preparing 

forecast models, and preparing the critical locations in the landscape.

Among major disturbance agents (insects and disease, fire, climate change), fire 

stands out as being most potentially influenced by human decisions because we 

can, up to a point, choose when and where fires burn. We can choose, for example, 

to allow a fire to burn or we can attempt to suppress it. While successful suppression eliminates potential short term ecological impacts, it may defer the “external” 

costs of fire management until a wildfire burning under more extreme climatic conditions occurs that is difficult or impossible to suppress. Given general uncertainty 

about how a fire is likely to burn and the potential post-fire effects or “severity” of 

a fire, it is not surprising that managers choose to suppress many fires. In response 

to the short-term uncertainty fire managers face, fire, ecology, and hydrologic researchers are building new tools to rapidly provide managers with information 

about risks on relatively short time scales. For example, predictions of the potential 

for fire spread that account for topographic influences on snowpack and fuel moistures may be used to better manage wildfires in regions of complex terrain (Holden 

and Jolly 2011). Having spatial information on fuel moistures available to the fire 

spread models reduces the uncertainties, and therefore the costs of choosing not to 

suppress when conditions actually allow such a choice (see text box on modeling 

topoclimatic variation).

Natural resources management, as a profession, has developed an eye toward 

the long view. The general principles of cyclic dynamics in ecosystems have long 

been understood and have been used as an underpinning for developing scientific 

ecological management of forest ecosystems. Death and regrowth over time scales 

longer than human lifespans and over landscapes of millions of acres are not new 

ideas at all, but formed the conceptual basis for the conservation movement well 

over a century ago. It paralleled a realization that forests were not just a place to 

take wood from once and move on, but places for cultivation of sustainable goods 

and services. The profession has thoroughly embraced principles of robustness 

in management, exploring stability through resilience and resistance for various 

aspects of ecosystems and coupled social and economic systems (reference the 

Sustained-Yield Forest Management Act of 1944 from about the same time as 

Smokey Bear Act). It may sound anathema to suggest that we turn to management 

techniques associated with short time scale management, but the lessons of the 

“surprises” associated with climatic change over the last few decades has been that 

long term persistence and conservation of species may, at some point, consist of 

surviving from one decision or one crisis period to the next. The learning model 

of the wildland firefighting community may provide a good example of how to 

improve other land management approaches in a dynamic and partially unpredictable future.
114 USDA Forest Service RMRS-GTR-290. 2012.

Modeling Topoclimatic Variation to Improve Wildland Fire Management Decision Support

Regional climate is mediated at finer scales by topography, particularly in areas of mountainous terrain like the western US.

Three primary topographic features influence topoclimate. Surface air temperatures generally decrease with elevation as a

function of the environmental lapse rate. Incident solar radiation (governed by aspect and shading by adjacent mountains) is

generally greater on south vs. north facing slopes, increasing air temperatures, decreasing humidity and increasing the surface

temperature of coarse woody debris and soils. The difference in radiation loading on south vs. north facing slopes influences

the timing of snowmelt, which is translated into variation in fuel moistures and soil moisture deficits later in the season.

Nighttime air temperatures are strongly influenced by topographic dissection. Depending on atmospheric conditions, Cold air

drains downhill, accumulating in valley bottoms and basins. Nighttime air temperatures in valley bottoms can be 10 degrees

celsius colder in valley bottoms than areas above the valley floor. Topographic position also effects wind speeds and valley

bottoms tend to be sheltered compared to ridgetops. Wind has a strong influence on evapotranspiration, because it strips still

boundary layer air from the surface of leaves, fuels and soil, accelerating drying.

Topoclimate

Figure T10-1. A simplified framework for

incorporating snowmelt timing and topoclimatic

variation into models of fuel moisture and fire

danger indices

Figure T10-3. Average minimum temperature and

maximum humidity across the 2009 fire season

combined with delayed snowmelt timing lead to

complex spatial patterns In fuel moisture and

Energy Release Component (ERC) 

Figure T10-2.

Modeled snowfree date by

aspect. Notice

the early snow

departure on

west and southfacing slopes 

Can We Use High Resolution Topographically-Informed Fire Danger Models To Aid in Management of Wildland Fires?

Topoclimatic variation has a huge influence on the evolution of fuel moistures during the fire season. Delayed snowmelt timing and

cooler, wetter conditions on North slopes would limit fire spread. By taking advantage of topographic variation in fuel moistures and

fire danger in areas of complex terrain, we could take advantage of wildfires early in the fire season and allowing more fires to burn

while using terrain as a fire break. Importantly for aquatic ecosystems, we see that the combination of cold air drainage and high

nocturnal relative humidity in valley bottoms leads to much lower fuel fuel moistures at peak fire danger in valley bottoms. This

suggests that some riparian areas could be buffered from severe fire behavior and effects by local topoclimatic variation.

Mean snow off date by aspect

Penman-Monteith ET/

Energy Release

Component (ERC)

Snow melt model

(or MODIS snow)

High resolution daily

climate/weather inputs

Tmln,

Tmax,

RH,

Preclp)

Solar

Wind

First snowfree day

National

Weather

Service 7 day

forecast

Gridded fuel moisture/fire

danger and water balance,

accounting for variation in

aspect and elevation

Modeling topoclimatic influences on fuel

moistures/water balance Figure T10-2. 

Modeled snow-free 

date by aspect. 

Notice the early 

snow departure on 

west- and southfacing slopes.

Figure T10-3. Average minimum temperature and maximum humidity 

across the 2009 fire season combined with delayed snowmelt timing 

lead to complex spatial patterns in fuel moisture and Energy Release 

Component (ERC).

Figure T10-1. A simplified framework for incorporating 

snowmelt timing and topoclimatic variation into models of fuel 

moisture and fire danger indices.
USDA Forest Service RMRS-GTR-290. 2012. 115

The suggestion for the next steps in adapting to climate change, and particularly 

with reference to robust management of forests and streams under dynamic conditions, is for natural resource managers to do what they are already doing a little 

better. Most real world decisions have time pressure stemming from some degree 

of unexpectedness. A few examples:

• A large beetle kill occurs; should salvage harvest be done or not?

• A flood washes out a road; should it be restored, decommissioned, or 

abandoned as is?

• A local population of an endangered species disappears after a fire/debris 

flow, and the creek goes dry or a severe heat wave passes. Reintroduce, 

wait, abandon?

• Invasive brook or rainbow trout are found in a native population of cutthroat 

or bull trout. Take action or ignore?

• A new dam is proposed or new reservoir operating rules are proposed. 

Propose conditions?

• A new mine is proposed; is it consistent with species conservation plans?

Most readers should recognize these as day-to-day business for a natural resources management agency, and many may also recognize that these kinds of 

decisions will increase in number or complexity as climate change advances. At 

the same time, almost every one of these issues is handled as a separate emergency, 

with a separate wind-up time, separate meetings, and separate teams to handle it. 

An important, introspective question, is to reflect on how well prepared you are to 

handle such questions on any part of the domain for which you have responsibility 

to help with these questions. Could you really handle most of these well and within 

the time constraints imposed by natural processes or other agencies? If you are a 

line officer without detailed technical knowledge, do you know you can, or just 

believe? How can you check your answer?

Reconciling the long view with the short term is mostly about growing a commitment to having no more “surprises.” Events and decisions like the ones listed 

above are commonplace at the scale of the western United States, but may be so 

rare on a given agency field unit that experience and data are limited for those who 

must ultimately write the technical analysis documents. As a consequence, valuable 

time is lost to building expertise and collecting the new relevant data. While timber inventories are pretty common (and supported by the USFS Forest Inventory 

Analysis Program), information about riparian forests and aquatic species distributions is rare, piecemeal, and incomplete (see, for example, trout distribution data 

used by Wenger and others 2011a; Wenger and others 2011b). The ability to make 

rational (and defensible) decisions quickly about salvage opportunities in the absence of such information is severely impaired. By the time necessary information 

is collected, the opportunities may have expired, and decisions without information could lead to appeals with an equivalent result. We could complain in defense 

that the fire, or outbreak, or other event was a surprise ... but in a profession dedicated to the long view, that answer is not, ultimately, satisfying.

Not having “surprises” is not about perfect prognostication; it is about recognizing the potential for many different events or outcomes to happen, possibilities 

rather than expectations. For land and aquatic resource managers, it may mean 

needing to consider and evaluate sensitivities to events that may not happen during 

their tenure. It also means keeping time and room in their budgets to prepare for 

and handle increasingly common events. While the business of the fire-fighting is 

essentially contracted out from the perspective of a field management unit, other 
116 USDA Forest Service RMRS-GTR-290. 2012.

events like the ones listed above and various other consequences are handled locally. Being nimble and prepared for decisions and action in response to a range 

of disturbances across districts and forests will consist largely of having current 

knowledge of the resources to be managed, so that when the time frame becomes 

tight, the best decisions can be made.

Geographic information is probably the most common information need identified in the preceding chapters. For example, for prioritizing forest and stream 

restoration efforts several maps are necessary:

• Mapping of fish habitat networks and distributions

• Mapping fuel status and potential fire severity based on fuel loading

• Mapping riparian vegetation cover and status

• Mapping of current aquatic habitat conditions

• Mapping road sediment and fragmentation issues

• Mapping streamflow patterns (e.g., flood timing, low flow magnitudes) and 

sensitivity

• Mapping stream temperature patterns and its climatic sensitivity

• Mapping the debris flow risks

The value of geographic information for decisions on fire and forest fuel management is well recognized and substantial outcomes for providing the information 

to managers were realized in the Landfire project (http://www.landfire.gov, and 

see Textbox on Mapping Future Fire Severity). The success of this kind of project 

relates to a strong need in a major program (fire and fuel management), persistent 

crises related to drought and fire occurrence, major advances in remote sensing 

science in the preceding decades, and a program dedicated to improving and collecting forest inventory (USFS Forest Inventory Analysis). Developing a parallel 

information base for aquatic and riparian management will benefit from an increasing capacity for remote sensing of stream and riparian characteristics (see, 

e.g., McKean and others 2008; McKean and others 2009 for geomorphic information; Isaak and others 2010 for stream shade information). Inexpensive data 

collection of stream temperature data may provide a wealth of spatial data across 

large areas, which will be particularly important with respect to climate change 

impacts (Isaak and others 2010; Rieman and Isaak 2010). Also, new models are 

emerging that simulate the genetic and habitat connectivity of fishes (Landguth 

and others 2011). Unfortunately water, riparian, and aquatic biota are not the primary missions of any land management agency, nor do they carry the same broad 

sense of urgency or crisis that fire does. The challenge will be in articulating the 

benefits of rapid access to both terrestrial and aquatic information for more punctual and better decisions.

Fine scale mapping of the heterogeneity of forest and aquatic habitats is a growing enterprise that will be useful in improving decisions as well. Data from GCM 

outputs, even some of the higher resolution downscaled projections now available, 

are inconsistent with the scale at which land managers ultimately carry out treatments. Tree harvest, thinning, and regeneration occurs at the scale of stands or 

hillslopes, yet most projections of tree species range shifts are made at 1 km scale 

or greater, ignoring the interactions between topography and climate and their influence on species occurrence. For example, valley bottoms, where streams occur, 

often experience remarkably different climates than areas upslope because of nocturnal cold air drainage, shading from adjacent terrain, and sheltering from wind. 

Such microclimates can be intensively sampled using inexpensive thermistors 
USDA Forest Service RMRS-GTR-290. 2012. 117

Mapping the Probability That Fires Will Burn Severely

Assessing the potential ecological effects of large fires is crucial to effective management before, during, and 

after fires. Great tools exist for assessing post-fire effects, including the burned area reflectance classification 

(BARC) maps produced by the Burned Area Emergency Rehabilitation (BAER) using satellite imagery (Parsons 

and others 2010) and field methods. Managers also need tools for easily and quickly forecasting the potential 

severity of future fires. Now, managers can use the maps, models, and other tools developed by scientists 

with the FIRESEV project (http://www.firelab.org/research-projects/fire-ecology/128-firesev) funded by the 

Joint Fire Science Program to aid management decisions at multiple planning stages, including pre-fire fuels 

treatments and strategic management of active fire incidents.

