Synthesis of Science to Inform Land 

Management Within the Northwest 

Forest Plan Area

Volume 1

United States Department of Agriculture

Forest 

Service

Pacific Northwest

Research Station

General Technical Report

PNW-GTR-966 Vol. 1

NORTHWEST

FOREST PLAN

June

2018
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Technical Coordinators

Thomas A. Spies is a senior scientist, U.S. Department of Agriculture, Forest Service, 

Pacific Northwest Research Station, 3200 SW Jefferson Way, Corvallis, OR 97331; Peter 

A. Stine is research program manager and biogeographer (retired), U.S. Department of 

Agriculture, Forest Service, Pacific Southwest Research Station, and is a research associate, 

John Muir Institute for the Environment, University of California–Davis, 1 Shields Avenue, 

Davis, CA 95616; Rebecca Gravenmier is a science coordinator, U.S. Department of Agriculture, Forest Service, Pacific Northwest Research Station, 1220 SW 3rd Ave., Portland, 

OR 97204; Jonathan W. Long is an ecologist, U.S. Department of Agriculture, Forest 

Service, Pacific Northwest Research Station, 1731 Research Park Dr., Davis, CA 95618; and 

Matthew J. Reilly is a postdoctoral researcher, Humboldt State University, Department of 

Biological Sciences, 1 Harpst Street, Arcata, CA 95521.

Cover photos: Upper left: prescribed burn operations on the Wallowa-Whitman National 

Forest, Oregon; photo by Wallowa-Whitman National Forest. Upper right: a northern 

spotted owl in the McKenzie River Basin in Oregon; photo by John and Karen Hollingsworth, U.S. Fish and Wildlife Service. Lower right: old-growth forest, Oswald West State 

Park, Oregon; photo by David Patte, U.S. Fish and Wildlife Service. Lower left: marbled 

murrelet; photo by Kim Nelson, Oregon State University.
Synthesis of Science to Inform 

Land Management Within the 

Northwest Forest Plan Area

Volume 1

Thomas A. Spies, Peter A. Stine, Rebecca Gravenmier, 

Jonathan W. Long, and Matthew J. Reilly, Technical Coordinators

U.S. Department of Agriculture

Forest Service

Pacific Northwest Research Station

Portland, Oregon

General Technical Report PNW-GTR-966 Vol. 1

June 2018
Abstract 

Spies, T.A.; Stine, P.A.; Gravenmier, R.; Long, J.W.; Reilly, M.J., tech. coords. 2018.

Synthesis of science to inform land management within the Northwest Forest Plan area. 

Gen. Tech. Rep. PNW-GTR-966. Portland, OR: U.S. Department of Agriculture, Forest 

Service, Pacific Northwest Research Station. 1020 p. 3 vol.

The 1994 Northwest Forest Plan (NWFP) was developed to resolve debates over oldgrowth forests, endangered species, and timber production on federal forests in the range 

of the northern spotted owl. This three-volume science synthesis, which consists of 12 

chapters that address various ecological and social concerns, is intended to inform forest 

plan revision and forest management within the NWFP area. Land managers with the U.S. 

Forest Service provided questions that helped guide preparation of the synthesis, which 

builds on the 10-, 15-, and 20-year NWFP monitoring reports and synthesizes the vast 

body of relevant scientific literature that has accumulated in the 24 years since the NWFP 

was initiated. It identifies scientific findings, lessons learned, and uncertainties and also 

evaluates competing science and provides considerations for management. 

This synthesis finds that the NWFP has protected dense old-growth forests and 

maintained habitat for northern spotted owls, marbled murrelets, aquatic organisms, and 

other species despite losses from wildfire and low levels of timber harvest on federal lands. 

Even with reductions in the loss of older forests, northern spotted owl populations continue 

to decline. Moreover, a number of other goals have not been met, including producing a 

sustainable supply of timber, decommissioning roads, biodiversity monitoring, significant 

levels of restoration of riparian and dry forests, and adaptation and learning through 

adaptive management. 

New conservation concerns have arisen, including a major threat to spotted owl 

populations from expanding populations of the nonnative barred owl, effects of fire 

suppression on forest succession, fire behavior in dry forests, and lack of development of 

diverse early-seral vegetation as a result of fire suppression in drier parts of moist forests. 

Climate change and invasive species have emerged as threats to native biodiversity, and 

expansion of the wildland-urban interface has limited the ability of managers to restore fire 

to fire-dependent ecosystems. 

The policy, social, and ecological contexts for the NWFP have changed since it was 

implemented. The contribution of federal lands continues to be essential to the conservation 

and recovery of fish listed under the Endangered Species Act and northern spotted owl 

and marbled murrelet populations. Conservation on federal lands alone, however, is likely 

insufficient to reach the goals of the NWFP or the newer goals of the 2012 planning rule, 

which emphasizes managing for ecosystem goals (e.g. ecological resilience) and a few 

species of concern, rather than the population viability of hundreds of individual species. 
The social and economic basis of many traditionally forest-dependent communities 

has changed in 24 years, and many are now focused on amenity values. The capacities 

of human communities and federal agencies, collaboration among stakeholders, the 

interdependence of restoration and the timber economy, and the role of amenity- or recreation-based communities and ecosystem services are important considerations in managing 

for ecological resilience, biodiversity conservation, and social and economic sustainability. 

A growing body of scientific evidence supports the importance of active management 

or restoration inside and outside reserves to promote biodiversity and ecological resilience. 

Active management to promote heterogeneity of vegetation conditions is important to 

sustaining tribal ecocultural resources. Declines in agency capacity, lack of markets for 

small-diameter wood, lack of wood processing infrastructure in some areas, and lack of 

social agreement have limited the amount of active management for restoration on federal 

lands. All management choices involve social and ecological tradeoffs related to the goals 

of the NWFP. Collaboration, risk management, adaptive management, and monitoring are 

considered the best ways to deal with complex social and ecological systems with futures 

that are difficult to predict and affect through policy and land management actions.

Keywords: Northwest Forest Plan, science, management, restoration, northern spotted 

owl, marbled murrelet, climate change, socioeconomic, environmental justice.
Preface

In 2015, regional foresters in the Pacific Northwest and Pacific Southwest Regions of the 

USDA Forest Service requested that the Pacific Northwest and Pacific Southwest Research 

Stations prepare a science synthesis to inform revision of existing forest plans under the 

2012 planning rule in the area of the Northwest Forest Plan (NWFP, or Plan). Managers 

provided an initial list of hundreds of questions to the science team, which reduced to them 

to 73 questions deemed most feasible for addressing through a study of current scientific literature. The stations assembled a team of 50 scientists with expertise in biological, ecological, and socioeconomic disciplines. At the suggestion of stakeholders, a literature reference 

database was placed online so the public could submit additional scientific literature for 

consideration. By spring 2016, writing was underway on 12 chapters that covered ecological and social sciences. 

The draft synthesis, which was ready for peer and public review by fall 2016, went 

through a special review process because it was classified as “highly influential science” in 

accordance with the Office of Management and Budget’s 2004 “Final Information Quality 

Bulletin for Peer Review.” The synthesis was classified as such because it fit the category of 

a scientific assessment that is novel, controversial, or precedent-setting, or has significant 

interagency interest. Per the bulletin, the two research stations commissioned an independent entity, the Ecological Society of America (ESA), to manage the peer-review process, 

including the selection of peer reviewers. 

The bulletin also stipulates that such an assessment be made available to the public 

through a public meeting to enable the public to bring scientific issues to the attention of 

peer reviewers. Accordingly, a public forum was held in Portland, Oregon, in December 

2016. For those who could not travel to Portland, the forum was accessible via live Web 

stream, and multiple national forests within the NWFP area hosted remote viewing. Written 

comments on the draft synthesis were collected for 2 months. This generated 130 public 

comments, totaling 890 pages, which were given to the peer reviewers for consideration 

in their review, as they deemed appropriate. The OMB guidelines further direct that the 

peer-review process be transparent by making available to the public the ESA’s written 

guidance to the reviewers, the peer reviewer’s names, the peer review reports, and the 

responses of the authors to the peer reviewer comments—all of which are available at 

https://www.fs.fed.us/pnw/research/science-synthesis/index.shtml. 

The peer reviewer comments, which were received in spring 2017 and informed by 

public input, resulted in substantive revisions to chapters of the synthesis. The result is this 

three-volume general technical report (an executive summary of the synthesis is available 

as a separate report). This document is intended to support upcoming management planning on all public lands in the Plan area, but is expected to serve primarily lands managed 

by the U.S. Forest Service. We hope it will be a valuable reference for managers and others 

who seek to understand the scientific basis and possible tradeoffs associated with forest 

plan revision and management decisions. The synthesis also provides an extensive list of 

published sources where readers can find further information.
We understand that the term “synthesis” can have many different meanings. For our 

purposes, it represents a compilation and interpretation of relevant scientific findings that 

pertain to key issues related to the NWFP that were identified by managers and by the 

authors of the document. Such a compilation not only summarizes science by topic areas 

but also interprets that science in light of management goals, characterizes competing 

science, and makes connections across scientific areas, addressing multilayered and interacting ecological and socioeconomic issues. In a few cases, simple analyses of existing data 

were conducted and methods were provided to reviewers. 

The synthesis builds upon the 10-, 15-, and 20-year NWFP monitoring reports, and 

authors considered well over 4,000 peer-reviewed publications based on their knowledge 

as well as publications submitted by the public and others suggested by peer reviewers. For 

some of the questions posed by land managers, there was ample scientific research from 

the Plan area. For many of the questions, however, little research existed that was specific 

to the area. In such cases, studies from other regions or current scientific theory were used 

to address the questions to the extent possible. In many cases, major scientific uncertainties 

were found; these are highlighted by the authors. 

The synthesis chapters characterize the state of the science but they do not develop 

management alternatives, analyze management tradeoffs, or offer recommendations as to 

what managers should do. The synthesis does identify ideas, facts, and relationships that 

managers may want to consider as they develop plans and make management decisions 

about particular issues. The final chapter attempts to integrate significant cross-cutting 

issues, e.g., ecological and socioeconomic interdependencies, compatibility of different 

management goals, and tradeoffs associated with different restoration actions. All the 

chapters identify where more research is needed to fill critical information gaps. 

We would like to acknowledge the peer reviewers who considered hundreds of public 

comments as part of the process of reviewing our lengthy draft manuscripts. We also thank 

the many contributors to the development of the synthesis in draft and final form, including 

those who provided editing, layout, database, and other support services. 
Contents

Volume 1

1 Chapter 1: Introduction

Peter A. Stine and Thomas A. Spies 

1 Background and Purpose of This Science Synthesis

3 Northwest Forest Plan History and Context

3 The Forest Ecosystem Management Assessment Team

4 Principal Elements of the NWFP

7 History of Reporting on the Research and Monitoring Within the NWFP Area

7 10-, 15-, and 20-Year Monitoring Reports 

9 Scope and Approach of This Science Synthesis

9 Rationale for Topics Covered

10 Sources of Information Considered

10 Dealing With Scientific Uncertainty

11 Role of Peer Review in This Document

12 The NWFP Area

12 Ecogeographic Variability of NWFP Area

13 Other Syntheses Reports Relevant to the NWFP Area

13 Role of Science in Supporting Land Management

14 Context of the NWFP and Forest Plan Revision Under the New Planning Rule

16 Emergent Issues 

20 Literature Cited

23 Appendix: Priority Management Questions to Guide the Northwest Forest Plan 

Science Synthesis As Defined by Pacific Northwest and Pacific Southwest Forest and 

Regional Staff and Edited by the Science Synthesis Team 

23 Priority Questions

27 Other Topics to Be Considered in the Integration Section of the Synthesis (Pulled From 

Region 5 and Region 6 Long List)

29 Chapter 2: Climate, Disturbance, and Vulnerability to Vegetation Change in the 

Northwest Forest Plan Area

Matthew J. Reilly, Thomas A. Spies, Jeremy Littell, Ramona Butz, and John B. Kim 

29 Introduction 

30 Guiding Questions

30 Background and Setting

35 Key Findings

35 Past Climate Change in the Northwest Forest Plan Area

36 Fire History

41 20th-Century Climate Change in the Northwest Forest Plan Area

41 Projecting Climate Change for the 21st Century 

42 21st-Century Climate Change Projections for the Northwest Forest Plan Area

43 Implications of Observed Climate Trends for Water Balance Deficit and Vegetation Change
46 Mechanisms of Vegetation Change

50 Abiotic Disturbances

55 Vegetation Models and Potential Future Vulnerability

56 Summary of Vulnerabilities to Climate Change

58 Other Vulnerabilities

59 Adaptation to and Mitigation of Climate Change

63 Research Needs, Uncertainties, Information Gaps, and Limitations

64 Conclusions and Management Considerations

65 Literature Cited

92 Appendix: Crosswalk of Simpson (2013) Potential Vegetation Zones With Existing 

Vegetation From the Classification and Assessment With Landsat of Visible Ecological Groupings (CALVEG) System 

95 Chapter 3: Old Growth, Disturbance, Forest Succession, and Management in the 

Area of the Northwest Forest Plan

Thomas A. Spies, Paul F. Hessburg, Carl N. Skinner, Klaus J. Puettmann, Matthew J. 

Reilly, Raymond J. Davis, Jane A. Kertis, Jonathan W. Long, and David C. Shaw 

95 Introduction

98 Guiding Questions

99 Key Findings

99 Vegetation Patterns and Classification

101 Disturbance Regimes

115 Forest Succession and Landscape Dynamics 

137 Effects of Fire Exclusion

143 Use of Historical Ecology in Conservation and Restoration

146 Ecosystem Function 

148 Conservation and Restoration Needs 

151 Timber Management and Old-Growth Conservation 

153 Reserves in Dynamic Ecosystems 

161 Connectivity and Fragmentation

161 Restoration Approaches

175 Invasive Plant Species and Pathogens

177 Postfire Salvage and Management

181 Research Needs, Uncertainties, Information Gaps, and Limitations

182 Conclusions and Management Considerations

183 Management Considerations Summarized

183 Guiding Questions

184 Ecology of Old-Growth and Other Vegetation Types (Questions 1 and 2) 

185 Value of Ecological History (Question 3)

186 Conservation and Restoration Needs (Questions 4 and 5)

186 Competing Science Related to Need for Restoration (Question 5)

187 Trends in Forests in the NWFP Reserve Network (Question 9)

187 Reserve Approaches in Dynamic Landscapes (Questions 8 and 9) 
187 Restoration Approaches (Questions 6 and 7)

189 Post-Wildfire Management (Question 10) 

189 Acknowledgments

189 Literature Cited

231 Appendix 1: Crosswalk of Simpson (2013) Potential Vegetation Zones With Existing Vegetation From the Classification and Assessment With Landsat of Visible 

Ecological Grouping (CALVEG)

232 Appendix 2: Fire Regime Mapping Method

239 Appendix 3: Summary of Fire History Studies in the Northwest Forest Plan

245 Chapter 4: Northern Spotted Owl Habitat and Populations: Status and Threats 

Damon B. Lesmeister, Raymond J. Davis, Peter H. Singleton, and J. David Wiens 

245 Introduction 

246 Guiding Questions

247 Key Findings

247 Population Status and Trends

250 Habitat Status and Trends

258 Barred Owls 

263 Spotted Owl Prey

264 Disturbance

270 Climate Change

271 Other Threats

272 Research Needs, Uncertainties, Information Gaps, and Limitations

272 Research Needs

277 Scientific Uncertainty

279 Conclusions and Management Considerations

280 Management Considerations

283 U.S. and Metric Equivalents

283 Acknowledgments

284 Literature Cited

301 Chapter 5: Marbled Murrelet

301 Martin G. Raphael, Gary A. Falxa, and Alan E. Burger 

301 Introduction 

304 Guiding Questions

305 Key Findings 

305 NWFP Expectations

306 NWFP Monitoring Results for Marbled Murrelets

316 Status of Marbled Murrelets Elsewhere in the Species’ Range

318 Nesting Habitat Relationships

329 Climate Change Considerations

332 Research Needs, Uncertainties, Information Gaps, and Limitations

334 Conclusions and Management Considerations
334 Are NWFP Assumptions Still Valid?

337 Management Considerations

338 Acknowledgments

338 Metric Equivalents

339 Literature Cited

354 Glossary

370 Acknowledgments

Volume 2 

371 Chapter 6: Other Species and Biodiversity of Older Forests

Bruce G. Marcot, Karen L. Pope, Keith Slauson, Hartwell H. Welsh, Clara A. Wheeler, 

Matthew J. Reilly, and William J. Zielinski 

461 Chapter 7: The Aquatic Conservation Strategy of the Northwest Forest Plan—A 

Review of the Relevant Science After 23 Years

Gordon H. Reeves, Deanna H. Olson, Steven M. Wondzell, Peter A. Bisson, Sean 

Gordon, Stephanie A. Miller, Jonathan W. Long, and Michael J. Furniss 

Volume 3

625 Chapter 8: Socioeconomic Well-Being and Forest Management in Northwest 

Forest Plan-Area Communities

Susan Charnley, Jeffrey D. Kline, Eric M. White, Jesse Abrams, Rebecca J. McLain, 

Cassandra Moseley, and Heidi Huber-Stearns 

717 Chapter 9: Understanding Our Changing Public Values, Resource Uses, and 

Engagement Processes and Practices

Lee K. Cerveny, Emily Jane Davis, Rebecca McLain, Clare M. Ryan, Debra R. 

Whitall, and Eric M. White 

809 Chapter 10—Environmental Justice, Low-Income and Minority Populations, and 

Forest Management in the Northwest Forest Plan Area

Susan Charnley, Delilah Jaworski, Heidi Huber-Stearns, Eric M. White, Elisabeth 

Grinspoon, Rebecca J. McLain, and Lee Cerveny 

851 Chapter 11: Tribal Ecocultural Resources and Engagement

Jonathan Long, Frank K. Lake, Kathy Lynn, and Carson Viles 

919 Chapter 12: Integrating Ecological and Social Science to Inform Land Management in the Area of the Northwest Forest Plan

Thomas A. Spies, Jonathan W. Long, Peter Stine, Susan Charnley, Lee Cerveny, Bruce 

G. Marcot, Gordon Reeves, Paul F. Hessburg, Damon Lesmeister, Matthew J. Reilly, 

Martin G. Raphael, and Raymond J. Davis 
xii

GENERAL TECHNICAL REPORT PNW-GTR-966

Koosah Falls, Willamette National Forest.

Photo by USDA Forest Service.
1

Chapter 1: Introduction

Peter A. Stine and Thomas A. Spies1

Background and Purpose of This 

Science Synthesis

We live in an era of information. Although this brings many 

benefits to society, it creates challenges for those responsible 

for understanding and applying new and older information 

to their day-to-day work. How does one keep up with the 

volume of relevant information that is published daily?

People who manage the 24 million ac (9.7 million ha) 

of public land within the area of the Northwest Forest Plan 

(NWFP, or Plan) depend on sound scientific knowledge 

about ecological systems and about how they function and 

how they respond to change. The Plan area stretches from 

Washington’s northern border to a significant portion of 

northern California, encompassing diverse geography, 

ecological systems, and human communities. The authors 

of the NWFP understood that scientific knowledge would 

be critical to the efficacy of the plan, both in preparation 

of plan guidance and in learning how affected forests and 

communities (i.e., socio-ecological systems) would change 

over time, with and without active management. Current 

direction to national forests that are undertaking forest 

plan revisions also specifically calls for sound scientific 

information to guide plan preparation and to make selected 

changes to how forests might be managed in the future. 

Land managers responsible for updating forest plans find 

it challenging to remain current with all the new scientific 

knowledge. For a geographic region as large, diverse, and 

complex as the Plan area, this presents one of the greatest 

challenges to plan preparation and execution. 

The majority of public lands within the NWFP area 

are managed by the U.S. Forest Service. This includes 

roughly 19.2 million ac (7.68 million ha) on 17 national 

forests (the Deschutes, Fremont-Winema, Gifford Pinchot, 

Klamath, Lassen, Mendocino, Modoc, Mount Baker–Snoqualmie, Mount Hood, Okanogan-Wenatchee, Olympic, 

Rogue River–Siskiyou, Shasta-Trinity, Siuslaw, Six Rivers, 

Umpqua, and Willamette National Forests). There are also 

roughly 2.5 million ac (1 million ha) of U.S. Department 

of the Interior Bureau of Land Management (BLM) lands 

and roughly 2.3 million ac (0 .92 million ha) of National 

Park Service lands within the Plan area. This synthesis is 

intended to support upcoming management work on all public lands, but is expected to serve primarily Forest Service 

lands and their impending forest plan revisions. In 2016, the 

BLM revised its resource management plans for its lands 

in western Oregon. Although the BLM and Forest Service 

are using distinct and separate planning processes to revise 

land use plans within the Plan area, the two agencies share 

common goals for long-term monitoring of the impacts of 

the implementation of their land use plans.

To help meet the challenge of forest plan revision, 

this science synthesis provides a comprehensive overview 

of the full body of relevant science accumulated in the 

24 years since the NWFP was initiated. The synthesis 

was developed at the behest of the Pacific Southwest 

and Pacific Northwest Regions (Forest Service Regions 

5 and 6). To accomplish this task, the Pacific Northwest 

(PNW) Research Station and the Pacific Southwest (PSW) 

Research Station assembled a team of scientists who are 

experts in a variety of biological, ecological, and socioeconomic disciplines. 

The term “synthesis” can have many different meanings. For our purposes, it is a compilation of relevant 

scientific findings that pertain to key issues around the 

NWFP. Such a compilation not only summarizes science 

by topic areas but also makes connections across scientific 

themes and addresses multilayered and interacting natural 

and socioeconomic resource issues. This report has been 

prepared to assist land managers in updating existing 

forest management plans and on-the-ground projects. 

Our hope is that it will serve as a reference that provides 

a condensed and integrated understanding of the current 

state of knowledge regarding the NWFP, as well as an 

extensive list of published sources, where readers can find 

further information. 

1 Peter A. Stine is a research program manager and biogeographer 

(retired), U.S. Department of Agriculture, Forest Service, Pacific 

Southwest Research Station, 1731 Research Park Drive, Davis, 

CA 95618, and a research associate, John Muir Institute for the 

Environment, University of California–Davis, 1 Shields Avenue, 

Davis, CA 95616; Thomas A. Spies is a research forester, U.S. 

Department of Agriculture, Forest Service, Pacific Northwest 

Research Station, 3200 SW Jefferson Way, Corvallis, OR 97331.
2

GENERAL TECHNICAL REPORT PNW-GTR-966

This synthesis is not a bibliography or an interpretation of all available science; and is not intended to direct 

management through recommendations or analysis of management alternatives. In contrast, the charge given to the 

scientists who served as members of the Forest Ecosystem 

Management Assessment Team (FEMAT) under the original NWFP (FEMAT 1993) requested that scientists assess 

the science and use their expert knowledge to develop a set 

of plan alternatives and corresponding management recommendations. President Bill Clinton selected and adapted 

one of these plan alternatives, which formed the basis of the 

standards and guides for the NWFP. This science synthesis 

provides a summary and interpretation of relevant science 

findings to support subsequent planning efforts under Forest 

Service regulations. 

Our approach largely follows the role of “science 

arbiters,” one of the four roles that scientists can play 

in policy arenas (Pielke 2003). Science arbiters answer 

questions from managers from a scientific perspective (e.g., 

What are the ecological differences between dry forests and 

moist forests, or what is known about the ecological effects 

of different restoration strategies?). But they do not develop 

or evaluate policy alternatives. We do not play an alternative 

role of “honest brokers of policy alternatives” who develop 

a wide range of policy alternatives and characterize their 

possible consequences using scientific findings and expert 

opinion. That was the role that the scientists in FEMAT 

played. Although this synthesis does not develop plan 

alternatives or evaluate them, it does characterize what is 

known about the ecological effects of various management 

practices (e.g., salvage logging or prescribed fire), and it 

identifies ecological and socioeconomic tradeoffs associated 

with different management goals (e.g., ecosystem integrity 

vs. single species) and practices. We also characterize how 

well the NWFP has met some of its original goals by using 

information from the monitoring programs and peer-reviewed published sources. 

The synthesis builds upon the 10-, 15-, and 20-year 

NWFP monitoring reports and it considered well over 

4,000 peer-reviewed publications. The authors of individual chapters have extensive knowledge of the scientific 

literature, and much of what was reviewed comes from 

their knowledge of the most relevant work. As part of this 

review process, we also established a Web portal to enable 

members of the public to offer appropriate literature that 

they wanted to ensure would be included in the review. 

We provided a comprehensive summary of the scientific 

literature that we considered salient to the key issues to 

be addressed by land managers as they begin considering 

forest plan revision. 

The breadth of topics and number of scientific papers 

that could be covered in this synthesis is enormous. At the 

direction of Regions 5 and 6, we focused on topics that had 

a direct bearing on activities that resulted from the NWFP 

and subsequent forest plan revision. Focal topics were 

distinguished from a large set of management questions 

identified by Forest Service management staff in the two 

regions. The core author team worked with Forest Service 

managers to condense the initial set of questions to 73 (see 

app. 1). The final list was established by removing questions 

that were outside the scope of this effort (including those 

that could not be addressed by published scientific information or were not relevant to the NWFP), then identifying 

only those topics that could be addressed by reviewing the 

evidence contained in the scientific literature (i.e., at least 

some scientific information exists that would enable some 

insight on the question). The final questions were grouped 

into four main categories (Vegetation/Forest Management, 

Terrestrial Species/Habitat Management, Aquatic/Riparian 

Management, and Social/Economic, including Timber 

Production), which formed the basis for the organization of 

the synthesis. Lead authors used these questions to build 

chapter outlines and provide useful information to support 

subsequent management planning efforts. 

The authors of the chapters address the management 

questions using a range of approaches. In some cases, there 

is ample scientific evidence from the Plan area to address 

the questions; however, in many cases, few research studies 

exist from the NWFP area. In such cases, studies from 

other regions or current scientific theory are used to address 

the questions to the extent possible. In many cases, major 

uncertainties are identified, while in others much uncertainty remains. The following chapters provide comprehensive reviews of the relevant scientific literature within their 
3

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

topic areas, but the authors do not evaluate tradeoffs among 

different resource management and planning objectives. 

Chapter 12, however, addresses the most significant integration issues as well as potential tradeoffs to identify where 

additional evaluation or more monitoring/research will be 

necessary in subsequent assessments and planning efforts to 

resolve potential or existing conflicts.

Northwest Forest Plan History and Context

The NWFP is rooted in the environmental history of the 

region and followed a series of ecological and socioeconomic 

triggers in the 1980s and early 1990s (Johnson and Swanson 

2009). Historically, the ecosystems of this region have been 

influenced by many tribes of native people for millennia (see 

chapter 11). More than two centuries ago, their civilizations 

and stewardship of the ecosystems of the region were greatly 

affected by visitors and settlers from the Eastern United 

States or from European countries, and the United States 

gradually seized or acquired lands from tribes, converting 

much of the forested area into farmlands, industrial timberlands, and other new land uses. By the beginning of the 20th

century, large tracts of forest lands in the Western United 

States were put into “forest reserves” and managed by the 

U.S. Forest Service to protect watersheds and ensure a 

continuous supply of timber. The initial reserve era gave way 

to the era of sustained-yield forestry to support economic 

growth (Steen 2004). These practices continued into the 

1970s, when three significant federal laws were passed: the 

National Environmental Policy Act (NEPA) of 1970, the 

Endangered Species Act (ESA) of 1973, and the National 

Forest Management Act (NFMA) of 1976. Collectively, 

these laws engendered an era of increasing environmental 

awareness and concern. During the next two decades, the 

stage was set for conflict between timber-focused policies 

and the emerging public concern over the environmental 

impacts of forest management practices in the Northwest. By 

1990, conservation of biodiversity had ascended to become a 

new priority for federal forests, and numerous organizations 

stepped in to initiate litigation, which ultimately led to establishment of the NWFP in 1994 (Johnson and Swanson 2009). 

The NWFP was a product of many social and ecological drivers, but the focal point of the deliberations was the 

protection of the old-forest ecosystems that provide habitat 

for northern spotted owls (Strix occidentalis caurina). 

The Plan also addressed the needs of the marbled murrelet 

(Brachyrampus marmoratus), anadromous fish, and other 

species associated with older forests, as well as stressing 

the importance of sustaining rural communities and 

economies through continued timber harvest (Charnley 

2006). There are many alternative views and definitions of 

“old growth” (chapter 3) (Haynes et al. 2006). For the sake 

of simplicity, we use only the term “old-growth forests” in 

this introduction. 

The 1980s were part of a transformative period for 

the Pacific Northwest and northern California (Johnson 

and Swanson 2009). For many years, timber harvest was 

extensive across the region, and concerns about the effects 

that the logging of old growth had on wildlife and riparian 

areas grew steadily into the early 1990s. The 1990 listing of 

the northern spotted owl as a threatened species precipitated 

numerous legal challenges regarding the cumulative impacts 

of federal timber management in the Pacific Northwest 

and northern California. When a federal court issued an 

injunction in 1991 on all timber sales on federal lands within 

the range of the northern spotted owl, the political and 

environmental landscape shifted substantially. The ensuing 

political crisis set the stage for the emergence of the NWFP. 

These dramatic events and emerging science precipitated federal government engagement, up to and including 

the White House, to seek a workable solution. Over the next 

2 years, beginning in earnest with the Northwest Forest 

Summit in 1993, the federal government forged a plan. The 

extensive involvement of the White House and principal 

land management agencies (i.e., the Forest Service and 

BLM) led to the 1994 adoption of the NWFP by the Clinton 

Administration (Pipkin 1998).

The Forest Ecosystem Management 

Assessment Team

President Clinton established three interagency working 

groups to build a foundation for what would ultimately 

become the NWFP. One of these groups was FEMAT, a 

team of scientists, resource managers, and technicians from 

many different universities and public agencies, charged 
4

GENERAL TECHNICAL REPORT PNW-GTR-966

with identifying management alternatives that could attain 

the greatest economic and social contribution from forests, 

while meeting all applicable laws and regulations (FEMAT 

1993). Specifically, FEMAT was asked to consider and 

develop conservation approaches, restoration actions, and 

adaptive management strategies to meet the following biological diversity goals: (1) habitat for the northern spotted 

owl and marbled murrelet, (2) habitat for other species 

associated with old growth, (3) spawning and rearing habitat for anadromous fish, and (4) maintenance of a connected 

old-growth forest reserve system on federal lands. 

FEMAT issued an extensive report (FEMAT 1993) that 

analyzed the ecological, social, and economic implications 

of 10 management options for the federal forests within the 

range of the northern spotted owl. The team used expert 

opinion to assess biophysical processes and disturbances, 

community capacity, and economic factors, and it estimated 

tradeoffs and risk to species associated with different levels 

of protection for biodiversity and timber production. This 

was, and may still be, the most extensive regional forest 

biodiversity and management assessment of its kind. Many 

of today’s persistent policy challenges were raised and 

considered 24 years ago in this report. The FEMAT report 

identified risk and uncertainties associated with the different conservation and management issues and recognized 

that monitoring and adaptive management would be needed 

to maintain a long-term, scientifically based and adaptive 

plan. This synthesis summarizes published research, 

monitoring and knowledge of plan implementation over the 

past 24 years, providing a current scientific foundation for 

forest planning. 

Principal Elements of the NWFP

Conservation and management of old-growth forests are 

central to the NWFP and the past 24 years of its implementation. As readers consider the various chapters in this 

synthesis, they will see that old-growth forests have both an 

ecological and a social dimension. These dimensions can be 

linked, but also can emerge in quite different contexts. We 

address and discuss these facets in the following chapters.

The principal tasks of the NWFP were to conserve and 

restore habitats for animals and plant species associated 

with old-growth forests and maintain and restore habitat 

for anadromous fish within the confines of existing laws 

and regulations (e.g., NFMA and ESA). Management of 

the affected 24 million ac (9.7 million ha) of land was 

altered significantly to meet these new biological diversity 

goals. At the time, relatively little was known about most 

species associated with late-successional and old-growth 

forests, and this is still the case. Although the biology and 

ecology of the northern spotted owl were relatively well 

understood, there were many gaps in our understanding 

of this long-lifespan species with a low reproductive rate. 

The major shift in federal forest management was part of 

a larger global trend toward increasing protection for the 

forest biodiversity through a process called “ecosystem 

management” (Grumbine 1994). As Chuck Meslow, then 

leader of the Oregon Cooperative Wildlife Research Unit at 

Oregon State University, explained, the NWFP originated at 

a time when many scientists were beginning to advocate for 

a more ecological approach to managing remaining oldgrowth forests (FEMAT 1993). 

The intent of ecosystem management, as it was 

initially envisioned at the time, was to sustain ecosystems 

by maintaining (1) viable populations of native species, (2) 

native ecosystem types, and (3) evolutionary and ecological 

processes over long time horizons (Grumbine 1994). In 

doing so, it was posited that such a management regime 

would accommodate human use and occupancy within 

the capacities of ecosystems. The NWFP changed federal 

management by giving priority to ecological sustainability; 

the team was directed to plan for social and economic 

values after meeting ecological objectives. The hope was 

that the Plan could find common ground through the right 

balance of biodiversity and timber management objectives 

(Charnley 2006).

The NWFP evolved out of three preceding efforts 

in the early 1990s to find a solution to the conflicts over 

federal forest management (Thomas et al. 2005): (1) a 

conservation strategy for the northern spotted owl (Thomas 

et al. 1990), (2) “Gang of Four” report on alternatives for 

management of Pacific Northwest late-successional forests 

for multiple species (Johnson 1997, Johnson et al. 1991), 

and (3) the Scientific Analysis Team (known as the SAT) 
5

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

report, which conducted a scientific analysis that added 

riparian protection and more species to the assessment. 

(Thomas et al. 1993). These efforts laid the foundation for 

much of the NWFP. FEMAT, established by the president, 

used this and other sources of information to develop 

options that would (1) consider human and economic 

dimensions of the problem; (2) protect the long-term health 

of forests, wildlife, and waterways; (3) be scientifically 

sound, ecologically credible, and legally responsible; 

(4) produce a predictable and sustainable level of timber 

sales and nontimber resources that would not degrade the 

environment; and (5) emphasize collaboration among the 

federal agencies responsible for management of these lands 

(Thomas et al. 2005).

FEMAT developed 10 options for the president and 

agency heads to consider. They selected option 9, which was 

based on both ecosystem- and species-level conservation 

and restoration strategies. This option was subsequently 

modified to meet viability requirements under NFMA 

during the final environmental impact statement process, 

and the final plan was set forth in the record of decision 

(ROD), with the following key elements:

• Adoption of a yet-to-be-defined ecosystem 

management approach

• Seven land allocations (see fig. 1-1) to address key 

conservation/management concerns, including:

• Congressionally reserved areas

 (7.3 million ac/2.95 million ha)

• New late-successional reserves 

(7.4 million ac/2.99 million ha)

• New adaptive management areas 

(1.5 million ac/607 000 ha) 

• New managed late-successional areas

• Administratively withdrawn areas

• New riparian reserves (2.6 million ac/

1 million ha)

• Matrix (for ecologically sensitive timber production) (nearly 4 million ac/1.6 million ha)

• An emphasis on effective consultation with more 

than 70 federally recognized tribes to avert conflicts with American Indian trust resources on public 

lands and exercise of tribal treaty rights.

• Standards and guidelines that provided detailed 

requirements describing how land managers would 

treat forest lands within the range of the northern 

spotted owl.

• A new monitoring program consisting of implementation monitoring (are the standards and guidelines being followed?) and effectiveness monitoring 

(is the plan having the desired effect?).

• “Survey and manage” measures to provide for 

other late-successional species that may not be covered under the conservation strategies for the spotted 

owl and marbled murrelet, and for aquatic ecosystems and old-growth forests.

Reserves are a key component of the terrestrial and 

aquatic components of the NWFP and are discussed at 

length in chapters 3, 4, 5, 7, and 12. Reserves were intended 

to provide immediate and wide-ranging benefits for target 

species (e.g., spotted owls) and target ecosystems (oldgrowth forests, streams). Reserves were carefully delineated 

across the Plan area with the intention of improving ecological conditions for key Plan elements such as spotted owls or 

anadromous fish. We use monitoring results to evaluate how 

those conditions have changed and how well the underlying 

goals of the Plan have been met. 

The ROD for the NWFP amended the planning 

documents for 19 national forests.2 It is important to 

recognize that, over the past 24 years, implementation of the 

Plan across the entire area has varied from location to 

location. This can be attributed to geography and variation 

in how planning standards and guidelines have been 

interpreted by different forests, districts, and personnel over 

time. This is inevitable given the challenges of implementing a complex land management plan across a broad and 

diverse geography. The monitoring data we used to evaluate 

2

 The Northwest Forest Plan area currently includes 17 national 

forests; in 2000, the Okanogan and Wenatchee National Forests 

administratively merged as the Okanogan-Wenatchee National 

Forest, and in 2002 the Fremont and Winema National Forests 

administratively merged as the Fremont-Winema National Forest. 

The Plan area also includes five Bureau of Land Management 

districts and one resource area (formerly six districts and one 

resource area), with extensive standards and guidelines that 

comprised a comprehensive ecosystem management strategy. 
6

GENERAL TECHNICAL REPORT PNW-GTR-966

Ownership and land use allocations (2013)

Congressionally reserved

Late-successional reserve

Managed late-successional area

Administratively withdrawn

Adaptive management area

Adaptive management area in reserve

Matrix

Private

State

Tribal

0 50 100 200

Miles

0 75 150 300 o Kilometers

Figure 1-1—Land allocation categories and original 12 physiographic provinces (outlined in 

black) for the Northwest Forest Plan area. Note that “matrix” includes riparian reserves and other 

unmapped buffers (e.g., Survey and Manage). 
7

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

the NWFP are regional in scale and may not capture 

variability in Plan effects. In addition, unlike the effectiveness monitoring program, the implementation monitoring 

program has not been continued, making it difficult in some 

cases to determine what has actually occurred. The limits of 

the monitoring programs mean that some of our characterization of the Plan may not be correct. 