We can map where future fires are most likely to burn severely based on statistical models of how climate, 

fuels, topography, and current season climate and fire weather influenced burn severity for fires that 

burned 1984-2006. Dillon and others (2011b) analyzed more than 7000 fires mapped by the Monitoring 

Trends in Burn Severity project (www.mtbs.gov) to develop statistical models based on past fires. They 

are currently extrapolating those models across each of 15 different ecoregions across the western United 

States to predict the likelihood of high severity fires in areas not yet burned. Dillon and others (2011b) are 

working with a very large random sample of pixels and an approach called Random Forests, thus expanding 

the methods Holden and others (2009) used to map the probability of severe fire occurrence based on 

topography and vegetation on the Gila National Forest. The resulting dynamic predictive maps will soon 

be available with 30-m spatial resolution for all lands across the western United States. Accuracy will be 

assessed based on field data and more recent fires. Though they will be imperfect, these spatially explicit 

predictions about the ecological effects of current and future fires can be incorporated into decision 

making. Managers will be able to access these maps directly off the Internet for immediate use. Or, they 

can use other predictive modeling tools, including the Wildland Fire Assessment Tool (WFAT, www.niftt.

gov) with local data to predict the potential for severe fire. WFAT provides an interface between ArcGIS, 

FlamMap3 (http://www.firemodels.org/index.php/national-systems/flammap) algorithms and First Order 

Fire Effects Model (FOFEM, http://www.firelab.org/science-applications/fire-fuel/111-fofem) algorithms to 

produce predicted fire behavior and fire effects map layers. This tool is useful to managers in prioritizing 

fuel treatments on the basis of predicted fire behavior and effects and in assessing the effectiveness of fuel 

treatment proposals in a geospatial context. Dillon and others (2011b) expect that the map layers and related 

tools can also be incorporated into existing decision support frameworks such as the Wildland Fire Decision 

Support System (WFDSS, https://wfdss.usgs.gov/wfdss/WFDSS_Home.shtml) and the Rapid Assessment of 

Values at Risk (RAVAR, http://www.fs.fed.us/rm/wfdss_ravar/ravar/index.shtml).

In the process, scientists are learning why fires burn more severely in some places than in others. Crossregional analyses of trends in burn severity are now possible, thanks to the MTBS data. Dillon and others 

(2011a) think that while area burned is greatly affected by climate (Littell and others 2009), local topography 

and fuels are relatively more important to the ecological effects of those fires, though this varies across 

vegetation types and ecoregions. Indeed, Dillon and others (2011a) found that topography strongly 

influenced burn severity. Including climate improved predictions some and more so for those years when 

fires were widespread and extensive areas burned. Climate and weather as “top down” influences on 

wildland fire (e.g., through fuel moisture, temperature, wind) that affect where and how fires burn at a broad 

scale (Dillon and others 2011a). In contrast, topography and fuels are “bottom up” controls that interact with 

climate and weather to alter fire behavior and effects locally. Topography is often a strong driver of general 

vegetation distribution, which in turn influences the distribution of fuels and patterns of severity.

Through the Fire Severity Mapping System, managers can access and create map products when and where 

they need them. By integrating LANDFIRE data layers, fire effects models, and new techniques for analyzing 

satellite-derived burn severity data into one comprehensive computer modeling package, managers can 

predict fire hazard and fire severity over a wide range of spatial scales. The system includes the digital 

maps as well as simulation models and other analysis tools to support decision making during wildfires, for 

targeting post-fire rehabilitation efforts, and for long‐term planning. This complements existing tools widely 

used. 
118 USDA Forest Service RMRS-GTR-290. 2012.

(< $20 USD), to build models relating broader climate patterns to local temperature 

patterns (Holden and others 2011). Further information on terrestrial ecosystems at finer scales will develop from the use of Light Detection and Ranging 

Data (LiDAR) to map the three-dimensional spatial distribution of aboveground 

biomass (Hudak and others 2006) over heterogeneous topography. Distributed 

temperature measurements along fiber optic cables allow the identification of upwelling groundwater creating thermal refugia in streams (Selker and others 2006). 

Understanding the role of fine scale heterogeneity in the landscape process is an 

emerging field of work that will better inform how micro-habitats influence the 

potential presence of species in an area and how fire will spread over a landscape.

Doing a better, more efficient, job of natural resources management will also 

include increasing use of seasonal scale climate forecasts to program activities 

and task resources. Such forecasts are commonly made based on knowledge of 

the states and anticipated states of climate phenomena such El Niño, the Pacific 

Decadal Oscillation, or the Atlantic Multidecal Oscillation. Forecasting at seasonal 

scales (e.g., winter to summer) has been embraced by the fire community and 

has been applied to positioning for upcoming fire seasons (Wells 2007). Water 

resource managers and farmers are also common users of seasonal forecasts. The 

importance of seasonal-scale weather forecasting is not as well recognized in either 

terrestrial or aquatic ecosystem management. Because the information is already 

being generated and provided, however, the diligent resource manager may want 

to consider ways in which it could improve the bottom line. Some examples may 

include:

• A forecast for a wetter summer would imply relatively low risk of later 

extreme fire danger when deciding not to suppress a fire early in the 

summer.

• Summers with greater impending fire risk might be a time to prioritize 

having field crews available to assess riparian or aquatic conditions after 

fires.

• Those summers will likely also need some effort in preparing post-fire 

management plans.

• Wet summers are a good time to prepare to gather the information that will 

be used in the dry summers.

• Understanding the triggers that water managers will have for altering 

reservoir schedules and helping guide appropriate or provide monitoring of 

migratory fish to ensure their success.

• An impending flood-prone winter may trigger a response to survey culvert 

conditions and perform the kinds of temporary maintenance that can reduce 

flood impacts.

• In places with distributed small and isolated habitats, an impending fireprone summer may trigger a response to survey current conditions and 

make any short term preparations (e.g., collect individuals for offsite 

protection).

Any skill developed in using seasonal scale forecasting to improve work flow, 

productivity, and conservation successes will translate into improved skill in using 

long-term climate projections as well.

The informational perspective has been a long-standing foundation of conservation biology. Aldo Leopold (1966) described keeping all the parts as the “first 

precaution of intelligent tinkering.” It has been more formally described as retaining genetic representation and diversity across the landscape. Retaining redundant 
USDA Forest Service RMRS-GTR-290. 2012. 119

examples is particularly important in the context of disturbance. The principle 

operating behind this is that evolutionary adaptation is a learning process. The 

learning occurs as a range of “hypotheses” are tested against the environment they 

are in, leaving behind a range of comparably suitable outcomes. Learning in this 

way can only occur from among the phenotypes presented; thus species with more 

limited diversity have less capacity to learn in the short term. Random innovation through mutation is a possibility, but learning is much slower than if relative 

advantages exist from within existing genetic makeups. Essentially, species with 

a greater diversity of phenotypes are more likely to carry information already that 

allowed them to survive in analog climates or circumstances in the past.

The details may vary but one fundamental principle emerges from the discussion:

The distinction between winners and losers in a changing climate will largely 

hinge on who has the best information.
120 USDA Forest Service RMRS-GTR-290. 2012.

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Advanced Topics on Fish 

Populations and Fire
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Fire and Fish: a Synthesis of Observation and Experience

Bruce Rieman1, Robert Gresswell2, and John Rinne3

1USDA Forest Service, Rocky Mountain Research Station (retired), Seeley Lake, MT

2U.S. Geological Survey, Northern Rocky Mountain Science Center, Bozeman, MT

3USDA Forest Service, Rocky Mountain Research Station (retired), Flagstaff, AZ

Introduction

The effects of wildfire on aquatic systems and fishes occurring in them has been 

linked to the direct or immediate influence of the fire on water quality and the indirect or subsequent effects on watershed characteristics and processes that influence 

water quality and quantity, stream channels, and aquatic biota (Gresswell 1999). 

Early research linking fire and aquatic systems focused on effects to soil, erosion, 

and water yield and quality, with a relatively limited temporal and spatial context 

(see Gresswell 1999 for a review). Wildfire generally was perceived as a destructive force threatening aquatic resources (e.g., fish populations) and related values 

(Rieman and Clayton 1997; Kaufman 2004).

Subsequent efforts have integrated physical and ecological processes related 

to fire, and these concepts have been temporally categorized as short-term (i.e., 

<1 year), mid-term (1-10 yr), and long-term (10s to 100s of years) (e.g., Minshall 

and Brock 1991; Minshall and others 1998, 2001; Mihuc and Minshall 2005). 

Concomitantly, spatial context is critical because heterogeneity in fire severity, 

stream or watershed characteristics, and ecological communities constrains subsequent events and ecological responses. A broader perception includes the role 

of wildfire as a fundamental agent of disturbance potentially shaping heterogeneity, diversity, and productivity in aquatic ecosystems (Reeves and others 1995; 

Gresswell 1999; Bisson and others 2003).

The effects of fire and fire related management have been of particular importance to those interested in, or responsible for, management of native fishes. Fish 

and associated fisheries often hold particular social and economic importance. The 

sometimes dramatic short-term effects of fire and postfire disturbance on stream 

channels, water quality, and mortality of individual organisms can be readily apparent. As a result, attempts to influence fire and its effects on aquatic systems and 

fish populations before and, particularly, during and after the fire, have consumed 

considerable resources, time, and energy, and engendered substantial debate (e.g., 

Dunham and others 2003; Rieman and others 2003; NMFS 2007; Rhodes and 

Baker 2008; Rieman and others 2010).

It is our intent to summarize the known effects of fire-related processes in forested biomes of the western United States, briefly review existing knowledge 

regarding direct and indirect effects to fish, and consider the implications for fish 

populations. We integrate earlier summaries (e.g., Gresswell 1999; Dunham and 

others 2003; Rieman and others 2003) with more recent information and conclude 

with a final synthesis, including implications for conserving or restoring the resilience of fish populations to wildfire.
160 USDA Forest Service RMRS-GTR-290. 2012.

Fire Effects

Direct Effects

Wildfires consume flammable materials, and in the process, shade and cover 

provided by vegetation and woody debris is altered (Gresswell 1999). If a severe 

fire burns near or across a stream, water temperature can increase substantially 

(Hitt and others 2003). Dissolution of smoke, ash, and volatile compounds can 

alter pH and concentrations of trace metals, nutrients, and other chemical constituents in streams (Cushing and Olsen 1963; Minshall and others 1989; Spencer and 

Hauer 1991; Earl and Blinn 2003; Spencer and others 2003).

Direct physical effects can produce mortality of aquatic organisms including fishes, amphibians, invertebrates, and periphyton (Rieman and others 1997; 

Gresswell 1999). The causes of mortality have not been definitively identified, but 

potential mechanisms include rapid increases in temperature and accumulation of 

toxic chemicals (e.g., ammonium, trace metals, and cyanide) (Minshall and others 

1989, 1997; Spencer and Hauer 1991; Barber and others 2003).

The observed spatial pattern of direct effects seems to depend on the extent and 

severity of fire and size of the watershed (Gresswell 1999). Minshall and Brock 

(1991), for example, suggested that fire would not directly influence temperature 

in third order or larger streams. Furthermore, high variability in postfire distribution of fishes in small streams appears to be associated with the heterogeneity of 

riparian fire severity (Rieman and others 1997). Mortality or displacement of fishes 

may be extensive, extending for several kilometers, but effects are often incomplete or patchy within and among stream reaches influenced by fire (Minshall and 

Brock 1991; Rinne and Neary 1996; Rieman and others 1997).

More recently, Howell (2006) found evidence of high to complete mortality or 

displacement of steelhead (Oncorhynchus mykiss), Chinook salmon (O. tshawytscha), and brook trout (Salvelinus fontinalis) in several moderate to severely burned 

reaches of three small (<2.6 m width) streams in eastern Oregon. Fish persisted at 

normal densities in stream reaches immediately downstream of those experiencing 

more severe fire. Based on observations from streams in northwestern Montana, 

Jakober (2001) reported that direct effects of several large fires were observable 

the following year. For example, the majority of fish in a drainage located in northwestern Montana were killed during a fire in 1996, and bull trout appeared to 

have been extirpated from headwater reaches of two small streams (Jakober and 

Dentino 2003). In contrast, fish distribution and density were relatively unaffected 

1 year after fires that burned in Lolo National Forest during 2000 (Jakober 2002). 