Decisionmakers considered monitoring to be an essential component of the selected alternative. Monitoring was 

intended to provide information to determine if standards 

and guidelines were being followed (implementation monitoring) and to verify if they were achieving desired results 

(effectiveness monitoring). In addition, a third type of monitoring, validation monitoring, was identified as a way to 

determine if underlying assumptions of the Plan were sound 

(this monitoring program was never formally established). 

The monitoring plan was subsequently cited by U.S. District 

Court Judge William Dwyer in his ruling upholding the 

Plan after challenges from the timber industry. The judge 

ruled that monitoring was a key element of the Plan and 

was essential to its success. Information obtained through 

monitoring, together with new research and experience 

gained through implementation, would provide the basis for 

adapting the Plan in the future (USDA 1994).

History of Reporting on the Research and 

Monitoring Within the NWFP Area

The NWFP involved the scientific community, through 

research and monitoring, in ways and to lengths not used 

before in Forest Service planning and management. The 

NWFP was driven, in large part, by a requirement to meet 

certain standards under the ESA and the viability clause 

of the NFMA, as well as by changes in land management 

related to three other federal laws (Thomas et al. 2006). 

These circumstances quickly triggered the need to engage 

scientists from the beginning, to provide both the planning and implementation process with robust, reliable 

scientific information.

The record of decision included the requirement of 

a detailed monitoring plan to ensure that management 

actions meet the prescribed standards and guidelines, and 

that actions complied with applicable laws and policies. 

Information obtained through monitoring, together with new 

research and information from adaptive management areas 

and studies, were intended to provide a basis for changes to 

the Plan, including changes to the standards and guidelines. 

Although a formal validation monitoring program was 

never established, research activities were conducted to help 

testing of hypotheses related to NWFP goals. 

10-, 15-, and 20-Year Monitoring Reports 

The NWFP was designed to include an adaptive management 

approach to enable “learning from doing.” The record of 

decision called for gathering information through an extensive monitoring effort, together with targeted new research 

and other new sources of information, to provide a basis for 

adaptive management and updating the selected alternative 

with new scientific knowledge. This set lofty aspirations 

for the scientific rigor of the Plan; however, there has been 

little adaptive management work done (i.e., actual designed 

experiments to test management strategies and assumptions 

in designated AMAs) since the Plan was initiated. 

Monitoring was designed for data collection at multiple 

scales, ranging from site-specific projects to the regional-scale planning area, to allow localized information to 

be compiled and considered in a regional context. Many 

but not all of the data sources used in the 20-year reports 

were initially developed and used for the 10- and 15-year 

monitoring reports. During each 5-year monitoring cycle, 

previously used data sources are updated to incorporate new 

research findings and other information, or to correct errors 

or previous misconceptions. So, to the extent possible, 

results are comparable between the two major reporting 

periods, but caution is suggested when examining topics 

that relate findings from one time period to the next because 

of minor analytical or reporting differences between 

monitoring reports.

Monitoring results have been evaluated and reported 

in 1- and 5-year intervals since the inception of the NWFP. 

The first comprehensive analysis of 10 years of NWFP 

monitoring data was published in a series of general technical reports (GTRs) summarizing what had been learned 

over that time. This was an important first step in adaptive 

management. The 10-year report synthesized the status and 
8

GENERAL TECHNICAL REPORT PNW-GTR-966

trends of five major elements of the plan: old-growth forests, 

old-growth forest species at risk, aquatic systems, socioeconomics, and adaptive management (Haynes et al. 2006). 

It also synthesized the new science that resulted from 10 

years of research related to the Plan. At this time, the cadre 

of researchers and managers also addressed four additional 

interconnected questions: 

1. Has the NWFP resulted in changes that are consistent with objectives identified by President Clinton? 

2. Are major assumptions behind the Plan still valid? 

3. Have we advanced learning through monitoring 

and adaptive management? 

4. Does the Plan provide robust direction for the 

future (Haynes et al. 2006)?

Based on the first 10 years of data collection, findings 

were ambiguous and conclusions hard to reach—perhaps 

unsurprisingly for a plan that was expected to take 100 

years to achieve its goals. It was clear that the complexity of 

ecosystem interactions and the effects of new drivers (e.g., 

encroachment of barred owls, climate change, and changes 

in social values) were far greater than had been envisioned 

10 years earlier. Nonetheless, insights into ecosystem 

response began to emerge, including circumstances and 

ecological interactions not contemplated at the time the Plan 

began. Rapp (2008) provided some highlights of the first 

decade of monitoring and research as follows:

• Nearly all existing old-growth forest on federal land 

was protected from timber harvest (although 100- 

percent protection was not part of the original plan). 

• Old-growth forest on federal land had an estimated 

net increase of roughly 1.2 million ac (~480 000 ha), 

increasing from 7.87 million ac (3.15 million ha) to 

9.12 million ac (3.65 million ha) in the first 10 years 

as a result of accretion by growth. 

• Despite protection of northern spotted owl habitat 

on federal land, spotted owl populations declined at 

a greater rate than expected in the northern half of 

their range, likely because of barred owl competition, and losses of habitat to wildfires. 

• Watershed condition improved slightly because of 

reduced harvest in riparian areas, tree growth, and 

increased emphasis on restoration. 

• Federal timber harvest in the NWFP area was only 

54 percent of the level set by the Plan’s goals. 

• In spite of mitigation measures, most local communities near federal lands suffered significant job 

losses and other adverse effects. 

• State, federal, and tribal governments worked 

together on forest management issues more effectively than in the past. 

• Increased collaboration with communities changed 

how the agencies get work done.

Recently, reports analyzing a full 20 years of monitoring data under the NWFP were released by the Regional 

Interagency Executive Committee and published as GTRs 

(Davis et al. 2015, 2016; Falxa and Raphael 2016; Grinspoon 

et al. 2016; Miller et al. 2017). These reports summarize 

the latest periodic monitoring data gathered since 1994, 

with a focus on the past 5 years. Some of the key findings 

contained in these new reports include:

• Overall late-successional and old-growth habitat area 

has decreased 3 percent on federal lands, with the 

biggest losses resulting from wildfires. However, this 

rate of loss was in line with expectations outlined in 

the FEMAT report during the design of option 9.

• Nesting habitat of the marbled murrelet showed a net 

decrease of about 2 percent on federal lands and 27 

percent on nonfederal lands.

• In Washington, there was an annual rate of decline 

of 4.6 percent in the population of marbled murrelets 

between 2001 and 2013; a cumulative decline over 

10 years of 37.6 percent. Populations had no detectable trends in Oregon and California. 

• The forest types suitable for nesting and roosting 

for northern spotted owls on federal lands decreased 

by 1.5 percent since inception of the NWFP. Forest 

succession is resulting in habitat recruitment that 

has compensated for losses resulting from wildfire, 

timber harvest, and insects and disease. However, 

suitable habitat (i.e., the full range of conditions 

necessary for a species to survive, persist, and 

reproduce) has declined more because of the influx 

of barred owls into forests with otherwise suitable 

forest vegetation throughout much of the range of 
9

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

spotted owls. Recent northern spotted owl research 

indicates that populations are declining throughout 

the range of the subspecies, and that annual rates of 

decline are accelerating in many areas. Dugger et 

al. (2016) observed strong evidence that barred owls 

negatively affected spotted owl populations, primarily by decreasing apparent survival and increasing 

local territory extinction rates. The amount of suitable owl habitat, local weather, and regional climatic 

patterns also appear to be related to demographic 

parameters, including survival, occupancy (via colonization rate), recruitment, and, to a lesser extent, 

fecundity (Dugger et al. 2016). 

• The attributes of watershed conditions (in-channel 

physical habitat, macroinvertebrates, and water temperature) showed slight improvements, but uncertainties in the trends of overall conditions remain. 

Upslope and riparian areas showed moderate, broadscale improvements in vegetation structure and 

larger score increases from road decommissioning 

in a number of watersheds. In the regional average, 

these increases were largely offset by declines in 

scores because of fires, particularly on congressionally reserved lands.

• Timber volume harvested has fluctuated over the 

past 20 years. The volume of timber offered has 

been on a general upward trend since 2000, with volume offered in 2012 at about 80 percent of probable 

sale quantity (PSQ) identified in the NWFP (based 

on revisions to the original PSQ of 1.1 billion board 

feet, as stated in the ROD, to a PSQ in 2012 of about 

805 million board feet).

• Rural communities are not all alike, forest management policies affect different communities differently, and the social and economic bases of many 

traditionally forest-dependent communities changed 

in the years since the start of the NWFP.

• Federal-tribal relations are more effective and 

meaningful when there is common understanding 

of consultation, tribal rights, federal trust responsibilities, and compatibility of tribal and federal 

land management.

Scope and Approach of This 

Science Synthesis

The PNW Research Station partnered with the PSW 

Research Station to prepare this synthesis, which was 

initiated at the request of Forest Service land managers. 

The two station directors guided this effort, and the 

day-to-day activities were led by Thomas Spies and Peter 

Stine. Other core team members included Matthew Reilly, 

Jonathan Long, and Becky Gravenmier. The core team, in 

consultation with the station directors, identified a group 

of experienced, knowledgeable scientists to serve as lead 

chapter authors. This put the responsibility for each chapter 

in one place and ensured that we would draw upon highly 

qualified sources. 

The public has expressed interest in this synthesis, 

given the importance of the NWFP in the management of 

Northwest forests and its influence on forest management 

approaches around the world. During listening sessions 

held in spring 2015 to gather feedback from the public 

about forest plan revisions, attendees provided suggestions 

relevant to the development and publication of this science 

synthesis. We heard many participants express a desire for 

continuous communication about the science, more access 

to scientific information, and participation in a greater variety of information-sharing venues. A number of steps were 

taken to enhance public input into this process, including 

a Web portal for submitting literature for consideration in 

the synthesis, and a public forum to accept oral and written 

public input to the peer review team.

Rationale for Topics Covered

Questions from managers guided the focus of the synthesis. 

The set of 73 management questions were grouped into the 

following major headings:

• Vegetation conditions, including forest management/climate change/ecological disturbance effects 

on old growth and other vegetation types.

• Terrestrial species, including habitat management 

for the northern spotted owl; marbled murrelet; and 

other plant, plant-ally, invertebrate, and vertebrate 

species, and conservation of the biodiversity associated with old-growth forests.
10

GENERAL TECHNICAL REPORT PNW-GTR-966

• Aquatic/riparian management, including aquatic 

and riparian species and ecosystems.

• Socioeconomic well-being, including timber production, collaborator and stakeholder attitudes, and 

tribal values and resources.

• Integrated topics: themes that cross over between 

chapters or separate management activities.

This synthesis is organized into 12 chapters, in three 

volumes, that include an introduction, 10 chapters addressing 

the primary topics of concern, and a final “integration” chapter 

that ties together what has been learned and reported in the 

various chapters and conveys how this synthesized knowledge 

bears on vital forest management activities. Each chapter 

provides a summary of the relevant scientific literature, lessons 

learned over the past 20 years, and the relevance of these 

findings to management. The synthesis does not provide management recommendations, nor does it conduct assessments of 

likely outcomes of different approaches to plan revisions. 

Sources of Information Considered

This science synthesis considered science published by 

peer-reviewed scientific or professional journals, or reviewed 

through an agency-sponsored, third-party process that meets 

the general criteria for competent and credible peer review. 

This process collected material from many sources, including an extensive body of original research and monitoring 

activities). In addition, academic theses, government reports, 

symposium proceedings, and the like may have been used to 

support certain topics that were not adequately covered in the 

peer-reviewed literature. Most of the literature considered was 

compiled by the authors based on their experience with the 

subject matter. In some cases, especially in chapter 3 (“Old 

Growth, Disturbance, Forest Succession, and Management in 

the Area of the Northwest Forest Plan”), some simple analyses of existing data were conducted to illustrate key ideas. 

Through a Web portal developed specifically for this purpose, 

we also provided opportunities for the public to suggest 

literature sources that we may not have already considered. A 

“Science Synthesis Literature Database” (https://www.fs.fed.

us/pnw/research/science-synthesis/literature-database.shtml) 

for the NWFP area lists all publications reviewed in this 

report, including many recommended by the public. 

Dealing With Scientific Uncertainty

There is always some degree of uncertainty embedded in 

scientific findings, especially related to our understanding of large and complex socio-ecological systems. The 

scientific literature in the fields covered by this synthesis 

does not necessarily address specific questions that land 

managers posed. Accordingly, chapter authors selected from 

a wider range of published research in an effort to reduce 

this uncertainty. To do so, we made judgments based on 

scientific consensus about how the findings of different 

scientific reports related to management questions, what 

the uncertainties are within published reports, and what the 

uncertainties are related to our interpretation of multiple 

reports. We report what is known about these topics with 

high confidence whenever possible, and describe what 

issues remain uncertain. 

In the FEMAT report, an expert evaluation process 

was used to address gaps in the scientific literature, as 

well as limits to our understanding, to better estimate 

the likely outcomes and risks to biodiversity associated 

with different conservation and management options and 

practices. FEMAT convened panels of scientific experts to 

rate the probabilities of viability outcomes for components 

of the Plan (such as northern spotted owls and aquatic 

functions) for the different Plan options. Although the 

FEMAT results and recommendations represented a consensus of scientific knowledge at the time, they contained 

considerable uncertainties, thus monitoring and adaptive 

management were regarded as being critical to the Plan’s 

scientific basis. This synthesis does not rely on an expert 

judgment process to fill large information gaps related to 

management questions or Plan trends. For example, we do 

not rate the probability of the long-term viability of the 

northern spotted owl in light of threats from barred owls 

or climate change. Although we use expert knowledge 

to interpret existing science, we avoid speculation about 

outcomes related to management effects, climate change, 

or other drivers or threats for which there is no published 

science. In this sense, the synthesis is more limited in 

scope than FEMAT was in the interface between science 

and policy. The process of assessing Plan alternatives, 

developing revisions to the standards and guidelines, 
11

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

or choosing actions in the face of uncertainties will be 

handled by federal land managers in subsequent steps 

of the upcoming planning precess. We report what is 

known to apprise managers of the best available scientific 

information and allow them to apply that information to 

their management concerns.

Role of Peer Review in This Document

Unlike FEMAT, the science synthesis has been subject to 

external peer review and revision based on those reviews. 

The Office of Management and Budget (OMB) explained 

the importance of peer review in its Information Quality 

Bulletin for Peer Review3 as follows:

Peer review is one of the important procedures 

used to ensure that the quality of published 

information meets the standards of the scientific 

and technical community. It is a form of deliberation involving an exchange of judgments about the 

appropriateness of methods and the strength of 

the author’s inferences. Peer review involves the 

review of a draft product for quality by specialists 

in the field who were not involved in producing 

the draft.

The OMB guidelines require that influential scientific 

information developed by a federal agency be subjected 

to formal, independent, external peer review to ensure its 

objectivity. Scientific knowledge is cumulative, building 

upon previous findings; therefore, safeguarding this trust 

is essential. Peer-reviewed science does not guarantee that 

what is presented is true or factual, because new information may overturn, refute, or refine previous findings. 

Peer-reviewed science is also not necessarily definitive 

because of the limitations of knowledge, current perspectives, and available studies. However, peer review is the 

standard within the scientific community for determining 

which findings meet and exceed adequate thresholds of 

scientific scrutiny. For these reasons, this science synthesis focused on material that has been peer reviewed and 

published in print or online. 

3

 https://www.gpo.gov/fdsys/granule/FR-2005-01-14/05-769.

Peer-reviewed published literature, however, is limited 

for some topics. For example, some social, economic, 

health, cultural, or highly specialized ecological topics 

tend to have less coverage in the peer-reviewed literature. 

To address such gaps, authors were given latitude to 

incorporate relevant scientific information from academic 

theses and other research subjected to some form of 

committee review. In some cases, analyses were done 

using existing data and with data sources identified and 

methods of analysis provided. For example, in chapter 3, 

we developed a new classification and map of NWFP fire 

regimes by synthesizing existing data on climate, lightning ignitions, potential vegetation types, and fire-history 

studies. In contrast, forest management strategies and 

plans such as the NWFP are generally not peer reviewed or 

based only on peer-reviewed information. National forest 

managers consider a host of other sources of information 

to inform their plan revisions and involve the public in 

forest plan development.

In general, the authors focused on peer-reviewed 

research that occurred in the synthesis area or in forest ecosystems with highly similar ecological or social conditions. 

Ecological and social research is always context-specific, 

thus we attempted to guard against use of overgeneralizations applied to areas apart from where the research was 

conducted. This can be especially true of the ecologically 

and socially diverse region of the NWFP. Scientific studies 

are often published with caveats about their spatial and 

temporal scale. However, many basic ecological processes 

are universal, thus we can apply some findings to other 

locations. Obviously, basic research cannot be conducted 

everywhere, so it is important to make prudent application 

of scientific findings from a given location to other areas. 

To address this challenge, the synthesis notes the extent and 

limitations of available information, especially by highlighting various research gaps.

This science synthesis has been identified as a “highly 

influential scientific assessment,” in accordance with the 

OMB’s 2004 peer-review bulletin (see footnote 3), which 

means that the information contained therein could have 

a large impact on the public or private sector, or be of 
12

GENERAL TECHNICAL REPORT PNW-GTR-966

significant interest to multiple agencies, or be controversial. 

For this report, we have employed an external peer-review 

process that includes multiple reviewers with relevant 

expertise and experience assigned to each of the chapters, 

and three reviewers who reviewed the entire document. The 

review was managed by the Ecological Society of America, which selected the review team from scientists with 

extensive experience and strong credentials, and managed 

the review process independently. 

The peer-review team, led by the Ecological Society of 

America’s director of scientific programs, Clifford Duke, 

was given basic instructions for conducting peer review in 

accordance with OMB direction for peer review of highly 

influential scientific assessments developed by federal 

agencies (USOMB 2002). Peer-review comments were 

delivered to the author team in March 2017, and authors 

used them to develop the final document. Authors also prepared reconciliation documents for each chapter explaining 

how all comments were used.

The NWFP Area

The establishment and implementation of the NWFP was 

unprecedented in many ways. Its geographic scope, breadth 

of topic areas, and long-term investment in monitoring and 

research all combined to set a new standard for large-scale 

land management.

The NWFP area covers 24 million ac (9.7 million ha) 

of federally managed land, extending from the Mendocino 

National Forest and Ukiah District of the BLM near the 

coast of northern California to the northern boundaries of 

the Mount Baker–Snoqualmie and Okanogan-Wenatchee 

National Forests on the Canadian border. The area spans 

almost 10 degrees of latitude and ranges from coastal 

rain forest landscapes to dry east-side pine forests. This 

expansive and diverse footprint created significant challenges for establishing management guidance and the 

scientific foundation needed to support it. By recognizing 

and embracing the variability of this landscape, NWFP 

managers intended for management efforts to be more 

nuanced and thus more effective at addressing particular 

features in any given area.

Ecogeographic Variability of NWFP Area

Efforts to classify and partition the natural world into 

component parts have been directed at many different levels 

of biological or ecological organization, from genes and 

species to communities and ecosystems (Grossman et al. 

1998). The NWFP area spans many biological community 

and ecosystem types and disturbance regimes, and the Plan 

goals include conservation strategies that focus on ecosystems as well as individual species. It is vital that the application of scientific findings within the Plan area recognize this 

broad geographic and ecological diversity. This concern 

is addressed in several chapters in which ecogeographic 

variation is central to careful treatment of management 

challenges (e.g., chapter 2 on climate, chapter 3 on oldgrowth forest, and chapter 5 on northern spotted owls). 

Climate, geology, disturbance, and topography all play 

important roles in controlling forest community patterns at 

regional scales in the Pacific Northwest (Barbour et al. 2007, 

Franklin and Dyrness 1973, Ohmann and Spies 1998). The 

relationships among environment, the biota, and disturbance 

differ across the region, making it precarious to extrapolate 

findings from one ecoregion to another. Kennedy et al. 

(2012) highlighted the importance of understanding the finer 

grain patterns of forest ecosystems within the NWFP area 

and their response to disturbances. This understanding is 

critical for delivering effective management insights across 

the many, sometimes subtly different, forest conditions distributed within the Plan area. The authors made a concerted 

effort to address this subject, as in chapter 12, “Integrating 

Ecological and Social Science to Inform Land Management 

in the Area of the Northwest Forest Plan.”

The NWFP area was originally partitioned into 12 physiographic provinces (see fig. 1-1) based on recognized landscape 

subdivisions exhibiting different physical and environmental 

features (Thomas et al. 1993). The resulting breakdown of 

provinces reflected the regional distribution of major forest 

types (and state boundaries for management purposes).

A number of qualitative approaches to classifying 

geographic variation have been used, including Ecoregions 

of the United States (Bailey 2009) and the Holdridge life 

zones, as discussed in Lugo et al. (1999). Quantitative 

ecoregionalization approaches are also available (e.g., 

Hargrove and Hoffman 2004, Hessburg et al. 2000), but 
13

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

are less often adopted by land managers because of the 

long-standing habit of using the more qualitative schemes. 

It is noteworthy that the quantitative schemes show highly 

intuitive, spatially disjunct patterns of ecoregions, which are 

largely absent in the qualitative approaches, suggesting that 

early delineations of ecoregional boundaries are inadequate. 

The various qualitative methods for identifying ecological 

regions use macroclimatic conditions (climate unaffected by 

landform), and prevailing plant formations as the means for 

classification (Bailey 2009). 

Vegetation classifications are a critical part of regional 

ecological characterizations. Vegetation can be classified 

based on successional potential (e.g., the late-successional 

vegetation that would develop in the absence of disturbance 

for a particular environment), or on current vegetation 

structure and composition. Both types of vegetation 

classifications are needed. The two Forest Service regions 

use different vegetation classification schemes (Region 

6 uses potential vegetation, and Region 5 uses actual or 

current vegetation [cover types]) (chapter 3), which makes 

it challenging to conduct a seamless ecological assessment 

across the entire Plan area. For this synthesis, we used the 

Region 6 potential vegetation classification and developed a 

crosswalk for linking the two types of classifications. 

We also now have access to ecological delineations that 

are more data-driven, using data models based on machine 

learning. An example is the habitat modeling developed for 

the northern spotted owl and contained within the recent 

recovery plan for this taxon (USFWS 2011). The effort, 

aimed at partitioning habitat in the range of the spotted owl 

(essentially the same as the NWFP area), used machine 

learning via MaxEnt (Phillips et al. 2006) to predict relative 

existing habitat suitability. Results of this data-driven effort 

provide a delineation of 11 “modeling” regions as oppossed 

to the 12 ecoregions originally described for the NWFP 

area. It is unclear how accurate these habitat suitability 

models are for predicting actual habitat suitability of different vegetation conditions for northern spotted owls. Barred 

owls, a significant component of current northern spotted 

owl habitat through much of its range, drastically complicate 

our ability to assess habitat suitability. Further work will be 

needed to understand spotted owl response in the different 

habitat regions delineated by this modeling work. 

Regardless of how this large Plan area is dissected, it is 

increasingly clear from recent scientific work that geography matters. The diversity of the NWFP landscape is both 

stark and subtle. We draw more specific attention to this 

issue throughout the following chapters. 

Other Syntheses Reports Relevant to the 

NWFP Area

The effectiveness of the NWFP was originally evaluated 

through a set of reports produced 10 years after its initiation 

(Haynes et al. 2006). This set included a series of status 

and trends reports, a synthesis of all regional monitoring 

and research results, a report on interagency information 

management, and a summary report. Although some 

existing science was synthesized in the 2006 report, it 

was not a comprehensive characterization of the literature 

and did not address a special set of questions posed by 

managers. Updated monitoring reports were produced in 

2009 and 2015 that evaluated the first 15 and 20 years of 

monitoring data developed under the NWFP (Davis et al. 

2015, and others). Each of these monitoring reports included 

key summaries of the results for each monitoring module, 

methods, and a set of recommendations for monitoring 

into the future. These monitoring reports did not include a 

broader evaluation of the scientific literature. 

Other efforts have been made in recent years to 

consolidate relevant scientific information within the Plan 

area. Notably, the Forest Service published The Ecology 

and Management of Moist Mixed-Conifer Forests in Eastern Oregon and Washington: a Synthesis of the Relevant 

Biophysical Science and Implications for Future Land 

Management (Stine et al. 2014). This synthesis overlapped 

with the NWFP area along the east Cascades of both 

Oregon and Washington and addressed some similar land 

management issues. 

Role of Science in Supporting Land Management

This synthesis will inform the development of revised 

land and resource management plans for 17 national 

forests by synthesizing relevant information on key topics 

and management questions across the NWFP area. The 

synthesis will directly support land managers’ ability to 
14

GENERAL TECHNICAL REPORT PNW-GTR-966

make decisions grounded in the best available science, 

and will provide managers with the needed foundation 

for assessments as required under the 2012 planning rule 

(USDA FS 2012). 

Context of the NWFP and Forest Plan Revision 

Under the New Planning Rule

The 2012 National Forest System Land Management 

Planning Rule brought forth a wide range of changes to 

the forest planning process through the most collaborative 

rulemaking effort in agency history. The agency’s goal 

was to implement an adaptive land management planning 

process that was inclusive, efficient, collaborative, and 

science-based, and that would promote healthy, resilient, 

diverse, and productive national forests and grasslands. This 

new rule is currently being used by national forests to revise 

forest plans that, in many cases, are 30 or more years old.

The 2012 planning rule, like the 1982 planning rule, 

sets a broader goal framework and direction for the NWFP 

revision. The National Forest Management Act requires 

the Forest Service to “provide for a diversity of plant 

and animal communities...to meet overall-multiple-use 

objectives” (Schulz et al. 2013). The 1982 rule required that 

this regulation be met by “maintaining viable populations 

of existing native and desired nonnative species in the 

planning area.” As a result, the 1994 NWFP emphasized 

viability of all species as a goal. This requirement imposed 

an administrative burden on the agency and proved quite 

difficult to accomplish and provided controversial results.

(Schultz et al. 2013). Consequently, the 2012 rule does not 

use viability of all species as a basis for conservation of 

biological diversity, but instead directs that maintenance 

of species be met through “coarse filter” (ecosystem) 

approaches that maintain ecological integrity, ecological 

functions, and habitat connectivity. The 2012 rule acknowledges that ecosystem-scale strategies do not necessarily 

provide for all species, and that a few species may require 

special attention as “species of special concern.” We do 

not make recommendations on how to revise the NWFP, 

given the changes in planning rule direction since the Plan 

was developed. However, the NWFP contained specific 

objectives pertaining to conservation strategies for both 

ecosystems (coarse filter) and particular species (fine filter) 

and how these were intended to meet biological diversity 

goals. In several places in this synthesis, we discuss the 

published scientific findings that convey the advantages 

and shortcomings of employing these different conservation tactics.

Another change in the 2012 planning rule, compared 

to the 1982 rule, is its emphasis on using planning that is 

adaptive, as well as to more fully base Forest Service land 

management on scientific findings. The rule acknowledges 

that the body of science that can inform land management 

planning in such areas as conservation biology and ecology 

has advanced considerably since the 1982 planning rule 

was drafted. The new 2012 rule thus calls for planning to 

include three phases: assessment, plan development/amendment/revision, and monitoring (fig. 1-2). The assessment 

phase prepares the staff on a national forest for subsequent 

efforts to consider a full range of options for plan revision, 

including evaluation of existing information about relevant 

ecological, economic, and social conditions, trends, and 

sustainability, and their relationship to the land management 

plan within the context of the broader landscape. Assessment, including landscape assessments and other supporting 

science, can include local or traditional sources of information in addition to peer-reviewed science. This framework is 

intended to support an integrated approach to the management of resources and uses, incorporates the landscape-scale 

context for management, and ideally will help the Forest 

Service adapt to changing conditions, while improving 

management based on new information and monitoring.

The assessment process is conducted and managed by 

a responsible official, usually the forest supervisor, who 

has the discretion to determine the scope, scale, and timing 

of an assessment. Importantly, this synthesis is intended 

to be available to responsible officials in time to support 

their plan revision process. It also will support subsequent 

monitoring efforts, which are also required under the 

new planning rule. Monitoring information is intended 

to enable planners to change plan components or other 

content as needed. 
15

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Given the pivotal role of science in the new planning 

rule, and the breadth and complexity of potential decisions 

in the NWFP area, development of this science synthesis 

was deemed essential to the entire plan revision process. 

The 17 national forests within the NWFP’s footprint are 

expected to revise their land and resource management 

plans in the near future under the guidance of the new rule. 

The regional foresters in Regions 5 and 6 have been charged 

with following the new rule’s detailed requirements, including the enhanced role of science in forest plan revisions. 

The new rule requires that: 

[the] responsible official shall determine what 

information is the most accurate, reliable, and 

relevant to the issues being considered. The 

responsible official shall document how the best 

available scientific information was used to inform 

the assessment, the plan decision, and the monitoring program as required in §§ 219.6(a) (3) and 

219.14(a) (4). Such documentation must: Identify 

what information was determined to be the best 

available scientific information, explain the basis 

for that determination, and explain how the information was applied to the issues considered.

Accordingly, the Regions 5 and 6 regional foresters 

have asked that this science synthesis provide a thorough, 

up-to-date review of the relevant scientific literature 

pertaining to key resource management topics within the 

NWFP area.

Figure 1-2—The science synthesis is part of the preassessment phase in forest plan revision and will inform the assessment phase of the 

planning process. NEPA = National Environmental Policy Act.
16

GENERAL TECHNICAL REPORT PNW-GTR-966

Emergent Issues 

Much has changed in the arenas of land management and 

science in the past 20-plus years. New issues have arisen that 

those designing or implementing the NWFP did not face at its 

inception. Going forward, some of these issues are particularly relevant to the fate of land management decisions within 

the NWFP area. The major considerations are summarized 

here briefly and amplified in subsequent chapters, particularly 

chapter 12, which explores various crosscutting themes and 

important implications for future forest plan revision.

Changing climate—

We devote an entire chapter (chapter 2) to the significance 

of climate change and the many ramifications it has on 

environmental conditions and on options that land managers 

have to achieve natural resource objectives. This issue has 

precipitated many shifts in conservation science and land 

management. Today, land managers are confronting difficult challenges and an uncertain future as they endeavor to 

mitigate climate effects through innovative management of 

forested landscapes. This development will continue to have 

a major impact on land management decisions throughout 

the NWFP area. Chapter 2 of this report is intended to lay 

a foundation for more indepth discussions of the realized 

and potential impacts of climate change on the other topics 

discussed in this synthesis. Although some core issues 

related to climate change are considered in chapter 2, 

additional chapters more specifically characterize climate 

change effects and concerns. 

Single-species and multispecies conservation strategies—

The NWFP revolved around a select number of species 

at risk within the overall Plan area. Conservation of the 

northern spotted owl and the marbled murrelet were 

principal objectives for the Plan, and much NWFP management direction revolved around their species-specific needs. 

Additional focus was placed on conservation of aquatic 

ecosystems that support the many taxa of anadromous fish 

throughout the planning area. These include 15 species of 

salmon and steelhead formally listed as threatened, and one 

listed as endangered, since the Plan was initiated.

Although these particular taxa remain a vitally important focus in the Plan area, there has been much discussion 

and contemplation in the scientific literature about land 

management strategies aimed at single species, as reflected 

in changes in the 2012 planning rule described above. 

Management strategies aimed at individual endangered 

species may not always be in alignment with strategies to 

conserve ecosystem function. There is no single path to 

resolve this dilemma; it is a matter of much scientific debate 

and a subject we explore in more detail in chapter 12.

Successional and disturbance dynamics—

Succession, disturbance, and other ecosystem processes 

create a wide array of structural and compositional conditions within any given vegetation type. A primary focus of 

the NWFP was to manage for the continued existence of 

“old-growth forests” and their associated species. Succession and disturbance are continuously operating to shape 

forests, both independently and in concert. These topics are 

addressed in great detail in chapter 3. 

The concept of ecological succession has been considered by ecologists for almost 200 years. More recently, 

however, the specific role of periodic disturbances (e.g., 

fire, windstorms, flooding) has been recognized as a critical 

element in shaping forests and promoting biological diversity by maintaining a variety of seral stages on landscapes. 

Disturbance ecology, especially fire ecology and the 

historical and contemporary role of fire within the NWFP 

area, has emerged in the past 30 years as a foundational 

science around which ecosystem management can be based. 

In many dry forests, simple models of successional change 

that were developed for moist forests do not apply because 

frequent fire regulated vegetation change in dry forests. 

Even within wetter forest areas, the effects of different 

historical disturbances, including fire, are important to consider in the conservation of important values (see chapters 3 

and 11). This means that strategies to conserve and restore 

biological diversity across the diverse NWFP area may 

differ strongly between forest types, especially between dry 

and moist forests. After 150 years of Euro-American land 

use, the effects of anthropogenic disturbances, both obvious 

and subtle, have altered forest ecosystems and plant and 

animal communities. Knowledge of human influences on 

disturbance regimes is fundamental to sustaining biological 

diversity and ecosystem resilience. 
17

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Historical range of variability—

In the early developmental stages of the NWFP, the 

concept of historical range of variability (HRV) and its 

use in ecosystem management was just emerging in the 

scientific literature for the Pacific Northwest (Cissel et al. 

1994). In the original discussions, this concept was useful 

for developing management goals for ecosystems that were 

based on inherent dynamics and processes rather than 

static structure targets. Although HRV is not explicitly 

referenced in the 2012 planning rule, the idea is addressed 

in directives for the rule in terms of “natural range of 

variability,” which is essentially equivalent (Wiens et al. 

2002). The rule does require forest plans “... to maintain 

or restore the ecological integrity of terrestrial and aquatic 

ecosystems and watersheds in the plan area,” where 

ecological integrity depends in part on the functioning of 

natural disturbance regimes, which typically occur within 

some natural range of variation for a given climatic period. 

This is especially relevant in considering the significant 

role of fire in many different forest types throughout 

the NWFP area. For example, managing for ecological 

integrity in forest types subject to moderate- to high-frequency fire is quite different than in forest types where fire 

occurs infrequently. The complexity of land management 

becomes more apparent as we consider not just a simple 

dichotomy of wet and dry forests, but instead a spectrum 

of precipitation and fire regimes as well as the importance 

of fine-scale heterogeneity. 

Research on changing climates has also emerged in the 

past 20 years, with a profound impact on our view of the 

HRV and its implications for management. We now face 

new scientific challenges in the restoration of degraded ecosystems, while managing for ecosystem resilience to climate 

change during the “Anthropocene,” a proposed term for the 

geological and ecological epoch in which human activity 

has been the dominant influence on landscapes, invasive 

species, and climate change. These new impacts make 

maintaining some historical ecological patterns and processes difficult or impossible to reestablish (Corlett 2015). In 

chapters 3, 4, and 12, we assess this dilemma by describing 

scientific findings about the resilience of a variety of forest 

types to climate change, and consider what the implications 

are for maximizing suitable habitat for northern spotted 

owls. The notion of HRV and its potential consequences on 

other topics is also considered in other chapters.

Invasion of the barred owl and use of the term “habitat”—

The term “habitat” is widely used in natural resources publications and popular literature to describe the environmental area inhabited by a particular species of plant or animal. 

However, the many variations on the precise meaning of 

this term can lead to confusion. In common usage, “habitat” typically focuses primarily on the forest cover type 

chosen to depict the age and structure of a forest, or, more 

generally, the vegetation type that typifies the structure 

and composition of vegetation preferred by a given species. 

We note this because such definitions of habitat typically 

miss features believed to be important in conveying the 

full array of conditions suitable for a species. In particular, 

we identify the influence of an array of ecological factors, 

especially the role of nonnative species. Their impact has 

prompted much discussion as to what people generally 

consider to be habitat for any given indigenous species. In 

this report, we define habitat as follows:

An area with the environmental conditions and 

resources (e.g., vegetation structure, food/prey, 

water, etc.) necessary for individuals of that species 

to survive and reproduce.

This definition specifically intends to draw attention 

to the phrase “environmental conditions,” which includes 

potential effects of competitors or predators, including those 

that may be nonnative species. Clearly, competition between 

spotted owls and invasive barred owls represents a profound 

impact on the suitability of habitat for spotted owls.

Landscape ecology and management—

For many decades, forest management was conducted at the 

stand scale. The stand was traditionally an operational unit 

used by forest managers to target local forest management 

objectives, largely around local timber production goals. 

However, social and scientific trends over the past 25 

years have led to broader scale silvicultural objectives and 

appreciation of more complex forest structures and nested 

scales for understanding forest dynamics. 
18

GENERAL TECHNICAL REPORT PNW-GTR-966

Landscape ecology has emerged as a discipline that 

embraces the inherent spatial variation in landscapes, 

expressed at a variety of scales. We now more thoroughly 

appreciate the relationship between pattern and process in 

landscapes; the relationship of human activity to landscape 

pattern, process, and change; and the effects of scale and 

disturbance on the landscape. Above all, we now understand 

and intentionally incorporate the biophysical and societal 

causes and consequences of landscape heterogeneity as part 

of a landscape management philosophy. Several chapters in 

this report give consideration to the emergence of a landscape point of view.

Changes in agency capacity and workforce—

Federal agency budgets, number of employees, and number 

of field offices in the NWFP area have dropped substantially 

since the Plan was implemented, in large part because of 

shrinking timber programs and related budget allocations. 

These reductions have been most pronounced in Forest 

Service Region 6, and least pronounced on BLM lands. 

Declines in budgets and staffing have decreased the capacity of agencies to accomplish forest management goals, 

including forest restoration. Community-based organizations, local business partners, environmental and recreation 

organizations, and other groups have helped fill critical gaps 

by raising money and providing labor to accomplish forest 

management goals on federal lands in the face of declining 

agency capacity. But communities must have means to play 

this role. Title II funding from the Secure Rural Schools 

and Community Self-Determination Act has also played 

a vital role in helping pay for ecosystem management and 

forest restoration work on federal forests. However, the 

future of this law is uncertain given that this law expired in 

2015 and it requires Congressional reauthorization. Thus, 

the issue of how to accomplish ecosystem management and 

forest restoration amidst reductions in agency capacity will 

continue to be a challenge.

Changes in wood processing infrastructure—

Wood processing infrastructure in Plan-area communities began declining in the 1980s. This decline has 

continued into the 2000s because of reduced demand for 

wood products from the Pacific Northwest, and in the 

supply available from federal forests, as well as because 

of changes in wood processing technology. Supply and 

demand of wood products is also influenced by a complex 

set of international market forces. Local supply is affected 

by changes in timber management resulting from policies 

and regulations that constrain available volume. Supply 

available to local markets is also significantly affected by 

international timber markets, which are entirely independent of federal forest policy. However, a decline in locally 

provided supply has had a profound impact on the local 

timber-processing industry, and its capacity to maintain its 

infrastructure. 

This current lack of infrastructure makes the sale of 

timber, small-diameter wood, and biomass less economical, 

owing to longer haul distances and reduced demand for 

wood products, factors that reduce stumpage prices. Not 

only does this create a financial barrier to accomplishing 

forest management goals on federal forests; it also poses 

financial challenges for private forest owners who face 

declining markets for their wood products. For mills to stay 

in business, or for investments in new infrastructure development to occur, a reliable supply of raw material is needed. 

Private lands may be unable to increase wood product 

production and still ensure sustainable harvest levels. Thus 

federal lands have an important role to play in providing a 

sustainable supply of wood products to keep existing wood 

processing infrastructure operating, and to expand it if 

desired through new investments. To date, federal forests 

in the NWFP area have not met the goal of ensuring a 

predictable supply of timber, nor have they met the probable 

sale quantity established by the Plan. This topic is treated in 

detail in chapter 8.

Evolving public values and public policies around 

natural resources—

Social scientists and policy analysts studying environmental 

values and attitudes in the United States documented a shift 

away from the predominantly commodity-oriented view of 

forest management, common prior to the 1980s, to a more 

mixed or balanced perspective that includes commodity and 

noncommodity uses. This shift in public values followed a 

series of policies initiated in the 1960s that placed greater 

attention on protection of wildlife, wilderness, air, and 
19

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

water, as well as a desire for improved relationships with 

tribal governments, to name a few concerns. 

Longitudinal studies conducted both on a national scale 

and in subregions of the United States indicate a gradual 

shift in public attitudes. Since the 1990s, attitudes about 

public lands have shifted from a sole focus on economic 

values, outputs, and commodities toward a greater diversity 

of values that includes noneconomic values, especially 

protection of ecosystems and aesthetic values. Sometimes 

this transition is described as a shift from an exclusively 

anthropocentric perspective to a balance of anthropocentric 

and biocentric perspectives. Residents of the NWFP area 

echoed this national trend. 

In reflection of this value shift, the Forest Service was 

one of the first public land management agencies to adopt 

an ecosystem management approach in the 1990s, one that 

aimed to conserve ecological services and restore resources 

while meeting the needs of current and future generations. 

In more recent years, public recognition of the dual focus 

of producing goods and services while protecting resources 

has gained ground, and the challenges in achieving this 

balance in a complex ecological system appear to be more 

widely understood. 

Ecosystem services—

The concept of ecosystem services was originally characterized by economist E.F. Schumacher as “natural capital” 

in 1973. Only recently has the concept become widely 

recognized as relevant to land and resource management. 

The 2005 Millennium Ecosystem Assessment (MEA 

2005) provided a simple definition of ecosystem services 

as “the benefits people obtain from ecosystems.” Historically, management efforts focused on the provision 

of such resources as water and timber. Currently, policy 

and management efforts have increased the appreciation 

and importance of the full suite of services derived from 

ecosystems, including nonprovisioning services such as 

spiritual and cultural heritage values. Our understanding 

of the full scope of ecosystem services and attendant 

societal values associated with Northwest forests is still 

emerging. Our aptitude for quantifying these values, 

particularly in monetary terms, will continue to evolve as 

methods improve.

Attitudes toward land management agencies—

Public lands management is an important element of public 

discourse in the national environmental policy arena. Some 

recent issues have been controversial in the public eye. The 

number of appeals and litigation of forest decisions provides clear evidence that social views about forest management are often polarized. Effective public engagement can 

help provide accessible processes for public deliberation. 

Studies have shown that public dissatisfaction with opportunities to participate has led to more appeals of agency 

decisions, and that participants desire public processes that 

are more collaborative.

An important factor shaping natural resource management outcomes is the degree of trust between land management agencies and the public. A lack of public trust in 

government is cited as a primary barrier in natural resource 

planning (see chapter 9) that potentially can lead to litigation 

or noncompliance, and, ultimately, to managerial impasse. 

Furthermore, trust has been shown to be correlated with 

social acceptability of forest management actions, although 

the actual causes of social acceptability are likely far more 

nuanced. There are two basic kinds of trust: institutional 

trust (trust in agencies to represent and serve the public), and 

interpersonal trust (trust cultivated based on personal relationships). When social trust is improved, there is greater 

support for land management policies. The assumption held 

by many is that trust can be built (and conflict reduced) 

through fair participation processes or transparent decisionmaking. Trust building occurs when stakeholders engage in 

meaningful dialogue in a context of shared power and high 

levels of substantive knowledge. Collaborative processes 

represent opportunities to build iterative experiences and 

develop relationships among multilateral stakeholders and 

between stakeholders and public land management agencies. 

Examples of how collaborations between the Forest Service 

and tribal governments and communities are facilitating 

cross-boundary management and pursuit of integrated social 

and ecological objectives are featured in chapter 11. These 

examples illustrate how local units and communities are 

working to fulfill the many goals for public lands management as reflected in the NWFP and the new planning rule, 

as well as the many challenges in that pursuit.
20

GENERAL TECHNICAL REPORT PNW-GTR-966

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B.G. 2000. Ecological subregions of the interior 

Columbia Basin, USA. Applied Vegetation Science. 

3(2): 163–180. doi:10.2307/1478995.

Johnson, K.N. 1997. Science-based assessments of the 

forests of the Pacific Northwest. In: Kohm, K.A.; 

Franklin, J.F., eds. Creating a forestry for the 21st

century. Washington, DC: Island Press: 397–409. 

Johnson, K.N.; Franklin, J.F.; Thomas, J.W.; Gordon, 

J. 1991. Alternatives for management of latesuccessional forests of the Pacific Northwest. Report to 

the Agricultural Committee and the Merchant Marine 

Committee of the U.S. House of Representatives. 

Corvallis, OR: Oregon State University, Department 

of Forest Resources. 59 p.

Johnson, K.N.; Swanson, F.J. 2009. Historical context 

of old growth forests in the Pacific Northwest—policy, 

practices, and competing world views. In: Spies, T.A.; 

Duncan, S.L., eds. Old growth in a new world: a Pacific 

Northwest icon reexamined. Washington, DC: Island 

Press: 12–28.

Kennedy, R.E.; Yang, Z.; Cohen, W.B.; Pfaff, E.; 

Braaten, J.; Nelson, P. 2012. Spatial and temporal 

patterns of forest disturbance and regrowth within the 

area of the Northwest Forest Plan. Remote Sensing of 

Environment. 122: 117–133. doi:10.1016/j.rse.2011.09.024.

Lugo, A.E.; Brown, S.L.; Dodson, R.; Smith, T.S.; 

Shugart, H.H. 1999. The Holdridge life zones of the 

conterminous United States in relation to ecosystem 

mapping. Journal of Biogeography. 26(5): 1025–1038. 

doi:10.1046/j.1365-2699.1999.00329.x.

Millennium Ecosystem Assessment [MEA]. 2005.

Ecosystems and human well-being: synthesis. 

Washington, DC: Island Press. 137 p.

Miller, S.A.; Gordon, S.N.; Eldred, P.; Beloin, R.M.; 

Wilcox, S.; Raggon, M.; Andersen, H.; Muldoon, A. 

2017. Northwest Forest Plan—the first 20 years (1994–

2013): watershed condition status and trend. Gen. Tech. 

Rep. PNW-GTR-932. Portland, OR: U.S. Department of 

Agriculture, Forest Service, Pacific Northwest Research 

Station. 74 p.

Ohmann, J.L.; Spies, T.A. 1998. Regional gradient 

analysis and spatial pattern of woody plant communities 

of Oregon forests. Ecological Monographs. 68(2): 

151–182. doi:10.1890/0012-9615(1998)068[0151:rgaasp]

2.0.co;2.

Phillips, S.J.; Anderson, R.P.; Schapire, R.E. 

2006. Maximum entropy modeling of species geographic 

distributions. Ecological Modelling. 190(3–4): 231–259. 

doi:10.1016/j.ecolmodel.2005.03.026.

Pipkin, J. 1998. The Northwest Forest Plan revisited. 

Washington, DC: U.S. Department of the Interior, Office 

of Policy Analysis. 117 p. https://reo.gov/documents/

reports/NFP_revisited.htm. (20 September 2017).

Rapp, V. 2008. Northwest Forest Plan—the first 10 years 

(1994–2003): first-decade results of the Northwest Forest 

Plan. Gen. Tech. Rep. PNW-GTR-720. Portland, OR: 

U.S. Department of Agriculture, Forest Service, Pacific 

Northwest Research Station. 42 p. doi:10.2737/pnwgtr-720.
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Schultz, C.A.; Sisk, T.D.; Noon, B.R.; Nie, M.A. 2013.

Wildlife conservation planning under the United States 

Forest Service’s 2012 planning rule. The Journal of 

Wildlife Management. 77(3): 428–444. doi:10.1002/

jwmg.513.

Spies, T.A.; Duncan, S.L. 2009. Old growth in a 

new world: a Pacific Northwest icon reexamined. 

Washington, DC: Island Press. 360 p.

Steen, H.K. 2004. The U.S. Forest Service: a history. 

Seattle, WA: Forest History Society in association with 

University of Washington Press. 356 p.

Stine, P.; Hessburg, P.; Spies, T.; Kramer, M.; Fettig, 

C.J.; Hansen, A.; Lehmkuhl, J.; O’Hara, K.; Polivka, 

K.; Singleton, P.; Charnley, S.; Merschel, A.; White, 

R. 2014. The ecology and management of moist mixedconifer forests in eastern Oregon and Washington: 

a synthesis of the relevant biophysical science and 

implications for future land management. Gen. Tech. 

Rep. PNW-GTR-897. Portland, OR: U.S. Department of 

Agriculture, Forest Service, Pacific Northwest Research 

Station. 254 p. doi:10.2737/pnw-gtr-897.

Thomas, J.W.; Franklin, J.F.; Gordon, J.; Johnson, K.N. 

2006. The Northwest Forest Plan: origins, components, 

implementation experience, and suggestions for change. 

Conservation Biology. 20(2): 277–287. doi:10.1111/j.1523-

1739.2006.00385.x.

Thomas, J.W.; Raphael, M.G.; Anthony, M.G.; 

Forsman, E.D.; Gunderson, A.G.; Holthausen, R.S.; 

Marcot, B.G.; Reeves, G.H.; Sedell, J.R.; Solis, 

D.M. 1993. Viability assessments and management 

considerations for species associated with late-succession 

and old-growth forests of the Pacific Northwest—the 

report of the Scientific Analysis Team. Washington, DC: 

U.S. Department of Agriculture, Forest Service, National 

Forest System. 530 p.

U.S. Department of Agriculture, Forest Service [USDA 

FS]. 2012. National forest system land management 

planning. Subpart A. 36 CFR Part 219. Federal Register. 

77(68): 21260–21276.

U.S. Department of Agriculture, Forest Service [USDA 

FS]. 2016. Ecosystem restoration. Forest Service 

Manual. Chapter 2020. Washington, DC.

U.S. Department of Agriculture, Forest Service; 

U.S. Department of the Interior, Bureau of Land 

Management [USDA and USDI]. 1994. Record of 

decision for amendments to Forest Service and Bureau 

of Land Management planning documents within the 

range of the northern spotted owl. [Place of publication 

unknown]. 74 p. 

U.S. Department of the Interior, Fish and Wildlife 

Service [USFWS]. 2011. Revised recovery plan for 

the northern spotted owl (Strix occidentalis caurina). 

Portland, OR. 258 p. 

U.S. Office of Management and Budget [USOMB]. 

2003. Ensuring the quality of data disseminated by the 

federal government. Federal Register. 67(36): 8452–

8460. https://www.whitehouse.gov/omb/fedreg_final_

information_quality_guidelines. (20 September 2017).
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Appendix: Priority Management Questions 

to Guide the Northwest Forest Plan 

Science Synthesis As Defined by Pacific 

Northwest and Pacific Southwest Forest 

and Regional Staff and Edited by the 

Science Synthesis Team 

The Northwest Forest Plan (NWFP) science synthesis was 

constructed based on a set of questions submitted by Forest 

Service land managers. The questions addressed concerns 

that developed from 24 years of experience in implementing 

the Plan, as well as new issues that have emerged since the 

Plan was initiated. The Science Synthesis Team reviewed 

an initial list of 190 questions submitted by Forest Service 

land managers and suggested additional questions that 

they believed were relevant and could be addressed in the 

synthesis. The team then removed redundant questions and 

grouped others to arrive at the final list of 73 questions delineated below. This list is sorted into four general topical areas 

that are covered in one or more of the 12 synthesis chapters. 

Based on available information, the synthesis attempted to 

fully or partially address all the questions. Although the 

chapters do not necessarily address these questions directly, 

they were organized to be consistent with the scientific 

understanding of the issues that these questions address. 

In each chapter, the management considerations section 

endeavored to more directly link the science to management 

issues related to these questions. To the extent possible, the 

synthesis addressed how the science differs by physiographic 

province, vegetation type, and disturbance regime.

Priority Questions

Vegetation/forest management/climate change/ecological 

disturbance (old-growth and other vegetation types)—

1. What is the latest science on active management, 

including “ecological forestry,” to protect and 

restore late-successional forests and maintain ecological diversity? 

2. How do the effects differ by treatment (mechanical and prescribed fire) in terms of key ecosystem 

components (structure, composition, connectivity, 

and function)? What are the associated costs and 

commodity outputs? 

3. What is the latest science on the dynamic landscape approach versus a fixed reserve system in 

terms of providing sustainable amounts and adequate distribution and connectivity of late-successional forest across the landscape? 

4. How does each approach allow us to adapt in 

response to large-scale disturbances? 

5. What is the relationship between amount and configuration of old growth and potential to sustain a 

variety of disturbance regimes and late-successional-dependent species?

6. How might management and conditions on other 

ownerships affect the above relationship with the 

understanding that old growth is likely to persist 

only on federal lands?

7. What is the latest science on treatments in stands 

greater than 80 years of age when the objective 

is to accelerate the development of late-successional habitat? 

8. Similarly, what is the latest science on limiting harvest of large trees (usually >21 inches diameter at 

breast height when conducting restoration activities?

9. What are the latest estimates for historical/natural 

range of variation (HRV/NRV)? What is the proportional mix of seral stages and special habitats 

(e.g., hardwoods, meadows, etc.)? 

10. What are estimates of patch and gap size, connectivity, disturbance (fire, insect and disease, 

drought), habitat, and within-patch heterogeneity?

11. What are important differences between “dry forests” vs. “wet forests” and how can these distinctions be used to prioritize restoration activities?

12. What does the latest science tell us about the 

concept about using HRV/NRV to inform ecological restoration, in terms of the mix of structural 

conditions, species composition, patch size, etc.? 

Does HRV/NRV help inform landscape-level patch 

dynamics and within-stand heterogeneity?

13. What are the effects, if any, on invasive species 

on old-growth forests and succession following 

disturbance?
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GENERAL TECHNICAL REPORT PNW-GTR-966

14. What is the competing science on restoration of 

Pacific Northwest forest systems? For example, we 

need to have an upfront discussion of differing viewpoints in the science on the need for restoration of 

late-successional/old growth (LSOG) in dry forests.

15. What is the relationship between retention of dead 

wood, including dead and damaged trees, and 

potential for disturbance in dry forests with a frequent fire regime?

16. How does dead wood affect our ability to maintain 

LSOG?

17. What is the relationship between retention of green 

trees in harvest units and ecological diversity and 

species viability? 

18. What is the relationship between green tree retention potential and insect and disease epidemics 

(especially dwarf mistletoe) in post-harvest or 

post-wildfire situations?

19. How does each approach allow us to adapt in 

response to large-scale disturbances? 

20. How do green tree retention effects differ by physiographic province and vegetation type?

21. What is the latest science on the connectivity of 

late-successional and other key habitats (fixed corridors versus landscape permeability)?

22. What does the current body of science suggest 

about postfire recovery options, including the social 

license and economics associated with salvage?

23. What are the ecological features associated with 

early-successional vegetation, and what is the role 

of early-successional vegetation in ecosystem function and biodiversity?

24. What are the potential conservation and restoration 

needs related to early-successional vegetation?

25. What are our most vulnerable ecosystems, species, 

and resources due to climate change? 

26. What are the key adaptation strategies that could 

mitigate these vulnerabilities? 

27. What different management strategies might be needed 

for forests and terrestrial and aquatic ecosystems?

28. How do we deal with uncertainty in our restoration 

efforts, models, and predictions?

29. What are the anticipated changes in climate within 

the NWFP area, and what are the potential impacts 

to disturbance processes (insect, disease pattern, 

drought, fire, etc.), vegetation, species habitats, 

aquatic ecosystems, and the provision of goods and 

services (timber, values, etc.) within the area?

30. What resources and components of a regional planning framework require analysis and consideration 

at the regional scale?

Terrestrial species/habitat management (northern spotted owl, marbled murrelet, other species associated with 

older forests)—

1. What is the latest science surrounding the effects 

of various treatments (silviculture, fuels) and 

wildfire on LSOG and plantations and what are the 

effects on terrestrial wildlife species, with particular attention on northern spotted owl (NSO), barred 

owl (BAOW), marbled murrelet (MAMU), and 

survey and manage (S&M) species? 

2. How or do these species use these treated habitats 

post-treatment, and are there ways to modify treatment to benefit these terrestrial species? 

3. How do these treated habitats compare to 

untreated habitat in terms of habitat use and reproductive success? 

4. How does use of treated and untreated areas compare to use of postfire habitats, including salvage? 

5. How do the risks of fire compare in treated and 

untreated habitats, and are the impacts of treatments by the risk of habitat loss due to fire?

6. What is the latest science on the interaction of 

barred owls and spotted owls and the impact to 

recovery of the spotted owl?

7. What is the relationship of fires to barred owl 

encroachment?

8. What is the current scientific understanding about 

the rarity of survey and manage species, and how 

effective are the management recommendations 

for habitat buffers in retaining these species across 

treated landscapes?
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

9. Is forest management under the NWFP providing 

habitat for rare and uncommon species as planned?

10. Are rare and uncommon species maintaining populations under NWFP management?

11. Have we accumulated enough information to 

change status of these species? Are there species 

originally ranked as having low potential for persistence that are now of less concern, particularly 

with the reduction in harvest levels of old growth 

we’ve seen under the NWFP?

12. Has the Interagency Special Status/Sensitive 

Species (ISSSP) program benefitted these species?

13. What is the effect of prescribed fire and wildfire on 

rare and uncommon species (S&M)?

14. Are known site buffers as effective as landscape 

scale habitat management in ensuring species persistence, dispersal and habitat connectivity? 

15. Does the current S&M species list truly represent 

currently rare species with population persistence 

questions dependent upon LSOG habitat?

16. Does the current NSO critical habitat better represent late-successional forest and provide for a higher 

level of assurance of persistence for NSO, MAMU, 

and S&M species when compared to the current 

NWFP late-successional reserve (LSR) network? 

17. Is there a difference in persistence in treated vs. 

untreated LSRs or LSOG habitat in the face of 

wildfire, insects and disease, and climate change?

18. What role and importance are riparian reserves 

and various buffer widths as terrestrial species 

(including mollusks) habitat, including dispersal 

and connectivity, and how does riparian reserve 

management impact the terrestrial species that 

utilize them?

19. How can we manage a riparian area for the variety 

of habitats needed? 

20. What is the status of other species of concern (not 

included as survey and manage species) within the 

footprint of the NWFP? 

21. What is the effect of pesticide use associated with 

cannabis cultivation or species viability (i.e. fisher)?

22. How can we manage for viable populations of 

snag-dependent species when snags are not present 

long-term on the landscape?

23. How can we identify important biological refugia? 

What are they and where are they?

Aquatic/riparian management (aquatic and riparian 

species and ecosystems)—

1. What is the current thinking/science on riparian 

thinning/management? Has it produced the desired 

results, including contributions toward recovery of 

listed fish species, impaired waters, and reduction 

of fire risk? 

2. What are the effects of common silvicultural 

treatments/prescriptions with respect to Aquatic 

Conservation Strategy (ACS) goals and objectives 

(especially riparian microclimate and stream temperature, wood recruitment, diversity in riparian 

species structure and composition, fish populations, terrestrial processes)? 

3. What are the effects of not managing previously 

harvested stands in riparian reserves (RRs)? What 

is the risk of severe wildfire in untreated riparian 

corridors, and do/how do various types of treatment reduce this risk?

4. What does the current science indicate regarding 

the value of woody material in second-growth 

riparian reserves? When and where should the creation of large wood be a purpose and need driving 

silvicultural treatment in riparian reserves? 

5. What does the current science indicate about the 

role of vegetation management in affecting ground 

water flows and temperatures, and how do those 

changes affect surface water?

6. Does current science indicate that the ACS is 

needed to achieve Plan goals of maintaining and 

restoring the ecological health of watersheds and 

aquatic ecosystems on public lands? 

7. Are all components (riparian reserves, key watersheds, watershed restoration, watershed analysis, 

ACS objectives, standards and guidelines, monitoring and evaluation) necessary to achieve these goals? 
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GENERAL TECHNICAL REPORT PNW-GTR-966

8. Does the current science indicate that refinements 

to the ACS may be needed to increase its efficacy?

9. Does ACS provide appropriate levels of connectivity or does it need to be refined?

10. What are the effects of interbasin water transfers 

and water diversions?

11. What does the current science indicate about 

where in the NWFP area the greatest potential 

for conflicts exist over water supply and demand 

for additional storage based on the current water 

supply and demand situation, projected changes 

in supply due to climate change, and projected 

changes in demand due to climate change and 

population growth.

12. How well have RRs met their intended objectives? 

13. Does current science support or refine Forest 

Ecosystem Management Assessment Team 

(FEMAT) conclusions regarding the role and 

function of RRs? If so, how? 

14. What have we learned since FEMAT that should be 

incorporated into RR designation and management 

in plan revisions?

15. What is the latest science on the effectiveness of 

treatments within riparian reserves, and implementation of varying riparian reserve widths?

16. Is the type, scope and scale of watershed restoration that has occurred over the life of the NWFP 

consistent with FEMAT and Plan assumptions? 

17. How effective are instream restoration treatments 

(e.g., large woody debris [LWD] augmentation, 

channel reconstruction) in achieving ACS objectives at multiple spatial and temporal scales? Fish 

passage restoration? Road decommissioning and 

improvements? Riparian restoration treatments 

(e.g., reforestation, thinning, gaps)? 

18. What does the current science indicate about 

potential short-term impacts to aquatic and riparian ecosystems when managing for long-term 

restoration of aquatic and ecosystem processes 

and functions (e.g., short-term stream temperature 

increases to achieve long-term large wood recruitment and normal disturbance processes)?

19. What are the consequences of the current road management regime on water and aquatic resources? 

Consider (a) the status and trends in the size of the 

road system on NFS and other federal lands, (b) the 

amount of the current system that poses a high risk 

to aquatic resource, and (c) the amount of the system that is being maintained or improved.

Social/economic (including timber production) (socioeconomic well-being, timber harvest; collaboration and 

stakeholder attitudes; tribal values and resources)— 

1. What does social science tell us about how stakeholders’ attitudes, beliefs, and values (ABV) have 

changed over the past 20 years, and how those ABV 

are associated with resource management (including 

recreational experience, resource use or protection)? 

2. How have stakeholders’ relationships to landscapes 

and natural resources changed in the Northwest 

Forest Plan area? 

3. What value do people place on cultural ecosystem services from public lands, including outdoor recreation?

4. What are the general conditions of and influences 

upon values of special concern to tribes (including 

first foods such as salmon, elk, huckleberry, camass root) in the NWFP area? 

5. What management strategies does science suggest would enhance these values of special concern to tribes? 

6. What does the body of science indicate are important factors contributing to successful collaboration 

in forest management?

7. Where are our most successful examples of such 

collaboration?

8. What are the most important factors in successful 

collaboration?

9. What strategies are suggested by science for 

engaging communities in forest plan revision in the 

NWFP area?

10. What are implications for forest management from 

trends in the size and socioeconomic status of lowincome, minority, and tribal populations (i.e., environmental justice populations) in the NWFP area? 
27

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

11. Are these populations growing?

12. What are the drivers of change related to socioeconomic well-being in rural communities? 

13. What are the implications for forest management of 

trends in socioeconomic well-being in rural communities in the NWFP area? 

14. How does the body of science inform sustainable 

recreation and social interest in valuing place (as 

required under the 2012 planning rule)? 

15. What does the science infer about the contribution 

of outdoor recreation across the region to social 

and economic sustainability?

16. What are the trends in outdoor recreation use and 

visitor satisfaction on public lands? 

17. What are the drivers for change related to recreation? 

18. What are the implications for forest management 

of changes in land use and ownership in the past 

20 years?

Other Topics to Be Considered in the Integration 

Section of the Synthesis (Pulled From Region 5 

and Region 6 Long List)

1. Influence of illegal marijuana cultivation on federal 

lands on resources (this was noted under terrestrial 

biological resources question #15, but effects on 

resources other than fisher will also be considered).

2. Effects of invasive species on forest succession and 

habitats (this topic is noted under vegetation question #10 in the context of old growth)

3. Salvage logging 

4. Conservation of nonfederally listed species (noted 

under terrestrial biology question #5)
28

GENERAL TECHNICAL REPORT PNW-GTR-966

Prescribed burn operations on the Wallowa-Whitman National Forest, Oregon. 

Photo by USDA Forest Service.
29

Chapter 2: Climate, Disturbance, and Vulnerability to 

Vegetation Change in the Northwest Forest Plan Area

Matthew J. Reilly, Thomas A. Spies, Jeremy Littell, 

Ramona Butz, and John B. Kim1

Introduction 

Climate change is expected to alter the composition, 

structure, and function of forested ecosystems in the 

United States (Vose et al. 2012). Increases in atmospheric 

concentrations of greenhouse gases (e.g., carbon dioxide 

[CO2]) and temperature, as well as altered precipitation 

and disturbance regimes (e.g., fire, insects, pathogens, and 

windstorms), are expected to have profound effects on 

biodiversity, socioeconomics, and the delivery of ecosystem services within the Northwest Forest Plan (NWFP, or 

Plan) area over the next century (Dale et al. 2001, Franklin 

et al. 1991). The ecological interactions and diversity of 

biophysical settings in the region are complex. The effects 

of climate change on ecological processes will occur 

through a variety of mechanisms at a range of spatial 

scales and levels of biological organization, ranging 

from the physiological responses of individual plants to 

the composition and structure of stands and landscapes 

(Peterson et al. 2014a). Understanding and incorporating 

how climate change projections and the potential ecological effects and uncertainties differ within the region (e.g., 

Deser et al. 2012) is essential for developing adaptation 

and mitigation strategies. 

Climate change has the potential to affect all ecological and socioeconomic components of the NWFP, as 

well as other objectives for federal forest managers in this 

region. However, climate change is only one factor that 

managers must consider when addressing conservation and 

other goals for the NWFP region. The overarching goal 

of this chapter is to lay a general foundation of current 

knowledge and understanding of climate change for the 

subsequent chapters in this synthesis report, and not to 

analyze and report the projected effects of climate change 

on all the different components of the Plan in detail. The 

chapters that follow address the role of climate change in 

the context of their particular topics (e.g., northern spotted 

owls, aquatic ecosystems). This chapter focuses on the 

following topics:

• Regional climate setting, including an introduction to the major vegetation zones and disturbance 

regimes of the region (see chapter 3 for a more 

detailed discussion of disturbance regimes)

• Climate history of the region from the Holocene 

through the 20th century

• Overview of climate modeling approaches and limitations

• Projected changes in climate and how these vary 

across the region

• Mechanisms of vegetation change and potential climate change vulnerabilities 

• Projected effects on vegetation at regional scales 

• Uncertainties associated with models and knowledge 

of climate change effects

• Management considerations and strategies for adaptation and climate change mitigation goals. (See 

chapters 3 and 12 for a more complete discussion of 

management options)

This chapter does not address broader issues of NWFP 

ecological and socioeconomic goals in the context of 

climate change. These topics are covered in chapter 12, in 

which climate change is considered along with other factors 

(e.g., nonnative species, ecosystem vs. species approaches 

to conservation, and tradeoffs) in a discussion of the science 

underlying the goals of the NWFP and the 2012 planning 

rule. This chapter is also guided by questions from managers, as follows: 

1 Matthew J. Reilly is a postdoctoral researcher, Humboldt State 

University, Department of Biological Sciences, 1 Harpst Street, 

Arcata, CA 95521; Thomas A. Spies is a senior scientist and John 

B. Kim is a biological scientist, U.S. Department of Agriculture, 

Forest Service, Pacific Northwest Research Station, 3200 SW 

Jefferson Way, Corvallis, OR 97331; Jeremy Littell is a research 

scientist, U.S. Department of the Interior, Geological Survey, 

Alaska Climate Science Center, 2160 Koyukuk Drive, Anchorage, 

AK 99775; Ramona Butz is an ecologist, U.S. Department of 

Agriculture, Forest Service, Pacific Southwest Region, 1330 

Bayshore Way, Eureka, CA 95501.
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GENERAL TECHNICAL REPORT PNW-GTR-966

Guiding Questions

This chapter addresses the following:

1. How did climate and vegetation change from the 

early Holocene to the late 20th century, and how 

did these changes vary across the NWFP area?

2. What are recent trends in climate change and how 

do they vary geographically across the NWFP area?

3. What are the major tools for projecting climate 

change and what are the associated uncertainties 

and limitations? 

4. What changes in climate are projected for the 

NWFP area and how do these projections differ 

across the region? 

5. What are the implications of recent and projected 

climate trends for vegetation change?

6. What are the mechanisms of vegetation change 

associated with climate change? 

7. Which ecosystems and species are most vulnerable 

to climate change? 

8. What are the key adaptation strategies that could 

reduce vulnerability to climate change? 

Background and Setting

The NWFP area covers approximately 24.4 million ac 

(9.9 million ha) and includes multiple physiographic 

provinces across Washington, Oregon, and northern 

California (fig. 2-1). These physiographic provinces 

encompass a variety of disturbance regimes (see chapter 

3 for more discussion and information) as well as a broad 

range of environmental and climatic gradients (fig. 2-2). 

Climate is cooler and wetter toward the north in the 

coastal and inland mountains, but transitions to a more 

Mediterranean climate with warmer, drier summers and 

greater interannual variability to the south (fig. 2-3). Most 

precipitation in the region falls during the winter months, 

often as snow at higher elevations. The Olympic Peninsula, Western Lowlands, and Coast Range are located 

in the western portion of the region. These receive the 

greatest annual precipitation and often experience a summer fog layer along the coast that can partially moderate 

summer moisture stress. The crest of the Cascade Range 

extends from northern Washington to northern California, bisecting much of the region and creating a steep 

gradient in precipitation from west to east. The western 

Cascades encompass a wide range of elevations, temperatures, and precipitation, which generally decreases 

toward the south. The eastern Cascades extend in a 

narrow band from Washington to the California border 

and are generally much drier than the western Cascades 

and most of the NWFP area. The Klamath Mountains, in 

southwest Oregon and northwest California, represent the 

most climatically and geologically diverse province in the 

area, with a strong west-to-east gradient in precipitation 

and summer moisture stress. The Willamette Valley 

makes up a relatively small portion of the NWFP area and 

is predominantly nonforested. 

The broad range of environmental and climatic gradients is reflected in the distribution of several potential 

vegetation zones across the region (figs. 2-1, 2-2, and 2-3) 

(Simpson 2013) (https://www.ecoshare.info/category/

gis-data-vegzones). Potential vegetation zones represent 

climax vegetation types that would eventually develop in 

the absence of disturbance; therefore, existing or current 

vegetation varies often within zones depending on seral 

stage (i.e., successional stage or stage of structural 

development) and time since disturbance. For example, 

the most abundant vegetation zone in the NWFP area, 

western hemlock (Tsuga heterophylla), is currently 

dominated by Douglas-fir (Pseudostuga menziesii). 

Vegetation zones provide an ecological framework for 

discussing climate and vegetation change across broad 

geographic extents (chapter 3). Vegetation zones have 

overlapping species pools but consist of unique plant 

community assemblages, as well as similar but internally 

variable biophysical conditions and historical disturbance 

regimes that differ geographically (Winthers et al. 2005; 

chapter 3). Vegetation zones have characteristic pathways 

of structural development that differ in complexity 

and reflect regional gradients in productivity as well as 

historical and contemporary disturbance regimes (Reilly 

and Spies 2015). 
31

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Figure 2-1—Geographic distribution of potential vegetation zones (Simpson 2013) and physiographic provinces within the 

Northwest Forest Plan area. 
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GENERAL TECHNICAL REPORT PNW-GTR-966

The major vegetation zones (figs. 2-1 and 2-4) of the 

region generally correspond to those presented by Franklin 

and Dyrness (1973) and were broken into moist and dry forests in the NWFP (chapter 3). This characterization is overly 

simplistic, as annual precipitation in any given zone varies 

geographically. Moist vegetation zones make up about 60 

percent of the region, and are primarily located in coastal 

areas and west of the Cascade crest. These include Sitka 

spruce (Picea sitchensis), redwood (Sequoia sempervirens), 

tanoak (Lithocarpus densiflorus), western hemlock, western 

redcedar (Thuja plicata), Pacific silver fir (Abies amabilis), 

and mountain hemlock (Tsuga mertensiana). Dry forest 

vegetation zones are located east of the Cascade crest, and 

also comprise a large portion of inland areas in southwest 

Oregon and northwest California. They include western 

juniper (Juniperus occidentalis), ponderosa pine (Pinus 

ponderosa), Douglas-fir, grand fir (Abies grandis) and white 

fir (Abies concolor), and subalpine forests dominated by 

subalpine fir (Abies lasiocarpa), Engelmann spruce (Picea 

engelmannii), and whitebark pine (Pinus albicaulis). A 

more detailed and comprehensive characterization of plant 

communities in individual vegetation zones can be found in 

Franklin and Dyrness (1973).

A B

14,413 ft

0 ft

4393 m

0 m

Elevation

Annual 

precipitation

719 cm 283 in

23 cm 9 in

Annual 

temperature

17 °C

-12 °C

63 °F

10 °F

C

Figure 2-2—Maps of (A) elevation, (B) annual precipitation, and (C) annual temperature in the Northwest Forest Plan area. Temperature 

and precipitation are derived from 30 arc-second (~800 m) PRISM (parameter-elevation regressions on independent slopes model) (Daly 

et al. 2008) grids averaged from 1971 to 2000, and were obtained from the Landscape Ecology, Modeling, Mapping and Analysis group 

at Oregon State University. Darker lines outline physiographic provinces shown in figure 2-1; lighter black lines show state boundaries.
33

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

More information on geographic variability and 

current vegetation in Oregon and Washington is available 

at Ecoshare (https://www.ecoshare.info/publications) and 

is discussed further in chapters 1, 3, and 12. Appendix 2-1 

provides a crosswalk for linking equivalent vegetation types 

between the Simpson (2013) vegetation zones and existing vegetation in northern California based on Regional 

76 °F

21 °F

24 °C

-6 °C

Summer 

temperature

Summer 

precipitation

43 in

3 cm 1 in

109 cm

Summer 

moisture

stress

High

Low

High

Low

Summer 

fog

A B C D

Figure 2-3—Maps of (A) mean summer temperature, (B) total summer precipitation, (C) summer moisture stress, and (D) summer fog 

in the Northwest Forest Plan area. Temperature and precipitation are derived from 30 arc-second (~800 m) PRISM (parameter-elevation 

regressions on independent slopes model) (Daly et al. 2008) grids averaged from 1971 to 2000, and were obtained from the Landscape 

Ecology, Modeling, Mapping and Analysis group at Oregon State University. Summer moisture stress was calculated by dividing 

summer temperature by summer precipitation for May through September. Summer fog is a proxy based on the optimal path length from 

coastline representing the easiest path of fog movement given topography and terrain blockage (Daly et al. 2008). Darker lines outline 

physiographic provinces shown in figure 2-1; lighter black lines show state boundaries.