In the southwestern United States, direct mortalities of fish appear to be less common than those following subsequent hydrologic events (Rinne and Jacoby 2005).

Direct effects may also result from fire suppression activities and the use of fire 

retardants (Gresswell 1999). Sodium ferrocyanide, a component of a commonly 

used fire retardant, is known to be toxic to aquatic organisms (Little and Calfee 

2002; Pilliod and others 2003; Angeler and Moreno 2006), but the temporal and 

spatial extent of these effects have been poorly understood. In fact, Crouch and 

others (2006) recently found no evidence that retardant increased the levels of any 

chemical constituents above those from wildfire alone. The authors suggested that 

ferrocyanide was elevated through pyrogenic sources (Crouch and others 2006). 

Nonetheless, the Forest Service eliminated use of retardants with sodium ferrocyanide in 2007 because of the potential toxicity to aquatic organisms (USDA 2007). 

The replacement, known commercially as PhosChek, was less toxic in laboratory 
USDA Forest Service RMRS-GTR-290. 2012. 161

studies (Little and Calfee 2003). The Forest Service concluded there was little 

risk for direct mortalities from PhosChek under prescribed operational use (Comas 

2007; USDA 2007), but that conclusion remains controversial (NMFS 2007).

Indirect Effects

The indirect effects of wildfire on streams will be influenced and constrained 

by direct effects, the subsequent supply of materials (e.g., water, sediment, woody 

debris, nutrients, biological propagules) and energy (insolation), stochastic events 

(e.g., storms), and the interactions of these factors on physical and biological succession (Gresswell 1999; Dale and others 2005; Milner and others 2007). We 

provide only a brief overview focused on water, sediment, wood, stream temperature, and stream food webs as an example of the dynamic, and often complex, 

nature of the interactions influencing habitats for fishes.

Water and Sediment.–Changes in supplies of water and sediment are commonly 

observed after wildfire (Legleiter et al 2002). Short- or mid-term changes reflect 

alteration of vegetation and soils associated with severity and extent of the fire, 

geology and geomorphology, and the duration, intensity, and timing of postfire 

precipitation (Swanson 1981; Rinne and Neary 1996; Wondzell and King 2003). 

Peak flows and erosion can increase substantially, but the magnitude of response is 

related to factors pertaining to the watershed and weather (Robichaud and others 

2005). For example, hydrophobic (water repellant) soils have been linked to floods 

and increased erosion in some cases, but effects are influenced by fire severity, soil 

texture, and vegetation type. Mass failures, debris flows, and flood catalyzed by 

fire have been important in the history of many watersheds and may be primary 

drivers in the long-term sediment supply for those systems (Reeves and others 

1995; Benda and others 2003; May and Gresswell 2003; Meyer and Pierce 2003; 

Moody and Martin 2009). Dramatic increases in erosion that follow some fires 

tend to decline within 10 years as vegetation is reestablished (McNabb and others 

1989; Burton 2005; Luce 2005; Robichaud and others 2009); however, the process 

may extend as root strength fails in fire-killed trees. The fire-hydrologic interaction 

has been characterized as an episodic pattern of disturbance and recovery that contributes to important variation of stream conditions in space and time (e.g., Reeves 

and others 1995; May and Gresswell 2003; Miller and others 2003).

Temperature.–Water temperatures commonly increase following fire, sometimes by several degrees, but increases are not universal (Gresswell 1999; Dunham 

and others 2007). Postfire increases in water temperature have been associated 

primarily with loss of forest and riparian shading (Gresswell 1999; Isaak and others2010), but channel simplification, topographic shading, hyporheic flow, and the 

hydrologic changes accompanying fire can both accentuate and ameliorate these 

changes (Amaranthus and others 1989; Dunham and others 2007). Subsequent declines in stream temperature have been predicted with recovery of riparian canopy 

(Rieman and Clayton 1997), but recent work in the Boise River basin suggests that 

elevated temperatures can persist for one to two decades in some cases (Dunham 

and others 2007). Changes might even become permanent if fires initiate a transition to new vegetation communities associated with climate change (Isaak and 

others 2010).

Woody Debris.–Wildfire can play a key role in the recruitment of woody debris 

to streams (Reeves and others 1995; May and Gresswell 2003; Miller and others 2003). Wood is important because it controls channel morphology, sediment 

and water routing, and the heterogeneity of structure that may be important habitat for fishes and other organisms (Keller and Swanson 1979). In contrast to the 
162 USDA Forest Service RMRS-GTR-290. 2012.

fire-related hydrological changes that can attenuate within 10 years, changes in 

the recruitment of woody debris can extend for decades (Reeves and others 1995; 

Gresswell 1999; Scheidt 2006). Postfire accumulation of wood is related to the 

prefire forest, severity of the fire, and processes associated with wood routing and 

storage in the channel (May and Gresswell 2003; Scheidt 2006). Woody debris 

may even decline following fire if a large proportion of the vegetation (including 

instream debris jams) is burned completely, or if remaining wood is transported 

out of the system during periods of elevated discharge (Swanson and Lienkaemper 

1978). Mid-term accumulation may occur with toppling of fire-killed trees, undermining of riparian trees with increased flow and bank erosion, or the transport 

of wood from upslope via debris flows or avalanches (Swanson and Lienkaemper 

1978; Scheidt 2006). Longer-term recruitment may depend on the rate that mature 

trees develop. The interaction of processes controlling the supply, accumulation, 

transport and storage of woody debris can lead to substantial variability within and 

among individual streams, or across time (e.g., May and Gresswell 1993; Young 

and others 2006). The capacity of a stream system to store and continually rework 

stored materials, however, may also lead to relatively stable or uniform conditions 

in some systems (Scheidt 2006).

Food Webs.–A handful of studies before 1989 provided a limited perspective 

on wildfire and responses in biological communities and food webs in streams 

(Gresswell 1999). In general, it is clear that fire and subsequent effects can disrupt 

invertebrate communities in segments of small streams, and that at longer time 

scales, effects depend on the severity and extent of the fire, subsequent hydrologic disturbance, and the characteristics and recovery of riparian vegetation. There 

is a substantial body of literature that initially defined the anticipated responses 

of stream communities and food webs to the extended effects of wildfire (e.g., 

Minshall and others 1989, 1997, 1998, 2001, 2004; Robinson and Minshall 1996; 

Minshall 2003; Earl and Blinn 2003; Mihuc and Minshall 2005). The generalized 

prediction emerging from this work can be summarized as an ecological succession in a temporal frame (Minshall and others1989; Gresswell 1999; Dunham and 

others 2003; Minshall and others 2004).

The short-term effects of postfire disturbances are related to the biophysical 

template following the fire, and responses can be dramatic, varying from virtually 

undetectable to the complete loss of invertebrates and algae. Where disruption is 

great, invertebrate and algal communities are often recolonized quickly, but abundance and diversity may continue to vary (Earl and Blinn 2003; Minshall and 

others 2004).

In the mid-term (1-10 yr), diversity in aquatic communities tends to increase and 

productivity can be high with increased sunlight, stream temperatures, and nutrient 

flux. Where the riparian canopy is substantially reduced, foodwebs are expected 

to shift from largely allochthonous to autochthonous sources of carbon, with a 

concomitant response in trophic guilds of macroinvertebrates (or shifts feeding 

strategies by individual species; Mihuc 2004) and detrital respiration. Recent research has provided additional evidence that severe wildfire can stimulate primary 

production and a shift to primary consumers that support a greater biomass of 

predatory insects, potentially fish, and even consumers in linked riparian communities (Malison 2008; Malison and Baxter2010).

In the longer term (10-300 yr), watershed processes are anticipated to interact 

with succession in terrestrial and riparian vegetation. Influx of wood generally 

leads to organic litter accumulation. Trophic pathways are anticipated to shift 

from autochthonous toward allochthonous carbon sources as riparian vegetation 

increases and the canopy closes (Minshall and others 1989). Recent research has 
USDA Forest Service RMRS-GTR-290. 2012. 163

generally supported these predictions and extended the perspective (Minshall and 

others 2004; Robinson and others 2005), but changing climate may constrain future riparian communities leading to unanticipated outcomes.

The long-term effects of wildfire on macroinvertebrate communities and food 

webs remain speculative, largely because direct evidence from empirical studies of 

that temporal extent simply does not exist. The results clearly depend, however, on 

the myriad interactions between disturbance, terrestrial succession, and watershed 

process that directly and indirectly influence the legacy of materials and linkages 

within, and between terrestrial and aquatic systems (Dale and others 2005; Milner 

and others 2007, 2008; Pettit and Naiman 2007).

Fish Population Responses

Stream environments can change quickly during or following a fire (e.g., a single 

flood event) and may be catastrophic in ecological terms (including local extirpation of organisms in individual stream reaches). Concomitantly, indirect effects of 

wildfire often initiate or constrain processes, and responses can extend for decades 

and even centuries. Wildfire clearly plays an important role in supply of food and 

materials and the heterogeneity of channel conditions that contribute to the mosaic 

and productivity of habitats (Reeves and others 1995; May and Gresswell 2003; 

May and Lee 2004). Although direct and indirect effects of wildfire can induce fish 

mortality, the long-term consequences for fish populations and assemblages will 

ultimately depend on the legacies of material, biota, and the associated physical 

and ecological processes that shape them. Understanding these changes ultimately 

requires a population-level perspective.

The response of fish populations to the direct and indirect effects of wildfire has 

been a focus of considerable interest in the last two decades. Because fires are hard 

to predict, most of the early research was opportunistic. In some cases, wildfires 

burned in watersheds where pre-existing population data facilitated before-after 

comparisons (e.g., Novak and White 1990; Rinne and Neary 1996; Rieman and 

others 1997), but inference was also gained from comparisons among burned and 

unburned streams, or the temporal trajectory of populations associated with the 

immediate effects of fire (Minshall and Brock 1991; Rieman and others 1997). 

Initial research focused principally on description of changes in distribution and 

abundance of extant populations or segments of populations (Rieman and others 

1997; Gresswell 1999), but subsequent syntheses considered a broader context of 

population and ecological processes (Rieman and Clayton 1997; Dunham and others 2003; Heck 2007).

Although historic information suggested that direct effects of fire could produce substantial fish mortality (Minshall and Brock 1991; Rinne and Neary 1996;

Rieman and others 2007), we know of no examples of population extirpation associated with immediate effects of wildfires. In general, population level implications 

of wildfire appear to depend on longer-term processes.

Reductions in abundance, contraction in distribution, and even local extirpation have been reported with the indirect effects of large fires (Bozek and Young 

1994; Rinne and Neary 1996; Gresswell 1999; Rieman and others 1997; Rinne and 

Carter 2008). Extreme effects, including local extirpations, have most often been 

observed in the southwestern United States (Rinne and Neary 1996). Negative 

consequences are related to diminished water quality associated with postfire ash 

flows, loss of habitat connectivity during periods of drought and intermittent flows, 

violent postfire flooding, and loss of food base and habitat (Rinne 2003, 2004; 

Rinne and Carter 2008). Fish populations in the Southwest may be especially 
164 USDA Forest Service RMRS-GTR-290. 2012.

vulnerable because the spring and early summer fire season is followed by monsoon in late summer (Rinne 2004). When heavy rains follow severe fire, and heavy 

ash or “slurry flows” have been reported (Rinne and Neary 1996).

The potential for a severe disturbance is coupled with the fact that many of 

remaining native fish populations in the Southwest are limited to small streams 

isolated from any source of immigration (Brown and others 2001). In fact, many 

fishes in the region have been formally listed as threatened or endangered under 

ESA, and therefore, there is substantial concern that wildfires could negatively 

influence persistence. In some cases, rescue efforts have been initiated to capture surviving fish during, or immediately following, a fire and move them to a 

secure environment (e.g., hatchery) until habitat conditions in the burned watershed stabilized (Brooks 2006; A. Unthank, USDA Forest Service Regional Office, 

Albuquerque, NM personal communication).