Dominance 1 in the Pacific Southwest Region (Region 

5) CALVEG database. This crosswalk provides a means 

of interpreting the Simpson vegetation zones in terms 

of existing vegetation in California. More details on the 

CALVEG database are available at https://www.fs.usda.gov/

detail/r5/landmanagement/resourcemanagement/?cid=stelprdb5347192. 
34

GENERAL TECHNICAL REPORT PNW-GTR-966

Matthew Reilly Justin Braaten Andrew Merschel

Matthew Reilly Matthew Reilly

Tom Iraci

A B

C

E

D

F

Figure 2-4—Examples of forests from several vegetation zones illustrating the broad range of environmental and biophysical settings in 

the Northwest Forest Plan area: (A) western hemlock, (B) redwood, (C) mountain hemlock, (D) subalpine fir, (E) grand fir/white fir, and 

(F) ponderosa pine.
35

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Key Findings

Past Climate Change in the Northwest Forest 

Plan Area

The climate and vegetation of the NWFP area went through 

continuous change over the past 11,700 years during the 

Holocene. During this time, complex interactions between 

a fluctuating climate and fire drove vegetation change at 

millennial scales (Bartlein et al. 1998, Marlon et al. 2009, 

Walsh et al. 2015, Whitlock 1992, Whitlock et al. 2008). 

Species responded individualistically to changes in climate, 

sometimes forming assemblages that lack contemporary 

analogs (Whitlock et al. 2003). Species ranges expanded and 

contracted over time, with some species persisting in refugia where local conditions allowed persistence in regions 

where climate was generally unsuitable (Gavin et al. 2014). 

Refugia likely provided an important role in the persistence 

of populations through the numerous climatic transitions 

that occurred in the region since the last glacial maximum 

(Bennett and Provan 2008, Hampe and Jump 2011). 

Knowledge of vegetation changes during the Holocene 

is particularly rich in the NWFP area, and a number of 

paleoecological studies document change across the region. 

The Holocene is commonly divided into different periods 

that can be distinguished by climate and fire activity. We 

follow the divisions of Walsh et al. (2015) in a recent review, 

though other studies use different dates to delineate periods, 

and the timing of changes in climate and vegetation differ 

across the NWFP area (Whitlock et al. 2003).

Paleoecological studies use charcoal and pollen found 

in sediment cores from lakes, as proxies for past climatic 

conditions, and to reconstruct changes in vegetation composition over time (Whitlock et al. 2003). These studies are 

limited in terms of their spatial and temporal precision, but 

offer important historical context and insight on climate and 

vegetation change by broadening our understanding of the 

historical range of variability at millennial time scales. 

The early Holocene—approximately 12,000 to 8,000 

years before present (BP)—was a time of rapid vegetation 

change, with assemblages that include current subalpine 

and lower elevation species that lack modern analogs 

(Whitlock 1992). Increased summer insolation during 

this period led to higher summer temperatures and drier 

conditions than the present, while lower winter insolation 

led to cooler and wetter winters, likely amplifying seasonality and summer drought compared to present-day 

climate (Bartlein et al. 1998, Whitlock et al. 2001). 

Fire activity was relatively low at the beginning of the 

early Holocene, but increased and remained high until 

approximately 8,000 years BP (Briles et al. 2005, Walsh 

et al. 2015). Nonforested areas and open woodlands were 

replaced by forests as glaciers receded early in this period, 

and xerophytic species increased at many low-elevation 

sites across western Oregon and Washington as summers 

warmed (Walsh et al. 2015). 

As the climate warmed during the early Holocene, 

species responded individualistically and became distributed along elevational and latitudinal gradients (Whitlock 

et al. 2003). Douglas-fir, red alder (Alnus rubra), and oak 

(Quercus spp.) replaced spruce and pine at lower elevations 

in the Coast Range and western Cascades (Cwynar 1987, 

Grigg and Whitlock 1998, Long et al. 1998, Sea and Whitlock 1995, Walsh et al. 2008). On the Olympic Peninsula, 

herbaceous tundra was replaced by subalpine fir (Gavin et 

al. 2001). Mid-elevations of the eastern Cascades of Oregon 

were dominated by open pine (Pinus spp.) forests, initially 

with an understory of Artemesia, which likely transitioned 

into a closed-forest environment with a greater abundance 

of Abies spp. Mid-elevations of the Klamath Mountains in 

Oregon and California were dominated by open woodlands 

composed of Pinus spp., Quercus spp., and incense cedar 

(Calocedrus decurrens) (Briles et al. 2005, Daniels et al. 

2005, Mohr et al. 2000).

Cooler, wetter conditions were associated with decreasing summer isolation during the middle of the Holocene 

(~8,000 to 4,000 years BP) (Bartlein et al. 1998). During 

this time, fire activity decreased (Briles et al. 2005, Walsh 

et al. 2015), and modern species assemblages were formed 

in some parts of the region (Whitlock et al. 1992). Redcedar 

and western hemlock increased during this period across 

low- and middle-elevation forests of the Coast Range, 

the Cascade Mountains, and the Puget Trough (Cwynar 
36

GENERAL TECHNICAL REPORT PNW-GTR-966

1987, Prichard et al. 2009, Walsh et al. 2008). Species 

composition shifted toward silver fir, mountain hemlock, 

and Alaska yellow-cedar (Callitropsis nootkatensis) on 

the Olympic Peninsula (Gavin et al. 2001). In the Klamath 

Mountains, expansion of Pinus spp., Cupressaceae, and 

Abies spp. also indicated cooler, wetter conditions during 

this period (Briles et al. 2005, Daniels et al. 2005, Mohr 

et al. 2000). With the exception of lower elevations, fire 

activity increased again approximately 5,500 years BP 

(Walsh et al. 2015).

Fire activity continued to increase during most of the 

late Holocene (~4,000 years BP to present) despite evidence 

that this period remained cool and moist (Bartlein et al. 

1998, Walsh et al. 2015). There is little evidence in the 

pollen record to suggest major changes in the composition 

of vegetation assemblages across most of Oregon and Washington during this time (Walsh et al. 2008, 2015; Whitlock 

1992). Modern forest assemblages in the Douglas-fir and 

white fir zones established approximately 2,000 years 

ago in the Klamath Mountains, where fire activity also 

increased during this time despite cool and moist conditions 

(Briles et al. 2005, 2008; Daniels et al. 2005; Mohr et al. 

2000). Climate and fire fluctuated during the past 1,000 

years. The warmest temperatures occurred during the 

Medieval Climate Anomaly (MCA) (900–1250 CE) and 

the coldest temperatures during the Little Ice Age (LIA) 

(1450–1850 CE) (Steinman et al. 2012). Precipitation also 

varied during this time, but there is less consensus about 

this in the literature. Cook et al. (2004) argued that a period 

of drought occurred during the MCA, but more recent 

evidence suggests a wet MCA and dry LIA (Steinman et 

al. 2014). Fire frequency increased during the MCA in the 

Klamath Mountains (Daniels et al. 2005, Mohr et al. 2000) 

as well as the rest of the region in Oregon and Washington 

(Walsh et al. 2015). Many of the currently existing oldgrowth forests in moist vegetation zones established at this 

time (chapter 3).

Climate fluctuations associated with surface temperatures in the Pacific Ocean also became more apparent 

over the past 1,000 years (Nelson et al. 2011). Warming 

and cooling of sea surface temperatures in the equatorial 

Pacific Ocean, referred to as the El Niño Southern Oscillation (ENSO), result in periodic (2 to 7 years) anomalies 

that affect regional air temperature and precipitation. 

During the El Niño phase, winter and spring conditions 

are generally warmer and drier than average (McCabe 

and Dettinger 1999). During the opposite La Niña phase, 

winter and spring are generally wetter and cooler, leading 

to a deeper than average snowpack (Gershunov et al. 

1999). The Pacific Decadal Oscillation (PDO) is defined 

by fluctuations in sea surface temperature in the Pacific 

Ocean and has longer characteristic periodicity of 20 to 

30 years (Mantua et al. 1997), although the PDO is not 

consistent over time at these frequencies (McAfee 2014) 

and has exhibited variable regime transitions in the pre-instrumental period (Gedalof and Smith 2001). Newman et 

al. (2016) pointed out that the PDO is not an independent 

phenomenon, but a combination of multiple processes 

that include ENSO. The relationship between ENSO and 

PDO is weaker in northern California where the respective 

controls of ENSO and PDO on climate are less predictable 

(Wise 2010). 

Fire History

Regional drought driven by teleconnections with sea 

surface temperature anomalies (e.g., PDO, ENSO) resulted 

in synchronous occurrence of fires in the NWFP area (Hessl 

et al. 2004, Trouet et al. 2006, Weisberg and Swanson 2003, 

Wright and Agee 2004), as well as elsewhere in the Pacific 

Northwest and other regions of the Western United States 

(Heyerdahl et al. 2008, Kitzberger et al. 2007, Schoennagal 

et al. 2005). Several fire history studies document fire 

frequency over the past 400 years (table 2-1). Historical 

fire regimes differed among individual vegetation zones as 

well as geographically within vegetation zones (see chapter 

3 for an indepth discussion). Fire was generally infrequent 

in most moist vegetation zones but ranged from about 

50 years to >200 years, with synchronous, regional fire 

episodes occurring across the region from the 1400s to the 

mid 1600s, and again from the early 1800s to approximately 

1925 (Weisberg and Swanson 2003). Fire was far more 

frequent in dry vegetation zones, where return intervals 

were shorter, generally ranging from 10 to 50 years until the 

late 19th and early 20th century. 
37

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Table 2-1—Fire history studies in the Northwest Forest Plan area by vegetation zone (continued)Vegetation zone Study

Extent

(time period) Method

Frequency/ 

return interval Rotation

Low/moderate/

high

High-severity 

patch size

Hectares Years Years Percent Hectares

Redwood:

Stuart 1987 300

(1898–1940)

Scars 7.8 — — —

Finney and Martin 1989 ~600

(1300–1860)

Scars 10.1 — — —

Brown and Swetnam 1994 <1000

(171–1962)

Scars 9.9 — — —

Brown et al. 1999 Unknown Age, scars 7–13 — — —

Brown and Baxter 2003 20 316

(1550–1930)

Scars 6–20 — — —

Stephens and Fry 2005 ~1000

(1615–1884)

Scars 12 — — —

Western hemlock:

Means 1982 Unknown Scars 100 — —Fahnestock and Agee 1983 Western Washington

(pre-1934)

Age class from 

historical survey 

records

— 598 — —

Stewart 1986 <1

(~1200–1982)

Age, scars 50a — — —

Yamaguchi 1986 Unknown

(post-1480)

Age, scars 40–150 — — —

Teensma 1987 11 000

(1482–1952)

Age, scars 114 78 — —

Agee et al. 1990 3500

(1573–1985)

Age, scars 137 — — —

Morrison and Swanson 

1990

1940

(1150–1985)

Age, scars 96 95 0–86/0–60/

0–100

<110 ha

Garza 1995 3540

(pre-1910)

Age, scars 93–158 134 24–41/9–23/

25–54

Impara 1997 ~140 000

(1478–1909)

Age, scars 85 271 — —
38

GENERAL TECHNICAL REPORT PNW-GTR-966 Table 2-1—Fire history studies in the Northwest Forest Plan area by vegetation zone (continued) Vegetation zone Study Extent (time period) Method Frequency/ return interval Rotation Low/moderate/ high High-severity patch size Hectares Years Years Percent Hectares Wetzl and Fonda 2000 2500 (1400–1985) Age, growth release 21.3b — — — Agee and Krusemark 2001 26 000 (pre-1900) Age, live residual structure from air photos — 296 7–9/18–31/ 62–90 — Robbins 1995 ~1562 km2 (1700–1990) Age, scars 49 (2–191) — — — Olson and Agee 2005 ~7000 (1650–1900) Age, scars 2–167 — — —

Weisberg 2009 14 504

(1550–1849)

Age, scars — 162 — —

Wendel and Zabowski 2010 1873

(1568–2007)

Age, scars 127 140 — —

Silver fir:

Hemstrom and Franklin 

1982

~53 000

(1200–1850)

Age — 465 — —

Fahnestock and Agee 1983 Western Washington

(pre-1934)

Age class from 

historical survey 

records

— 834 — —

Agee et al. 1990 3500

(1573–1985)

Age, scars 108–137 — — —

Morrison and Swanson 

1990

1940

(1150–1985)

Age, scars 239 149 0–80/0–78/

0–100

<50

Garza 1995 3540

(pre-1910)

Age, scars 154–246 — 24–57/

20–22/45–50

—

Mountain hemlock:

Dickman and Cook 1989 18 000

(post-1400)

Age — — — >3200

Fahnestock and Agee 1983 Western Washington

(pre-1934)

Age class from 

historical survey 

records

— 598 — —
39

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Table 2-1—Fire history studies in the Northwest Forest Plan area by vegetation zone (continued)Vegetation zone Study

Extent

(time period) Method

Frequency/ 

return interval Rotation

Low/moderate/

high

High-severity 

patch size

Hectares Years Years Percent Hectares

Agee et al. 1990 3500

(1573–1985)

Age, scars 137 — — —

Subalpine:

Fahenstock and Agee 1983 Western Washington

(pre-1934)

Age class from 

historical survey 

records

— 800 — —

Agee et al. 1990 3500

(1573–1985)

Age, scars 109 — — —

Douglas-fir and grand fir/white fir:

Leiberg 1903 Southern Oregon

(~1900)

Historical land survey — — ~14 000 

Weaver 1959 Unknown Scars 47 — — —

Agee et al. 1990 3500

(1573–1985)

Age, scars 52–93 — — —

Agee 1991 197

(1760–1930)

Age, scars 16 37–64 — —

Bork 1984 ~100

(pre-1900)

Scars 8 — — ~400

Wills and Stuart 1994 ~20

(1745–1849)

Age, scars 10.3–17.3 — — —

Taylor and Skinner 1998 1570

(1627–1849)

Age, scars 14.5 19 59/27/14 —

Van Norman 1998 45 000

(1480–1996)

Age, scars 123 — — —

Brown et al. 1999 2000

(1820–1945)

Age, scars 7.7–13 — — —

Everett et al. 2000 3240–12 757

(~1700–1860)

Scars 6.6–7 11–12.2 — 2.4–40

Stuart and Salazar 2000 ~120

(1614–1944)

Age, scars 27 (12–161) — — —
40

GENERAL TECHNICAL REPORT PNW-GTR-966 Table 2-1—Fire history studies in the Northwest Forest Plan area by vegetation zone (continued) Vegetation zone Study Extent (time period) Method Frequency/ return interval Rotation Low/moderate/ high High-severity patch size Hectares Years Years Percent Hectares Taylor and Skinner 2003 2325 (pre-1905) Age, scars 11.5–16.5 19 — — Wright and Agee 2004 ~30 000

(1562–1995)

Scars 19–24 — — 10–100

Hessburg et al. 2007 ~72 000

(~1930)

Historical aerial photos — — 18/58/24 ~10 000

Baker 2012 140 400

(~1770–1880 

Live structure from 

historical inventory

— 496c 18/59/23 —

Ponderosa pine:

Weaver 1959 Unknown Scars 11–16 — — —

Soeriaatmadja 1966 (1500–5000)

Unknown

Scars 3–36 — — —

West 1969 Unknown Age — — — <0.26

Bork 1984 ~100

(pre-1900)

Scars 4–7 — — —

Morrow 1985 2

(pre-1900)

Age — — — <0.35 

Hessburg et al. 2007 ~106 000

(1930–1940)

Live structure from 

historical aerial 

photos

— — 30/58/12 —

Baker 2012 123 500

(~1770–1880) 

Live structure from 

historical inventory

— 705c 40/44/16 —

a Stewart noted 15 fires over a 750-year period.b Estimated at a 200-ha scale.c Rotation for high severity only.

Note: Most fire history studies are based on fire scars or identification of cohorts of trees with similar establishment dates. Fire frequency or return interval are the most commonly reported metrics of fire activity in fire history studies. Another metric related to fire frequency is fire rotation, or the time it takes to burn an area equal to the size of the area of interest. Relatively few studies report fire severity.
41

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

20th-Century Climate Change in the Northwest 

Forest Plan Area

Increases in temperature and precipitation across the NWFP 

area during the 20th century exceeded average global 

increases and vary across the region as well as among 

seasons (Abatzoglou et al. 2014b, Mote 2003). Most of the 

research examining 20th-century climate in the Plan area has 

been aggregated to the scale of individual states (i.e., California, Oregon, and Washington), or summarized for the entire 

Western United States, and there is less work that focuses 

specifically on the Plan area. There is evidence supporting 

both strong human-caused climate change (Abatzoglou et 

al. 2014a, 2014b) and temperature increases associated with 

ocean/atmospheric variability (Johnstone and Mantua 2014a, 

2014b). However, Abatzoglou et al. (2014a) demonstrated that 

natural factors alone cannot explain warming in the region.

Average annual temperature in western Oregon and 

Washington increased by 1.6 °F (0.91 °C) during the 20th

century, with the greatest increase of 3.3 °F (1.83 °C) 

occurring during winter (Abatzoglou et al. 2014b, Mote 

2003). Likewise, precipitation during the same period also 

increased by 13 percent, with the greatest increase of 37 

percent during spring (Abatzoglou et al. 2014b, Mote 2003). 

California also experienced accelerated warming since 1970 

(Cordero et al. 2011) and recently experienced the hottest, 

driest period (2012 to 2014) in the observational record 

(Mann and Gleick 2015). This same period also includes the 

lowest precipitation in recorded history (Diffenbaugh et al. 

2015) and potentially in the past 1,200 years (Griffin and 

Anchukaitis 2014). In northwestern California, Rapacciuolo 

et al. (2014) estimated that mean temperature increased by 

0.3 °F (0.18 °C). The same study estimated that minimum 

temperature increased by 0.9 °F (0.47 °C) and maximum 

temperature decreased by 0.4 °F (0.24 °C) during the 

20th century, although these trends were calculated using 

temporal differencing rather than traditional slope-based 

trends, and do not necessarily account for differences in the 

density of weather stations used in the study (Rapacciuolo 

et al. 2014). Twentieth-century trends in precipitation 

differed across northern California with evidence of overall 

increases (Killam et al. 2014) as well as slight decreases in 

some parts of the NWFP area (Rapacciuolo et al. 2014). 

Climate trends across the region are similar to those 

reported from studies across the Western United States. 

These studies indicate changes in several characteristics of 

weather relevant to forest and vegetation dynamics. Spring 

(March to May) temperature increased approximately 1.8 °F 

(1 °C) from 1950 to 1998 (Cayan et al. 2001) and snowpack 

declined during the latter half of the 20th century (Knowles 

2015, Mote et al. 2005). Increases in winter temperature are 

linked with decreases in snowpack (Mote 2006) and earlier 

snowmelt, which have altered streamflow timing (Hamlet et 

al. 2005; Jung and Chang 2011; Stewart et al. 2004, 2005).

Decreases in the proportion of annual precipitation falling 

as snow (Klos et al. 2014), the amount of water contained in 

spring snowpack (i.e., the depth of water if the snow were to 

melt) (Hamlet et al. 2005), and increased evapotranspiration 

from longer growing seasons increased soil water deficits 

since the 1970s (Abatzoglou et al. 2014b). A longer freezefree season, an increase in the temperature of the coldest 

night of the year, and increased potential evapotranspiration 

during the growing season also occurred during this period 

(Abatzoglou et al. 2014b). Fog frequency along the coast of 

northern California declined by 33 percent during the 20th

century (Johnstone and Dawson 2010), as has low summertime cloudiness (Schwartz et al. 2014). Most recently, northern California experienced a dramatic shift with extreme 

drought conditions from 2012 to 2016 followed by extreme 

precipitation events and severe flooding (Wang et al. 2017). 

Remote-sensing studies indicate that most vegetation zones 

accross the NWFP area have already experienced moisture 

stress associated with drought and high temperatures during 

the early 21st century across the entire NWFP area (Asner et 

al. 2016, Cohen et al. 2016, Mildrexler et al. 2016).

Projecting Climate Change for the 21st Century 

Atmosphere-ocean general circulation models (GCMs) are 

the primary tools for projecting future climate scenarios 

(e.g., IPCC 2014). GCMs incorporate interactions among 

several important components of the Earth’s climate system, 

including atmosphere, land, ice, and ocean to simulate past 

and future climate at relatively coarse spatial scales (~0.25 

to 14 mi2 [~0.65 to 36.3 km2]) based on different scenarios of increasing greenhouse gas concentrations in the 
42

GENERAL TECHNICAL REPORT PNW-GTR-966

atmosphere. Because of differences in model formulation 

and sensitivity to forcing from physical influences on the 

atmosphere (e.g., greenhouse gases), GCM projections using 

the same initial conditions and emissions scenario differ 

(Lynn et al. 2009), as do projections from the same GCM 

owing to natural climate variability within a region (Deser 

et al. 2014). 

An ensemble of projections (combinations of projections from multiple GCMs) is commonly used in climate 

change studies to capture the range and patterns of 

variability among projections. Ensemble averages appear 

to provide the best estimates of observed climate (Pierce 

et al. 2009, Rupp et al. 2013). The range of projections 

in an ensemble also provides a measure of the amount of 

uncertainty, which increases as projections extend farther 

into the future (Tebaldi and Knutti 2007). Uncertainty 

in climate change projections can be attributed to three 

main factors: (1) climate change-scenario uncertainty, (2) 

model-response uncertainty, and (3) natural variability in 

climate (Hawkins and Sutton 2009). For a given climate 

change scenario, uncertainty in the warming estimates 

arises from differences in GCM formulation and parameterization. Natural climate variability presents the greatest 

uncertainty in the near to mid term for projecting climate 

change for the first half of the 21st century (Hawkins and 

Sutton 2009) and poses a major challenge for analyzing and 

communicating climate change variability within a region 

(Deser et al. 2012).

For its fifth and most recent assessment (AR5), the 

Intergovernmental Panel on Climate Change published a 

set of future scenarios that describe estimated trajectories 

of greenhouse gas concentrations. These scenarios are 

called representative concentration pathways (RCP), and 

each scenario is named after the increase in radiative 

forcing relative to preindustrial levels. Each pathway is 

the result of plausible future trends in human population 

growth, economic and technological development, and 

energy systems, as well as social beliefs and values that 

affect human behaviors influencing emissions and climate 

warming (van Vuuren et al. 2011). Climate change scenarios 

(e.g., climate changes that are likely given a specific RCP) 

are considered to be plausible and do not have probability 

distributions associated with them (Collins et al. 2014). 

Current rates of greenhouse gas emissions have exceeded 

previously anticipated concentrations, thus there is currently 

insufficient information to rule out any scenario (Manning 

et al. 2010, van Vuuren et al. 2010). All scenarios project 

increases in global mean temperatures, but there is a large 

range among the scenarios bracketing the low and high ends 

of potential greenhouse gas concentrations. Under the RCP 

2.6 scenario, which represents strong mitigation action, 

global mean temperatures are projected to increase by 2.9 

°F ± 0.7 °F (1.6 °C ± 0.4 °C) by the end of the century, while 

under RCP 8.5, the no-mitigation, high-growth scenario, the 

degree of warming is projected to be 7.7 °F ± 1.3 °F (4.3 °C 

± 0.7 °C) (Collins et al. 2014). Changes in global precipitation are projected to increase 0.5 to 4 percent/°C under RCP 

2.6 and by 1 to 3 percent/°C under other scenarios (Collins 

et al. 2014). 

Many relevant studies, especially in northern California, use an earlier generation of climate change scenarios 

published in the Special Report on Emissions Scenarios 

(Nakicenovic and Swart 2000). In this set of scenarios, 

the A2 scenario represents a very heterogeneous world 

with continuously increasing global population. The B1 

scenario represents a convergent world in which population peaks mid-century, then declines, transitioning 

to resource-efficient technologies. The B2 scenario 

is intermediate between A2 and B1, with population 

growth lower than the A2 and a less rapid transition to 

resource-efficient technologies.

21st-Century Climate Change Projections for the 

Northwest Forest Plan Area

Analysis of GCM projections for Oregon and Washington 

(Mote et al. 2014) and northern California (Cayan et al. 

2008, 2016; Garfin et al. 2014) depict a future with significant warming by the end of the 21st century, although 

the magnitude of warming varies at finer scales across the 

region. In Oregon and Washington, Dalton et al. (2013) 

projected increases in annual average temperature of 4.3 

°F (2.4 °C) and 5.8 °F (3.2 °C) by the middle of the century 

(2041 to 2070) under RCP 4.5 and RCP 8.5 scenarios, 

respectively. By the end of the century (2070 to 2099), 
43

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

average annual temperature is projected to warm by 5.9 

°F (3.3 °C) to 17.5 °F (9.7 °C), depending on the scenario 

(Mote et al. 2014). Warming is projected to occur across all 

seasons, with the greatest temperature increases occurring 

during summer months (Dalton et al. 2013). 

Projected changes in precipitation are more uncertain 

in Oregon and Washington. Some models project a 10 

percent decrease in annual precipitation by the end of the 

century (2070 to 2099) while others project as much as 

an 18 percent increase in precipitation (Mote et al. 2014). 

GCMs generally project wetter winters and drier summers 

(Dalton et al. 2013). Under the A2 and B2 scenarios, no-analog temperature conditions are projected by 2100 across 

much of the western Cascades and Klamath Mountains 

compared with those occurring in the recent past (Saxon et 

al. 2005). Under RCP 8.5, most of Oregon and Washington 

are projected to depart from their historical climate regime 

by 2050, when the mean annual temperature of a given 

location will exceed the 20th-century range of variability 

(Kerns et al. 2016). 

In northern California, under the mitigation-oriented 

B1 scenario, annual temperature is projected to increase 

by 2.7 °F (1.5 °C) by 2100, and, under the high-growth 

A2 scenario, the increase is projected to be 8.1 °F (4.5 °C) 

(Cayan et al. 2008). Simulations depict drier futures under 

the B1 and A2 scenarios, with total annual precipitation 

decreasing by 18 percent in the more extreme A2 scenario 

(Cayan et al. 2008). Increases in temperature are projected 

for all seasons across northern California, with the greatest 

increases occurring during summer months (Cayan et 

al. 2008). Projected decreases in summer precipitation 

range from 4 to 68 percent, whereas projected changes in 

precipitation during winter months range from a 9 percent 

decrease to a 4 percent increase. More recent projections of 

increases in winter precipitation using the RCP 8.5 scenario 

show a high degree of agreement among models (Neelin 

et al. 2013). Interannual variability is expected to increase 

with the occurrence of greater wet and dry extremes during 

the wet season (October to March) (Berg and Hall 2015). 

Most of northern California is projected to depart from its 

20th-century climate by the year 2040 (Kerns et al. 2016). 

The projected future climate in the Klamath Mountains 

represents conditions of temperature and precipitation not 

experienced in the recent past by 2100 under the A2 and B2 

scenarios (Saxon et al. 2005). Temperature is projected to 

depart the 20th-century range of variability between 2046 

and 2065 under the A2 scenario (Klausmeyer et al. 2011).

Implications of Observed Climate Trends for 

Water Balance Deficit and Vegetation Change

Changes in the magnitude and seasonality of temperature 

and precipitation patterns will most likely affect vegetation 

by altering the availability of water in the soil. Cumulatively, these are expected to be experienced ecologically 

through hotter periods of drought and greater deficits 

in water balance. Water-balance deficit for vegetation is 

defined as the difference between potential evaporation and 

actual evapotranspiration (Stephenson 1998). Ecologically, 

the water-balance deficit equates to the difference between 

the atmospheric demand for water from vegetation and the 

amount of water that is actually available to use. Even if 

precipitation remains similar to 20th-century levels, projected increases in temperatures could reduce the amount of 

soil moisture available for plants. 

Projections for changes in water-balance deficits differ 

among models (Littell et al. 2016) and across the region (fig. 

2-5). The majority of the region is projected to experience 

an increased summer (June, July, August, and September) 

water-balance deficit during the middle part of the 21st

century. The eastern Cascades, Klamath Mountains, and 

southern portion of the western Cascades in Oregon will 

likely experience the greatest increases in water-balance 

deficit, as well as the southeastern portion of the Oregon 

Coast Range and the northern portion of the California 

Coast Range. The least amount of change is projected in 

the northern portions of the Coast Range along the Pacific 

Ocean. Higher elevations of the Olympic Peninsula and the 

northern portion of the western Cascades in Washington 

are projected to experience less summer water-balance 

deficit in the future.

Although trends in average temperature and precipitation provide some context for vegetation change in 

the future, individual weather events are also expected 

to be important drivers of future dynamics (Jentsch et 
44

GENERAL TECHNICAL REPORT PNW-GTR-966

al. 2007). Climate extremes (e.g., acute drought) related 

to changes in the variability of temperature and precip-

itation may have disproportionate effects on vegetation 

and result in rapid vegetation change (e.g., Allen and 

Breshears 1998). Increased frequency and intensity of 

heat waves and extreme temperatures are predicted across 

North America by the end of the 21st century (Meehl and 

Tebaldi 2004). Prolonged heat waves (Bell et al. 2004), as 

well as dry daytime and humid nighttime heat waves, are 

projected in northern California (Gershunov and Guirguis 

2012). Models project increases in the number of both dry 

days and very heavy precipitation days during the wet 

season in northern California (October to March) (Berg 

and Hall 2015). This is consistent with an intensified 

water cycle characterized by shifts from extreme drought 

to years with anomalously high precipitation (Wang et 

al. 2017). Increases in peak flow magnitudes also sug-

gest greater potential for flooding in portions of inland 

northern California (Das et al. 2013), where floods may 

be more frequent and severe (Dettinger 2011, Salathé et 

al. 2014). Heavy precipitation events from warming and 

shifts in seasonal precipitation patterns may also increase 

flooding in most of Oregon and Washington (Tohver et 

al. 2014) and the northern California Coast Range (Kim 

2005). Rain-on-snow events may also be more common 

given warmer winter and spring temperatures, which are 

also projected to alter the timing of seasonal streamflow 

(Elsner et al. 2010). The availability of regional climate 

model outputs provides the climatic basis for better sim-

ulating physically consistent extremes relevant to forests 

processes (e.g., McKenzie et al. 2014, for fires), but these 

outputs are also subject to the constraints of GCMs used 

as boundary conditions.

Water balance 

deficit (inches)

-2 to -1.9

-1.9 to -1

-0.9 to 0

0 to 0.9

1 to 1.9

2 to 2.9

3 to 3.9

4 to 4.9

Figure 2-5—Projected changes in summer (June, July, August, and 

September) water-balance deficit across the Northwest Forest Plan 

area for 2030–2059 from a composite of the 10 best general circula-

tion model projections based on the CMIP3/AR4 scenarios following 

Littell et al. (2016). Higher water-balance deficit (browns) means 

decreased water available for plant uptake. Change is compared 

to the water-balance deficit from 1916 to 2006. Map boundaries 

correspond with the physiographic provinces in figure 2-1.
45

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Considering the coarse resolution of climate projections 

(~0.25 to 14 mi2 [~0.65 to 36.3 km2]), it is important to recognize the potential for landscape-scale variability in future 

climate and vegetation change. Differences in vegetation 

structure and topography can drive fine-scale variation 

in temperature extremes, with differences in maximum 

and minimum temperatures of similar magnitude to those 

projected at a broader scale in different climate change 

scenarios (Suggitt et al. 2011). Spatial variability in bedrock 

geology also has the potential to mediate seasonal changes 

in groundwater availability associated with increased 

temperature (Tague et al. 2008). Complex topography 

and cold air pooling may decouple climate conditions in 

mountain valleys from the surrounding landscape (fig. 2-6) 

(Daly et al. 2009), and snow may persist later in the season 

in canopy gaps and topographic depressions (Ford et al. 

2013). Temperature is generally lower and soil moisture 

higher in interior late-successional forests than in clearcuts 

or edges (Chen et al. 1993), and denser canopies can attenuate warming by providing shade to the forest floor (De 

Frenne et al. 2013). Recent findings also indicate that dense, 

old-growth forests in moist vegetation zones of the region 

have the potential to provide cooling effects at local scales 

(Frey et al. 2016). Thus, the actual changes in future climate 

experienced by an organism may differ depending on their 

tolerances or habitat preference. 

Figure 2-6—Projected changes in maximum December temperatures in response to a 2.5 °C regional temperature increase and changes 

in atmospheric circulation patterns in the western Cascade Range of Oregon. Source: Daly et al. (2010).
46

GENERAL TECHNICAL REPORT PNW-GTR-966

The potential for relatively stable climatic conditions 

at finer scales in some landscapes (e.g., topographically 

complex, mountainous terrain) suggests an important role 

for climatic refugia to contribute to the future persistence 

of some species (Noss 2001). Despite the conceptual appeal 

and historical importance of refugia, identification of refugia 

has proven difficult and has been largely descriptive, and 

refugia are likely to be species- and process-specific (Keppel 

et al. 2012). Refugia will most likely be found in topographically complex landscapes where microclimates differ 

because of differences in aspect, shading and insolation, 

and cold-air drainages (Dobrowski 2011). These areas may 

provide potential for species persistence through unfavorable 

climatic conditions, as well as sources for future recolonization provided that suitable conditions return in the future.

Mechanisms of Vegetation Change

Climate change is expected to alter vegetation through 

direct effects (e.g., from CO2 and climate on vegetation 

processes) and indirect effects (e.g., from disturbance 

processes). The direct effects of climate change and 

increasing CO2 on vegetation are expected to be expressed 

through changes in mortality, growth, and reproduction, all 

of which may be sensitive to altered phenology and biotic 

interactions within and among species (Peterson et al. 

2014a). The indirect effects of climate change are expected 

to be expressed through increases in the frequency, severity, and extent of disturbances, particularly drought, fire, 

insects, and pathogens. These have the potential for rapid 

ecological change at landscape scales, and are predicted to 

be a greater driver of ecological change than direct effects 

(Dale et al. 2001, Littell et al. 2010). The relative importance 

of these drivers, however, is likely to vary geographically 

across the region among species, seral stages, physiographic 

provinces, and disturbance regimes. Species are expected to 

respond individualistically to future changes in climate as 

they have in the past (Whitlock 1992).

Direct effects of climate change: demographic responses—

Tree mortality from higher temperatures and drought stress 

has already occurred in many forests of the Western United 

States, and is expected to increase in the 21st century (Allen 

et al. 2010, 2015). Warmer temperatures and increased 

frequency and duration of droughts projected for the NWFP 

area are likely to increase climate-induced physiological 

stress on plants (Adams et al. 2009). Drought-related stress 

can lead to two separate, but not mutually exclusive, mechanisms of tree mortality including hydraulic failure (irreversible desiccation and collapse of water transport structures) 

and carbon starvation (McDowell et al. 2008). Although 

there has been much recent work on the physiological mechanisms associated with tree mortality, a greater understanding of these mechanisms is needed to assess vulnerability 

among species and enhance our ability to predict mortality 

(Hartmann et al. 2015). Furthermore, a better understanding 

of the ecological consequences of mortality in terms of 

community-level change (i.e., structure and composition) 

and ecosystem function is needed (Anderegg et al. 2012). 

Mortality rates in old-growth forests in the Plan area 

have increased above most published rates (>1 percent/year) 

since the mid 1970s (van Mantgem et al. 2009). A regional 

study on mortality rates on Forest Service lands in Oregon 

and Washington corroborated the occurrence of elevated 

mortality rates in old-growth forests across all vegetation 

zones from the mid 1990s to mid 2000s during regionwide 

drought (Reilly and Spies 2016). However, Acker et al. 

(2015) found that mortality rates in old-growth forests on 

National Park Service lands (Olympic National Park, North 

Cascades National Park) in western Washington were lower 

than those reported by van Mantgem et al. (2009) and Reilly 

and Spies (2016). Lower mortality rates could be due to 

geographic variation not represented in van Mantgem et al. 

(2009) and Reilly and Spies (2016), but may also be indicative of decreasing stress-related mortality following a period 

of elevated mortality. Consistent with this idea, Cohen et al. 

(2016) found that remotely sensed forest decline peaked in 

the mid 2000s during the warmest decade in the past 100 

years (Abatzoglou et al. 2014b), then decreased. 

Increasing tree mortality rates have been documented 

in young stands of other regions, and some researchers 

suggest that they may be more vulnerable to changes in 

climate than old-growth stands (Luo and Chen 2013). 

However, Reilly and Spies (2016) found that mortality 

rates in early- and mid-seral stages from the mid 1990s to 

mid 2000s were lower than rates in young forests in the 
47

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

western hemlock and silver fir zones of the western Cascades (Larson et al. 2015, Lutz and Halpern 2006). With 

the exception of old-growth forests, in which increased 

mortality led to cumulative losses in basal area and 

density (van Mantgem et al. 2009), there is generally poor 

understanding of the effects of recent mortality on stand 

structure and composition, as well as how these effects 

differ around the region. 

The potential response of tree growth to climate 

change differs substantially among species depending on 

the factors that limit growth such as water and length of 

growing season (Littell et al. 2010, Peterson and Peterson 

2001). Growth in Douglas-fir is predicted to decrease under 

climate change where it currently is water limited (Restaino 

et al. 2016), but growth may increase where Douglas-fir is 

limited by growing-season length or lower than optimal 

temperatures (Albright and Peterson 2013; Creutzburg et 

al. 2017; Littell et al. 2008, 2010). In species of high-elevation forests where growth is limited by temperature 

and growing-season length (e.g., subalpine fir, mountain 

hemlock), growth increased during the 20th century 

because of warmer winter temperatures and longer growing 

seasons (McKenzie et al. 2001, Nakawatase and Peterson 

2006, Peterson et al. 2002). Warmer winters and earlier 

snowmelt may also increase potential for drought and water 

stress in higher elevation forests, especially toward the 

southern portion of their distribution in southern Oregon 

and northern California. However, these effects are not yet 

well documented or understood, and increased growth is 

expected to continue in the future (Albright and Peterson 

2013). The effects of projected climate change on ponderosa 

pine is uncertain as wetter fall seasons may increase growth 

while drier summers decrease growth (Kusnierczyk and 

Ettl 2002). These effects may differ across the landscape as 

ponderosa pine and western juniper may be more sensitive 

to drought at lower elevations (Knutson and Pyke 2008). 

The response of these species also depends on the potential 

for CO2 to enhance growth by increasing water-use efficiency (Soule and Knapp 2006). However, some evidence 

suggests that any benefits of CO2 fertilization will be 

outweighed in the future as the climate warms and water 

becomes a more limiting factor (Gedalof and Berg 2010, 

Restaino et al. 2016). Increased levels of CO2 also have the 

potential to accelerate maturation and increase seed production (LaDeau and Clark 2001, 2006), but little information 

is available on the effects of climate change on reproduction 

in species of the region. 

The ability of a species to respond to changes in climate 

(e.g., earlier warming and drying) with shifts in phenology 

will be an important factor in determining responses to 

projected climate change. Altered seasonality may affect 

growth and reproduction in some plant species. A major 

concern in the NWFP area associated with warmer winters 

and earlier springs is the requirement for many species 

(e.g., Douglas-fir, western hemlock, Pinus spp., Abies spp.) 

to experience chilling for the emergence of new leaves, or 

budburst (Harrington and Gould 2015). Douglas-fir may 

experience earlier budburst in some portions of its range 

because of warming, but reduced chilling may cause later 

budburst in the southern portion of its range (Harrington 

and Gould 2015). Earlier growth in northern and higher 

elevation portions of Douglas-fir’s range may lead to earlier 

growth initiation, but reduced chilling in the southern and 

lower elevation portions of its range are likely to lead to 

delayed growth initiation (Ford et al. 2016).

Climate change may also affect interactions among and 

within species in complex ways, but the effects are currently 

poorly understood. However, several recent studies from 

higher elevation moist forests in the silver fir vegetation 

zone of Washington provide some insights. For example, 

the negative effect of competition on growth is likely to be 

greater for saplings than for adults, and climate change may 

have less effect on closed-canopy forests at lower elevations 

than at higher elevations (Ettinger and HilleRisLambers 

2013). Individual growth is likely to increase most in lower 

density stands as trees may show little response to climate 

at higher density (Ford et al. 2017). Little is known about 

the effects of climate change on positive species interactions 

(e.g., facilitation), though they are known to be important in 

stressful subalpine environments elsewhere in the Western 

United States (Callaway et al. 2002), and are thought to play 

a role in early stand development in dry and cold vegetation 

zones (e.g., ponderosa pine, subalpine, mountain hemlock) 

in the NWFP area (Reilly and Spies 2015).
48

GENERAL TECHNICAL REPORT PNW-GTR-966

Indirect effects of climate change: disturbance—

The indirect effects of climate change will likely be 

expressed through increases in the frequency, severity, and 

extent of disturbance, and are predicted to be the primary 

mechanisms of ecological change in the future (Dale et al. 

2001, Littell et al. 2010). Disturbances include discrete events 

that alter the structure and function of ecosystems (Pickett 

and White 1985), but may also include prolonged droughts or 

multi-year epidemics of pathogens and insects. Disturbance 

agents are commonly characterized as biotic (e.g., pathogens, 

insects) or abiotic (e.g., fire, wind, volcanoes), and differ 

considerably in terms of their prevalence and severity (i.e., 

tree mortality) across the region and among vegetation zones 

(Reilly and Spies 2016) (chapter 3). There is great concern 

that interactions among climate change, forests, and disturbance regimes may result in disturbance effects outside of 

the natural range of variation (Dale et al. 2000). 

Of particular concern are multiple, successive, or 

compound disturbances (e.g., Paine et al. 1998). Interactions 

among multiple disturbances may result in multiplicative 

effects on the structure and function of ecosystems that 

differ from the cumulative effects of both individual 

disturbances. The effects of compound disturbances are 

difficult to predict, but may amplify disturbance severity, 

cause changes between ecological states (e.g., forest to 

nonforest transitions), and decrease forest resilience (Buma 

2015). However, despite growing recognition and interest in 

interactions among disturbances, the effects of compound 

disturbances remain poorly characterized and difficult to 

predict (Buma 2015, Seidl et al. 2017).

Biotic disturbances—

Biotic disturbances (e.g., insects and pathogens) elevate 

stand-scale mortality above what are considered normal 

“background mortality rates” associated with competition 

and stand development, but may also erupt into epidemic 

outbreaks that result in high levels of tree mortality (e.g., 

Raffa et al. 2008). Insects and pathogens do not always 

result in immediate tree mortality. However, the resulting 

decline in tree growth and vigor (Hansen and Goheen 2000, 

Marias et al. 2014) may initiate a long process of mortality 

(Manion 1981), making trees less resistant to wind disturbance and predisposing them to stem breakage (Larson and 

Franklin 2010). Although mortality rates associated with 

insects are generally much lower than those associated with 

fire in this region (Reilly and Spies 2016), insects resulted 

in greater loss of live carbon (Berner et al. 2017) and greater 

canopy mortality (Hicke et al. 2016) than fire in recent years 

at the regional scale. 

Native insects and pathogen activity is expected to 

increase as trees experience more stress associated with 

growing-season drought; however, the implications and 

magnitude of their effects are likely to be variable and differ 

geographically as well as among species (Chmura et al. 

2011, Kolb et al. 2016a, Sturrock et al. 2011). In addition to 

affecting host species, climate change will also affect population dynamics and geographic distributions of pathogen 

and insect species. Pathogen activity is likely to increase in 

areas where they typically infect drought-stressed host species, while the effects of climate change on pathogens that 

proliferate under moist conditions may be more variable and 

difficult to predict (Sturrock et al. 2011). Warmer winters 

and hotter droughts are expected to enable insects to move 

into previously unsuitable habitat (Bentz et al. 2010, 2016), 

and some regions in the Western United States experienced 

what are considered unprecedented outbreaks of insects in 

the past few decades (e.g., Raffa et al. 2008). Drought and 

insects may also interact to further stress trees and predispose them to mortality, but these dynamics are complex and 

are just beginning to be understood (Anderegg et al. 2015).

Native pathogens play a prominent but variable role in 

the disturbance regimes of both moist and dry vegetation 

zones of the region (Goheen and Willhite 2006, Hansen 

and Goheen 2000) (see Shaw et al. 2009 and chapter 3 for 

more information on insects and pathogens). Most native 

pathogens affect small, localized areas at low levels of 

tree mortality, but are pervasive and generally widespread 

across the region (Reilly and Spies 2016). Pathogens often 

initiate forest canopy gaps and can accelerate successional 

dynamics in old-growth Douglas-fir-dominated forests of 

the western hemlock vegetation zone (Holah et al. 1997). 

Laminated root rot (Phellinus sulphurascens) (formerly

weirii) affects Douglas-fir, true firs (Abies spp.), and 

mountain hemlock. Armillaria (Armillaria ostoyae) affects 

Douglas-fir, hemlocks (Tsuga spp.), pines (Pinus spp.), and 
49

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Engelmann spruce. Annosus root disease (Heterobasidion 

annosum) affects firs, pines, hemlocks, and Engelmann 

spruce. Black stain root disease (Leptographium wageneri) 

affects Douglas-fir and ponderosa pine. Several other types 

of pathogens are also present, including rusts (Cronartium

spp.) and mistletoes (Arceuthobium spp., Phoradenron spp.).

In the Coast Range, Swiss needle cast (Phaeocryptopus 

gaeumannii) is a disease specific to Douglas-fir that has 

increased since the early 1990s (Hansen et al. 2000b). 

Ritóková et al. (2016) found that the area affected by Swiss 

needle cast more than tripled between 1996 and 2015, with 

growth reductions of 23 percent in the Oregon Coast Range. 

Swiss needle cast is predicted to increase in the Oregon 

Coast Range in response to warmer and wetter conditions 

in the future (Stone et al. 2008), although an increase in 

drought conditions may inhibit spread of the disease (Rosso 

and Hansen 2003). High-density Douglas-fir plantations 

near the coast, where Sitka spruce and western hemlock 

were historically dominant, are thought to be particularly 

vulnerable to Swiss needle cast (Black et al. 2010, Hansen 

et al. 2000, Manter et al. 2003, Rosso and Hansen 2003). 

An extensive list of research studies of Swiss needle cast is 

available at http://sncc.forestry.oregonstate.edu/publications. 

Several species of insects, including bark beetles and 

defoliators, are also native to the NWFP area. Insects are 

more prevalent in drier vegetation zones and affected large 

areas east of the Cascade Range in recent decades (Hicke 

et al. 2016, Meigs et al. 2015). In Oregon and Washington, 

recent mountain pine beetle outbreaks were positively 

associated with warmer winter temperatures and negatively 

associated with drought stress and precipitation in the current 

and previous year of outbreak (Preisler et al. 2012). Mountain 

pine beetle has the potential to cause extensive mortality in 

lodgepole pine (Pinus contorta) and also affect other species 

of pines, including ponderosa pine, sugar pine (Pinus lambertiana), western white pine (Pinus monticola), and whitebark 

pine. Defoliating insects are also common, and though they 

often do not result in mortality, they may reduce growth 

and make trees more susceptible to other insect infestations. 

Several species of pine are susceptible to outbreaks of pandora moth (Coloradia pandora), and ponderosa pine is also 

susceptible to pine butterfly (Neophasia menapia). Spruce 

budworm (Choristoneura occidentalis) is a major concern 

east of the Cascade Range and affects Douglas-fir and true 

firs. Williams and Liebhold (1995) projected decreases 

in the area defoliated by spruce budworm with increased 

temperature alone, but the area increased with increases in 

temperature and precipitation. Douglas-fir is also susceptible 

to Douglas-fir beetle (Dendroctonus pseudotsugae), which 

operates on small patches of trees, especially after blowdown 

from wind events (Powers et al. 1999).

Several nonnative pathogens and insects are of particular 

concern in the NWFP area. White pine blister rust (Cronartium ribicola) is a major threat to whitebark pine (Goheen 

et al. 2002, Ward et al. 2006) as well as both western white 

pine and sugar pine (Goheen and Goheen 2014). Decline of 

Pacific madrone (Arbutus menziesii) related to multiple fungal 

diseases has been reported over the past 30 years, with larger 

older trees experiencing the most mortality (Elliott et al. 

2002). Balsam woolly adelgid (Adelges piceae) has affected 

subalpine fir and especially grand fir at lower elevations west 

of the Cascades (Mitchell and Buffam 2001). In southwest 

Oregon and northwest California, sudden oak death (caused 

by Phytophthora ramorum) has the potential to spread through 

air, water, and infected plant material (Peterson et al. 2014b, 

Rizzo and Garbelloto 2003) and may affect tanoak, various 

species of oak (e.g., California black oak [Quercus kelloggii]), 

other hardwood species (e.g., Pacific madrone and bigleaf 

maple [Acer macrophyllum]), and several species of shrubs 

(e.g., Rhododendron spp.) (see chapter 3). Warmer, wetter winters intensify risk of infection (Haas et al. 2015), and the area 

affected by sudden oak death is predicted to increase tenfold 

by the 2030s under projected warmer and wetter conditions 

(Meentemeyer et al. 2011). Sudden oak death is also associated 

with increased fire severity on soils in northwest California 

(Metz et al. 2011). Port Orford cedar (Chamaecyparis lawsoniana) is susceptible to a lethal, nonnative root pathogen 

(Phytophthora lateralis) that can be spread over long distances 

via organic matter carried on boots, vehicles, and animal 

hooves, and by water (Hansen et al. 2000a, Jules et al. 2002). 

Recent work suggests that despite rapid initial spread and 

colonization of Phytophthora lateralis, the rate of spread has 

slowed greatly since 2000 (Jules et al. 2014). 
50

GENERAL TECHNICAL REPORT PNW-GTR-966

Abiotic Disturbances

Abiotic agents of disturbance in the NWFP area include 

windstorms, fire, volcanic eruptions, landslides, and avalanches. These disturbances result in much higher levels of 

tree mortality than biotic disturbances, and are the primary 

natural agents of stand-replacing disturbance (Reilly and 

Spies 2016). Abiotic disturbances can create forest gaps 

and patches of mortality that range in size depending on 

the disturbance agent (Spies and Franklin 1989). Smaller 

gaps created by abiotic disturbances may increase stand and 

landscape heterogeneity, while large, infrequent disturbances may have effects on landscape composition and 

structure that may persist for centuries (Foster et al. 1998) 

and are qualitatively different from smaller disturbances 

(Romme et al. 1998). More details on abiotic agents of 

disturbance can be found in chapter 3.

Windstorms arising from extratropical cyclones off the 

Pacific Ocean have the potential to produce hurricane-force 

winds and extensive damage to forested ecosystems, and 

large storms affected parts of the NWFP area several times 

in recorded history (Mass and Dotson 2010). These events 

are generally characterized by southwesterly winds and 

occur during the winter when soils are saturated. Coastal 

areas, particularly the Coast Range in Oregon and Washington, as well as the Olympic Peninsula, were subject to 

multiple synoptic winds events during the 20th century. 

Some of these storms also affected inland areas and caused 

substantial tree mortality in portions of the western Cascades, particularly near the Columbia River Gorge (Sinton 

and Jones 2002). The most intense of these events, the 

Columbus Day Storm of 1962 (Lynott and Cramer 1966), 

killed approximately 11 million board feet of timber in 

Oregon and Washington (Teensma et al. 1991). High-wind 

events are positively associated with neutral to warm PDO 

conditions, and their influence has shifted northward over 

the past 120 years (Knapp and Hadley 2012), but we are 

currently unaware of any published literature including 

future projections of the frequency or intensity of windstorms in the region.

Fire played an important role in the historical dynamics 

of the region (Agee 1993), but a long period of fire exclusion 

reduced fire activity during the mid-20th century (Littell et 

al. 2009). However, increases in the frequency and extent of 

fire across the Western United States since the mid-1980s 

have been attributed to longer fire seasons associated with 

earlier snowmelt and warmer spring and summer temperatures (Jolly et al. 2015, Westerling et al. 2006) as well as 

drought (Gedalof et al. 2005, Littell et al. 2009). A recent 

study also linked increasing fire activity to human-driven 

climate change, which is contributing to a more conducive 

fire environment by increasing fuel aridity (Abatzoglou and 

Williams 2016). Annual area burned has increased since the 

mid 1980s (Miller et al. 2012, Reilly et al. 2017). However, 

recent fire activity differs substantially depending on spatial 

scale and geographic location across the region (Davis et 

al. 2015, Reilly et al. 2017), and there is growing consensus 

that the region experienced less fire than would be expected 

under historical conditions (Marlon et al. 2012, Miller et al. 

2012, Parks et al. 2015, Reilly et al. 2017).

The effects of recent fires have been extremely variable 

across the region, with most recent fire activity occurring 

in the Klamath Mountains, eastern Cascades, and western 

Cascades of Oregon (fig. 2-7). The annual area burned 

increased in most vegetation zones since the mid-1980s, but 

dry vegetation zones, including ponderosa pine, Douglas-fir, 

and grand fir/white fir, experienced less fire than they would 

have during presettlement times because of fire suppression 

(Miller et al. 2012, Reilly et al. 2017) (see chapter 3 for 

more discussion). Mean and maximum fire size from 1910 

to 2008 increased in northwest California (Miller et al. 

2012). Cold and moist vegetation zones (silver fir, mountain 

hemlock, and subalpine zones, but with the exception of 

western hemlock) experienced the greatest proportions of 

high-severity in recent fires, and most of the area burned 

in the previously mentioned dry vegetation zones has been 

at low and moderate severity (Miller et al. 2012, Reilly and 

Spies 2016, Reilly et al. 2017, Whittier and Gray 2016). Fire 

severity has been related to climate and drought at broad 

spatial scales since the mid 1980s (Abatzoglou et al. 2017, 

Keyser and Westerling 2017, Reilly et al. 2017). Although 

the area burned has increased in all major vegetation zones 

during this time, there is little evidence that the proportion 

burning at high severity has increased across the region 

(Law and Waring 2015, Miller et al. 2012, Reilly et al. 
51

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Fire severity

Low

Moderate

High

Northwest Forest Plan area

600,000

400,000

200,000

0

Acres

19851986198719881989199019911992199319941995199619971998199920002001200220032004200520062007200820092010

Year

Eastern Cascades

120,000

80,000

40,000

0

Acres

198519861987198819891990199119921993199419951996199719981999200020012002200320042005200620072008

20092010

Year

600,000

400,000

200,000

0

Acres

19851986198719881989199019911992199319941995199619971998199920002001200220032004200520062007200820092010

Year

Klamath Mountains

Western Cascades

120,000

80,000

40,000

0

Acres

1986198719881989199019911992199319941995199619971999200020012002200320042005200620072008

20092010

Year 1998

Figure 2-7—Geographic patterns of burn severity from 1985 to 2010 in the Northwest Forest Plan area. Burn severity is derived from the 

relativized version of the difference in the normalized burn ratio and is based on the percentage of basal area mortality as follows: low 

(<25 percent), moderate (25 to 75 percent), and high (>75 percent) (Reilly et al. 2017). Map boundaries correspond with the physiographic 

provinces in figure 2-1. 
52

GENERAL TECHNICAL REPORT PNW-GTR-966

2017). Although they found no increase in the proportion of 

high-severity fire, Reilly et al. (2017) found that increases 

in high-severity patch size during this time were associated 

with more area burned during drought years in all major 

vegetation zones.

Despite concern that insect outbreaks may exacerbate 

fire effects by altering fuel structure (Hicke et al. 2012), 

there is a growing body of literature within the region and 

across the Western United States indicating that the two 

disturbances are not positively linked (Hart et al. 2015, 

Meigs et al. 2015), and that prefire insect activity does not 

make fires more severe (Agne et al. 2016, Meigs et al. 2016, 

Reilly and Spies 2016). These findings are also consistent 

with several other studies in other regions of the Western 

United States (Black et al. 2013, Bond et al. 2009, Donato et 

al. 2013, Harvey et al. 2013, Simard et al. 2011). 

Hessl (2011) outlined a framework proposing three 

major pathways through which future fire activity may 

respond to climate change. Most studies to date have 

assumed that the major pathway to change will be based 

on alteration of fuel conditions as the relationships among 

weather, fuel moisture, and fire activity are well established. 

Fewer studies have focused on changes in the second 

pathway, alteration of fuel amount, though this may be 

of particular concern given its relation with severity. The 

least is known about the third pathway, changes in sources 

of ignition. This pathway will be subject to changes in 

lightning frequency as well as changes in human ignitions 

and fire-suppression efforts. 

A number of studies using different techniques project 

increases in a variety of metrics of fire activity (i.e., area 

burned, fire size, fire severity, fire interval) during the 21st

century, although projections differ considerably across the 

NWFP area (table 2-2). Most studies report coarse-scale 

projections (i.e., individual states), and few include details at 

geographic variablity within study areas (i.e., east vs. west). 

Stavros et al. (2014) found that the probability of very large 

fires will increase based on climate projections for Oregon 

and Washington, but increases will be minor in northern 

California. McKenzie et al. (2004) used statistical models 

and found that an increase in temperature of 3.6 °F (2 °C) 

will increase fire extent by 1.4 to 5 times for many Western 

states, including Oregon, Washington, and California. 

Using a similar statistical approach, Littell et al. (2010) 

found that area burned is likely to increase by 2 to 3 times 

across Washington by the end of the 2040s. They also found 

that area burned in the western Cascades of Washington 

is expected to increase by more than eight times, but on 

average will still affect only a small extent (9,100 ac) of 

the ecoregion by the 2080s. Liu et al. (2013) projected 

increases in fire potential associated with warming and 

drought from 2014 to 2070. Turner et al. (2015) projected 

an increase in area burned by 3 to 9 times in a portion of 

the central western Cascades of Oregon. Krawchuk et al. 

(2009) also predicted increases in fire probability in the 

western Cascades. Barr et al. (2010) projected an increase 

in annual fire extent of 11 to 22 percent in the Klamath 

River basin by 2100. Davis et al. (2017) projected increases 

in fire suitability across multiple provinces in Oregon and 

Washington during the 21st century (under RCP 4.5 and 8.5, 

respectively), including the Klamath Mountains (18 to 48–58 

percent), the western Cascades (1 to 13–18 percent), and the 

eastern Cascades (11 to 40–45 percent). Although projections differ geographically, all studies predict increased fire 

activity during the 21st century.

There are few statistical predictions for moist maritime 

forests (i.e., Sitka spruce, redwood, western hemlock) 

because there has been very little area burned near the coast 

in the past several decades (Littell et al. 2010). Davis et 

al. (2017) found no increase is fire suitability in the Puget 

Trough and only minor increases (<1 to 2 percent) in the 

Coast Range. Creutzburg et al. (2017) projected very little 

increase in area burned by 2100 compared to the period 

from 1959 to 2009 in the Oregon Coast Range. Fried et al. 

(2004) suggested a decrease of 8 percent in area burned by 

fires along the north coast of California over the 21st century 

under continued fire-suppression efforts. Liu et al. (2013), 

however, predicted an increase in fire potential (measured 

as Keetch-Byram Drought Index) from 2.5 to 5 times 

owing to changes in fire weather in coastal forests by 2070. 

Westerling et al. (2011) projected 300 percent increases in 

area burned in northwest California. Krawchuk et al. (2009) 

projected little change in fire potential in coastal forests, but 

increased potential across the rest of the region. Rogers et 
53

Table 2-2—Projections for future fire activity in the Northwest Forest Plan area from published studiesStudy Method Geographic extent Scenario Time period

Projected change 

from current

Suppression 

effects Variable

Percent

Stavros et al. 2014 Statistical Oregon, Washington, 

northern California

RCP 4.5, RCP 8.5 2031–2060 + No Very large fire 

occurrencea

McKenzie et al. 2004 Statistical Oregon, Washington, 

northern California

A2, B2 2070–2100 + No Area burned

Littell et al. 2010 Statistical Washington A1B 2020–2080 +200 to 300 No Area burned

Turner et al. 2015 Process Willamette Valley, 

Oregon

RCP 4.5, RCP 8.5 2100 +300 to 900 No Area burned

Krawchuk et al. 2009 Statistical Global A2, B1 2070–2090 + No Fire probabilityb

Fried et al. 2004 Statistical Northern California 2× CO2 N/Ac -8 Yes Area burned

Spracklen et al. 2009 Statistical Oregon, Washington, 

northern California

A1B 2050 +78 No Area burned

Barr et al. 2010 Process Klamath Basin, Oregon, 

and northern California

A2 2075–2085 +11 to 22 No Area burned

Liu et al. 2013 Statistical Continental United States A2 2041–2070 No No Fire potentiald

Westerling et al. 2011 Statistical Northern California A2 2085 +100 No Area burned

Rogers et al. 2011 Process Oregon, Washington A2 2070–2099 +76 to 310/29–41 Yes Area burned/burn 

severitye

Sheehan et al. 2015 Process Oregon, Washington RCP 4.5, RCP 8.5 2071–2099 -82 to 14 Yes Mean fire interval

Creutzburg et al. 2017 Statistical Oregon RCP 8.5 2100 Negligible Yes Area burned

Parks et al. 2016 Statistical Western United States RCP 8.5 2040–2069 No change to decrease No Fire severityf

Davis et al. 2017 Statistical Oregon, Washington 2071–2100 No change to increase No Suitability for large 

wildfiresg

Note: Most studies project area burned or other variables associated with increased area burned (fire suitability, large fire occurrence), and there are relatively few projections for fire severity.a Very large fires are defined as those >50,000 ac.b Fire probability is the probability of fire occurrence. c This study does not project to an explicit time period in the future, but rather conditions based on 2× current CO2 scenario.d Fire potential is measured by the Keetch-Byram Drought Index y.e Burn severity is based on combustion of biomass.f Burn severity is based on a postfire composite burn index (CBI) based on changes in multiple strata, including soil and rock, litter and surface fuels, low herbs and shrubs, tall shrubs, and trees.g Large wildfires are defined in this study as >40 ha.N/A = not available.
54

GENERAL TECHNICAL REPORT PNW-GTR-966

al. (2011) used a mechanistic vegetation model (MC1) that 

integrates fire and suppression efforts, and found increases 

in area burned in Oregon and Washington from 76 to 310 

percent by 2070 to 2099. Although this increase may seem 

high, it is important to note that the recent extent of fire in 

moist forest is very low, and a tripling of fire may still be a 

relatively small amount in absolute terms. 

Although several studies have projected future increase 

in fire activity, far less work has been done on future fire 

severity. This component of fire regimes is less well studied 

and understood (Hessl 2011, Parks et al. 2016), potentially 

because of the complexities of incorporating feedbacks 

from fire and climate on fuel structure and arrangement at 

stand and landscape scales. Previous fires have the potential 

to inhibit the spread of subsequent fires occurring within 

a limited time window (Parks et al. 2014), and increased 

area burned in the future may provide a feedback related to 

decreased fuel availability. Rogers et al. (2011) used a process model (MC1) and suggested increases in burn severity 

of 29 to 41 percent that related to increases in productivity 

and biomass during non-summer months. However, a recent 

study incorporating changes in vegetation type, fuel load, 

and fire frequency predicted either no change or potential 

reductions in fire severity across the entire NWFP area for 

2040–2069 under the most extreme climate change scenario (RCP 8.5) (Parks et al. 2016). The authors attributed 

decreases in fire severity to greater water deficits, decreased 

productivity, and less available fuel. 

The wide range of projections of climate change effects 

on fire within the NWFP area are likely the result of several 

factors. These factors include differences in emissions 

scenarios, spatial and temporal scale, model structure 

(e.g., statistical vs. process), and variability in how models 

project precipitation. In addition, McKenzie and Littell 

(2017) showed that differences in climate-fire relationships 

among physiographic provinces are likely to be substantial, and further analysis is required to put differences in 

methodological and regional future projections of fire into 

context. At coarser regional scales, dynamical and statistical 

approaches to projecting future fire activity may agree, 

but the mechanisms operating at more local scales require 

careful interpretation. 

Cumulative effects of climate change on tree species 

distributions and range shifts—

The cumulative effects of changes in mortality, growth, 

and recruitment will ultimately be manifest in shifts in 

species distributions and ranges. These effects will also 

depend on the size and degree of connectivity within 

populations. Range expansion occurs through migration 

and colonization at the outer limits, or “leading edge,” of 

a species’ distribution where climate is becoming more 

favorable. Range expansion at the leading edge is controlled 

by fecundity and dispersal (Thuiller et al. 2008). More 

vagile species that produce greater amounts of seeds and 

have a greater ability to disperse will have more potential 

to track climate change than those with poor dispersal 

ability. At the lower limits or “trailing edge” of a species’ 

distribution where climate is becoming less favorable, range 

contraction and progressive isolation will occur through 

local extirpation. Range contraction is related to the ability 

of a species to persist in refugia that experience less change 

than the surrounding landscape. Individuals at the trailing 

edge may thus play an important role in the maintenance 

of genetic diversity for some species (Hampe and Petit 

2005). Although local extirpation may occur throughout the 

range of species, small, isolated populations at the trailing 

edge may be particularly vulnerable as the climate changes 

rapidly (Davis and Shaw 2001). 

It is likely that species that are more adapted to cold 

environments will be more sensitive to warming at their 

lower limits of elevation or latitude, while expansion of 

species adapted to warmer conditions is expected at upper 

range limits at high elevation or latitude (HilleRisLambers 

et al. 2015). Range limits may also be altered at the eastern 

limits of the range of some species as a result of increasing 

aridity. Warmer temperatures are likely to lead to range 

expansion at the leading edge for some species at the upper 

tree line, but not necessarily for species in closed-canopy 

forests at lower elevations (Ettinger and HilleRisLambers 

2013, Ettinger et al. 2011). However, expansion at upper 

range limits may be limited by dispersal and low abundance 

of adult trees that produce seed (Kroiss and HilleRisLambers 2015). Warmer temperatures may increase germination 

and survival of seedlings provided adequate water, as well 
55

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

as increase sapling growth rates (Ettinger and HilleRisLambers 2013, Ettinger et al. 2011, HillesRisLambers et al. 

2015), but many tree species are long lived and may exhibit 

lagged responses to climate change in terms of range shifts 

(Kroiss and HillesRisLambers 2015). 

A common approach to detecting range shifts is comparing current distributions of mature trees and seedlings. 

Juveniles (and seedlings specifically) with limited root systems and smaller reserves of carbon are more vulnerable to 

mortality from drought and temperature extremes (Jackson 

et al. 2009). Monleon and Lintz (2015) provided evidence of 

range shifts for common tree species in California, Oregon, 

and Washington where the range of seedlings extended 

to temperatures 0.22 °F (0.12 °C) colder than that of adult 

trees, and seedlings were found at higher mean elevations 

and latitudes than mature trees for most species during 

the period from 2001 to 2010. Results also suggested that 

overall distributions of individual species remained relatively stable, but most species were more abundant toward 

the colder edge of their range and distributions changed 

the least at the warm end of their range. Some of the more 

common tree species with seedlings found at significantly 

colder temperatures included western redcedar, silver fir, 

western hemlock, grand fir, and mountain hemlock. 

Thus far, individual tree species have shown differential responses to recent warming, and it is likely that tree 

species will respond differently to projected future changes 

in climate. Lintz et al. (2016) examined recent changes in 

basal area and density of 22 tree species on unburned Forest 

Service lands in Oregon and Washington from the mid 

1990s to mid 2000s. Several species had stable populations 

in terms of density and basal area, including noble fir (Abies 

procera), western redcedar, western hemlock, ponderosa 

pine, and Douglas-fir. These findings are consistent with 

HilleRisLambers et al. (2015), who suggested that compositional change in the near term will be slow in higher elevation forests of the silver fir vegetation zone. The greatest 

levels of mortality in Lintz et al. (2016) occurred in western 

white pine, whitebark pine, Pacific madrone, subalpine fir, 

lodgepole pine, grand fir, Engelmann spruce, and western 

yew (Taxus brevifolia). Although this study suggested only 

slight mortality-related declines of Alaska yellow-cedar 

(Callitropsis nootkatensis), this species has experienced 

recent mortality across large areas in southeast Alaska associated with a warming climate (Krapek and Buma 2015).

Recent work from the Klamath Mountains and eastern 

Cascades in northern California suggests that multiple 

species, including red fir, Jeffrey pine, lodgepole pine, 

and white fir, experienced recent increases in mortality 

(Mortenson et al. 2015). Results from this study indicated 

that mortality rates for all species were generally higher 

in smaller size classes. Despite increases in the number of 

recently dead red fir associated with dwarf mistletoe and 

drought, the population structure of this species was stable. 

Vegetation Models and Potential Future 

Vulnerability

Several climate change vulnerabilities have been identified 

either explicitly in the literature, or may be inferred based 

on knowledge of long-term vegetation change in the region, 

distribution and dynamics of current vegetation, and projected 

changes across the region. Increases in temperature, as well as 

altered precipitation and disturbance regimes, are expected to 

alter vegetation across the region (see “Summary of Vulnerabilities to Climate Change” on next page). Several types of 

simulation models are commonly used to predict vegetation 

responses to potential future climate scenarios, each with 

their own unique set of assumptions, strengths, and weaknesses (see Peterson et al. 2014a for a more indepth review). 

Models simplify the complexity of ecological processes by 

making assumptions that are ideally based on empirical 

measurements. However, because empirical data are often 

only available for a few species at a few geographic locations, 

models are most often based on applications of theory on 

how species interact and respond to environmental gradients. 

As a result, the best use of models may be for understanding 

variability in the magnitude of effects as opposed to predicting specific outcomes (Jackson et al. 2009, Littell et al. 

2011). Some of the most common models used to project the 

effects of climate change can be generally characterized as 

species distribution models (SDM), dynamic global vegetation models (DGVM), and landscape models. These models 

have their own unique assumptions and relative strengths 

and weaknesses, which should be carefully considered when 

interpreting results.
56

GENERAL TECHNICAL REPORT PNW-GTR-966

Species distribution models are statistical models based 

on empirical observations of the relationship between a 

species occurrence and the observed range of environmental or bioclimatic conditions. SDMs are commonly used due 

to their simplicity, but generally do not represent ecological 

processes (e.g., biotic interactions, dispersal, adaptation) 

that constrain species distributions (Ibáñez et al. 2006), 

and are problematic when extrapolating to future climates 

that have no modern analogs (Bell and Schlaepfer 2016). 

Despite these limitations, SDMs provide a basic understanding of how suitable bioclimatic conditions constrain 

the current distribution of a species, as well as how this 

distribution might change under any number of different 

climate change scenarios.

DGVMs are a type of process model that predict ecosystem processes along with the distribution of specific biomes or 

plant function groups. These models (e.g., MC1) incorporate 

biogeography and ecophysiology of vegetation types (e.g., 

coniferous forests, grasslands, woodlands) as well as climate 

and disturbance to project broad-scale vegetation changes. 

Biogeochemistry models are also process models, but focus 

more specifically on carbon, water, and nutrient cycles and 

are often used to investigate the effects of climate change on 

productivity and carbon storage. Both types of models are 

capable of incorporating some of the important ecological processes affecting vegetation response to climate change (e.g., 

disturbance, CO2, site water balance), but have generally been 

applied at broad regional scales with coarse spatial resolution.

Summary of Vulnerabilities to Climate Change

General vulnerabilities to climate change include 

increased wildfire and insect activity driven by drought 

and extreme weather events, ongoing and new invasions 

of nonnative species, and loss of some high-elevation 

species. Fragmented populations at range margins (e.g., 

Alaska yellow-cedar), as well as narrowly distributed 

species and species with poor dispersal, are vulnerable 

to declines from losses of climate-suitable habitat, 

especially in areas that lack topographic conditions that 

foster the potential for long-term persistence in relatively 

climate-stable refugia. 

The greatest vulnerability to climate change 

exists in the drier and colder portions of the region in 

the eastern Cascades, southern portion of the western 

Cascades of Oregon, coastal and inland areas of the 

Klamath Mountains, and the California Coast Range. In 

dry vegetation zones of these regions, increases in area 

burned during drought conditions may result in larger 

patches of high-severity fire and drive landscape-scale 

change. In general, there is good model agreement that 

subalpine forests are likely to be reduced everywhere 

except in the northern portion of the eastern Cascades. 

Several tree species in both wet and dry vegetation zones 

are vulnerable to nonnative pathogens whose effects 

may be exacerbated by climate change. These include 

whitebark pine, subalpine fir, sugar pine, western white 

pine, Port Orford cedar, tanoak, and multiple species of 

oak. Old-growth forests may also be vulnerable to periods 

of elevated mortality rates associated with insects and 

pathogens during drought. Along the coast, decreases in 

summer fog may substantially reduce suitable climate for 

redwood and other coastal species that depend on it to 

mitigate summer drought. 

Much of the coastal and inland area toward the 

central and northern part of the region show either less 

potential increase or decreases in water-balance deficit 

during the summer months. However, high-elevation 

areas may see reduced snowpacks with more precipitation 

falling as rain. Warmer, wetter conditions may also promote native and nonnative pathogen activity, especially 

Swiss needle cast on Douglas-fir near the coast. Some of 

these areas may be vulnerable to a continued northward 

shift of high-wind events, particularly near the coast in 

Washington. Although they have been rare in the past 

century, these areas have historically experienced large 

fires driven by synoptic warm, dry wind events from the 

east during drought conditions projected for the future. 
57

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Landscape models (e.g., LANDIS-II) (Scheller et al. 

2007) generally focus explicitly on simulating processes 

(e.g., dispersal, growth, mortality) and can represent 

interactions among vegetation, disturbance, climate 

change, and management scenarios at a variety of different 

spatial and temporal scales. Landscapes are represented as 

gridded cells in which individual cohorts of trees compete 

for resources, grow, and die. Although some ecological 

processes are represented in landscape models, many 

processes that will be sensitive to climate change (e.g., 

CO2 fertilization, phenology, biotic disturbances) are 

not incorporated in these or other models for projecting 

vegetation change.

Model projections—

DGVMs generally project persistence of cool, maritime 

forests in the western hemlock and Sitka spruce vegetation 

zones of the Coast Range in western Oregon and Washington (Creutzburg et al. 2017, Rogers et al. 2011, Shafer 

et al. 2015, Turner et al. 2015). SDMs project persistence 

of western redcedar, Sitka spruce, and western hemlock 

across 55 to 82 percent of their current distributions by 

2080 (DellaSala et al. 2015). However, most species in lower 

elevation, moist vegetation zones are predicted to have less 

suitable climatic conditions than currently by the mid-21st

century (Saxon et al. 2005). One DGVM-based study 

projected losses of conifer forest across much of the Coast 

Range in Oregon with increases in cool mixed forests under 

the RCP 4.5 scenario, and increases in warm mixed forests 

under the RCP 8.5 scenario (Sheehan et al. 2015). Although 

western redcedar is thought to be moderately vulnerable 

to climate change, bigleaf maple is considered to be one of 

the least vulnerable species in the region (Case et al. 2016). 

Consistent with a potential decrease in summer fog (Johnstone and Dawson 2010), DellaSala et al. (2015) projected 

a decrease in suitable climate for redwood of almost 25 

percent by 2080.

SDMs project some of the greatest changes for the 

southern and southwestern part of the NWFP area, with 

less change in the north and in the western Cascades 

(Crookston et al. 2010; DellaSala et al. 2015; Hargrove and 

Hoffman 2004; McKenney et al. 2007, 2011; Rehfeldt et al. 

2006). Using a DGVM, Turner et al. (2015) projected the 

dominant vegetation type in a portion of the central western 

Cascades of Oregon to remain forest by 2100, but that the 

forest would transition from evergreen needleleaf forest to 

a mixture of broadleaf and needleleaf growth forms. An 

SDM-based study by Latta et al. (2010) suggests annual 

growth increases of 2 to 7 percent in moist vegetation zones 

west of the Cascade Mountains depending on scenario. 

However, projections from mechanistic models differ, with 

some projecting moderate to extreme decreases owing to 

increases in fire activity (Rogers et al. 2011), and others 

projecting slight to small decreases in growth (Coops and 

Waring 2011b). Shafer et al. (2015) suggested that growth 

will decrease in the southwestern part of the region based 

on projections from a DGVM.