Despite the potentially deleterious consequences, population collapses have not 

been ubiquitous in the aftermath of even very large fires. In some cases, fire had 

little apparent effect (Gresswell 1999; Riggers 2001; Jakober 2002). In others, 

abundance was seriously depressed, and reach-level extirpation occurred; however, populations rebounded relatively quickly (e.g., 1-6 years; Novak and White 

1990; Rieman and others 1997; Jakober and Dentino 2003). The emerging context 

of disturbance ecology and metapopulation dynamics and documented population 

resurgence resulted in a hypothesis of recovery strongly mediated by the expression of migratory life histories and by dispersal from refugia not influenced by the 

fire and subsequent events (Rieman and Clayton 1997; Rieman and Dunham 2000; 

Dunham and others 2003).

Recent research has sustained this view. In an extensive study of 32 watersheds that burned in large fires on the Boise National Forest between 1986 and 

1994, Burton (2005) found that even where reach or broader-scale extirpation 

of fishes had occurred, recolonization through dispersal was complete. Howell 

(2006) reported rapid (within 4 years) recovery of rainbow and steelhead populations through immigration from habitats not influenced by postfire disturbances. 

Debris flows actually provided new habitat for Chinook salmon by destroying a 

culvert and promoting access to new areas above the former barrier (Howell 2006). 

Similarly, fire-flood generated debris fans in the Middle Fork Salmon River created new habitat used by spawning Chinook salmon within months of the events (R. 

Thurow USDA Forest Service, Rocky Mountain Research Station, Boise Idaho, 

personal communication of unpublished data). Examples of rapid responses now 

extend beyond the North American continent, and recovery of native fishes within 

24-36 months of extirpation following extensive postfire sediment flows has been 

documented in Australia (Lyon and O’Conner 2008).

In fact, evidence that fishes exhibit resilience to fire is unmatched by recovery 

from anthropogenic disturbance (Neville and others 2009). Dunham and others 

(2007) contrasted two levels of postfire disturbance (burn-and-debris-flow and 

burn-only) with unburned streams in the Boise River basin, but severity of disturbance did not have substantial influence on the distribution of either rainbow trout 

(O. mykiss) and tailed frogs (Ascaphus truei) within 10 years following fire. It appeared that even if effects from wildfire (e.g., elevated summer water temperature) 

continued for a decade or more, human activities that reduce the capacity of organisms to respond to disturbance may be a greater threat to persistence than wildfire 

(Dunham et al 2007). Similarly, Neville and others (2009) found no evidence of 

genetic bottlenecks associated with fire related disturbance in 55 streams affected 

by the Boise fires, but genetic diversity decreased in relation to human caused migration barriers (i.e., impassable road culverts).
USDA Forest Service RMRS-GTR-290. 2012. 165

The potential spread of nonnative species facilitated by fire-related disturbance 

is another important issue for some dwindling native fish populations (Dunham 

and others 2003). Because stream temperature can affect the relative distribution 

and interaction of native and nonnative species (e.g., McHugh and Budy 2005; 

Rieman and others 2006; Benjamin and others 2007; McMahon and others 2007), 

it has been suggested that warming of streams affected by wildfire may lead to 

expansion of nonnative species (e.g., brook trout or brown trout Salmo trutta) and 

concomitant contraction of native species, like bull trout S. confluentus and cutthroat trout (Dunham and others 2007; Isaak and others 2010). Results of a recent 

study in the Bitterroot River watershed where bull trout and westslope cutthroat 

trout co-occur with nonnative brown trout, brook trout, and rainbow trout suggested that native salmonids were resilient to fire (Sestrich and others 2011). Although 

nonnative trout invaded 4 of 17 stream sections within the burned area, the invasion rate was similar in sections that did not burn. Habitat conditions declined, but 

native trout populations actually exceeded pre-disturbance levels 3 years following the fire (Sestrich and others 2011).

To this point, the emphasis has been on dispersal and migration as key mechanisms behind population-level responses of fishes to fire, but research examining 

sublethal effects of habitat changes provides some evidence that phenotypic plasticity and adaptation to the effects of fire are mechanisms that may contribute 

to population resilience as well. In one recent study, habitat characteristics and 

growth of coastal cutthroat trout (O. clarkii clarkii) were evaluated in severely 

and moderately burned watersheds and an unburned control (Heck 2007). Stream 

temperature increased following the removal of riparian vegetation during the fire, 

and growth of fish in burned watersheds was greater than in the unburned control. 

Increased water temperatures and a longer growing season may have triggered a 

bottom-up response resulting in an increased growth of cutthroat trout, but longevity decreased (Heck 2007). Apparently, survival in these headwater streams was 

limited by the maximum size of the individual trout, and faster growing individuals reached the critical point at an earlier age. In a study of nine tributaries of the 

Boise River (i.e., three each of reference, burned, and burned with channel reorganization), there was a direct relationship between growth and maturation rate, and 

level of disturbance 10 years following fire (A. Rosenberger, University of Alaska, 

personal communication of unpublished data). Such adaptation could be critical to 

persistence of some isolated populations (Letcher and others 2007).

Summary, Implications, and Conclusions

Wildfire can have dramatic effects on streams and on populations of native fishes. Mortalities have been associated with the direct effects of severe fires, but these 

consequences have been most frequently observed in relatively small streams and 

over limited extent (1-2 km). Indirect effects linked to the physical and ecological 

process occurring after a fire can extend over 10s to even 100s of km of stream, 

however, and effects may last for decades, and potentially even centuries. Wildfire 

and associated disturbances have been, and likely will continue to be, a proximate 

cause of extirpation for small populations of native fishes. In general, however, the 

effects of wildfire are far less dramatic. In some cases, the effects of wildfires have 

been difficult to measure, and in others, populations that were initially depressed 

have rebounded dramatically, even increasing in abundance or extent relative to 

prefire conditions.
166 USDA Forest Service RMRS-GTR-290. 2012.

Gresswell (1999) suggested that over much longer time scales (decades to 

millennia), fish were well adapted to disturbances occurring at the frequency of 

wildfire. Recent evidence suggests that effects of disturbance in individual habitat 

patches are greatest on individuals and local populations that are least mobile, and 

early speculation to that effect (Warren and Liss 1980; Rieman and Clayton 1997; 

Gresswell 1999) has been supported by recent findings. Because habitat fragmentation is directly linked to the ability of fish to move among disparate portions of a 

stream network, this factor alone may be much more critical to the persistence of 

western salmonids than disturbance related to wildfire. There is a growing body of 

evidence that watershed characteristics including size, complexity (number and arrangement of tributaries), connectivity, and management history play a major role 

in the population structure and resilience at multiple levels of biological organization across disturbance-prone landscapes (Wofford and others 2005; Neville and 

others 2006; Guy and others 2008).

Clearly, context shapes the effects of wildfire on populations and communities 

of native fishes. Fish are most vulnerable to both direct and indirect effects of fire 

where populations are restricted to relatively small areas of habitat, and risk is 

greatest in isolated stream segments or small networks in steep, confined drainages where severe fires are likely to burn a large proportion of the headwaters and 

riparian corridor. Where populations are relatively large, have access to diverse, 

well-connected habitats and/or the capacity to adapt to changing environments, 

vulnerability is lessened; in many cases, the capacity and even the productivity of 

habitat can even be improved following wildfire (e.g. Malison and Baxter 2010).

The differences in context are evident among individual populations and watersheds across the landscape. Some populations within a basin may be highly 

vulnerable to any form of disturbance, but others are much less so (e.g., Guy and 

others 2008; Peterson and others 2008; Neville and others 2009). Populations of 

Chinook salmon, bull trout, and other species that migrate over large expanses of 

habitat in the Pacific Northwest and northern Rocky Mountains have fared well in 

response to the major fires of recent decades. Managers in these systems may tend 

to worry less about wildfire than the more chronic effects of watershed development. They may even see wildfire as beneficial for creating or expanding new 

habitats (e.g., Howell 2006). In contrast, managers concerned with remnant populations of native fishes isolated by invasive species and dewatered stream channels 

in the central Rocky Mountains and the Southwest may view the next wildfire as 

the catalyst for local extirpation, or even species level extinction (e.g., Brown and 

others 2001; Rinne 2004; Brooks 2006).

Large disturbances will likely continue in all of these regions. The watersheds of 

the Rocky Mountains and interior west have been shaped by disturbances linked to 

wildfire, large storms, and the flux, routing, and storage of water (in groundwater, 

snow and ice), sediment, and wood. On evolutionary time scales, large fires, mass 

erosion events, floods, and even local extinctions have likely been common (e.g., 

Bennett 1990; Reeves and others 1995; Meyer and Pierce 2003), and yet many 

species and populations of fishes have persisted and adapted (Gresswell 1999). 

Management might alter the frequency and magnitude of particular disturbance 

processes, but it likely cannot eliminate large and intense watershed events (e.g., 

Kirchner and others 2001; Istanbullouglu and others 2002). In fact, frequency and/

or magnitude of some disturbances may even increase with changing climate (e.g., 

McKenzie and others 2004; Westerling and others 2006). Furthermore, the extent 

of thermal habitats for coldwater salmonids and the connectivity among refugia 

are expected to decline, further increasing the vulnerability of some populations 
USDA Forest Service RMRS-GTR-290. 2012. 167

(Rieman and others 2007; Williams and others 2009; Haak and others 2010; Isaak 

and others, 2010).

So how can managers respond to the challenges posed by changing fire regimes 

and the conservation of native fishes and diverse aquatic ecosystems? In general, 

options can be categorized as management before, during, and after fire (Dale and 

others 2001; Dunham and others 2003), but a framework for monitoring and adaptation is critical in all cases (Dunham and others 2003). Management before 

fire includes maintaining or restoring the resistance or resilience to disturbance 

before the next disturbance occurs. Conditions contributing to resistance and resilience in populations emerge repeatedly in the discussion above and elsewhere 

(e.g., Gresswell 1999; Dunham and others 2003; Bisson and others 2003; Rieman 

and Isaak 2010). Conservation or restoration of relatively large networks of habitat 

and the physical and ecological processes that maintain them, the broad expression and adaptation of life histories, and the potential for connections within, and 

among populations are all critical. In essence, resilience will depend on a spatial 

and temporal structure and diversity in populations and habitats that can absorb or 

benefit from the effects of fire.

In some cases only remnant habitats exist and reconnection or expansion to support resilience and adaptive potential is not possible or desirable (e.g., the potential 

invasion of nonnative species). The only alternative may lie in attempts to change 

the character of disturbance. For example, by reducing the extent or grain of severe 

fire, even small populations may have some chance of persisting. Considerable 

discussion has focused on the management of fuels through prescribed fire or thinning as a tool for altering the severity of fire, if not its extent (Reinhardt and others 

2008; Rieman and others 2010). Continued debate about the potential utility of 

such efforts is related to the tradeoffs between the potentially damaging effects of 

management relative to the effects of the fires (Rhodes and Baker 2008; Rieman 

and others 2010) that may occur with, or without that management (Littell and 

others 2009).

If the conditions supporting the resilience and adaptation of populations cannot 

be created before a large fire occurs, managers are left with limited options. They 

might attempt to suppress or mitigate effects during or after the fire. Although 

extraordinary measures could make the difference between extinction and persistence (e.g., Brown and others 2001), there is some question whether postfire 

watershed remediation can even influence the kinds of events that might actually 

threaten populations (Backer and others 2004). Regardless of their efficacy, these 

actions can be extraordinarily expensive and necessarily will be limited in the extent or number of populations, streams, or watersheds that might be considered. 

Concomitantly, there are other activities (e.g., salvage logging) that can be detrimental to particularly vulnerable populations in the postfire environment (Beschta 

et al 2004; Karr and others 2004; Reeves and others 2006).

Clearly, better information is needed to understand where and how the conditions in forests, watersheds, and native fish populations transition from those that 

may benefit from and are resilient to the effects of wildfire, to those that are vulnerable to those effects. At what point is a population too small or too isolated to 

persist? Can we actually influence the character of fire to alter that balance? How 

do we weigh the threats and benefits of aggressive management? Peterson and others (2008), Dare and others (2009), and Rieman and others (2010) offer some tools 

that may help refine this discussion and offer some approaches that can begin to 

answer these questions. Further collaboration between researchers and managers 

could extend this process. In the interim, the basic concepts of resilience outlined 

repeatedly throughout this review provide a critical foundation for the future.
168 USDA Forest Service RMRS-GTR-290. 2012.

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Genetic Variation Reveals Influence of Landscape Connectivity 

on Population Dynamics and Resiliency of Western Trout in 

Disturbance-Prone Habitats

Helen M. Neville1, R.E. Gresswell2, and J.B. Dunham3

1 Trout Unlimited, Boise, ID

2 U.S. Geological Survey—Northern Rocky Mountain Science Center, Bozeman, MT

3 U.S. Geological Survey—Forest and Rangeland Ecosystem Science Center, Corvallis Research Group, Corvallis, 

OR

Introduction

Salmonid fishes have evolved and persisted in dynamic ecosystems (Waples 

and others 2008) where disturbance events vary in frequency, magnitude, timing, 

and duration (Gresswell 1999; Dale and others 2001), as well as the specific nature 

of associated effects (e.g., changes in thermal or flow regimes, geomorphology, 

or water chemistry; Reeves and others 1995; Benda and others 2004; Bisson and 

others 2005). In the western United States, one of the major drivers of disturbance 

in stream ecosystems is fire (Reeves and others 1995; Rieman and Clayton 1997; 

Gresswell 1999). Although there is a growing consensus that fish populations 

can ultimately benefit from the productive and heterogeneous habitats created by 

fire (Reeves and others 1995; Benda and others 2003; Minshall 2003; Rieman 

and others 2003), to persist they obviously have to withstand the immediate and 

shorter-term effects of fire, which can reduce or even extirpate local populations 

(Rieman and Clayton 1997; Brown and others 2001; Burton 2005; Sestrich 2005). 

Movement among interconnected stream habitats is thought to be an important 

strategy enabling persistence during and following fire, and there is mounting 

concern that the extensive isolation of salmonid populations in fragmented habitats is reducing their resiliency to fire (Gresswell 1999; Dunham and others 2003; 

Rieman and others 2003).

In spite of this concern, there are few direct observations of salmonid responses 

to fire. In fact, guidance is based largely on a broader understanding of the influences of landscape structure and disturbance in general on salmonid fishes (Dunham 

and others 2003; Rieman and others 2003), and there is considerable uncertainty 

about how best to manage for salmonid resilience to wildfire. Studies are limited 

by the difficult logistics of following fish responses in the face of unpredictable 

events such as wildfires. Therefore, BACI (Before-After-Control-Impact) study 

designs are nearly impossible, and replication is similarly challenging because 

fires are often low-frequency events. Furthermore, conventional ecological study 

approaches (e.g., studies of fish distribution, abundance, life histories, and movement) are logistically difficult to implement (but see Dunham and others 2007). 

Overall, a major challenge to understanding resilience of salmonid populations in 

fire-prone environments is related to moving beyond localized case studies to those 

with broader applicability in wildfire management (Dunham and others 2003).

Genetic data can be useful for overcoming many of the limitations inherent in 

ecological studies (Neville and others 2006a; Schwartz and others 2006). Here 
178 USDA Forest Service RMRS-GTR-290. 2012.

we review several case studies of western trout where population genetic data 

have provided insight about fish responses to fragmentation and disturbance more 

generally, and specifically in relation to fire. Results of these studies confirm the 

importance of movement and landscape connectivity for ensuring fish persistence 

in fire-prone landscapes, and highlight the usefulness of genetic approaches for 

broad-scale evaluation and monitoring of population responses to fire and related 

management actions.

General Effects of Fragmentation on Trout Populations

We begin with a brief review of several mechanisms by which habitat fragmentation may reduce the likelihood of persistence for salmonid populations. General 

consequences of isolation include increased risk of extinction from random demographic and genetic processes due to small population size (McElhany and others 

2000), reduced potential for re-colonization or repopulation following disturbance 

(“metapopulation dynamics”; Gilpin and Hanski 1991), and decreased habitat diversity and “internal dynamics,” which can affect population persistence (Pickett 

and Thompson 1978). Isolated populations are typically smaller, and so are likely 

to have greater temporal fluctuations or other stochastic effects that increase variation in individual reproductive success and lead to smaller effective population 

sizes (Waples 1990, 2002) and frequent population bottlenecks (Neville and others 2006a). The consequent loss of genetic diversity from stronger genetic drift 

and increased frequency of inbreeding in these small isolated populations lowers 

adaptive capacity (Bijlsma and Loeschcke 2011) and can increase the risk of extinction (Frankham 2005). Additionally, in smaller, simplified habitat patches fish 

cannot move to use a full complement of alternate local habitats that might bolster recruitment, provide refugia, or facilitate metapopulation dynamics (Dunning 

and others 1992; Schlosser 1995; Ebersole and others 2001; Ebersole and others 

2003). Small populations are thus more vulnerable to physical disturbance because 

of the increased likelihood that a single event will affect the entire population. 

Finally, isolation of western trout populations has greatly reduced the occurrence 

of a migratory life history (Young 1995; Fausch and others 2009). Where present, migratory individuals using alternate habitats can boost reproductive capacity 

(Jonsson and others 2001; Morita and others 2009) and potentially buffer populations from environmental variability and disturbances, such as fire (Dunham and 

others 1997; Rieman and others 1997b; Dunham and others 2003; Neville and 

others 2006a).

Genetic Characterization and Monitoring of Populations

Because fragmentation and disturbances such as fire are expected to affect 

the genetic characteristics of populations, genetic data can be used as an indirect 

measurement of these processes. In fact, for many questions (e.g., measurements 

of abundance or dispersal rates), genetic assessments requires less time and less 

expense than traditional methods such as mark-recapture and radio telemetry 

(Schwartz and others 2006). Additionally, genetic approaches can provide information about temporal dynamics or cryptic influences that may not be captured by 

traditional methods. In such cases, evaluations of population characteristics are often more accurate than traditional methods (Neville and others 2006a). For instance, 

population fluctuations or founder events following fires may not be reflected in a 
USDA Forest Service RMRS-GTR-290. 2012. 179

current census if a population has recovered demographically (Dunham and others 2007), but these events may have had significant long-term effects on effective 

population size and can be captured using genetic approaches (e.g., they might 

be expected to show evidence of population bottlenecks, particularly if they are 

isolated). As another example, there is growing concern that invasion of nonnative 

trout and hybridization with native species may become more prevalent following 

fire-related disturbances (Dunham and others 2003). Detecting hybridization in 

the field can be difficult, but hybridization can be assessed with confidence using 

genetic techniques (Boecklen and Howard 1997). Genetic approaches should be 

particularly valuable for capturing impacts from fire or other disturbances on natural populations, then, because samples can be collected across large landscapes 

(Manel and others 2003), and a sample collected at one point in time can provide a retrospective view of influences that may have affected a population over 

decades (a “snapshot,” but one that provides information about past influences). 

After an initial assessment, genetic data can then be collected periodically to monitor changes in population characteristics through time (Schwartz and others 2006).

Applications of Genetic Data for Evaluating Effects of Fragmentation 

and Disturbance

Using genetic diversity as an indicator, several recent studies of trout populations demonstrate how movement and life history variability are related to 

population resilience in fragmented and disturbance-prone landscapes. Neville 

and others (2006b) evaluated genetic characteristics of Lahontan cutthroat trout 

Oncorhynchus clarkii henshawi residing in high-desert stream habitats vulnerable 

to different types of disturbances including fire. The study area encompassed one 

of the last interconnected and relatively pristine stream networks in the current 

range of this fish. Because some streams in the network were degraded and isolated by barriers, this system provided a rare example of diverse movement and 

life history variation in a complex habitat that could be contrasted with population characteristics in isolated habitats within the same catchment. The mainstem 

river, a hypothesized migratory corridor connecting several tributaries, was found 

to maintain high genetic diversity and contain a mixture of genotypes representing 

fish from different streams that likely moved into mainstem river habitats when 

not spawning in the tributaries. In contrast, samples from physically connected 

but high-elevation tributaries where fish were thought to express a ‘resident’ life 

history showed moderate levels of differentiation from other samples and lower 

genetic variability within samples, confirming a certain degree of demographic 

and geographic segregation from migratory forms (Northcote and Hartman 1988; 

Varley and Gresswell 1988; Northcote 1992). Fish from habitats isolated by culverts or natural barriers had the lowest effective population sizes and levels of 

genetic diversity, often exhibited severe genetic bottlenecks, and were highly 

differentiated from other populations (patterns observed in other trout and charr 

populations isolated above barriers, Taylor and others 2003; Yamamoto and others 

2004; Wofford and others 2005; Morita and others 2009). In one case, a population 

above a small natural waterfall that was thought to have been extirpated by drought 

(Dunham 1996) was later recolonized. Therefore, fish from interconnected larger 

habitats in this study seemed to maintain spatially segregated life history forms and 

functioning metapopulation dynamics, both of which are likely to contribute to the 

overall viability of this network of populations in a harsh desert environment (see 
180 USDA Forest Service RMRS-GTR-290. 2012.

Rieman and Dunham 2000). Populations in isolated habitats exhibited negative 

genetic effects of isolation, likely resulting from the small and demographically 

unstable nature of these populations. It is not surprising that many isolated populations across the historic range of the subspecies have disappeared in recent decades 

(Elliott and others 1997).

Other studies have also highlighted the importance of habitat connectivity and 

complexity in disturbance-prone environments. Disturbances such as landslides 

and debris flows are common throughout the range of coastal cutthroat trout 

Oncorhynchus clarkii clarkii. In western Oregon, however, these fish exist in 

several ecoregions differing significantly in landscape attributes such as stream 

gradients, topology (arrangement of tributary branching), and connectivity, all of 

which might influence the effect of disturbances on local populations (Guy and 

others 2008). For example, the interior Cascades ecoregion is characterized by 

steep streams with little branching structure, but the Coast Range ecoregion is 

lower gradient and has highly dendritic stream systems with few instream barriers to movement. Accordingly, Guy and others (2008) found patterns of genetic 

diversity to reflect these landscape characteristics in each ecoregion. The authors 

had hypothesized that Coast Range populations would show genetic resilience to 

disturbance, given that a disturbance in any single stream would affect only a portion of a dendritic system, where the capacity for dispersal from other parts of 

the network was high. As expected, they found the Coast Range populations exhibited greater genetic diversity than populations from the Cascades ecoregion, 

where landslides and debris flows in steep single-channel streams likely affected 

whole populations directly with little possibility for subsequent gene flow. Other 

patterns related to the physical distances separating habitats suggested that genetic 

structure in the Coast Range ecoregion were driven by within-watershed dispersal, 

whereas in the Cascades ecoregion genetic drift and disturbance-associated population bottlenecks seemed to be the dominant factor influencing genetic patterns. 

The study design could not differentiate between influences of disturbance regimes 

versus historical influences related to colonization of the two ecoregions; however, 

Guy and others (2008) suggest that the effect of disturbances may be greater in 

habitats with reduced connectivity and complexity where fish have little ability 

to respond to these events by movement (see also Gresswell 1999). These results 

should inspire further investigation into these connections.

Another recent genetic study of 55 stream populations of rainbow trout 

(Oncorhynchus mykiss) in the Boise and Payette river basins of Idaho suggested 

similar influences of isolation on population resilience in the face of disturbance 

(Neville and others 2009). This study was designed specifically to compare 

the effects of wildfire and related disturbance and habitat fragmentation, and 

incorporated the natural variability in fire history across these watersheds over the 

last several decades. The authors compared genetic diversity in samples without 

a history of recent wildfire to those with a history of stand-replacing wildfire, as 

well as those that had experienced both wildfire and a severe channel-reorganizing 

disturbance that may have extirpated or greatly reduced fish populations. Stream 

habitats also varied in size (catchment basin area) and isolation caused by humanconstructed road culverts. Surprisingly, there was no evidence that fish from streams 

with different fire histories had different levels of genetic diversity; overall, fish 

in watersheds that had experienced wildfires and associated channel reorganizing 

events maintained similar levels of genetic variability as fish in streams with 

no known disturbance. Genetic diversity did increase, however, with increasing 

habitat size, and was lower in populations found above culvert barriers. Recent 

demographic assessment of a subset of these trout populations sheds detailed light 
USDA Forest Service RMRS-GTR-290. 2012. 181

on the rapid recolonization (Rieman and others 1997a; Burton 2005; Dunham and 

others 2007) and flexible life history strategies that enable populations to recover 

quickly after fire in interconnected stream networks (Dunham and others, this 

volume). Furthermore, results of genetic analyses suggest that human influences 

such as barriers to dispersal may pose greater threats to populations of native trout 

than wildfire (Neville and others (2009).