All types of models project that high-elevation forests 

will experience the greatest change within the region, with 

moderate to total reductions in suitable climate by the end 

of the 21st century (Crookston et al. 2010; Halofsky et al. 

2013; Hargrove and Hoffman 2004; Mathys et al. 2016; 

McKenney et al. 2007, 2011; Rehfeldt et al. 2006; Shafer et 

al. 2015). Case et al. (2016) suggested that western white 

pine and whitebark pine have relatively high vulnerability 

to climate change, while noble fir and silver are moderately 

vulnerable. Mechanistic models project that suitable climate 

for subalpine fir will be available only in the northern Cascade Range (Coops and Waring 2011b, Rogers et al. 2011), 

although climate suitability may increase for mountain 

hemlock in Oregon (Coops and Waring 2011a). Two additional studies also using mechanistic models also predicted 

large decreases in the distribution of lodgepole pine by the 

2100s (Coops and Waring 2011a, Mathys et al. 2016). SDMs 

project reduction of 15 to 39 percent by 2080 for several 

species occurring in high-elevation wet vegetation zones, 

including silver fir, grand fir, Alaska yellow-cedar, and 

mountain hemlock (DellaSala et al. 2015). In general, there 

is more model agreement for subalpine forests than for other 

vegetation zones, and most suggest that suitable climate 

is likely to be reduced everywhere except in the northern 

portion of the eastern Cascades.
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GENERAL TECHNICAL REPORT PNW-GTR-966

Model projections for vegetation change in dry coniferous forests in the southern and eastern parts of the region 

show little agreement. Species distribution models suggest 

decreases in suitable climate for ponderosa pine, while some 

DGVMs project increases or only slight changes in temperate 

coniferous forests (Coops et al. 2005, Halofsky et al. 2013, 

Rogers et al. 2011, Sheehan et al. 2015) and others projected 

decreases (Coops and Waring 2011a). Halofsky et al. (2014) 

projected that while the area of dry mixed-conifer forest 

is expected to increase from 21 to 26 percent by 2100, the 

area of moist mixed-conifer forest is expected to decrease 

36 to 60 percent in the grand fir/white fir vegetation zone of 

the central eastern Cascades. Shafer et al. (2015) projected 

expansion of woodland vegetation during the 21st century. 

Case et al. (2016) suggested that grand fir will only be 

moderately sensitive to climate change. Given the lack of 

agreement among model projections for vegetation change in 

dry coniferous forests, these results should be used cautiously 

in planning and management (Peterson et al. 2014a). 

In northern California, the projected changes in most 

scenarios include losses of evergreen conifer forests and 

increases in mixed evergreen forest primarily because of 

increased fire activity (Lenihan et al. 2008). A mechanistic 

model projects that Douglas-fir will be stressed across 

almost all of northern California (Mathys et al. 2016). 

Increases are projected in the hardwood component, 

shrublands, and grasslands, particularly throughout 

the eastern and drier areas, while maritime evergreen 

needleleaf forests are expected to contract (DellaSala et al. 

2015). Barr et al. (2010) projected that the upper Klamath 

River basin will support primarily grassland in place of 

sagebrush and juniper by 2100. In the lower Klamath River 

basin (California), conditions suitable for hardwood forests 

(oaks, tanoak, madrone, etc.) are projected to expand, while 

those suitable for conifer-dominated forests are projected 

to contract. Results from Kueppers et al. (2005) primarily 

suggest range expansion and persistence of currently existing populations of valley oak (Quercus lobata). Expansion 

and persistence of blue oak (Q. douglasii) is projected in 

the northern part of its range, but projections primarily 

suggest range contraction toward the southern portion of 

northern California.

Other Vulnerabilities

Invasions of nonnative plant species have the potential to 

alter vegetation dynamics, soil properties (Caldwell 2006, 

Slesak et al. 2016), and disturbance regimes (Brooks et al. 

2004) (see also chapter 3). Most nonnative plant species 

were initially introduced for horticultural uses and erosion 

control, or as contaminated crop seed (Reichard and 

White 2001). Gray (2008) used a systematic inventory of 

forest health monitoring plots and found that more than 

50 percent of plots in almost all physiographic provinces 

in the NWFP area had nonnative species present. Most 

common nonnative plants are associated with management 

(e.g., clearcuts, thinning), though there is potential for 

the spread of some nonnative, shade-tolerant shrubs in 

undisturbed forests (Gray 2005). There is also evidence 

from the region that roads facilitate the spread of nonnative 

plants (Parendes and Jones 2000, Rubenstein and Dechaine 

2015). Little information is available on temporal trends in 

the abundance of nonnative plants, but increasing temperatures may favor exotic species, especially grasses in 

California (Sandel and Dangremond 2012). Warm, dry sites 

with increased topographic exposure may be particularly 

vulnerable to exotic species, especially annual grasses, 

following high-severity wildfire (Dodson and Root 2015). 

Gray et al. (2011) provided a field guide and prioritized 

list of nonnative plants along with range maps that cover 

the entire Plan area. More information on management 

of nonnative species is also available in Harrington and 

Reichard (2007).

Many species that depend on climate-sensitive habitats will also likely be sensitive to climate change (Case 

et al. 2015). Narrowly distributed species (e.g., rare and 

threatened, endemics) that specialize in uncommon or 

sparsely distributed habitats (e.g., serpentine soils, montane meadows) are expected to have difficulty responding 

to changing climatic conditions. Increases in Alaska 

yellow-cedar mortality in southeast Alaska associated 

with warmer climatic conditions and projections of future 

decreases in habitat suitability (DellaSala et al. 2015) 

suggest that this species may be particularly vulnerable 

to loss. Damschen et al. (2010) found decreases in the 
59

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

richness and cover of endemics on serpentine soils in 

southwest Oregon from the 1950s to early 2000s that 

were consistent with a warming climate. Harrison et al. 

(2010) found changes in forest herb communities in the 

Klamath Mountains of Oregon that were also consistent 

with expectations of a drier climate during the second 

half of the 20th century, including lower cover of species with northern affinities and greater compositional 

similarity to communities on southerly aspects. Loarie et 

al. (2008) projected decreases in the richness of endemic 

plant species by 2100 for those that cannot disperse, 

but potential increases if plants can disperse to suitable 

areas. If dryer growing season conditions accompany 

projected warming trends, cool, mesic topographic 

refugia are likely to become increasingly important for 

species persistence (Dobrowski 2011, Olson et al. 2012, 

van Mantgem and Sarr 2015). Montane wetlands may be 

especially at risk from reductions in water levels, shorter 

hydroperiods, and increased probability of drying out 

(Lee et al. 2015). 

Adaptation to and Mitigation of Climate Change

Adaptation and mitigation are essential to strategic planning for the effects of climate change (Millar et al. 2007). 

Adaptation options include management actions at stand 

and landscape scales to reduce vulnerabilities to climate 

change. Mitigation includes efforts to increase carbon 

sequestration in forest ecosystems and provide new energy-efficient products and technologies for society. Halofsky 

and Peterson (2016) provided a summary of an extensive list 

of vulnerabilities and corresponding strategies and tactics 

that were identified and developed through a series of 

science-management partnerships across the northwestern 

United States (http://adaptationpartners.org/library.php). 

Strategies for adaptation and mitigation have been identified 

for forests in the Pacific Northwest, including drier forests 

of southwest Oregon (Halofsky and Peterson 2016; Halofsky 

et al. 2016, 2017). Here, we highlight general management 

actions that could promote adaptation to climate change. We 

summarize these options in table 2-3. For a broader discussion of conservation options (including reserves) in a period 

of climate and other landscape changes and their specific 

relevance to NWFP goals, see chapter 12. 

Table 2-3—Summary of adaptation options for climate change vulnerabilities in the Northwest Forest Plan area

Vulnerability Strategy Tactics

Increased drought stress Increase resilience Thinning

Favor drought-resistant species/genotypes

Foster genetic and phenotypic 

diversity

Protect trees adapted to water stress

Collect seed for future

Maintain connectivity for natural species migration

Increasing area affected by 

fire, insects, and pathogens

Increase stand resilience Thinning and prescribed fire

Increase stand heterogeneity

Favor fire-tolerant species

Increase landscape resilience Increase landscape heterogeneity

Increase diversity of patch sizes 

Use topography to guide treatments

Loss of forest cover Monitoring of change Use existing data and add more where needed

Planting/assisted migration

Maintain connectivity for natural species migration

Exotic species Increase control efforts Early detection/rapid response/frequent inventory

Interagency coordination

Source: Halofsky and Peterson 2016.
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GENERAL TECHNICAL REPORT PNW-GTR-966

Adaptation—

Several adaptation options to reduce climate change vulnerability are available (table 2-3). These range from manipulation of stand and landscape structure to foster resistance and 

resilience to future disturbance, to protection of intact areas 

and climate change refugia that provide connectivity,and 

facilitate species migration to more favorable habitats. In the 

case of disturbance, managers may choose to take actions 

prior to and in anticipation of disturbance to reduce vulnerability, or after a disturbance to affect the ongoing process of 

recovery (Dale et al. 1998). 

Manipulation of stand and landscape structure with 

management tools (i.e., thinning, prescribed fire) is thought 

to increase resistance and resilience to future vulnerabilities 

associated with drought and disturbance (e.g., fire, insects) 

in drier forests that may be subject to moisture stress and 

fire (Hessburg et al. 2015, Spies et al. 2010). Findings from 

dry forests in other regions support the use of thinning as 

an option to increase soil water availability, reduce growing-season moisture stress, and improve vigor in older trees 

(Bradford and Bell 2017; McDowell et al. 2003, 2006), but 

the NWFP area is lacking specific studies on this topic. 

Prescribed fire has also been found to increase resistance 

to drought in dry forests of the Sierra Nevada of California 

(van Mantgem et al. 2016). Thinning has effectively been 

used and reduced fire severity in dry Douglas-fir of Washington’s eastern Cascades (Prichard et al. 2010), and other 

regions in the Western United States (Wimberly et al. 2009). 

Fuel treatment may be effective at reducing fire behavior and 

burn severity during moderate burning conditions; however, 

treatments may not be effective during large, weather-driven 

fires (Lydersen et al. 2014, Reinhardt et al. 2008). 

A general principle for thinning to reduce fire severity at the stand scale includes maintaining older trees of 

fire-tolerant species, reducing understory density, and 

increasing height to live crowns (Agee and Skinner 2005). 

Given that these actions will likely increase surface fuels, 

thinning followed by prescribed fire may help reduce 

surface fuels. Landscape-scale treatments that restore structural heterogeneity in places where historical fire regimes 

have been interrupted are proposed as a way to reduce 

vulnerability to high-severity fire and extensive pathogen 

and insect outbreaks in the future (Hessburg et al. 2015). 

Topography can provide a physical template to consider 

when designing and implementing landscape-scale treatments (e.g., thinning on dry ridges). Increasing landscape 

heterogeneity is thought to impede the spread of contagious 

disturbances (e.g., fire, insects), but empirical evidence 

supporting this is currently lacking. 

There is relatively little research on the use of thinning 

in moist forests as a climate change adaptation strategy. 

These forests were relatively dense historically. Thinning, 

specifically variable-density thinning, can help the growth 

and survival of the residual trees, as well as improve the 

adaptive capacity and ecological diversity of stands (Neill 

and Puettmann 2013) (see chapter 3). In drier parts of moist 

vegetation zones, where fire was more frequent, thinning 

and prescribed fire could be used to mimic low- and moderate-severity fire and promote landscape diversity, which 

in turn could promote landscape-scale resilience to climate 

change (chapter 3). The use of thinning in moist forests is 

generally focused on plantations and younger forests and 

would have to be balanced against landscape-level goals for 

maintaining high canopy cover in older forests, which can 

buffer climatic changes as described above (Frey et al. 2016). 

Assisted migration of genotypes and species that are 

adapted to future climate scenarios may improve resilience 

of species that are not be able to migrate, but this option 

is controversial and poorly understood (Marris 2009). 

Coastal Douglas-fir populations in particular are considered 

genetically “maladapted” to future climates in Oregon and 

Washington (St. Clair and Howe 2007). Bansal et al. (2015) 

found that populations of Douglas-fir from cooler climates 

had greater resistance to drought than those from warmer climates, contrary to expectations. Populations from areas with 

relatively cool winters and dry summers were more tolerant 

to drought and cold and may be the best adapted to warmer 

future climate conditions (Bansal et al. 2016). There is little 

information available from other species from the NWFP 

area, though a study from Arizona found that ponderosa 

pine seedlings that originated from low-elevation, drier sites 

survived the longest during drought (Kolb et al. 2016b). 

An alternative to assisted migration involves increasing 

connectivity by establishing large blocks of forest managed 
61

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

change was improved when refugial areas were incorporated 

into the reserve design of the NWFP. 

Protection of climate change refugia based on physiography, soils, and vegetation are a key part of climate change 

adaptation strategies (fig. 2-8), but identification of refugia has 

proven difficult (Keppel et al. 2012, Morelli et al. 2016). Most 

studies of refugia have been ad hoc or descriptive and primarily conceptual, and multiple lines of evidence using different 

approaches from across disciplines (e.g., SDMs, downscaled 

climate models, genetics) may be necessary to further understanding of refugia (Keppel et al. 2012). Refugia will most 

likely be found in topographically complex landscapes where 

microclimates vary from differences in aspect, shading and 

insolation, and cold-air drainages (Dobrowski 2011). McRae 

et al. (2016) mapped potential landscape resilience based on 

topoclimate diversity and regional connectivity for the Pacific 

Northwest and northern California. Many of the areas of 

highest resilience occurred in mountainous areas of federal 

Figure 2-8—Examples of the physiographic and vegetation-based refugia that may experience reduced rates of climate change. Source: 

Morelli et al. (2016).

for biodiversity and resilience to climate change. Where 

forests are more fragmented by land use and past management, corridors can facilitate the flow of organisms through 

the matrix of unsuitable habitat (Krosby et al. 2010, Nuñez et 

al. 2013). Linking contemporary climates with future climate 

analogs is one approach to promote connectivity in the future 

and facilitate movement of species in the future (Littlefield 

et al. 2017). Vos et al. (2008) suggested the following to mitigate projected climate changes: (1) linking isolated habitats 

to nearby climate-proof reserves, (2) increasing colonization 

capacity of reserve networks that are projected to remain 

suitable in the future, and (3) optimizing reserve networks in 

which climate remains relatively stable (e.g., refugia). In the 

only biodiversity-climate resiliency study of the NWFP area, 

Carroll et al. (2010) found that reserves based on spotted owl 

conservation criteria overlapped areas of high localized-species richness, but poorly captured core areas of localized 

species’ distributions. They found that resilience to climate 
62

GENERAL TECHNICAL REPORT PNW-GTR-966

Total carbon (tons/acre)

0 to 49

50 to 99

100 to 149

150 to 199

200 to 249

250 to 299

Figure 2-9—Total forest carbon density in the Northwest 300 to 349

Forest Plan area (2000–2009). Carbon estimates are 

from Wilson et al. (2013). Map boundaries correspond 

with the physiographic provinces in figure 2-1.

lands (e.g., Olympic Peninsula and the Klamath Mountains 

region). Morelli et al. (2016) presented a synthesis and review 

of literature pertaining to climate change refugia for climate 

adaptation. They provided a framework for identifying, 

mapping, and conserving climate change refugia to meet 

management objectives. This involves consideration of valued 

resources and vulnerabilities, identification of climate change 

refugia, and prioritization of refugial areas. 

Increasing connectivity may be insufficient for those 

species that are unable to migrate as rapidly as the climate 

changes (Dobrowski et al. 2013). Connectivity considerations 

would likely need to be species-specific because each species 

experiences the same landscape in different ways (Betts et 

al. 2014). Refugia should also be large enough to support 

populations they are aimed at conserving (Stewart et al. 

2010). Planning and monitoring are also essential for adaptation and can help identify microclimatic settings that may 

provide suitable refugia in the future, coordinate planning 

across jurisdictions and ownerships, and revise management 

goals and objectives to be consistent with the uncertainty 

that accompanies climate change (Spies et al. 2010). For a 

broader discussion of refugia and connectivity related to the 

reserve network of the NWFP, see chapters 3 and 12. 

Mitigation—

Mitigation includes efforts to increase carbon sequestration 

in forest ecosystems and provide new energy-efficient 

products and technologies for society. Of these, we focus on 

the former, which has been proposed as a means of climate 

mitigation (Depro et al. 2008, Law and Harmon 2011, Ryan 

et al. 2010), and then discuss how management practices have 

the potential to affect carbon sequestration in the NWFP area. 

Forests in the NWFP area have great potential to store 

large amounts of carbon in both live and dead biomass 

(Smithwick et al. 2002). Total carbon storage levels differ 

among physiographic provinces (fig. 2-9) as a result of 
63

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

in losses resulting from combustion in fire compared with 

losses from fuel reduction are unlikely to make fuel reduction a viable mitigation strategy (Ager et al. 2010, Campbell 

et al. 2012, Kline et al. 2016, Mitchell et al. 2009, Restaino 

and Peterson 2013, Spies et al. 2017). As the amount of 

fire on the landscape increases, the difference in carbon 

sequestration between untreated and treated landscapes 

declines and the likelihood that thinning will pay off in 

respect to the overall carbon balance increases (Loudermilk 

et al. 2014). 

Research Needs, Uncertainties, 

Information Gaps, and Limitations

Despite the accumulating scientific information that 

supports increased warming, considerable uncertainty 

surrounding the effects of climate change on precipitation, 

vegetation response, and disturbance remains a significant 

challenge to forest management (Halofsky and Peterson 

2016, Millar et al. 2007). Many of these research needs are 

mentioned throughout this chapter, but we identify several 

specific information gaps here.

1. Future role of climate extremes and weather 

events as disturbances (e.g., heat waves, floods, 

windstorms).

2. Clarification of the effects of future changes in 

CO2, temperature, and water deficit on growth and 

mortality, and how these effects differ geographically across the region within and among species 

and seral stages.

3. Effects of recent tree mortality on composition and 

structural development across seral stages in all 

vegetation zones.

4. Role of drought on future patterns of disturbance 

occurrence and severity (e.g., fire, insects, pathogens) in all vegetation zones.

5. Role of interactions among multiple disturbances 

(e.g., compound and linked disturbances, including 

insects and fire). 

6. Effects of climate change on demographic processes related to migration (e.g., fecundity, dispersal) and how these differ among species in 

different vegetation zones. 

productivity and disturbance (Law et al. 2004). Recent 

findings suggest that forests on Forest Service lands in 

Oregon and Washington currently store about 63 percent 

of their potential maximum carbon (Gray et al. 2016). At 

current rates, harvest and disturbance have little overall 

impact on carbon sequestration on federal lands in Oregon 

and Washington as a whole, but this differs at smaller 

scales among geographic areas (Gray and Whittier 2014). 

This is particularly true in areas in which dry forests have 

experienced substantial landscape change in recent fires. 

In the Oregon Coast Range, projected increases in productivity are associated with projections of increased carbon 

storage (Creutzburg et al. 2017), but gains could be offset 

by losses depending on harvest intensity (Creutzburg et 

al. 2016). Projections suggest future decreases in carbon 

storage from increases in fire activity in the eastern and 

western Cascade Range of Washington (Raymond and 

McKenzie 2012). In forests west of the Cascades where 

fire is less frequent, decreasing harvesting, increasing 

rotation age, and maintaining and increasing the extent 

of late-successional and old-growth forests are strategies 

to increase carbon storage toward theoretical maximum 

limits (Creutzburg et al. 2016, 2017; Hudiburg et al. 2009). 

Maintaining and increasing the area of dense old-growth 

forests with high biomass also has the potential to mitigate 

temperature changes in topographically complex mountainous environments (Frey et al. 2016).

Carbon stores in the more fire-prone drier eastern and 

southwestern part of the region are more unstable and less 

predictable owing to recent increases and future projections 

of increased fire activity (Restaino and Peterson 2013). 

Some studies from other regions in the Western United 

States (i.e., the Southwest and Sierra Nevada) suggest that 

thinning and fuel reduction can mitigate carbon loss from 

fire. Fuel reduction may reduce losses of carbon at stand 

levels compared with the consequences of high-severity 

wildfire burning in stands with high fuel loads (Finkral and 

Evans 2008; Hurteau and North 2009; Hurteau et al. 2008, 

2011, 2016; North and Hurteau 2011; North et al. 2009, 

Stephens et al. 2009). However, because the probability of 

treated areas burning is generally low (Barnett et al. 2016), 

and most biomass is not consumed by fire, slight differences 
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GENERAL TECHNICAL REPORT PNW-GTR-966

7. Limited understanding of the role of biotic interactions (e.g., competition with and among species) on 

vegetation response to climate change.

8. Reducing uncertainty surrounding projections in 

the amount and geographic distribution of species 

in dry vegetation zones (e.g., ponderosa pine).

9. Response of high-elevation forests to increased 

summer drought.

10. Effects of thinning on resilience to drought in all 

vegetation zones.

11. Effects of increasing landscape heterogeneity 

from fuel treatments (e.g., thinning and prescribed fire) and recent wildfires on future fire 

and insect activity.

12. Phenotypic responses of individual species to 

drought and warmer winter temperatures. 

13. The potential role and identification of climate and 

disturbance refugia in all vegetation zones.

14. Multiscale assessment (i.e., stand to landscape) of 

fuel treatment effects on carbon mitigation under 

increasing fire activity.

15. Potential of the current NWFP reserve network and 

management standards and guidelines to provide 

climate refugia, connectivity to facilitate migration of different species, and stand and landscape 

conditions that promote resilience to drought, fire, 

insects, pathogens, and nonnative species. 

Conclusions and Management 

Considerations

Despite the uncertainty surrounding projections of future 

climate, disturbance and vegetation change, several key 

vulnerabilities have been identified and are supported by a 

large body of scientific evidence (see box on page 56). Most 

models agree and project that the region will experience 

warmer, drier summers and potentially warmer and wetter 

winters. Conditions are projected to exceed the 20th-century range of variability around the 2050s, particularly in 

the Klamath and southern Cascade Mountains. Potential 

impacts in lower elevation, moist vegetation zones (i.e., 

western hemlock) include decreased growth and productivity, especially where species are already water limited 

during the growing season. The greatest vulnerability to 

climate change is in higher elevation forests, specifically 

in the subalpine vegetation zone. These forests are likely to 

experience large decreases in area and may potentially be 

limited to refugia in the Northern Cascade Range (Mote et 

al. 2014). Although a great deal of uncertainty surrounds 

future vegetation change in dry forests, most models consistently agree on an increased role of fire in the 21st century, 

which is likely to include more area burned and larger 

patches of high-severity fire. However, most models do not 

project fire severity or include fire/climate/fuel feedbacks 

that could be used to project severity. 

Projections for climate and vegetation change represent a range of outcomes that can be used to estimate the 

potential magnitude of effects across the region, but they 

do not predict specific outcomes. Recent scientific findings 

suggest several important management considerations for 

mitigation and adaptation in the face of ongoing climate 

change across the NWFP area. It is important to consider 

the potential variability in projections among physiographic 

provinces and even among landscapes and topographic 

settings within a physiographic province when planning 

management activities. 

1. Considering a variety of approaches may be helpful when managing in the face of uncertainty. “Bet 

hedging” strategies and multiple courses of action 

may help to minimize risk and enable further learning. One strategy for dealing with this uncertainty in 

a planning context is to use scenarios and risk analysis (Acosta and Corral 2017, Bizikova and Krcmar 

2015, Pasalodos-Tato et al. 2013) (see also chapter 12). 

Maintaining dense late-successional forests may help mitigate effects of climate change 

and have the potential to buffer warming at finer 

scales in moist vegetation zones where fires are 

infrequent. In addition to storing large amounts of 

carbon, late-successional forests may also provide 

refugia for species that depend on cooler, mesic 

habitats. In dry forest landscapes, maintaining 

large areas of dense, multilayered older forests 

would be inconsistent with a strategy for increasing resilience to drought and fire (chapter 3). 
65

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

2. Landscape-scale treatments to reduce fuels with 

thinning, prescribed fire, and managed wildfire 

may promote heterogeneity in dry forests where 

historical fire regimes were interrupted during 

the 20th century. These activities can also reduce 

vulnerability to high-severity fire during moderate weather conditions, as well as to extensive 

pathogen and insect outbreaks. Topography can 

provide a physical template to consider when 

designing and implementing landscape-scale 

treatments (e.g., thinning on dry ridges and 

around sheltered refugia). 

Maintaining and increasing connectivity 

may facilitate migration of species experiencing 

unsuitable climatic conditions. However, connectivity needs are likely to differ among species, and 

generic connectivity measures may not be adequate for focal species. In situations in which species’ climatic envelopes are changing more rapidly 

than species are migrating, assisted migration can 

promote genetic and phenotypic diversity and may 

help maintain forest cover, although the net benefits of this practice are uncertain and controversial 

in the scientific literature. 

3. Monitoring of populations, species distributions, 

forest conditions, and disturbance are essential 

to inform management decisions and help prioritize objectives for adaptive management in 

response to changes. Most species are expected 

to respond individually to projected changes in 

climate and disturbance regimes, and future forest 

communities may not have contemporary analogs. Understanding the responses of an individual 

species and how they differ across its range can 

assist in developing strategies to promote species 

persistence and prioritize management efforts.

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Douglas-fir forests northeast of Mount St. Helens, 

Washington, following an A.D. 1480 eruption. Seattle, 

WA: University of Washington. 100 p. Ph.D. dissertation.
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GENERAL TECHNICAL REPORT PNW-GTR-966

Appendix: Crosswalk of Simpson (2013) Potential Vegetation Zones With Existing 

Vegetation From the Classification and Assessment With Landsat of Visible Ecological 

Groupings (CALVEG) System 

Values indicate the percentage of the potential vegetation zone 

that falls into the CALVEG class. Existing vegetation comes 

from the Regional Dominance Type 1 field in the CALVEG 

database and indicates the primary, dominant vegetation 

alliance. The listed existing vegetation alliances comprise 95 

percent of each potential vegetation zone in northern California. Current vegetation types with less than 2 percent cover in 

a potential vegetation zone are not shown. For information on 

CALVEG, see http://www.fs.usda.gov/detail/r5/landmanagement/resourcemanagement/?cid=stelprdb5347192.

Potential vegetation zone CALVEG regional dominance 1

Western hemlock Douglas-fir (40.3%), white fir (18.5%), Jeffrey pine (15.5%), tanoak (madrone) (9%), black oak 

(3.9%), ultramafic mixed conifer (3.7%), California bay (2.9%), red fir (2.4%) 

Tanoak Douglas-fir (40.3%), tanoak (madrone) (11.3%), Oregon white oak (6.2%), California bay (5%)

Shasta red fir Red fir (33.2%), white fir (10.1%), Jeffrey pine (10.1%), barren (10%), mixed conifer–fir (8.1%), 

alpine grasses and forbs (5.1%), pinemat manzanita (5%), subalpine conifers (4.9%), upper 

montane mixed chaparral (2.9%), perennial grasses and forbs (2.1%)

Port Orford cedar Douglas-fir (46.6%), ultramafic mixed conifer (24.8%), Douglas-fir–white fir (7.9%), tanoak 

(madrone) (2.9%), Douglas-fir–ponderosa pine (2.9%), mixed conifer–pine (2.2%), Oregon 

white oak (2%)

Other pine Lower montane mixed chaparral (16.5%), gray pine (10.1%), chamise (8%), Oregon white oak 

(7.1%), interior mixed hardwood (6.6%), canyon live oak (5.6%), blue oak (5.6%), annual 

grasses and forbs (4.8%), Douglas-fir–ponderosa pine (4.4%), scrub oak (3.6%), Douglas-fir 

(3.5%), mixed conifer–pine (3.3%), Sargent cypress (3.2%), black oak (2.5%), knobcone pine 

(2.2%), ponderosa pine (2%) 

Grand fir/white fir Mixed pine conifer (27.1%), white fir (19%), Douglas-fir–white fir (14%), Douglas-fir (10.6%), 

Douglas-fir–ponderosa pine (6.3%), red fir (5.9%), mixed conifer–fir (2.5%), upper montane 

mixed chaparral (2%) 

Douglas-fir Douglas-fir (29.3%), Douglas-fir–ponderosa pine (13.3%), Oregon white oak (12.7%), mixed 

conifer–pine (7.8%), lower montane mixed chaparral (5.3%), canyon live oak (4.6%), black oak 

(4%), interior mixed hardwood (3.8%), ponderosa pine (3.2%), annual grasses and forbs (2%) 

Juniper Annual grasses and forbs (45.3%), mixed conifer–pine (17.2%), barren (8.3%), Douglas-fir–

ponderosa pine (7%), upper montane mixed chaparral (4.3%), perennial grasses and forbs 

(2.9%), manzanita chaparral (2.8%), ponderosa pine–white fir (2.3%), Jeffrey pine (2%) 

Map available from http://www.ecoshare.info/category/gis-data-vegzones.

Source: Simpson 2013. 
93
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GENERAL TECHNICAL REPORT PNW-GTR-966

Old-growth forest, Oswald West State Park, Oregon. 

Photo by David Patte, U.S. Fish and Wildlife Service.
95

Chapter 3: Old Growth, Disturbance, Forest 

Succession, and Management in the Area of the 

Northwest Forest Plan

Thomas A. Spies, Paul F. Hessburg, Carl N. Skinner, Klaus 

J. Puettmann, Matthew J. Reilly, Raymond J. Davis, Jane A. 

Kertis, Jonathan W. Long, and David C. Shaw1

Introduction

In this chapter, we examine the scientific basis of the 

assumptions, management strategies, and goals of the 

Northwest Forest Plan (NWFP, or Plan) relative to 

the ecology of old-growth forests, forest successional 

dynamics, and disturbance processes. Our emphasis is 

on “coarse-filter” approaches to conservation (i.e., those 

that are concerned with entire ecosystems, their species 

and habitats, and the processes that support them) (Hunter 

1990, Noss 1990). The recently published 2012 planning 

rule has increased emphasis on land management rooted 

in ecological integrity and ecosystem processes, using 

coarse-filter approaches to conserve biological diversity 

(Schultz et al. 2013). Fine-filter approaches (e.g., species 

centric), which are also included in the 2012 planning 

rule, are discussed in other chapters. We synthesize new 

findings, characterize scientific disagreements, identify 

emerging issues (e.g., early-successional habitat and 

fire suppression effects) and discuss uncertainties and 

research needs. We also discuss the relevance of our 

findings for management. Climate change effects on vegetation and disturbance and possible responses (adaptation 

and mitigation) are addressed mainly in chapter 2 of this 

report. Although, our effort is primarily based on published literature, we bring in other sources where peer-reviewed literature is lacking, and we conduct some limited 

analyses using existing data. We are guided by the NWFP 

monitoring questions, those from federal managers and 

our reading of the past three decades of science. 

Old-growth forests can be viewed through many 

ecological and social lenses (Kimmins 2003, Moore 2007, 

Spies and Duncan 2009, Spies and Franklin 1996). Socially, 

old growth has powerful spiritual values symbolizing 

wild nature left to its own devices (Kimmins 2003, Moore 

2007), and many people value old growth for its own sake 

(“intrinsic” values, sensu Moore 2007). Old growth also has 

many “instrumental” or useful functions, including habitat 

for native plants or animals (e.g., the northern spotted owl 

[Strix occidentalis caurina]), carbon sequestration (Harmon et al. 1990), and other ecosystem services. No single 

viewpoint fully captures the nature of the old-growth issue 

as it relates to federal forest management. We focus here 

on ecological perspectives (Kimmins 2003, Oliver 2009, 

Ruggiero et al. 1991, Spies 2004, Spies and Franklin 1996), 

many of which are overlapping conceptually and in common parlance. Old growth is many things at the same time; 

for example, old growth is: 

• An ecosystem “distinguished by old trees and 

related structural attributes. Old-growth encompasses the later stages of stand development that 

typically differ from earlier stages in a variety of 

characteristics including tree size, accumulation 

of large dead woody material, number of canopy 

layers, species composition and ecosystem function” 

(USDA FS 1989). 

• An ecological state resulting from interactions among 

successional, disturbance, and ecosystem processes 

(e.g., nutrient and carbon cycles, microclimate). 

• A biological condition defined in terms of life histories and demographics of forest plant species. 

• A habitat for particular fauna, flora, and fungi. 

1 Thomas A. Spies is a senior scientist, U.S. Department of 

Agriculture, Forest Service, Pacific Northwest Research Station, 

3200 SW Jefferson Way, Corvallis, OR 97331; Paul F. Hessburg

is a research landscape ecologist, U.S. Department of Agriculture, Forest Service, Pacific Northwest Research Station, 1133 

N Western Ave., Wenatchee, WA 98801; Carl N. Skinner is a 

geographer (retired), U.S. Department of Agriculture, Forest Service, Pacific Southwest Research Station, 3644 Avtech Parkway, 

Redding, CA 96002; Klaus J. Puettmann is a professor, Oregon 

State University, Corvallis, OR 97331; Matthew J. Reilly is a 

postdoctoral researcher, Humboldt State University, Department of 

Biological Sciences, 1 Harpst Street, Arcata, CA 95521; Raymond 

J. Davis is a wildlife biologist and Jane A. Kertis is an ecologist, 

U.S. Department of Agriculture, Forest Service, Pacific Northwest 

Region, 3200 SW Jefferson Way, Corvallis, OR 97331; Jonathan 

W. Long is an ecologist, U.S. Department of Agriculture, Forest 

Service, Pacific Northwest Research Station, 1731 Research Park 

Dr., Davis, CA 95618; David C. Shaw is a professor, Oregon State 

University, Corvallis, OR 97331.
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GENERAL TECHNICAL REPORT PNW-GTR-966

We define old-growth forests based on live and 

dead structure and tree species composition (see below). 

Old-growth forests in the NWFP area differ with age, 

forest type, environment, and disturbance regime (Reilly 

and Spies 2015, Spies and Franklin 1991). The variability 

and complexity of site conditions, forest succession, and 

disturbance processes make defining old-growth difficult 

or impossible under a single definition. Under the U.S. 

Department of Agriculture (USDA), Forest Service 

(USDA FS 1989) definition (above), the only features 

distinguishing old-growth from other forests, across all 

forest types, are the dominance or codominance of old, 

large, live and dead trees (multiple canopy layers are not 

necessarily a defining characteristic). For example, in 

fire-frequent historical forest types, old-growth forests 

have large old live and dead trees, but amounts of deadwood are low, canopies are generally open, and areas with 

multiple canopy layers are uncommon (Dunbar-Irwin and 

Safford 2016, Safford and Stevens 2016, Youngblood et al. 

2004) (fig. 3-1). 

In the NWFP, “older forests” were defined as “late-successional/old-growth” based largely on stand developmental 

and successional patterns of Douglas-fir/western hemlock 

(Pseudotsuga menziesii/Tsuga heterophylla) forests (Franklin et al. 2002) (fig. 3-2). This multilayered closed-canopy 

old growth (e.g., canopy cover >80 percent) was the focal 

point of old-growth conservation during the development of 

the NWFP, but as we shall argue, old growth is far more 

diverse than that and functions quite differently across the 

range of the northern spotted owl. “Older forests” in the 

original NWFP includes mature forests, 80 to 200 years of 

age—a pre-old-growth stage, known somewhat confusingly 

as “late-successional”2

 in the Plan), and old-growth forests. 

Old-growth has been defined in the NWFP and elsewhere as 

forests containing large and old, live and dead trees, a 

variety of sizes of other trees, and vertical and horizontal 

heterogeneity in tree clumps, gaps, and canopy layering (see 

2

 Most of the time in this document, we use the term “late successional” to refer to vegetation that is in the later stages of forest 

succession where age, height, and biomass are near maximum and 

shade-tolerant species are the primary understory or overstory tree 

species. This broad class would include old growth according to 

classic definitions in textbooks (Barnes et al. 1998). 

O’Hara et al. 1996, Spies 2006, and Davis et al. 2015 for 

more discussion of old-growth or old-forest definitions). 

According to Spies and Franklin (1988), old-growth is part 

of a structural and compositional continuum of successional 

stages that varies by environment. According to O’Hara et 

al. (1996), speaking of frequently disturbed environments, 

old forest is a part of the successional continuum that varies 

by environment and disturbance processes, which have the 

ability to advance or retard succession.

To operationalize the successional continuum concept of old-forest development, Davis et al. (2015) created 

an old-growth structure index (OGSI) to characterize the 

degree of old-growth structure (“old-growthiness” calibrated by potential vegetation type) that occurs in a stand 

of any age or history, for use in mapping and monitoring 

in the Plan area. Two definitions for late successional/old 

growth were then created: OGSI 80 (structural conditions 

commonly found in forests that are 80 years and older) 

and OGSI 200 (structural conditions that are representative of forests containing trees that are more than 200 

years of age). These classes roughly correspond to the 

definitions used by FEMAT, the Forest Ecosystem Management Assessment Team (FEMAT 1993), for mature 

trees (80 to 200 years old) (e.g., “late- successional” in 

the NWFP) and old growth (>200 years) but have the 

advantage of being structure based and calibrated to different potential vegetation types. Also, given that this is a 

continuous index, other age/development thresholds (e.g., 

120 years) could be used for mapping and monitoring. 

We note that the structure index and definitions used 

in the monitoring program are based on current forest 

conditions from forest inventory plots, which means that in 

fire-frequent dry zone forests, the structure and composition 

of old growth is a product of 100 years or more of fire 

exclusion and highly altered forest development processes. 

Inventory definitions for dry, old forests based on densities 

of large-diameter fire-tolerant trees have been developed for 

the eastern Washington Cascade Range (Franklin et 

al. 2007a). However, definitions and indices of dry, firedependent, old-growth forest structure at stand and landscape scales are still needed for the larger NWFP area (see 

below for further discussion). 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Figure 3-2—Multilayered, old-growth Douglas-fir and western hemlock stand in the western Oregon Cascades.