In total, these recent studies yield substantial empirical evidence of a linkage 

between landscape structure and resilience as indicated by genetic diversity in 

salmonid fishes. Because genetic diversity can be used to characterize connectivity and population dynamics (Neville and others 2006a), these relationships 

underscore the importance of connectivity for both the persistence and long-term 

viability of salmonids (see McElhany and others 2000; Dunham and others, this 

volume). The genetic examples presented here support the growing consensus that 

salmonid fishes are able to withstand disturbances, including fire (Waples and others 2008), if given a habitat template for expression of various movement and 

life history strategies (Schlosser and Angermeier 1995; Hendry and Stearns 2004; 

Dunham and others this volume).

Management Implications

Given the link between stream network connectivity and population resilience, 

an obvious management strategy would be to reconnect trout habitats wherever 

physically possible to allow dispersal processes to occur naturally. But management decisions today are complicated by wide-spread invasions and hybridization 

with nonnative fishes which, it has been hypothesized, may be facilitated following fires (Dunham and others 2003). Although examples in this paper suggest that 

strategies promoting the isolation of populations above artificial barriers may not 

be successful, short-term isolation may be the only alternative for preserving the 

integrity of native populations in cases where invasion is certain (Fausch and others 2009).

This issue emphasizes the usefulness of genetic data for monitoring native trout 

populations in the context of fire management. Where isolation is deemed necessary, genetic techniques have potential for evaluating the probability of persistence 

of isolated populations over time (see Yamamoto and others 2004, for an example with white-spotted charr Salvelinus leucomaenis) by assessing the effective 

population size prior to and following isolation, and evaluating the effects of any 

observed perturbations, such as fire. Genetic data are a powerful tool for monitoring threats from hybridization where native populations are at risk of invasion by 

congenators (e.g., Neville and Dunham, in press), and can be helpful in prioritizing 

management resources by evaluating the genetic integrity (i.e., whether or not they 

have already been hybridized) of above-barrier populations before barrier removals are planned and executed.

Wherever possible, therefore, information concerning the genetic characteristics of a population should be integral to assessments of trade-offs associated 

with isolation or connectivity (Kruse and others 2001; Peterson and others 2008; 

Fausch and others 2009). Undoubtedly, consideration of these trade-offs will become even more complex as suitable habitat becomes further fragmented and 

degraded in association with climate change, and the need for improving resiliency through restoring and reconnecting habitats increases (Rahel and others 1996; 

Rieman and others 2007; Williams and others 2009). Long-term monitoring of 

the genetic ‘health’ and purity of populations (Dunham and others 1999) will be 
182 USDA Forest Service RMRS-GTR-290. 2012.

highly valuable in fire-prone habitats, especially when paired with insight from 

other ecological methods evaluating movement (e.g., telemetry or stable isotope 

analysis, see examples in Balkenhol and Waits 2009; Sepulveda and others 2009). 

This type of research will continue to improve our understanding of the complex 

mechanisms by which salmonids have adapted to the effects of fire (e.g., Dunham 

and others, this volume). Furthermore, genetic research on habitat connectivity 

and complexity underscores the need to develop a broad-based forest management 

strategy that focuses on protecting remaining fish populations and habitat from 

invasion and further anthropogenic degradation, while restoring degraded habitat 

and connectivity among habitats.

Acknowledgements

We thank Bruce Rieman, Doug Bateman, Phaedra Budy and an anonymous 

reviewer for helpful comments on earlier drafts of this manuscript.

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USDA Forest Service RMRS-GTR-290. 2012. 187

Fish Life Histories, Wildfire, and Resilience—A Case 

Study of Rainbow Trout in the Boise River, Idaho

Amanda E. Rosenberger1, Jason B. Dunham2, and Helen Neville3

1School of Fisheries and Ocean Sciences, University of Alaska Fairbanks, Fairbanks, AK

2U.S. Geological Survey, Forest and Rangeland Ecosystem Science Center, Corvallis, OR

3Trout Unlimited, Boise, ID

Introduction

In this short piece we address the question of how aquatic ecosystems and species can change in response to disturbances, such as those related to the influence of 

wildfire on stream ecosystems. Our focal species is rainbow trout (Oncorhynchus 

mykiss) in the Boise River, Idaho. Rainbow trout in this system have persisted 

in the face of widespread and often severe wildfires occurring since the 1990s 

(Rieman and others 1997; Burton 2005; Dunham and others 2007).

Wildfire can lead to a variety of changes in stream environments (Minshall 

2003). In the Boise River, recent wildfires ranged from light to severe burns 

(Dunham and others 2007), leading to variable changes in riparian and hillslope 

vegetation, and in some cases initiation of major channel reorganizing events 

(Benda and others 2003; Miller and others 2003). These events often involved 

massive erosion of stream channels, sometimes throughout an entire tributary. In 

the latter case, local populations of rainbow trout were reduced to undetectable 

levels (Burton 2005). In others, a patchwork of wildfire appeared to eliminate rainbow trout from some reaches of stream, but not others (Rieman and others 1997). 

Following these wildfires, water temperatures remained elevated for several years, 

particularly for streams influenced by channel reorganization (Dunham and others 

2007; Figure 1). In spite of these dramatic environmental changes, rainbow trout 

have remained widespread in streams within the Boise River (Dunham and others 

2007). What characteristics of this species’ biology have conferred resilience in the 

face of these massive disturbances? What constraints may limit resilience?

One key to species resilience may be the expression of diverse and flexible life 

histories (Bisson and others 2009; Healey 2009; Waples and others 2009; Greene 

and others 2010). Our work on rainbow trout suggests the species is flexible in 

terms of how individuals and populations respond to variable environments, and 

that responses can be constrained by human influences that alter natural variability. The two major constraints we focus on here are invasions of nonnative trout 

(Neville and Dunham, in press) and loss of connectivity caused by stream culverts 

that block fish passage (Neville and others 2009).

Our work in the Boise River focused on smaller streams because both physical (Benda and others 2003; Miller and others 2003; Wondzell and King 2003) 

and biotic responses to wildfire may be more evident (Dunham and others 2003; 

Gresswell 1999; Minshall 2003). We recognized three broad classes of streams 

(Dunham and others 2007): 1) those lacking a recent history of wildfire; 2) streams 

with a recent history of moderate to high severity wildfire; and 3) streams in watersheds with a recent history of moderate-high severity wildfire that were followed 

by massive channel reorganization from a debris flow or severe flood. These events 
188 USDA Forest Service RMRS-GTR-290. 2012.

(fires, channel reorganization) occurred from 8-10 years prior to when we sampled 

fish, allowing for at least 2-3 generations of local reproduction.

We examined patterns of population density among these three classes of streams 

(for methods see Rosenberger and Dunham 2005; Dunham and others 2007), 

and found that density of age 1+ fish differed among the three classes of streams 

(Kruskal-Wallis Analysis of Variance, X2=5.96, df=2, p = 0.02). In streams without 

a recent history of wildfire densities of age 1+ fish were highest (0.22 fish/m2), 

those in burned streams were intermediate (0.17 fish/m2), and densities in reorganized streams were lowest (0.05 fish/m2).

The sizes of fish at different ages were also different among streams. For the 

youngest three age classes (0+, 1+, 2+), fish were smallest in unburned streams 

and largest in reorganized streams (A. Rosenberger and J. Dunham, unpublished). 

This paralleled the pattern of water temperature in these streams, with unburned 

streams being coldest and reorganized streams being warmer (Dunham and others 

2007). Warmer water temperatures could lead to faster metabolism and growth, if 

sufficient food is available (Hughes and Grand 2000). Additionally, fish in warmer streams may experience a longer growing season, and thus more opportunity 

to gain a larger size. Even though summer temperatures in the warmest streams 

exceeded levels that can be stressful to individuals, availability of suitable temperatures in spring and fall can still lead to faster growth (e.g., Tattam 2007). Faster 

growth may also be attributed to lower population densities and less intraspecific 

competition (Grant and others 1998; Ward and others 2006), especially in reorganized streams.

In parallel with growth, age at maturity was youngest in the warmer reorganized 

streams (A. Rosenberger and J. Dunham, unpublished data). A few individuals in 

Figure 1. Immediate aftermath of the 2003 Hot Creek Fire in the Boise National Forest. Pictured stream is a fishbearing tributary to the upper Middle Fork Boise River (photo credit: Joseph Benjamin).
USDA Forest Service RMRS-GTR-290. 2012. 189

reorganized streams began maturing as early as their first summer (age 0+), with 

the frequency of mature individuals increasing in the 1+ and 2+ age classes. By 

age 1+, we observed that 27% of the total number of fish sampled were mature, 

with males predominating among mature individuals (83% of total mature individuals). For the age 1+ cohort, larger males and females were much more likely 

to be mature. By age 2+ the length differences were no longer significant, perhaps 

reflecting the influences of size-selective mortality or emigration. In other words, 

fish that grew faster and matured at age 1+ may have reproduced and died, thus 

leaving behind only slower growing fish in the 2+ age class, or older fish may have 

emigrated. The pattern of growth and maturity we observed has been observed 

repeatedly in other salmonids grown in captivity (Wysujack and others 2008) and 

in the field (Olsson and others 2006; McMillan and others, in press).

We also compared genetic responses of rainbow trout to changes in habitat 

associated with wildfire in the Boise and adjacent Payette River basins (Neville 

and others 2009). We hypothesized that disturbances associated with wildfire 

and channel reorganization should lead to losses of genetic variation within local 

populations. Given that populations can potentially be extirpated by disturbances 

associated with wildfires (Burton 2005), or reduced in distribution (Rieman and 

others 1997) or abundance (see above), it is reasonable to expect that genetic drift 

may lead to losses of genetic diversity.

We surveyed genetic variability in samples of rainbow trout collected from 55 

small streams representing gradients of disturbance that paralleled our fish population surveys (Dunham and others 2007). We found that variability was similar 

among streams representing unburned, burned, and reorganized classes, as described above. In contrast, in streams that were isolated by human-caused fish 

passage barriers (impassable road culverts), we found losses of genetic variability 

were more likely, as has been found consistently in many similar studies (Neville 

and others 2006). In some locations, we found unexpectedly high levels of hybridization with nonnative cutthroat trout (O. clarkii) and hatchery rainbow trout 

introduced historically for sport fisheries, although hybridization was not related 

to fire history (Neville and others 2009; see also Neville and Dunham, in press). 

Overall, genetic variability within populations was most strongly associated 

with habitat size, with greater diversity observed in larger habitats. These results 

highlighted the importance of habitat size and connectivity, both critical factors 

influencing the resilience of fish in dynamic environments (Rieman and Dunham 

2000; Dunham and others 2003; Bisson and others 2009).

In summary, we observed responses of individuals and populations in association with influences of wildfire (e.g., increased water temperatures), and in 

association with human influences, such as fragmentation of habitat and introductions of nonnative trout. Fish in streams most dramatically impacted by wildfires 

grew faster, but matured earlier in life with some evidence for shorter overall lifespan resulting from early reproduction. In non-scientific terms, these fish appeared 

to adopt a “live fast – die young” strategy, but the genetic basis of these responses 

is unknown. More recent changes, such as loss of genetic variability and changes 

in allele frequencies due to hybridization were linked to human influences.

Implications for Species Resilience In the Face of Disturbance

In the case of responses of rainbow trout to wildfire and other human influences 

in the Boise River, we find that wildfire changes both ecosystems and species. 

We also found that human influences have fragmented habitats (e.g., movement 
190 USDA Forest Service RMRS-GTR-290. 2012.

barriers at road crossings) and constrained opportunities for rainbow trout to express its full range of life histories (e.g., because of hybridization with nonnative 

trout). These constraints appear to operate in ecological time, but ultimately may 

also constrain the evolutionary potential of rainbow trout. For example, if isolated 

populations lose substantial amounts of genetic variability or if genomes become 

compromised by hybridization with nonnative cutthroat trout these fish may have 

less capacity for evolutionary responses to their environment. Maintaining opportunities for natural expression of diverse phenotypes may be critical to long-term 

productivity, persistence, and continued evolutionary resilience of rainbow trout 

(Bisson and others 2009; Healey 2009; Waples and others 2009; Greene and others 

2010).

We found human influences within local habitats to be important, but human impacts may stem from influences or constraints operating much further downstream. 

For example, major dams in place in the Boise River and downstream in the Snake 

River have prevented migration of rainbow trout to the sea to express a “steelhead” 

life history. Steelhead trout were common in the Boise River and other nearby 

streams prior to their extirpation by these large dams (Busby and others 1996). In 

locations where migratory fish are available to recolonize empty habitats, populations can be highly resilient to disturbance. Examples include migrants within 

freshwater systems (e.g., Rieman and others 1997; Dunham and others 1997), but 

also migrants from marine ecosystems, such as salmon and steelhead (Bisson and 

others 2005; Howell and others 2006; Bisson and others 2009).

Even without a steelhead life history present, freshwater resident rainbow trout 

appear resilient in the face of wildfire, except in the smallest and most isolated 

habitats, as would be expected in theory and from the few empirical data available 

for salmonid fishes (Dunham and others 2003). Resilience is shown in the ability 

of rainbow trout to rapidly recolonize habitats following disturbance, variable responses to changing environments (e.g., growth, age at maturity), and maintenance 

of large enough numbers of adults to avoid genetic drift in the face of disturbance. 

Lower resilience is evident when other constraints imposed by humans (e.g., 

movement barriers, nonnative trout) limit the species’ ability to respond to disturbance (Bisson and others 2009; Healey 2009; Waples and others 2009; Greene and 

others 2010).

Finally, results of this work point to the importance of evaluating multiple responses of a species to wildfire. If, for example, we had relied only on ecological 

responses, we would have missed critical threats posed by intra and inter-specific 

hybridization. Furthermore, evaluating only distribution or abundance of a species 

provides only limited insights into possible mechanisms contributing to resilience. 

This has important implications for evaluating potential effects of wildfire or fire 

management alternatives on aquatic species. Whereas it is obvious that we need to 

understand more about processes that contribute to species responses to wildfire, 

studies of species responses typically address only the net results of such processes: presence or abundance. Because both of these responses can be influenced 

by multiple underlying processes, presence or abundance can offer only limited 

insight into the complex interplay of the underlying genetic, environmental, and 

demographic drivers. In conclusion, to better understand species responses in the 

face of changes in aquatic ecosystems caused by wildfire (or any other disturbance) we need to better understand the processes that constrain or contribute to 

resilience.
USDA Forest Service RMRS-GTR-290. 2012. 191

Acknowledgements

Thanks to Bruce Rieman and David Pilliod for constructive comments on an 

earlier version of this manuscript. 

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Use of trade or firm names in this manuscript is for reader information only and 

does not imply endorsement of any product or service by the U.S. Government.
USDA Forest Service RMRS-GTR-290. 2012. 195

Aquatic Species Invasions in the Context of 

Fire and Climate Change

Michael K. Young

USDA Forest Service, Rocky Mountain Research Station, Missoula, MT

Introduction

This paper focuses on the nexus among native and nonnative fishes with respect 

to fire and climate change in the western United States. Although many taxa are 

involved, I emphasize native and nonnative salmonids because these are obligate 

coldwater species that might be expected to respond strongly to fire and because 

most research has been conducted on these fishes. Also, forested lands in federal 

ownership in the western United States represent the stronghold for many native 

coldwater taxa and are the subject of fire, fuels, and water management. I caution, however, that climate change, coupled with a growing human population and 

increasing demands for certain ecosystem services (e.g., water for domestic, agricultural, and industrial use) may alter the species of concern, their interaction, and 

the geographic focus of this issue.

Overview: Nonnative Aquatic Species

In waters of the western United States, invasions of nonnative aquatic species 

are both exceptionally frequent and frequently detrimental to native aquatic species. For example, in Oregon, Washington, and Idaho, 120 nonnative forms of 

fish, crustaceans, mollusks, and amphibians have become established, and nonnative species constitute about half of the taxonomic diversity of fishes in these 

three states (ISAB 2008; Sanderson and others 2009). Declines in native species 

have coincided with the establishment of these nonnative taxa. All inland forms 

of trout and charr in the western United States—the subspecies of cutthroat trout 

(Oncorhynchus clarkii), Gila trout (O. gilae gilae), Apache trout (O. g. apache), 

and bull trout (Salvelinus confluentus)—have been proposed for federal listing 

under the Endangered Species Act or are currently listed, and nonnative species 

invasions have been identified as a cause of decline in all petitions and recovery plans. Moreover, introductions of nonnative species are expected to continue 

because of the growing globalization in trade that intentionally or inadvertently 

includes many new species (Cox 2004). Vehicle-based aquatic recreation also facilitates their spread; nonnative mussels, aquatic plants, and pathogens are being 

distributed throughout the country amid debris on boats, boat trailers, and perhaps 

even on wading boots (Gates 2007; Idaho Invasive Species Council 2007; Benson 

and Raikow 2009). In the late 19th and 20th centuries, state and federal agencies 

introduced many nonnative fishes to promote fishing, but more recently illegal 

angler-assisted transport of sportfish or baitfish is the leading cause of the spread of 

nonnative fishes (Rahel 2004). Many of these species form the basis of economically valuable and socially prized fisheries, even in remote wilderness settings 
196 USDA Forest Service RMRS-GTR-290. 2012.

(Dunham and others 2004). Consequently, the enhancement, monitoring, or control of nonnative aquatic species is often a management priority.

What Facilitates Invasions?

Although invasions of nonnative species are common, they are not necessarily 

inevitable (e.g., Fausch and others 2001). Invasion success is in part attributable to 

propagule pressure, because the number of introductions and number of individuals introduced is positively correlated with establishment (Lockwood and others 

2005). A second characteristic leading to establishment is habitat matching, in 

which the receiving environment is suitable for an introduced species throughout 

its life history. For example, many high mountain lakes provide excellent conditions for the growth and maturation of introduced salmonids stocked for angling, 

but because they lack inlets or outlets with fluvial spawning habitat, these populations decline relatively rapidly unless they are refounded by stocking. Alternatively, 

where such spawning habitats are present, nonnative trout readily establish in 

many instances and these locations serve as springboards for invasions elsewhere 

in a watershed (Adams and others 2001). This points out another aspect of successful invasions: spread. Watershed connectivity permits nonnative fishes to move 

throughout a basin, sometimes with surprising speed. Brook trout (Salvelinus fontinalis) that were passed over a migration barrier in a small Wyoming stream were 

found up to 4 km upstream within a month (M. K. Young, unpublished data), and 

summer movements of 2-3 km by this species are common (Gowan and Fausch 

1996). Iowa darters (Etheostoma exile), native to the Mississippi River basin, were 

first detected in the Yampa River in the Colorado River basin in 2003; by 2007, 

they were observed 229 km downstream in the Green River in Utah (Walford and 

Bestgen 2008).

Nevertheless, invasions of nonnative species sometimes stall or remain relatively contained even when abiotic conditions appear favorable for their advance. 

Adams and others (2002) noted that the upstream extent of brook trout in several 

Idaho watersheds had remained unchanged over two decades despite the absence 

of migration barriers. In some cases this has been ascribed to introduced populations requiring extended periods of time to overcome small initial populations, as 

well as a delayed increase in fitness as natural selection operates on the genome 

of the founding individuals or new arrivals contribute genomic diversity (Carroll 

and others 2007; Roman and Darling 2007). It has been posited that strong populations of native fishes may prevent or delay nonnative invasions, but presently there 

appears to be little empirical support for this hypothesis (Moyle and Light 1996).

Some Examples of Invasive Species and Their Effects

The responses of native aquatic species to nonnative invasions are mixed. In 

some locations, native fishes have persisted in sympatry with nonnative fishes for 

extended periods, and in others nonnatives rapidly replaced the indigenous fauna. 

These responses have sometimes involved the same native-nonnative species pairs. 

Because the effects of introductions are specific to the native and the introduced 

taxa, I focus this discussion on the effects of three nonnative trout—brook trout, 

brown trout (Salmo trutta), and rainbow trout (O. mykiss)—on cutthroat trout native to the inland western United States.
USDA Forest Service RMRS-GTR-290. 2012. 197

Brook Trout and Brown Trout

Brook trout have been linked to the loss of headwater populations of cutthroat 

trout in many portions of the inland West (Behnke 1992; Young 1995; Dunham 

and others 1999). Often, the mortality of young cutthroat trout exposed to brook 

trout is high enough to produce recruitment failure (Dunham and others 2002; 

Shepard and others 2002; McGrath and Lewis 2007). Peterson and others (2004) 

demonstrated that age-0 cutthroat trout survival was near zero in the presence of 

brook trout, and that the suppression of brook trout led to large increases in survival of age-0 and age-1 fish. Yet how brook trout replace cutthroat trout remains 

unknown. Both competition and predation have been suggested as mechanisms 

that may work in concert with the tendency for brook trout to mature at younger 

ages and have greater size-specific fecundity than do cutthroat trout (Kennedy and 

others 2003).

Brook trout, however, are not universally successful at displacing cutthroat trout 

from small streams. Whereas cutthroat trout populations east of the Continental 

Divide in Montana appear particularly vulnerable to displacement, similar populations to the west have persisted in sympatry for decades. This difference may 

be attributable to differences in winter climate (Shepard 2004; Fausch and others 

2006). This geographic difference also coincides with the prevalence of large cutthroat trout with fluvial life histories that may provide a demographic buffer to the 

presence of brook trout. And at smaller scales, the distribution of habitats serving 

as demographic sources—such as warmer water or reaches in unconfined river 

valleys (Benjamin and others 2007)—and sinks could be controlling brook trout 

establishment and spread (Adams 1999; Castric and others 2001; Petty and others 

2005).

The effects of brown trout on inland cutthroat trout are less studied, but appear 

to be similar to those of brook trout. Brown trout have been shown to be competitively superior to cutthroat trout in laboratory trials (Wang and White 1994) 

and field enclosures (McHugh and Budy 2005). In addition, de la Hoz Franco and 

Budy (2005) felt that the presence of brown trout truncated the downstream distribution of Bonneville cutthroat trout (O. c. utah) in the Logan River, and McHugh 

and Budy (2006) observed an array of subtle but significant changes in the growth 

and behavior of cutthroat trout in the presence of brown trout. The influence of 

brown trout on cutthroat trout appears to have a geographic signature; the most 

pronounced effects are on Rio Grande cutthroat trout (O. c. virginalis) in New 

Mexico (Paroz 2005) whereas there are many examples of long-term sympatry between more northerly subspecies of cutthroat trout and brown trout. An important 

element that may be facilitating coexistence at higher latitudes is that brown trout 

tend to be found in larger streams at lower elevations and do not occupy many 

tributaries used by cutthroat trout for spawning, reducing interactions between 

brown trout and juvenile cutthroat trout. This downstream-upstream partitioning of stream habitats by brown trout and other salmonids (Weigel and Sorensen 

2001) and their absence from small, high-elevation streams in the central Rocky 

Mountains (Bozek and Hubert 1992; Rahel and Nibbelink 1999) is thought to be 

related to their poor recruitment at low water temperatures (Jensen and Johnsen 

1999; McHugh and Budy 2005). Because brown trout eggs and fry are also vulnerable to high flows during the latter part of incubation through the early stages 

of emergence (Lobón-Cerviá and Mortensen 2005), the timing of peak discharge 

in these environments may also present an obstacle to more-upstream invasions.
198 USDA Forest Service RMRS-GTR-290. 2012.