Figure 3-1—Open, old-growth ponderosa pine stand maintained by low-severity fire in central Oregon. 

Tom Iraci Tom Iraci
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GENERAL TECHNICAL REPORT PNW-GTR-966

Old growth has been the focal point for forest conservation and restoration on federal lands in the Pacific 

Northwest. However, the broad goals of forest biodiversity 

conservation would not be scientifically viable if they 

focused on only one stage of a dynamic system—all 

developmental phases and ecological processes must be 

considered (Spies 2004), including postdisturbance stages 

(fig. 3-3), nonforest vegetation, and younger forests that 

constitute the dynamic vegetation mosaics that are driven 

by disturbance and succession. These other stages and 

types contribute to biodiversity, and hence, are as important 

to any discussion of forest conservation or management 

for ecological integrity as is the discussion of old growth. 

Indeed, these other successional conditions become future 

old growth, so the successional dynamics of the entire 

landscape ought to be the broader focus of discussions. 

Consequently, our discussion includes these other stages of 

forest succession, in addition to old growth. 

Guiding Questions

This chapter characterizes the current scientific understanding of old-growth forest conditions and dynamics and other 

successional stages in the NWFP area, especially as they 

apply to conservation and restoration of forest ecosystems 

and landscapes. We give special attention to composition 

and structure of trees (live and dead) as dominant components of forests but acknowledge that other characteristics 

are also important, including age (or time since disturbance) 

and composition, and structure of shrub, herb, and grass 

communities. Our focus is on the broad landscape, which 

inherently is a mosaic of vegetation conditions; questions 

related to conservation and restoration of animal species in 

terrestrial habitats and riparian and aquatic ecosystems and 

their habitats are dealt with in other chapters. 

We address the following major questions in this 

chapter, though not directly given their breadth, complexity, and certain degree of overlap. See the conclusions 

section for bullet statements that are explicitly linked to 

these questions. 

1. What are the structures, dynamics, and ecological 

histories of mature and old-growth forests in the 

NWFP area, and how do these features differ from 

those of other successional stages (e.g., early and 

mid successional)? 

2. How do these characteristics differ by vegetation 

type, environment, physiographic province, and 

disturbance regime? 

3. What is the scientific understanding about using 

historical ecology (e.g., historical disturbance 

regimes and natural range of variation [NRV]) to 

inform management, including restoration?

4. What are the principal threats to conserving 

and restoring the diversity of old-growth types 

and to other important successional stages (e.g., 

diverse early seral), and to processes leading to 

Figure 3-3—Early-successional vegetation 8 years after a high- old growth? severity fire in multilayered old growth in southwestern Oregon.

Thomas Spies
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

5. What does the competing science say about needs 

for management, including restoration, especially 

in dry forests, where fire was historically frequent? 

6. How do the ecological effects of treatments to 

restore old-growth composition and structure differ 

by stand condition, forest age, forest type, disturbance regime, physiographic province, and spatial 

scale? 

7. What are the roles of successional diversity and 

dynamics, including early- and mid-seral vegetation, in forest conservation and restoration in the 

short and long term? 

8. What is the current scientific understanding concerning application of reserves in dynamic landscapes?

9. How do recent trends of forests in the NWFP 

reserve network relate to both original NWFP 

goals, those of the 2012 planning rule, and climate 

change adaptation needs? 

10. What is the current understanding of postwildfire 

management options and their effects?

11. What are the key uncertainties associated with 

vegetation under the NWFP, and how can they be 

dealt with?

We address these questions using an organization based 

on major forest regions, disturbance regimes, and potential 

and existing forest vegetation types. 

Key Findings

Vegetation Patterns and Classification

Drivers of regional variation in vegetation—

Forest ecosystems of the vast NWFP region are ecologically diverse and complex and do not lend themselves 

to simple generalizations (fig. 3-4). In this synthesis, we 

account for some of that diversity by classifying ecosystems based on potential vegetation types at the zone or 

series level (Henderson et al. 1989, Lillybridge et al. 1995, 

Simpson 2007) in a manner similar to Küchler (1964, 

1974). Potential vegetation types and disturbance regimes 

are somewhat correlated, although disturbance regimes 

can differ significantly within potential vegetation types 

(i.e., biological and physical environments) (Hessburg et 

al. 2007, Kellogg et al. 2007, Wright and Agee 2004,) and 

differences in potential vegetation types or forest composition do not necessarily mean differences in fire history 

(Taylor and Skinner 1998). 

The major biophysical driving variables (aka “drivers”) 

of structure, composition, and dynamics of old-growth 

forests (and forests in general) are climate, topography, 

soils, succession processes, and disturbance processes 

(Franklin and Dyrness 1973; Gavin et al. 2007; Hessburg et 

al. 2000a, 2015; O’Hara et al. 1996; Oliver and Larson 1990; 

Spies and Franklin 1996). In conjunction with landform and 

soil conditions, the geographic and historical variability of 

the regional climate set the stage for somewhat predictable 

biotic communities, pathways of forest development, levels 

of ecosystem productivity, and spatial patterns of disturbance regimes (Agee 1993, Gholz 1982, Hessburg et al. 

2000a, Reilly and Spies 2015, Weisberg and Swanson 2003, 

Whitlock 1992). Climatic variation over time and space 

exerts a strong control over fire frequency (Agee 1993, 

Gavin et al. 2007, Walsh et al. 2015), and forest dynamics 

is a product of the self-organizing interactions of climate, 

topography, disturbance, and plant communities (Scholl 

and Taylor 2010). Forest succession is the process of change 

in tree, shrub, and herb species composition, and structure 

(size, density, and age structure) over time. Disturbances 

can advance, arrest, or retard succession either slowly and 

imperceptibly, rapidly and abruptly, steadily, or in other 

complex and poorly understood ways (O’Hara et al. 1996, 

Spies and Franklin 1996). In combination, forest succession 

and disturbance processes can produce a wide range of 

forest conditions within the NWFP area. 

Classification of vegetation—

Ecological classifications of environment and succession 

are used to promote understanding and implementation of 

management objectives. One way that Oregon and Washington ecologists account for environmental differences in 

succession and in old-growth characteristics (Davis et al. 

2015, Reilly and Spies 2015) is to use potential vegetation 

type (fig. 3-4).

Potential vegetation type is named for the native, 

late-successional (or “climax”) plant community that would 
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GENERAL TECHNICAL REPORT PNW-GTR-966

Figure 3-4—Geographic distribution of potential vegetation zones (aka vegetation types) (Simpson 2013) and physiographic provinces 

across the Northwest Forest Plan area. 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

occur on a site in the absence of disturbances (i.e., wildfire, 

bark beetle outbreaks, root disease, weather events), and 

reflects the biophysical environment (climate, topography, 

soils, productivity) and composition of overstory and 

understory species (Pfister and Arno 1980). Stages along 

the continuum within a potential vegetation type may be 

binned or categorized into distinct successional stages, 

which are mileposts for visualizing forest development 

subjectively given that no clear thresholds in development 

are known (Franklin et al. 2002, Hunter and White 1997, 

O’Hara et al. 1996, Oliver and Larson 1990, Reilly and 

Spies 2015, Spies and Franklin 1988). This classification is 

often required to enable large-landscape analyses, which 

cannot efficiently deal with developmental conditions 

treated as continuous variables.

Not all ecologists and managers use potential vegetation to stratify or map vegetation for management or 

research purposes. For example, managers in California do 

not use potential vegetation but use existing or “actual” 

vegetation cover type instead to classify their forests for 

management (CALVEG)3 (http://www.fs.usda.gov/detail/r5/

landmanagement/resourcemanagement/?cid=stelprdb

5347192.) To help make our discussion more useful to 

managers in California, we provide a cross-walk table (app. 

1) that links the Pacific Northwest Region (Region 6) 

potential vegetation types (see chapter 2, fig. 3-1) to Pacific 

Southwest Region (Region 5) existing vegetation classes. 

We also note, where appropriate, what the CALVEG classes 

might be for a given potential vegetation type. Most of our 

discussions in the text use estimated potential vegetation 

types for California and the rest of the Plan area based on a 

provisional map prepared by Michael Simpson (ecologist, 

Deschutes National Forest) (fig. 3-4). 

3

 One reason given for doing this is that in California vegetation, 

historical fire frequencies were quite high and the time since fire 

exclusion has been too short (e.g., 100 years) to really know what 

the capacity (potential future vegetation) would have been in the 

absence of disturbance. For purposes of this document, we use 

potential vegetation types, because we have a classification and 

map of these that covers the entire NWFP area (e.g., Simpson 

2013), and there is no existing vegetation classification and map 

for Oregon and Washington. The lack of consistent vegetation data 

layers between the two regions makes it challenging to apply the 

findings from one Forest Service region to another.

Moist and dry forests—

At a broad scale, forests of the NWFP area can be classified into moist forests (including the western hemlock, 

Sitka spruce [Picea sitchensis], coastal redwoods, Pacific 

silver fir [Abies amabilis], and mountain hemlock [Tsuga

mertensiana] potential vegetation zones west of the crest 

of the Cascade Range in Oregon and Washington), and dry 

forests (mainly ponderosa pine [Pinus ponderosa], Douglas-fir, grand fir [A. grandis], and white fir [A. concolor] 

potential vegetation types) east of the Cascade Range and 

in southwestern Oregon and northern California (Franklin 

and Johnson 2012). We use this moist forest and dry forest 

classification to frame much of this chapter. 

Disturbance Regimes

Fire regime classification—

For most forest types, fire was and continues to be the major 

landscape disturbance agent that resets succession or shifts 

its course to a new pathway (Reilly and Spies 2016). Other 

disturbance agents are important as well, including wind 

and biotic agents, but most disturbance regime classifications and maps focus on fire. We characterize the ecology of 

multiple disturbances for moist and dry forests in sections 

below. In this section, we focus on approaches to classifying 

historical fire regimes. 

Most of our current understanding of historical fire 

regimes is based on frequency—empirical studies of severity 

proportions and spatial patterns at landscape scales are 

relatively few (Hessburg et al. 2007, Reilly et al. 2017). Fire 

disturbances occur along a continuum of frequency, severity 

(e.g., tree mortality), seasonality, spatial heterogeneity, and 

event sizes. While there is no single classification of disturbance regimes, they are often binned into regime types that 

are based on fire frequency and severity (Agee 1993, 2003). 

Average fire frequency interval classes of frequent (<25 years), 

moderately infrequent (25 to 100 years), infrequent (100 to 

300 years), and very infrequent (>300 years) (Agee 1993) are 

often used, but other frequency classifications exist as well: 

e.g., ≤35, 35 to 200, and >200 years (Hann and Bunnell 2001, 

Hann et al. 2004, Rollins 2009, Schmidt et al. 2002). 

A widely used classification of fire-severity regimes for 

vegetation uses three bins of basal area or canopy mortality: 
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GENERAL TECHNICAL REPORT PNW-GTR-966

low (<20 percent), mixed or moderate (20 to 70 percent), 

and high (>70 percent)4 (Agee 1993, Hessburg et al. 2016, 

Perry et al. 2011) (fig. 3-5). Other classifications have been 

used, often with higher thresholds for canopy cover loss or 

mortality (e.g., 75 to 95 percent) (Miller et al. 2012, Reilly et 

al. 2017). The classification of Agee (1993) was initially 

4

 Note that while individual patches can exceed 70 percent 

mortality, fires typically have such high levels of mortality in only 

a small fraction of their total area. For example, the high-severity 

area of the 1988 Yellowstone fires was 56 percent (Turner et al. 

1994), and the high-severity percentage of the 2002 Biscuit Fire 

in the Klamath of Oregon and California was 14 percent with an 

additional 23 percent at moderate severity based on a sample of 

inventory plots (Azuma et al. 2004).

developed for the stand or patch scale, but the metric has 

also been applied to larger regional areas (Agee 1993, 

Heinselman 1981, Reilly et al. 2017) or entire fire events, 

which can create confusion about the meaning of fire 

severity (Hessburg et al. 2016): Is it a fine-grained mix of 

severities, or coarse-grained mix of high and low severity, 

or both? Severity can also be characterized in terms of 

fire-induced changes to soils (i.e., soil burn severity); 

however, we focus on vegetative effects in this chapter. Soil 

burn severity is used in Burned Area Emergency Response 

analyses and is often confused with burn severity to 

vegetation (Safford et al. 2007). 

Figure 3-5—Conceptual diagram characterizing the proportions of low-, moderate-, and high-severity fires in three major fire regime 

classes. Inset panels represent idealized landscape dynamics associated with each regime based on proportions and size class distributions of patches at each of the three levels of severity. From Reilly et al. 2017, who modified it slightly from Agee (1993, 1998). 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

For management applications and regional planning, 

broad-scale regime classifications are typically used (Haugo 

et al. 2015), but fire history studies indicate that fire regimes 

can be relatively distinctive at topographic and landform 

scales (10° to 103 ac) (e.g., Taylor and Skinner 1998, Tepley 

et al. 2013). At landscape scales (ca. 103 to 106 ac), most 

fires occur as a mix of low, moderate, and high severity, 

driven by variation in topography, land forms, microclimate, surface and canopy fuels, soils, and vegetation, as we 

explore in later sections. 

Combining fire regimes into broad average frequency 

and severity types is useful for regional planning (e.g., 

Rollins 2009, USDA and USDI 1994), but it oversimplifies 

variability that exists at finer scales, which is important for 

landscape planning and management. In general, simplifying fire into a few regime classes can obscure ecological 

diversity associated with fire effects (Hutto et al. 2016). 

Note that fire-severity proportions for any particular 

landscape or landform is often more restricted than implied 

by the broad ranges used to define broad regime classes. For 

example, for some landscapes in the very high frequency, 

low-severity regime (see below), the historical range of 

high-severity fire may be in the low end of the 0 to 20 

percent5 range used to define this class. 

A new fire regime classification—

For national and regional planning and management purposes, managers often use the LANDFIRE (Rollins 2009) 

fire regime classification. Our review of recent science in 

the NWFP region suggests that the national-scale product 

oversimplifies the fire history within the NWFP area. Thus, 

we developed a new classification and map (table 3-1, fig. 

3-6) by synthesizing existing data on climate, lightning, and 

potential vegetation types (see app. 2 for methods) and fire 

history studies (app. 3). 

5 Odion et al. (2014) called for restricting definitions of historical 

low- and mixed-severity fires to regimes where crown fires and 

active or passive torching are generally absent. However, this 

classification would not be useful, as crown fires can occur in all 

fire regimes including low-severity regimes (Agee 1993), particularly when the regimes are intermixed, as they often are, where 

large landscape contain a range of topography, environmental, or 

vegetation conditions. 

This classification and map are meant to be a rough 

guide for understanding and visualizing ecological variation at regional scales and for framing a discussion about 

forest conservation and restoration science in the NWFP 

area (figs. 3-4 and 3-5). They reflect current understanding 

of fire ecology and geographic variability in the region. 

This typology is different from that used in the record 

of decision (USDA and USDI 1994) and FEMAT (1993) 

documents, which divided the NWFP region into moist 

and dry physiographic provinces but did not characterize 

variability in regimes within them. The physiographic 

provinces explained much of the variation in the physical 

environment, but they contain considerable subregional 

variations in vegetation types and fire regimes that are 

important to understanding the ecology of the forests 

in NWFP area. The potential vegetation types differ in 

distributions of fire regimes that occur within them (fig. 

3-7), and the distribution of potential vegetation types 

differs between fire regimes, though the differences are 

relatively small between regimes within the moist or dry 

forests (fig. 3-8). Almost all fires in these regimes have 

mixed-severity effects, but they typically differ in the 

proportion and distribution of the high-severity effects. The 

very frequent low-severity regime, for instance, contains 

some area in high-severity fire patches at the scale of acres 

to tens of acres. The recognition of a drier, more fire-frequent mixed-severity zone on the west side of the Cascade 

Range in Oregon (fig. 3-6) is based on a number of studies 

(Agee and Edmunds 1992; Dunn 2015; Impara 1997; Reilly 

and Spies 2016; Tepley et al. 2013; Weisberg 2004, 2009). 

This regime, which typically burns with mixed severity 

and includes medium to large patches of high-severity fire, 

was first identified by Agee (1993), based in part on the fire 

history work of Morrison and Swanson (1990) from the 

western Cascades in Oregon. 

Our classification also recognizes that the California 

portion of the NWFP area cannot be simply divided into a 

moist (Coastal province) and dry (Klamath and Cascades 

provinces) province for understanding succession and 

disturbance regimes. In fact, that area has relatively little 

of the “moist” forest that is characterized by historically 
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GENERAL TECHNICAL REPORT PNW-GTR-966

infrequent, high-severity fires. Rather, forests in the 

California Coastal province were dominated by frequent, 

mixed-severity regimes, while the eastern Klamath and 

California Cascades were dominated by historical regimes 

of very frequent, low-severity fire. 

Historical maps of high-severity burned forest patches 

from Washington and Oregon (data not available from 

California) (Plummer et al. 1902, Thompson and Johnson 

1900) provide an independent source of primary data to 

evaluate the regional regime map. These maps support the 

hypothesis that the largest patches and percentage of 

forest burned by high-severity fire occurred in the 

infrequent high-severity regime; whereas the smallest 

patches and lowest area of forest burned by high-severity 

fire occurred in the very frequent/low-severity regime (fig. 

3-9).6

 The relatively high percentage of area burned in the 

infrequent fire regime may reflect elevated ignitions from 

Euro-American settlement activities, because lightning 

densities in these areas are low (fig. 3-10) and these forests 

are not typically fuel limited (Agee 1993). American 

Indian burning practices would have also been a historical 

component in some parts of the region, but the importance 

would have varied considerably among regimes (see 

chapter 11). For example, several studies (app. 3) have 

6

 These early 20th century maps are our best snapshots of this time 

period but do not necessarily represent the range of variability 

in fire sizes that would occur in these regimes over time. This is 

especially true for the infrequent, high-severity regime where 

sample of historical fires is small and extremely large patches of 

fire may have occurred in past centuries.

Table 3-1—Characteristics of major historical fire regimes used in this report and in figure 3-6 

NWFP 

forest 

zone Regime and landfire group PVTs and cover types Spatial characteristics 

Moist Infrequent (>200-year return 

intervals), stand replacing; 

LANDFIRE group V

PVT: wetter/colder parts of western 

hemlock, Pacific silver fir, 

mountain hemlock 

Cover types: Douglas-fir, western 

hemlock, Pacific silver fir, noble 

fir, mountain hemlock

Area dominated by large to very large 

patches (103

 to 106) of high-severity fire; 

low- and moderate-severity fire also 

occurs. Small- to medium-size patches 

were most frequent. 

Moderately frequent to 

somewhat infrequent (50- to 

200-year return intervals), 

mixed severity; LANDFIRE 

regime group III

PVTs: drier/warmer parts of 

western hemlock, Pacific silver fir 

and others 

Cover types: Douglas-fir, western 

hemlock, Pacific silver fir, 

noble fir

Mixed severity in space and time, typically 

including large (103

 to 104 ac) patches of 

high-severity fire and areas of low- and 

moderate-severity fire. Small patches of 

high-severity would be common within 

lower severity areas.

Dry Frequent (15- to 50-year 

return intervals) mixed 

severity; LANDFIRE 

regime group I and III 

PVTs: Douglas-fir, grand fir, white 

fir, tanoak 

Cover types: Douglas-fir, white fir, 

red/noble fir, western white pine

Mixed-severity fire with medium to large 

(102

- to 103-ac ) patches of high-severity 

fire.

Very frequent (5- to 25-

year return intervals) low 

severity; LANDFIRE 

regime group I

PVTs: ponderosa pine, dry to moist 

grand fir, white fir 

Cover types: ponderosa pine, 

Douglas-fir, mixed pine, oak

Dominated by low-severity fire with finegrained pattern (<10° to 102

 ac) of highseverity fire effects; large patches of highseverity fire rare in forests except in earlier 

seral stages (e.g., shrub fields). 

NWFP = Northwest Forest Plan, PVT = potential vegetation type/zone used in the Pacific Northwest Region. Cover type = current vegetation 

classification used in the Pacific Southwest Region. LANDFIRE regime groups follow Rollins (2009). 
105

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Figure 3-6—Generalized fire regimes for the Northwest Forest Plan (NWFP) area based on climate 

and lighting density. Fire frequency, particularly in coastal areas of California, may be underestimated because historical ignitions by American Indians are not included in the model. See table 

3-1 for more information about the regimes and appendix 2 for methods. Moist forests are typically 

associated with the infrequent and moderately frequent regimes, while dry forests typically are 

associated with the frequent and very frequent regimes. 
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GENERAL TECHNICAL REPORT PNW-GTR-966

noted that burning by American Indians likely caused 

fires to be very frequent (<29 years) (app. 3) in the 

redwood (Sequoia sempervirens) forests of northern 

California, although the map based upon climate and 

incidence of lightning classifies those areas as moderate 

frequency, mixed-severity fire regimes. 

The lack of close correspondence of fire regime with 

major potential vegetation type or climate zone (figs. 3-4 

and 3-6) indicates that vegetation type at the zone (series) 

level (at climax) and fire regime do not necessarily respond 

in the same way or at the same scale to variation in the 

environment (Kellogg et al. 2007) (see discussion of the 

regimes for more information). If disturbance regime 

variation within subregions and landscapes is not taken 

into account, efforts to retain or restore biological diversity 

based on historical fire regimes may not be effective or may 

have undesirable effects. 

Disturbance regimes of moist forests—

Moist forests occur primarily west of the crest of the 

Cascades in Washington and Oregon, including the Coast 

Range forests, and on the west slope of the Cascades, 

they extend into high-elevation wet and cool forests (fig. 

3-4). Potential vegetation types are dominated by western 

hemlock, Pacific silver fir, and mountain hemlock (fig. 3-8). 

Sources of stand-replacement disturbance in this region 

included fire, wind, and volcanic eruptions. Insects and 

diseases, especially root diseases, typically created finer 

grained disturbances such as canopy gaps (e.g., 0.1 ac [0.04 

ha]) to several acres in size) (Dickman and Cook 1989, Spies 

et al. 1989). In California, moist forests with infrequent 

fire regimes are confined to relatively small areas along the 

coast and in some higher elevations. 

100

80

60

40

20

0

Infrequent–high severity

Moderately frequent–mixed severity

Frequent–mixed severity

Very frequent–low severity

Redwood

Tanoak

Sitka spruce

Western hemlock

Silver fir

Mountain hemlock

Subalpine fir

Douglas-fir

Grand fir/white fir

Ponderosa pine

Percent

Vegetation type

Figure 3-7—Percentage of major potential vegetation types (PVTs) in the four different fire regimes. Small percentages of a fire regime 

within a PVT may be a result of errors in the PVT maps, fire regime maps, or both. 
107

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

0

5

10

15

20

25

30

35

40

45

50

A

Redwood

Tanoak

Sitka spruce

Western hemlock

Silver fir

Mountain hemlock

Subalpine fir

Douglas-fir

Grand fir/white fir

Ponderosa pine

Redwood

Tanoak

Sitka spruce

Western hemlock

Silver fir

Mountain hemlock

Subalpine fir

Douglas-fir

Grand fir/white fir

Ponderosa pine

Redwood

Tanoak

Sitka spruce

Western hemlock

Silver fir

Mountain hemlock

Subalpine fir

Douglas-fir

Grand fir/white fir

Ponderosa pine

Redwood

Tanoak

Sitka spruce

Western hemlock

Silver fir

Mountain hemlock

Subalpine fir

Douglas-fir

Grand fir/white fir

Ponderosa pine

Percent

0

5

10

15

20

25

30

35

40

45

50

BPercent

0

5

10

15

20

25

30

35

40

45

50 CPercent

0

5

10

15

20

25

30

35

40

45

50

DPercent

Vegetation type Vegetation type

Figure 3-8—Distribution of major potential vegetation types (PVTs) within the (A) infrequent, high-severity regime; (B) moderately 

frequent, mixed-severity regimes of the moist forests; (C) frequent, mixed-severity regime; and (D) very frequent, low-severity regimes 

of the dry forests. Only major PVTs are shown. See appendix 1 for crosswalk to California vegetation types. Forests currently dominated 

by ponderosa pine would occur within the Douglas-fir, grand fir, and white fir PVTs.
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GENERAL TECHNICAL REPORT PNW-GTR-966

Figure 3-9— Historical (1900–1902) patterns of early-successional postfire patches (where “destruction of timber was nearly or quite complete...areas...with only a partial destruction are not here represented”) (Plummer 1902). Note how patches were fewer and smaller in the high-frequency/low-severity regime compared to the other regimes. Many of these fires would have been ignited by settlement and logging activities but would have burned before fire suppression was effective in most cases. Burned forest patches were digitized from Thompson and Johnson (1900) and Plummer et al. (1902). Data were not available for California. 
109

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Two major fire regimes can be recognized within moist 

forests: infrequent (>200-year return interval) and dominated by high severity; and moderately frequent to somewhat infrequent (50- to 200-year return interval) fire with 

mixed-severity patterns (table 3-1). The infrequent regime 

is characterized by relatively long fire-return intervals and 

dominance of high-severity fire in medium to very large 

patches. Historically, mean fire-free intervals averaged 

greater than 200 years with some areas not experiencing fire 

for more than 1,000 years (Agee 1998). Although most of 

the area in high-severity patches is contained within larger 

patches in this regime, individual fires could have high-severity (>70 percent mortality) patches ranging from quite 

small (1 to 25 ac [0.04 to 20 ha) to very large (>106

 ac [~400 

000 ha]) (Agee 1993, 1998). Given the historical infrequency 

of such fires and the tendency for high-severity fire to erase 

information about previous fires, there are few empirical 

studies based on actual fire occurrence (using fire scars), and 

most of our collective knowledge is derived from studies 

that used age-class data to reconstruct large-scale fire rotations (Hemstrom and Franklin 1982) and maps of historical 

fires (fig. 3-6). Climate variation at century scales controlled 

fire frequency and successional dynamics (Gavin et al. 

2007, Long et al. 1998, Walsh et al. 2015). Fire frequency, 

for instance, was relatively high during the Medieval Warm 

climate anomaly about 1,000 years ago, but declined during 

the Little Ice Age between 1400 and 1850 BP. The low fire 

frequency in these systems was due to chronically high 

fuel moistures and infrequent lightning ignitions (Agee 

1993) (fig. 3-10). Large high-severity fires would typically 

occur during unusually dry periods when synoptic weather 

patterns created strong hot and dry east or north winds 

(Agee 1993; Morrison and Swanson 1990; Weisberg 1998; 

Weisberg and Swanson 2001, 2003), but even those fires 

typically left patches with surviving live trees, which would 

contribute to regeneration and habitat diversity. As in other 

settings, the frequency-size distribution of fires followed a 

negative exponential distribution; i.e., the smallest fires were 

the most numerous, and the largest fires accounted for the 

majority of area burned (e.g., see Moritz et al. 2011). 

Humans have played a role in fire occurrence in these 

forests. American Indian use of fire would have contributed 

to fire regimes, especially in drier regions and in local areas 

near Indian settlements in western valleys and coastal areas 

(Agee 1993, Walsh et al. 2015) (see chapter 11). We did not 

adjust the mapping of fire regimes for potential effects of 

Indian burning. Scientific opinions differ regarding the contribution of Indian burning to these forests over evolutionarily relevant time scales. Clearly, the contribution of such 

Figure 3-10—Density of lightning-ignited fires per 25,000 ac 

on forest lands in the Northwest Forest Plan area for the period 

1992–2013. Black lines are physiographic provinces as delineated 

in figure 3-4.
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GENERAL TECHNICAL REPORT PNW-GTR-966

burning was locally important in many areas. Euro-American influence began around the time of settlement (early 

1800s) and coincided with warming and progressively drier 

weather patterns as the Little Ice Age began winding down, 

potentially exacerbating fire activity (see Weisberg and 

Swanson 2003). 

In the drier parts of the moist forest subregion, fires 

were more frequent and mixed in severity, although 

medium to large patches of high mortality were present 

(table 3-1). The moderately frequent to somewhat infrequent 

regime (Morrison and Swanson 1990, Van Norman 1998) 

occurred across a range of potential vegetation types (fig. 

3-8), along the eastern slopes of the Olympic Mountains 

and Coast Ranges, and the interior valleys extending to the 

western slopes of the Cascades in Oregon (fig. 3-6). The 

climate there is warmer and drier than in the infrequent 

fire regime, and lightning ignitions are more frequent 

(fig. 3-10). Patches of high-severity fire could be highly 

variable and were probably somewhat smaller than in the 

infrequent high-severity regime (Morrison and Swanson 

1990) (fig. 3-9). Mixed-severity fires likely affected many 

older forests (Weisberg 2004). For example, many of the 

existing old-growth trees in the southern western Cascades 

of Oregon and interior parts of the Coast Range in Oregon 

showed evidence of low-severity fire occurrence (fig. 3-11). 

Severe windstorms also played a role in forest dynamics 

Figure 3-11—Percentage of fire-resistant mature and old trees with evidence of fire (scars or charred bark) in the western Cascades and 

Oregon Coast Range in relation to latitude. Line is smoothed running average in 0.5° bins. The increase in evidence of fire on tree boles 

around latitude 44.5° N in Oregon (about the latitude of Corvallis) indicates a shift from infrequent, high-severity to moderately frequent, 

mixed-severity fire regimes moving from north to south (right to left). Data source: Spies and Franklin (1991).
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

west of the Cascade crest (Knapp and Hadley 2012). Wind 

occasionally created large stand-replacement patches and 

frequently small gap disturbances across all forest types 

in the region. While the frequency of wind disturbance 

is greatest near the coast (Harcombe et al. 2004) and in 

the Columbia Gorge (Sinton et al. 2000), infrequent large 

regional-scale wind events, such as the 1805 “perfect storm” 

experienced by Lewis and Clark (Knapp and Hadley 2011), 

the 1962 Columbus Day windstorm (Lynott and Cramer 

1966), and the 1981 Big Blow of November 14th can affect 

forests across the west side of Oregon and Washington. The 

1962 storm may be the largest natural disturbance event in 

regional forest history, blowing down 11 billion board feet 

of timber across Washington and Oregon, in concentrations 

of over 80 ac/mi2 (12.5 ha/km2) in some areas (Teensma 

et al. 1991). The frequent occurrences of large windstorms 

in coastal areas control tree growth, forest structure, and 

successional patterns (Knapp and Hadley 2012). More 

frequently, windthrow disturbances are typically related 

to patterns of topographic exposure, which can concentrate windflow (Harcombe et al. 2004, Sinton et al. 2000, 

Wimberly and Spies 2001), root disease, or edges of older 

and younger patches of forests (Franklin and Forman 1987, 

Sinton et al. 2000) created by clearcutting or other stand-replacement disturbances. 

Biotic disturbance agents play important roles in 

succession, and in ecosystem processes and patterns of 

moist forests (table 3-2). They also play important roles in 

producing dead and damaged trees that serve as wildlife 

habitat (Bull 2002). These agents primarily include root 

diseases and bark beetles, although foliage diseases, 

defoliators, heart rots, rust diseases, and dwarf mistletoes 

can also be quite important. Root disease fungi and 

related organisms cause root death, heart rot of large roots 

and tree butts, reduced tree productivity, top dieback, 

and tree mortality, while interacting with bark beetles 

or other mortality agents to influence gap dynamics and 

stand structure (Hansen and Goheen 2000, Lockman and 

Kearns 2016). Phellinus sulphurescens (syn Poria weirii 

or P. weirii in the older literature) clones are thought to 

occur on about 5 to 16 percent of the landscape in the 

moist forests (Lockman and Kearns 2016, Washington 

State Academy of Sciences 2013), for example. Root 

rot diseases are often called, “diseases of the site” in 

the sense that once established in a stand, the fungi can 

persist for decades on belowground wood depending on 

management or compositional changes (Hadfield et al. 

1986, Shaw et al. 2009). 

Foliage disease fungi can be major disturbance agents 

that influence competitive relationships and tree productivity potentially throughout a climatic region (Bednářová 

et al. 2013). However, foliage diseases in Pacific Northwest 

forests are best known in young plantation forests, and 

are poorly studied in natural, or especially, older forests 

(Shaw et al. 2011). Swiss needle cast, caused by the native 

fungus Phaeocryptopus gaeumannii, is currently causing an 

epidemic in managed Douglas-fir coastal forests of Oregon 

and Washington state, within about 35 mi (56.3 km) of the 

Pacific Ocean, reducing plantation productivity an average 

of 23 percent within a study area of the northwest Coast 

Range of Oregon (Maguire et al. 2002, 2011, Navarro and 

Norlander 2016, Ramsey et al. 2016, Ritóková et al. 2016). 

The disease is particularly associated with lower elevations 

of the infrequent–high-severity fire regime (fig. 3-6). The 

role of foliage diseases in the development of forest stands, 

and in particular, old-tree crown dynamics, remains elusive. 

It is generally thought that maintaining tree species diversity, canopy complexity, and adherence to site compatible 

seed zones reduces the threat of foliage diseases to forest 

health (Shaw et al. 2009).

Bark beetles and wood borers are diverse, but major 

disturbance from mortality is mostly associated with 

climatic events such as drought, ice/snow breakage, and 

windthrow (Furniss and Carolin 1977). Two particularly 

important species are the fir engraver (Scolytus ventralis

(LeConte)) in true firs (Ferrell 1986) and the Douglas-fir 

beetle (Dendroctonus pseudotsugae (Hopkins) in Douglas-fir (Furniss and Kegley 2014). Mortality from both 

insects is associated with root diseases and drought, and, in 

the case of the Douglas-fir beetle, with windthrow events 

(Furniss 2014a, 2014b; Goheen and Willhite 2006). Typically, flareups of mortality from this beetle persist for a few 

years and then abruptly subside (Furniss and Carolyn 1977, 

Goheen and Willhite 2006). 
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GENERAL TECHNICAL REPORT PNW-GTR-966

Table 3-2—Major biotic disturbance groups, effect on trees, and ecological influences in forests of the 

Northwest Forest Plan area

Disturbance groupa Tree effects Ecological influences

Root diseases Major mortality agent Alters stand composition/structure

Growth reduction Creates snags, down wood

Root death Wildlife cavities

Creates ant/termite habitat

Attracts bark beetle mass attack

Root/butt heart trots Increases surface fuels

Live tree decays Wood volume reduction Wildlife cavity creation

Increased windsnap Reduced carbon sequestration

Creates ant/termite habitat

Foliage diseases Reduce foliage retention Less competitive in stands

Reduced growth Reduced carbon sequestration

Carbon starvation Alters stand composition/structure

Cankers and rusts Branch, top, tree death Reduced carbon sequestration

Foliage loss Reduce host species abundance

Tree deformation Wildlife habitat

Dwarf mistletoe Growth reduction Alters forest structure/composition

Top, branch, and tree death Encourages passive crown fire

Branch and tree deformation Wildlife habitat platforms

Increased susceptibility to other agents Influence with fire

Bark beetles Major mortality agent Alters composition/structure 

Patch attacks on bole Increases forest fuels

Top and branch death Wildlife habitat

Defoliators Growth loss Alters composition/structure

Top dieback Reduces canopy density

Mortality Wildlife habitat impacts

Aphids, adelgids and scale insects Growth loss Alters forest structure

Leaf, branch, and tree death Reduced carbon sequestration

Terminal and branch insects and pitch moths Tree leader death

Stunted growth Forest structure

Tree deformation Reduced competitive ability

a

 Groups from Shaw et al. (2009).

Source: Furniss and Carolin 1977, Goheen and Willhite 2006, Scharpf 1993, Shaw et al. 2009, Wood et al. 2003.
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Other important biotic agents include the hemlock 

dwarf mistletoe (Arceuthobium tsugense Rosendahl), which 

is the only known moist forest dwarf mistletoe, and can 

dramatically influence forest structure (Muir and Hennon 

2007). The plant occurs localized in western hemlock-dominated forests, where it is estimated to infect 10.8 percent 

of the western hemlock trees in Oregon (Dunham 2008). 

Hemlock dwarf mistletoe has a strong connection to fire 

history (Shaw and Agne 2017); more frequent fires favor 

less mistletoe. 

Disturbance regimes of dry forests— 

This region includes the mid to lower elevations of the 

eastern Cascades from Washington to California, southwestern Oregon, in the Klamath region, and inland portions 

of the California Coast Range. It spans a range of dry forest 

potential and current vegetation types, including ponderosa 

pine, Douglas-fir, and white fir (figs. 3-4 and 3-6; table 

3-1). Fire is the major stand-replacement disturbance in this 

region followed by outbreaks of major forest insects. 

The more moist and productive part of this region 

experienced a frequent, mixed-severity regime with fire-return intervals of 15 to 50 years (Agee 1991, Agee et al. 1990b, 

Stuart and Salazar 2000, Taylor and Skinner 1998, Van Norman 1998, Whitlock et al. 2004, Wright and Agee 2004). Fire 

events contained medium to large patches of high-mortality 

and extensive areas of low- and moderate-severity fire. The 

2002 Biscuit Fire is an example of such a fire (Halofsky et al. 

2011, Thompson and Spies 2009) (fig. 3-12). The occurrence 

of mixed-severity fire even at short fire-return intervals (e.g., 

<25 years) probably reflects the higher moisture conditions 

and site productivities in parts of this regime in comparison 

to the very frequent, low-severity dominated regime in 

Figure 3-12—Mosaic of high-severity burn patches in a portion of the 2002 Biscuit Fire in southwest Oregon in an area classified as 

historically supporting a frequent, mixed-severity fire regime (fig. 3-6). A large portion of the area with surviving tree canopies experienced low-severity surface fire.