Rainbow Trout

Rainbow trout from coastal streams in the western United States have been 

widely introduced into inland waters originally containing cutthroat trout (Behnke 

1992). Despite that these species tend to spawn at different times—cutthroat trout 

usually spawn after peak flows and rainbow trout before them—and in different 

places—cutthroat trout more often use small tributaries (De Rito 2004; Muhlfeld 

2008)—hybrids frequently appear where the species co-occur (Shepard and others 

2005). These hybrids are fertile and are capable of backcrossing with genetically 

pure fish of either source and with other hybrids, leading to introgression of nonnative genes into the cutthroat trout population. The spread of nonnative genes 

appears to rely primarily on straying by hybrid cutthroat trout rather than pure 

nonnative trout (Hitt and others 2003; Rubidge and Taylor 2004), although feral 

populations of rainbow trout in river main stems or a reservoir have also been 

implicated (Weigel and others 2003; Rubidge and Taylor 2005). First-generation 

hybrids do not appear to be at a selective disadvantage relative to parental stocks, 

but the fitness of later-generation hybrids is greatly reduced (Muhlfeld 2008). 

Despite this, introgression appears to progress via occasional reproductive bonanzas from hybrids and from rainbow trout parental stock (Muhlfeld 2008), and 

eventually rainbow trout genes may pervade virtually all remaining individuals to 

produce a hybrid swarm within a particular area (Allendorf and others 2001).

Although the probability of hybridization tends to rise with the number of 

introductions (Colautti 2005), spatial patterns in the distribution of nonnative 

genes are complex. Weigel and others (2003) reported that proximity to a source 

of stocked fish was only weakly related to hybridization between cutthroat trout 

and rainbow trout. Instead, larger streams at lower elevations were more likely 

to contain hybridized populations regardless of stocking location. They argued 

that isolating mechanisms, such as different spawning and emergence times or 

habitat preferences between adults of each species, were preventing hybridization from developing in headwater populations of cutthroat trout (Henderson and 

others 2000; Ostberg and Rodriguez 2006). In contrast, Hitt and others (2003), 

Rubidge and Taylor (2005), and Muhlfeld (2008) detected the upstream progression of hybrids between rainbow trout and cutthroat trout over time, and concluded 

that hybrid swarms would eventually occupy most of these waters except those 

with migration barriers.

Do Fires Favor Nonnative Fishes?

The immediate and long-term effects of wildfire on aquatic ecosystems and fish 

populations are the subject of this volume (Rieman and others, this volume) and 

of previous syntheses (Gresswell 1999; Young and others 2003), thus are reviewed 

only briefly here. The thermal or water chemistry changes associated with severe 

wildfire have led to direct mortality of trout populations (Rinne 1996; Howell 

2006), although many populations have endured such fires with few or no ill effects or showed only temporary declines (Rieman and Clayton 1997; Burton 2005; 

Sestrich 2005). More problematic may be post-fire floods, blackwater events, and 

debris torrents triggered by summer thunderstorms that have reduced or eliminated salmonid populations (Bozek and Young 1994; Brown and others 2001). 

Nevertheless, the majority of events are in small (<2.6 km2), steep (>20%) basins 

(Parrett and others 2003), and these would usually not threaten entire populations 

(Rieman and Clayton 1997; but see Brown and others 2001). The probability of 

such events often declines rapidly in subsequent years (Cannon 1999).
USDA Forest Service RMRS-GTR-290. 2012. 199

Secondary effects of fire may depress or boost populations. Channel stability 

often declines following fire because of the loss of riparian vegetation, increases in 

stream flow produced by reductions in evapotranspiration, and increases in sediment, but regrowth and resprouting of vegetation eventually attenuate off-channel 

contributions of sediment and water. Infall of fire-killed snags will spike in the first 

few decades following fire (Lyon 1984; Bragg 2000), which may lead to extremely 

complex channels with an array of complementary habitats. Growth rates of salmonids in burned streams have been observed to increase (Dunham and Rosenberger, 

this volume). Although this may in part be attributable to temporary changes in 

water chemistry, it is more likely that decreased shading increases primary productivity that leads to greater macroinvertebrate abundance and food availability 

(Wilzbach and others 2005). Because water temperatures are directly related to 

the amount of solar radiation reaching the water surface (Johnson 2004), post-fire 

temperatures tend to increase, sometimes by several degrees (Dunham and others 2007; Mahlum and others 2011). Warming may harm or bolster population 

growth rates depending on the pre-fire temperature regime (Dunham and others 

1999; Harig and Fausch 2002; Schrank and others 2003). Streams previously too 

cold for consistent reproduction or rapid growth may support larger, more stable 

populations, whereas those that are already relatively warm may become seasonally uninhabitable. Both conditions will gradually revert to their original state as 

canopy cover and stream shading increase (Luce and others, this volume), given 

that climatic patterns remain relatively constant.

It is sometimes thought that habitat degradation and disturbance in fluvial systems favors nonnative species either by creating conditions more suitable for them 

or by depressing the abundance of native species (ISAB 2008). Consequently, it has 

been hypothesized that post-fire habitats with increased water temperature, light, 

sediment transport and deposition, and channel instability may favor nonnative 

salmonids (Dunham and others 2003). At present, there is little evidence to refute 

or support this hypothesis. Similar to native salmonids, nonnative brook, brown, 

and rainbow trout have occasionally been eliminated from headwater streams during or after severe fires (Rinne 1996; Burton 2005). Furthermore, connectivity 

with unaffected portions of a watershed has facilitated the rapid recolonization of 

post-fire habitats (Novak and White 1990; Howell 2006), as has been observed for 

native species (Rieman and Clayton 1997).

Evaluations of the post-fire responses of sympatric populations of native and 

nonnative salmonids are more informative, but few such studies have been done 

(but see Howell 2006). One example involved monitoring the response of several native and nonnative species of salmonids in small watersheds throughout the 

Bitterroot River basin following fires in 2000 (Sestrich 2005; Sestrich and others 

2011). Declines in or the extirpation of all salmonid species were observed in some 

but not all of the watersheds that experienced high-severity fire. Where such declines were observed, westslope cutthroat trout (O. c. lewisi) and bull trout tended 

to recover rapidly and sometimes exceeded pre-fire abundances within 3 years. In 

contrast, brook trout were slower to respond and in some cases their abundance 

remains depressed (M. Jakober, Bitterroot National Forest, unpublished data). 

Brown trout, however, appeared in several monitoring reaches for the first time 

after the fires. Whether this was attributable to suitable post-fire environments or 

their ongoing population growth in river main stems (C. Clancy, Montana Fish, 

Wildlife and Parks, unpublished data) is uncertain.

In summary, there is little evidence that nonnative fishes show greater resistance to short-term habitat changes wrought by severe fire than do native species. 

Moreover, some native species may exhibit greater resilience to fire-related 
200 USDA Forest Service RMRS-GTR-290. 2012.

disturbance, in part because this represents a portion of the disturbance regime with 

which they evolved (Lytle and Poff 2004; Waples and others 2008). Nonetheless, 

I caution against generalizing the results of the limited research on this issue 

because the severity of habitat alteration, the environmental context of affected 

streams (e.g., their position in the watershed and arrangement of habitats), and the 

aquatic community (e.g., the abundance, diversity, and life histories of native and 

nonnative species) will influence the outcome. Because post-fire habitat changes—

particularly warmer water temperatures—may persist for decades or centuries, 

long-term shifts to nonnative species remain possible.

Climate Change, Fire, and Nonnative Fishes

Climate change is expected to alter global patterns in the distribution of flora and 

fauna (Wilson and others 2005). Based on recent modeling (IPCC 2007), climate in 

the western United States is projected to warm substantially before the end of this 

century. In montane environments, this warming is expected to cause more winter 

precipitation to fall as rain rather than snow (Knowles and others 2006), leading 

to smaller winter snowpacks (Mote and others 2005). In many streams in this region, earlier runoff and a longer low flow period are already evident (Stewart and 

others 2005), and greater flow intermittency in summer is anticipated (Boughton 

and others 2009). Besides discharge, the primary change in physical habitat is 

expected to be an increase in water temperature. As a consequence, populations of 

obligate coldwater species such as salmonids are predicted to retreat upstream to 

smaller waters to track suitable thermal environments (Rieman and others 2007; 

Wenger and others 2011a,b) and more temperature tolerant species will likely 

spread upstream as well (Rahel and Olden 2008). Initially, these may include nonnative coldwater species that prefer slightly warmer temperatures e.g., brown trout 

(McHugh and Budy 2005) and rainbow trout (Sloat and others 2002; Bear and 

others 2007), but additional warming may favor coolwater species such as smallmouth bass (Micropterus dolomieu; Sharma and Jackson 2008), which are already 

widely distributed and expanding in rivers in the inland West (LaVigne and others 

2008; Walford and Bestgen 2008). Mobile life history forms of native species that 

migrate downstream will encounter a gauntlet of potentially hostile nonnative species as well increasing areas of thermally unsuitable habitat and larger numbers of 

barriers to movement; selection against such forms may strongly favor resident life 

history strategies. Thus, the collective consequences of these changes will be to 

reduce overall occupied habitat and exacerbate the current patterns of population 

isolation for native coldwater fishes (Neville and others, this volume).

Climate change is also expected to increase the frequency, severity, and extent of fires (Westerling and others 2006). Combined effects of climate change 

and fire would place fish populations at even greater risk of extirpation during 

or shortly after severe wildfire because of low or no connectivity, fewer refuges, 

and smaller occupied habitats (Isaak and others, this volume). Thereafter, the loss 

of shading from stand-replacing fire in riparian zones and from associated debris 

torrents may further increase water temperatures to the extent that some sites may 

become uninhabitable for coldwater species. Moreover, climate warming coupled 

with severe fire may cross an ecological threshold that results in state changes in 

some forested basins in the western United States. This might include a shift from 

forest types adapted to infrequent severe fire to those tolerant of more frequent fire 

(Keane and others 2008) or from forested areas to grasslands (Schoennagel and 

others 2008). Such wholesale changes in the terrestrial vegetation and disturbance 

dynamics would alter stream conditions, probably leading to additional upstream 
USDA Forest Service RMRS-GTR-290. 2012. 201

incursions of nonnative species and the further diminution of habitats occupied by 

native coldwater fauna.

Climate change, fire, and nonnative species invasions are likely to redirect future management efforts. Whereas the present emphasis is on in situ habitat and 

population protection and restoration, in the future population salvage (Brooks 

2006), assisted migration (cf. Millar and others 2007), anthropogenic refounding, 

thermal mitigation (e.g., by increasing and retaining shading), and refuge designation (e.g., finding those areas most likely to provide suitable habitats for native 

species despite climate change) may assume importance as management tools in 

pre- and post-fire environments. The changing management of other resources 

may also provide some unforeseen opportunities for aquatic species management. 

The predicted decline in surface water available for human use under many climate 

change scenarios has led to proposals to build additional high-elevation reservoirs 

to store water for late summer release (Bates and others 2008). Such waters could 

become de facto refuges of native species that tolerate or require lentic habitats, 

and reservoir releases could increase (and sometimes cool) streamflow adequately to maintain coldwater species downstream. This should not be regarded as a 

panacea for native fish; reservoirs are often the target for nonnative species introductions that then spread elsewhere (Havel and others 2005).

There is substantial uncertainty associated with the rate of change and locations 

most likely to experience the greatest effects from climate change, in part because 

of the low predictability of whether, when, and how much major drivers of global 

climate, such as the North Atlantic circulation and El Niño-Southern Oscillation, 

will contribute (Kriegler and others 2009). Nevertheless, all trends and forecasts 

suggest a substantially altered and warmer climate in the western United States in 

this century. Retaining ecosystem services and providing habitats suitable for the 

persistence and evolution of the native aquatic fauna (Dunham and Rosenberger, 

this volume) under the constraints imposed by a changing climate, altered disturbance regime, and array of exotic newcomers represents a critical challenge in the 

not-so-distant future.

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Management 178.
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