Thomas Spies
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GENERAL TECHNICAL REPORT PNW-GTR-966

California or the eastern Cascades. Patterns of mixed-severity 

patches were historically shaped by prevailing topographic 

features (Beaty and Taylor 2001; Hessburg et al. 2015, 2016; 

Taylor and Skinner 1998; Weatherspoon and Skinner 1995) 

with variable proportions of both surface and crown fires 

accounting in part for tree mortality in mixed-severity fire 

regimes (Perry et al. 2011, Stephens and Finney 2002). 

The very frequent (<25 years) low-severity regime occurs 

in the driest forests7 of the NWFP area in a variety of pine, 

dry Douglas-fir, dry grand or white fir, and oak potential and 

current vegetation types (figs. 3-4 and 3-6, table 3-1, app. 1). 

Historically, fires burned very frequently, with average fire 

intervals between 5 and 25 years (Bork 1984; Everett et al. 

2000; Sensenig et al. 2013; Soeriaatmadja 1965; Taylor and 

Skinner 1998, 2003; Weaver 1959), although for many forests 

the range was much narrower. Overall, tree mortality from 

fire was low, with typically <20 percent of the trees killed in 

fires, and most high-severity effects occurring in very small 

patches (<1 ac [<0.40 ha]). Fire severity was primarily 

influenced by fine-scale patterns of surface fuels and topography (Churchill et al. 2013, Larson and Churchill 2012). Fuels 

were reduced frequently enough that active crown fire was 

infrequent. Frequent fires often created multicohort stands 

with low tree density and canopy cover (Hagmann et al. 2013, 

2014; Sensenig et al. 2013). Larger patches (>250 ac [>101 ha]) 

of high severity could occur but were uncommon in most 

areas (Agee 1993, Rollins 2009; Skinner 1995; Taylor and 

Skinner 1998, 2003) and were linked to topography (Taylor 

and Skinner 1998, 2003). The forested landscape was 

dominated by open forests with islands of denser vegetation, 

including clumps of trees of various sizes (Churchill et al. 

2013, Hessburg et al. 2007, Larson and Churchill 2012, 

Lydersen et al. 2013, Perry et al. 2011). Some scientists (e.g., 

Baker 2012) dispute the idea that these dry forests experienced a regime dominated by frequent, low-severity fire, and 

argue instead that they commonly experienced larger patches 

of high-severity fire (see section on alternative viewpoints 

below for more discussion of this). 

7

 In the Klamath and southern Cascades of California, these 

regimes occur where the climate is characterized by long warm/

dry seasons but relatively high precipitation, which is concentrated 

in the winter months. 

Wind is not a major disturbance agent in drier forests 

of the region that are typically inland from coastal areas, 

and south of areas where the strongest windstorms occur. 

Coastal California is south of most of the mid-latitude 

cyclones that affect the Oregon and Washington coast 

(Lorimer et al. 2009). Coastal redwood forests experience 

winter storms and high winds, but effects appear to be 

limited to canopy damage and scattered blowdown of trees 

on high ridges (Hunter and Parker 1993, Lorimer et al. 

2009). Drier ponderosa pine, Douglas-fir, and mixed-conifer forests experience scattered windthrow that creates 

canopy gaps and fine-scale pit and mound microtopography 

(Weaver 1943), but we are not aware of studies that document occurrence of larger patches of windthrow. Reilly and 

Spies (2016) report that between the 1990s and mid 2000s, 

wind was a very small component of all natural sources 

of mortality in dry forests of the Pacific Northwest. Agee 

(1994) reported similar results for the dry interior forests.

Major biotic disturbance agents in dry forests include 

several root diseases and host specialized dwarf mistletoes 

as chronic long-term stand influences that are associated 

with creating complexity in forest patches by killing and 

deforming trees, creating snags and gaps, and influencing 

fuels and fire (Goheen and Willhite 2006, Hadfield et al. 

1986, Hawksworth and Wiens 1996, Lockman and Kearns 

2016, Shaw and Agne 2017) (table 3-2). Major bark beetle 

and defoliator disturbances tend to be episodic, although 

individual old-tree death caused by bark beetles is chronic 

in some forests. Large outbreaks are more common in the 

eastern slope of the Cascades than in northern California, 

where tree species diversity, complex terrain, geological 

diversity, and contrasting site microenvironments may 

reduce the potential for widespread outbreaks. Heart rots, 

rust diseases, cankers, as well as foliage and tip diseases 

and insects may be locally significant, especially heart rots, 

which create cavities for wildlife (Bunnell 2013). 

Root diseases are widespread in dry forests (Filip and 

Goheen 1984, Hadfield et al. 1986, Lockman and Kearns 

2016), where they play an integral part in forest stand 

dynamics and canopy gap formation. In northwestern 

California, Hawkins and Henkel (2011) found that root 

diseases caused more mortality and gap formation in white 
115

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

fir than Douglas-fir, which in the absence of fire, allowed 

Douglas-fir to better persist in forest stands. This is not 

always the case in the dry forests. 

Dwarf mistletoes are host specialized parasitic seed 

plants that are a major influence on dry forest structure. 

Host-specialized mistletoes infest nearly all species, where 

they create structures such as witch’s brooms, dead tops, 

dead branches, and fuel ladders (Hawksworth and Wiens 

1996, Mathiasen and Marshall 1999, Shaw et al. 2004). A 

key ecological function of dwarf mistletoes is the creation 

of wildlife habitat structures via their large witch’s brooms, 

which provide nesting and roosting platforms for a variety 

of forest birds and other small mammals (Shaw et al. 2004). 

Douglas-fir dwarf mistletoes can provide the majority 

of nesting sites for the spotted owl in dry interior forests 

(Buchanan et al. 1995, Forsman et al. 1984). Dwarf mistletoe distribution and abundance is related to fire history; 

with more regular fire there is less dwarf mistletoe because 

heavily infested trees are prone to torching or passive crown 

fire initiation (Shaw and Agne 2017). Although fire influences dwarf mistletoe, dwarf mistletoe also influences fire 

behavior by creating complex fuels structures, contributing 

to surface fuels, increasing ladder fuels, decreasing canopy 

base height, and increasing canopy bulk density. 

 Bark beetles are associated with most mortality events 

in dry forests, however, determining whether the beetles are 

to blame for individual tree mortality can be a challenge. 

Drought, dwarf mistletoe, root diseases, defoliators, and 

other biotic or abiotic factors can all predispose weakened 

trees to bark beetle mass attack. Bark beetle outbreaks can 

also be initiated by long-term drought events, and these 

outbreaks can last well over a decade. Bark beetles are also 

host specialized, and they influence forest stand structure 

and development by killing specific tree species. In the 

aftermath, tree mortality associated with beetle outbreaks 

can contribute significantly to forest fuels, but it can take 

more than a decade or two for the snags of the former forest 

structure to fall down and accumulate on the forest floor. 

Major bark beetle outbreaks typically occur in dry forests 

east of the Cascade crest where expansive stands of lodgepole 

pine (Pinus contorta) have been hit very hard by mountain 

pine beetle (MBP) (Dendroctonus ponderosae) (Gibson et al. 

2009). Recent large bark beetle mortality events associated 

with periods of extended drought in the southern and central 

Sierra Nevada of California suggest that the potential for 

major climate change-driven outbreaks is ongoing and may 

result in species conversion in some areas (Moore et al. 2017). 

The interaction of fire with prior MPB events has become 

a significant research emphasis following large outbreaks 

throughout western North America. Following MPB mortality, canopy fuels decrease drastically within a few years, 

and depending on composition of the stand, surface fuels will 

significantly increase with time (Hicke et al. 2012).

Defoliators on the east side of the Cascade Range are 

a major disturbance agent in forest stands, with the western spruce budworm, Douglas-fir tussock moth (Orgyia 

pseudotsugata), pine butterfly (Neophasia menapia), and 

Pandora moth (Coloradia pandora) potentially able to 

defoliate large regions (Furniss and Carolin 1977, Goheen 

and Willhite 2006). Outbreaks of the western spruce 

budworm (Choristoneura occidentalis) have not occurred 

in dry forests of California and southwestern Oregon 

(Brookes et al. 1987), although the Douglas-fir tussock moth 

may defoliate true firs, and the Pandora Moth may affect 

ponderosa pine (Brookes et al. 1978, Wood et al. 2003). 

Defoliators have the potential to shift composition of stands 

to nonhosts owing to reduced growth and mortality effects, 

as well as increased potential for bark beetle infestation in 

defoliated trees (Brookes et al. 1978, 1987). The interactions 

of fire with forest defoliators suggest a negative association 

of fire and defoliated stands (Meigs et al. 2015). 

Forest Succession and Landscape Dynamics 

Moist forests—

Succession—Our synthesis of this regime is primarily based 

on studies from Douglas-fir and western hemlock forests (i.e., 

the western hemlock potential vegetation type) (Franklin et al. 

2002, Oliver and Larson 1996, Reilly and Spies 2015, Spies et 

al. 1988). Patterns of postfire and postwind stand-replacement 

succession for other potential vegetation types in this fire regime, which have received less study (e.g., mountain hemlock 

in Oregon and Washington, Pacific silver fir potential vegetation types) may have been generally similar, but they differ 

in a number of ways, including species composition, varied 
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GENERAL TECHNICAL REPORT PNW-GTR-966

pathogen and insect associations, and slower rates of structural and compositional development. These potential vegetation 

types also likely have lower levels of total biomass relative to 

Douglas-fir/western hemlock forests in late-successional stages, owing to shorter and cooler growing seasons. 

The archetypal or standard model of forest succession 

in this forest region and under these disturbance regimes 

has been characterized in many papers but is developed in 

greatest depth by Franklin et al. (2002), and most recently 

by Franklin and Johnson (2017) and Franklin et al. (2018) 

(fig. 3-13). Simply stated, after a stand-replacement disturbance such as wildfire or windstorm (1) considerable dead 

and live legacies of the disturbance remain for decades; 

(2) new shade-intolerant and tolerant plants and early-seral 

associated wildlife colonize a site; and (3) a dynamic mix 

of nonforest and forest plant species develops and persists 

until conifer canopy closure, which may take between 30 

and 100 years. The forest then goes through a process of 

structural and compositional changes and stages driven 

by growth, competition, immigration of shade-tolerant 

species, and fine- to moderate-scale mortality events that 

create canopy gaps of various sizes (Bradshaw and Spies 

1992, Spies et al. 1990). These canopy gaps can promote 

growth of shade-tolerant trees growing in the understories 

of densely shaded forests. This is not the only successional 

pathway that forests followed in this large and ecologically 

diverse region, but it is a common one, especially in wetter 

and northern parts of the western hemlock potential vegetation type in cover types characterized by Douglas-fir and 

western hemlock (Winter et al. 2002a, 2002b), and a lack 

of fire between stand-replacement events. We characterize 

this model of succession further below and describe its 

variations and other successional pathways that can occur. 

Early post-stand-replacement fire vegetation in the 

western hemlock–Douglas-fir forests of the western hemlock zone typically occurred as heterogeneous mosaics of 

grasses, herbs and shrubs, and hardwoods often with high 

levels of dead snags and down wood, and high species richness (Donato et al. 2011, Reilly and Spies 2015, Swanson et 

al. 2011) (fig. 3-3). Species compositional change, which can 

be rapid over the first 20 years as a function of the relative 

importance of invading and residual plant species groups, 

differs with time, the availability of propagules, disturbance 

characteristics, and properties of the environment (Halpern 

1988, 1989). Standing dead tree structure and decay states 

are also dynamic within western conifer forests during the 

first decade or two following fire (Russell et al. 2006). Studies of post-wildfire conifer forests in the Western United 

States indicate that wildlife use of early-seral vegetation 

following fire and logging can change rapidly with timesince disturbance, with some species appearing in the first 

few years before disappearing later and others increasing in 

abundance as snag conditions and plant species composition 

changes (Saab et al. 2007, Smucker et al. 2005). Gashwiler 

Figure 3-13—A common stand developmental pathway for a Douglas-fir and western hemlock forest following stand-replacement 

wildfire (from Franklin et al. 2018). 
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(1970) found that small mammal communities were quite 

dynamic in the first 10 years following clearcutting of an 

old-growth forest in the western Cascades of Oregon. The 

general pattern seems to be that while the “pre-forest” or 

early-seral stage can persist for many decades, the plant and 

animal communities are dynamic within that stage, and 

some species and communities are ephemeral. 

Dead wood levels were especially high where prefire 

forests were late successional or old growth (Spies et al. 

1988). Where fires burned early-successional and younger 

forest stand conditions, dead wood legacies were typically few and composed of smaller down logs (Nonaka et 

al. 2007, Spies et al. 1988). In contrast, where fires burned 

in forests containing large trees, levels of down wood 

were high, and individual pieces of large down wood may 

have persisted for several centuries while undergoing 

decomposition. Charcoal deposits from fires lasted in soil 

for up to one or more millennia (DeLuca and Aplet 2008). 

Scientific and conservation interest in early-successional 

vegetation has increased in recent years as scientists 

learned about ecosystem responses to severe disturbance 

from studies of the eruption of Mount St. Helens (Dale et 

al. 2005) and high-severity wildfires that have occurred 

in the Western United States in recent decades (e.g., 

Donato et al. 2011; Hessburg et al. 1999a, 1999b; Hutto 

et al. 2016). Post-high-severity and mixed-severity 

disturbance ecosystems are generally understood to 

support unique biodiversity and ecosystem functions 

(Donato et al. 2011; Franklin et al. 2017; Hessburg et al. 

2016; Swanson et al. 2011, 2014) relative to closed-canopy forests. This understanding is based largely on 

studies of clearcuts (e.g., Halpern 1988, Harr 1986) and 

volcanic eruptions (Dale et al. 2005) in the Northwest 

Forest Plan area, and few studies have been conducted in 

early-seral vegetation following wildfire or windstorms 

(e.g., Fontaine et al. 2009, Larson and Franklin 2005). 

Early-successional stages following natural disturbances 

are rich in biological legacies that include surviving 

organisms and organic matter such as dead trees. With 

tree canopies gone or greatly reduced, other life forms, 

including shrubs, grasses, and herbs often dominate the 

site, taking advantage of higher resource levels in light, 

water, and nutrients. These legacies clearly influence 

postdisturbance succession, stand development, and 

ecosystem function, though the variability in these relationships over time is not well understood. Variation in 

disturbance severity and predisturbance forest conditions 

has strong influence on legacy patterns, and subsequent 

forest succession that can persist for hundreds of years 

(Donato et al. 2011, Dunn and Bailey 2016, Spies et al. 

1988). In sum, early-seral stages are important when 

managing for conservation of native biodiversity and 

resilience in forested ecosystems and landscapes.

Given new scientific perspectives on early-seral 

vegetation, some have proposed that new terminology 

be used to describe it. For example, Franklin et al. 2018 

suggest that early-seral vegetation be termed “pre-forest” 

because trees are not the dominant life form, although 

they are often present as seedlings. They also suggested 

that the term “early-seral forest,” which has been used 

to define this stage, is not correct because this stage 

is not forested and introduces a “tree-centric” bias 

to discussions about conservation and management 

(Franklin et al. 2018). Other terms that have been used 

to describe this stage include grass-forb, shrub-seedling, 

stand initiation, and cohort establishment. Terminology 

to describe successional stage, structural or developmental stage, or seral stage can be confusing and not 

interchangeable (Powell 2012). For example, some trees 

such as Douglas-fir and red alder are characterized as 

“early-seral” species (Franklin and Hemstrom 1981, 

Klinka et al. 1996), which can form early-seral stands 

or forests. The ambiguity of the terminology around 

postdisturbance changes in vegetation (including later 

successional stages) makes it important to define how 

terms are used (e.g., Powell 2012), and in the case of 

early-seral or pre-forest vegetation to clearly identify the 

ecological characteristics (life forms, species, structures) 

and functions (habitat, nutrient cycling, productivity) that 

reflect the underlying meaning and use of those terms.
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The timing, composition, and structure (including 

cover thresholds) of tree canopy cover closure (e.g., canopy 

cover >70 percent (Yang et al. 2005) would have differed 

regionally by site conditions, disturbance characteristics, 

and seed source availability (Freund et al. 2014, Yang et al. 

2005). Canopy closure may have occurred as early as 20 to 

30 years following fire in moist productive sites, or where 

seed sources persisted in a canopy seed bank (Larson and 

Franklin 2005), but could have taken almost 100 years 

on other sites, after very large fires and with limited seed 

sources. These observations are based on studies of mature 

forests from the western Cascades (Freund et al. 2014). 

Tree establishment ended as the forest floor was covered 

by shrub and herbaceous vegetation, and tree canopies 

eventually closed (Freund et al. 2014, Tepley et al. 2014). 

Not all stands or patches followed the same pathway to 

older forest structure. Multiple successional pathways would 

have occurred that varied in timing of composition and 

structural change over the first 100 to 200 years or longer (fig. 

3-14) (Spies 2009). In riparian areas and moist coastal upland 

forests, shrubs and hardwood trees would often become established immediately after fire, limiting the establishment of 

conifer trees for many decades, and creating patches of hardwoods and shrubs with scattered conifers (Spies et al. 2002). 

Ultimately, those shorter lived hardwoods would die, leaving 

lower density conifer stands (or stands with variable-canopy 

dominance) with large dominant trees and well-developed 

crowns. For example, Spies and Franklin (1991) found that 

some 100-year-old stands of Douglas-fir and western hemlock 

that developed along with shrubs and hardwoods in the 

Oregon Coast Range had structural diversity that approached 

that of 400-year-old stands. Variability in seed sources, productivity, competition with shrubs and hardwoods, and partial 

stand replacement disturbances would have led to low-density 

relatively open younger forests where conifer canopy closure 

never occurs. These processes and pathways may actually 

be a faster route to complex older forest structure in some 

places than pathways that go through stages characterized 

by a higher density of conifers and conspecific competition 

(Donato et al. 2011, Tappiener et al. 1997). 

Where closed-canopy forests developed, succession 

was driven by processes of growth, competition, understory 

development, maturation, and small- to moderate-size 

canopy disturbances from wind, insects, disease, fire, 

hydrologic, or geomorphic processes (Franklin et al. 2002). 

Somewhat arbitrarily, 80 years after conifer forest establishment has been used as the onset for “mature” (e.g., OGSI 

80) Douglas-fir forests, and 150 to 200 years for the onset 

of multilayered old-growth forests (OGSI 200), depending 

on environment and disturbance history (Franklin et al. 

2002, Spies and Franklin 1991). Eighty years was used as 

the threshold for late-successional/old growth in the NWFP 

(USDA FS 1994) because that is about the earliest time 

when such stands begin to resemble maturing forests in the 

moist forest (does not apply to the dry forest zone). Analyses 

of chronosequences indicate there is considerable variation 

in forest structure around these age breaks (Spies and 

Franklin 1991) (fig. 3-15) likely driven by multiple successional pathways, legacies, and time since disturbance. The 

stands (i.e., sample plots) in figure 3-1 would have followed 

individual development pathways, some pathways may be 

sigmoid shaped in the case of stands developing after a 

nonforest condition, other pathways may have been more 

U-shaped in the case of stands developing with significant 

live or dead legacies of the predisturbance old-growth forest 

(Spies and Franklin 1988). 

The variability in structure with stand age indicates 

that at a regional scale, age or time alone is only a partial 

predictor of forest structure. The structural features 

of mature and old-growth forests would have included 

medium- to large-size (e.g., >40 inches) shade-intolerant 

tree species; smaller shade-tolerant trees of similar and 

lesser age in the mid to lower canopy layers; large standing 

and down dead tree boles; and horizontal and vertical structural heterogeneity of live and dead trees. Not all stands 

would have grown for centuries without stand-replacement 

fire—sometimes reburns within a few decades of a fire 

would occur consuming decayed dead wood and restarting 

succession (Donato et al. 2016, Gray and Franklin 1997, 

Nonaka 2003, Tepley et al. 2013). 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Figure 3-14—Multiple pathways of succession that could occur in the moist forests. Pathway A occurs when Douglas-fir canopy 

closure occurs within 50 years after a fire and western hemlock establishes early in succession. Pathway B occurs when the pre-forest 

shrub-dominated stage persists for many decades and hemlock is slow to establish. Pathway C occurs where shrubs and hardwood trees 

dominated early-successional development and reduced conifer densities so that conifer trees would not go through a self-thinning phase 

and large-diameter conifers and complex older forest structure would develop well before 200 years. Pathway D occurs where a partial 

stand-replacement fire occurs periodically in older forests and creates patches of dead trees, initiating new age cohorts of Douglas-fir or 

western hemlock trees beneath the surviving canopy and in openings created by the fire.
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GENERAL TECHNICAL REPORT PNW-GTR-966

Successional and landscape dynamics in the drier, 

southern part of the western hemlock zone, where fire frequency was 50 to 200 years (fig. 3-4), would have included 

some of the same pathways as would have occurred in the 

infrequent fire regime, but with different frequencies of 

those pathways across landscapes. At the scale of large 

patches and small landscapes (e.g., 102

 to 104 ac or ~40 to 

4000 ha), these forests would have had more age, structural 

and compositional heterogeneity than equivalent areas for 

the moister parts of the region where an infrequent fire 

regime occurred (fig. 3-16). For example, reanalysis of data 

from Spies and Franklin (1991) from the old-growth forests 

in the southern western Cascades of Oregon indicated that 

stand ages (age of the oldest Douglas-firs in the stand) 

were younger (~270 years) and basal area, proportion of 

shade-tolerant trees, and density of large snags and volume 

of down wood were all much lower than in old-growth 

stands in the northern Cascades of Oregon and the Cascades 

of Washington (400 to 500 years), after controlling for 

topography and aspect. Ares et al. (2012) found that snag 

densities in older forests in western Oregon also varied by 

aspect, with lower densities on south-facing slopes and in 

the foothills of the Cascades, where fire frequencies are 

higher than in the Coast Range. The mature and old-growth 

stages probably have more age classes of Douglas-fir than 

in the infrequent, high-severity regime forests as a result of 

more frequent partial stand-replacement fire (Dunn 2015, 

Tepley et al. 2013) (figs. 3-16 and 3-17). For example, Tepley 

et al. (2013) found that 85 percent of the older forest in their 

central western Oregon Cascades study area (primarily 

western hemlock potential vegetation type with some 

areas of Douglas-fir potential vegetation type) experienced 

non-stand-replacing wildfire during its centuries-long 

development (fig. 3-14D). These fires killed a portion of the 

Figure 3-15—An old-growth forest habitat index (OGHI) (Franklin et al. 2005) in relation to stand age for forest 

inventory and research plots in the Oregon Coast Range. The index is based on number of large trees, large snags, 

volume of down woody debris, and tree size diversity, which is a surrogate for canopy layering. Age was not 

used to develop the index. The index is similar to the structure index used in Davis et al. 2015. FIA/CVS = Forest 

Inventory and Analysis/Continuous Vegetation Survey.
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Scale: 1 km

Figure 3-16—Mosaic of fire severity patches in a Douglas-fir and western hemlock landscape in the western Cascade Range of Oregon. 

Black = a high mortality area (>70 percent), vertical lines = moderate mortality (30 to 70 percent), and stippled = low mortality areas 

(<30 percent). From Morrison and Swanson 1990.
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GENERAL TECHNICAL REPORT PNW-GTR-966

Figure 3-17—Conceptual model of stand-development pathways in Douglas-fir/western hemlock (current vegetation) forests in the 

moderately frequent, mixed-severity fire regime of the central western Cascade Range of Oregon. Dashed arrows represent stand development in the absence of fire, and solid arrows represent nonstand-replacing fire. Percentages indicate the percentage of the sample plots 

found in each structure type. SR = stand-replacing, NSR = non-stand-replacing. From Tepley et al. (2013). 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

overstory and established new cohorts of shade-tolerant 

or intolerant trees. Given the long time period that often 

occurred between fires, these landscapes of the infrequent 

and somewhat infrequent regimes would have typically 

been dominated by mature and old-growth forests. 

Historical landscape dynamics—Many of the current oldgrowth stands of the wetter portions of the moist forests 

date to around 400 to 500 years ago (Spies 1991), a period 

with widespread fire (Tepley 2010, Weisberg and Swanson 

2003) associated with positive phase of the Pacific Decadal 

Oscillation, which produced warmer conditions and drought. 

This warm period with many fires was followed by the Little 

Ice Age when cooler temperatures caused a reduction in both 

lighting- and human-ignited fires (Walsh et al. 2015) that may 

have allowed stands that established during the warm period 

to develop into older, multistoried forests. Empirical estimates of the amount or variation in old-growth forests or of 

any successional stage that occurred prior to Euro-American 

settlement are not available from any historical studies. Maps 

from the early 1900s can be used to approximate the amount 

of old forest present in the mid-20th century, suggesting that 

about 50 percent of all forest lands in this regime were covered by older forest (defined then in terms of large dominant 

and codominant trees), but that number varied widely across 

landscapes and watersheds (Davis et al. 2015). However, it 

is not clear how earlier mapping criteria related to current 

definitions of old growth, and by the 1930s, significant areas 

of older forest had already been lost to land clearing for settlement and agriculture, logging, and human-set wildfires. 

Empirical studies of fire frequency and severity can be 

used with statistical models and other simplifying assumptions to estimate the age-class distributions that might have 

been present in a historical landscape (Agee 1993, van 

Wagner 1978). For example, Fahnestock and Agee (1983) 

used historical maps and statistical models to estimate fire 

cycles in western Washington. They found the proportion 

of large trees to be 0.6 in Douglas-fir, 0.82 in western 

hemlock, and 0.87 in mountain hemlock forest cover types. 

Spies and Turner (1999) estimated that on average, 61 

percent of a given landscape would be old growth (>150 

years since stand-replacing fire) if fire frequencies were 300 

years. They assumed a constant climate and fire frequency, 

equal flammability of successional stages, and high-severity 

fire—assumptions that are violated in real landscapes. For 

example, temperature and precipitation has varied considerably over the Holocene (past 11,700 years), including 

the past several thousand years when the current forest 

community assemblages developed (chapter 2). Susceptibility of successional stages often differ depending on fuel 

conditions and microclimate, and old forests can be less 

flammable than younger ones (Kitzberger et al. 2011). 

Wimberly et al. (2000) used estimates of fire frequencies from lake cores in the Oregon Coast Range (Long et 

al. 1998) to estimate that fire rotation8 varied from about 

150 to 300 years during the past 3,000 years. They then 

used a spatial landscape simulation model to estimate that 

the mean amount of old-growth (>200 years) and late-successional forests (>80 years) (including old growth) could 

have varied from 39 to 55 percent and 66 to 76 percent, 

respectively, during the 3,000 years prior to Euro-American 

settlement. The model indicated that the minimum and 

maximum amount (i.e., the historical range of variation 

[HRV]) of old-growth and late-successional forest in the 

Coast Range during this period was 24 to 73 percent and 

49 to 91 percent, respectively. The range of variation was 

also a function of the scale of observation, with larger 

ranges for smaller areas, e.g., at the scale of a NWFP 

late-successional reserve (LSR) (~100,000 ac [~40 470

ha]) the range of late-successional forest would have been 

0 to 100 percent. These analyses suggest that older forest 

conditions would have dominated forests of the region, but 

large areas of dynamic early-seral vegetation and younger 

forest would occur episodically as evidenced by the large 

blocks of old-growth forest that would have originated after 

fire. LANDFIRE9

 (https://www.landfire.gov/NationalProductDescriptions24.php) estimated that the amount of “late 

8

 Fire rotation refers to the time required to burn an area equal to 

a defined landscape area (e.g., 1,000 ac [404.7 ha]). The entire area 

may not burn during this period; instead, some sites may burn 

several times and others not at all, but the summed area is equal to 

the defined area. Fire rotation = fire cycle.

9

 LANDFIRE is an interagency geospatial data development 

program that used expert opinion to model historical amounts of 

vegetation stages for potential vegetation types based on published 

literature. The estimate of amounts of vegetation classes do not 

include historical ranges.
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development” closed-canopy forest for the western hemlock 

zone was 70 percent, and the amount of open “early development” vegetation was 5 percent. Estimates of the HRV in 

successional stages are still needed for the NWFP area. 

At the scale of regional landscapes or ecoregions, 

models suggest that early-successional patches occupied 

<20 percent of the area on average but may have reached as 

high as 30 percent over the span of several thousand years 

(Wimberly 2002). At the scale of LSRs, some watersheds 

may have been entirely composed of early-seral conditions 

after wildfires. Studies from Washington and southwest 

British Columbia (Dunwiddie 1986, Hallett et al. 2003) 

indicate that fire-return intervals were much longer in the 

northern part of this regime, so periods when early-successional conditions were abundant in these ecoregions were 

probably less than in the Oregon Coast Range. Moreover, 

the amount of fire and early-successional forest probably 

varied considerably over the past several thousand years in 

resonance with climatic variation. 

The HRV in old-growth and other successional stages 

in the drier part of the western hemlock and other potential 

vegetation zones is less well known. It is also more difficult 

to estimate their abundance with statistical or simulation 

models given that many fires were non-stand replacing 

(Weisberg 2004) and resulted in multiaged patches and a 

large range of stand structures with a wide range of large 

live and dead tree densities, and tree species compositions 

(fig. 3-17). Estimates of historical amounts of old growth 

(i.e., areas of older trees with canopy layering) have been 

made from a few localities in the drier parts of the region. 

In the eastern part of the Oregon Coast Range, Wimberly 

(2002) estimated that the amount of this type of old growth 

over the 1,000 years prior to 1850 would have been less 

than 30 percent, where the fire-return interval was about 

75 years, and many fires were non-stand-replacing (Impara 

1997). The LANDFIRE estimates of these classes of historical amounts of “late” and early-development forest in drier 

parts of the western hemlock zone were 60 percent and 15 

percent, respectively (https://www.landfire.gov/NationalProductDescriptions24.php). The amount of dense old growth 

without a history of non-stand-replacing wildfire, was probably less in these types, however, while the amount of other 

types with old trees would have been more common (Tepley 

et al. 2013) (fig. 3-17). The ecological functions and broader 

ecological significance of this diversity of old-growth forest 

conditions have not been studied, but Tepley et al. (2013) 

suggest this structural and composition diversity of older 

forests may have promoted resilience of large old-growth 

forest structures to disturbances and climate changes. 

Dry forests— 

As fire-return intervals decrease from over 200 years 

in the wetter forests to less than 25 years in the driest 

forests, the role of fire shifts from resetting succession 

and creating large patches of early-seral vegetation to 

regulating forest structure and dynamics altogether, and 

creating fine to mesoscale mosaics of different vegetation 

conditions, including early seral (fig. 3-18). At the shortest 

fire-return intervals, the simple model of succession and 

stand dynamics—i.e., a stand-replacement fire followed 

by long intervals of vegetation change without fire—no 

longer applies. In fact, the entire concept of succession and 

stand development toward multilayered old-forest structure 

in fire-dependent systems becomes problematic where 

fires are very frequent (O’Hara et al. 1996). A pathway of 

stand-replacement disturbance followed succession toward 

multilayered, closed old-growth forests still applies to some 

sites within the frequent, mixed-severity regime dry forests 

(Camp et al. 1997, Merschel et al. 2014), but not so much in 

the very frequent, low-severity regime where fire was more 

of an intrinsic ecological process than an external disturbance event. Forest structural stages (e.g., stem exclusion, 

old-forest multistrata, old-forest single stratum) can still be 

classified and identified in two dry forest fire regimes, but 

the structural conditions can be quite variable and complex, 

and pathways of change can be multidirectional owing to 

the interplay of fire severity, time since last disturbance, 

seed sources, and environmental heterogeneity (Reilly and 

Spies 2015). We discuss the two regimes separately below 

but recognize that for many landscapes and existing forest 

history studies, the two regimes may intermingle or have 

been lumped together. 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Frequent, mixed-severity fire regimes—The potential 

vegetation types of the frequent, mixed-severity regime 

(15- to 50-year return interval) include Douglas-fir, grand 

fir, and white fir, and oak woodlands (fig. 3-4). The cover 

types of this regime include Douglas-fir, white fir, red/

noble fir (Abies procera), and western white pine (Pinus 

monticola). Ponderosa pine can still be a component of 

some of these forests (Merschel et al. 2014). Forests of 

this type were characterized by multiaged cohorts of seral 

dominants and landscape mosaics created by medium to 

large patches of high-severity fire (fig. 3-12), but the landscapes were probably dominated by areas of moderate- to 

low-severity fire. In a Douglas-fir-dominated landscape of 

northern California, Taylor and Skinner (1998) found older 

stands with diverse age structure, but fire-return intervals 

were shorter (e.g., ~15 years), severities were lower, and 

large severe fires were uncommon compared to Douglas-fir 

forests of the western Cascades of central Oregon. Many of 

“mixed-severity” areas of the drier eastern part of northern California have been mapped in our classification into 

the very high frequency, low-severity regime (fig. 3-6). 

Stands with the most diverse age structure in the Taylor 

and Skinner (1998) study experienced the greatest number 

of fires, whereas stands with fewer age cohorts had experienced fewer fires. Those with the most diverse age structure were those most closely exhibiting late-successional 

structure. However, in landscapes where fires were mostly 

low severity, the age-class/fire association was unclear 

(Taylor and Skinner 2003). 

Mixed-severity regimes in dry forests would likely 

result in higher diversity of plant and animal communities and patch (area that differs from its surroundings) 

Figure 3-18—Aerial photo of Beaver Creek Pinery showing spatial heterogeneity that can develop with frequent burning on a productive 

site in the southern Cascade Range of California.

Eric Knapp
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GENERAL TECHNICAL REPORT PNW-GTR-966

heterogeneity compared to high-severity regimes or very 

frequent low-severity regimes (Hessburg et al. 2016, Perry 

et al. 2011). Areas of passive and active tree torching, 

mostly associated with clumps or groups of small understory trees with low limbs, would have created patches 

of tree mortality that would function as canopy gaps of 

various sizes in older forests. Subsequent fires, either 

by torching or girdling, would in turn thin these patches 

diminishing the even-aged group to a few individuals. 

Shade-intolerant tree regeneration would be more likely 

to establish in larger (e.g., >1 ac [0.04 ha]) high-severity 

patches. A prominent hardwood component was often 

associated with conditions emerging after mixed-severity 

fires. These hardwoods may play a pivotal role in continued mixed-severity fires (see discussion below). 

The ecological importance of forests shaped by 

mixed-severity regimes (in both dry and moist forests) is 

widely recognized (DellaSala and Hanson 2015, Hessburg 

et al. 2016, Perry et al. 2011), but fine-scale studies that 

document how microclimate, wildlife, and fire respond 

to different expressions of vegetative heterogeneity, and 

different types of mixed-severity regimes have not been 

conducted. Our understanding of the mixed-severity regime 

in dry forests comes from patch- and landscape-scale reconstructions. That understanding is further complicated by 

lack of consistency in defining mixed-severity fire regimes 

across studies and lack of historical information about their 

spatial and temporal characteristics (app. 3). Several studies 

have characterized the spatial heterogeneity of patches dominated by this regime, especially for the eastern Cascades 

provinces (Hessburg et al. 1999a, 2000b, 2004, 2007; Perry 

et al. 2011). 

The stand-development trajectories of high-severity 

patches could initially follow the pathway described by 

Franklin et al. (2002), but where shrubs or seed source 

limitations occurred, stand development might not proceed through the stem-exclusion closed-canopy stage. In 

addition, some elements of complex older forest structure 

(e.g., large-diameter trees and heterogeneous understories) 

might develop more rapidly than in the wetter forest types 

(Donato et al. 2011), which often have to develop following a relatively uniform and dense self-thinning phase. 

The trajectory of development of a low-density tree patch 

can be altered if the area is severely burned again before 

trees are mature (Coppoletta et al. 2015, Lauvaux et al. 

2016, Tepley et al. 2017). 

Topography would have been an important driver of 

the mosaic pattern. Ridges and south-facing aspects with 

more frequent fire would tend to support more open-canopy 

stands of multicohort shade-intolerant and fire-tolerant 

trees, while valley bottoms, benches, and more northerly 

aspects with less frequent fire would have tended to support 

more complexly structured closed-canopy, multilayered 

stands of shade-tolerant and fire-intolerant trees (Agee 1998, 

Hessburg et al. 2016, Tepley et al. 2013). 

For the eastern Cascades of Washington, Agee (2003) 

used historical fire-return intervals and simple mathematical models to estimate range of variation in forest structure 

classes. This region would contain both the frequent 

mixed-severity and very frequent low-severity regimes 

(fig. 3-4). The proportion of medium to large trees (>15 in 

[40 cm]) in dry to moist forest vegetation types (ponderosa 

pine, Douglas-fir, grand fir warm and cool mesic), regardless of canopy cover, ranged from 38 to 64 percent of the 

landscape. Agee (2003) found that late-successional forest 

(containing shade-tolerant tree species and multilayered 

canopies) was not present in ponderosa pine, warm-dry 

and cool dry Douglas-fir, or warm grand fir forest types, 

and present in about 10 to 16 percent in the “cool-mesic 

grand-fir” type. The amount of early-successional vegetation in these potential vegetation types in this region 

ranged from 6 to 15 percent (Agee 2003, Hessburg et al. 

2000b). Hessburg et al. (2007) used aerial photography 

from the 1930s to 1940s to estimate that old, multistoried 

forests ranged from less than 5 percent to about 20 percent 

or more of dry coniferous forest watersheds, while the area 

of multistoried late-successional forest ranged from 17 to 

68 percent in mixed-severity-regime forests. The estimates 

of forest conditions from this period would have been 

affected by logging, fire exclusion, and fires associated 

with Euro-American settlement around the turn of the 

century (e.g., the widespread fires of 1910), but Hessburg 

et al. (2017) used methods that reduced the impact of these 

anthropogenic effects. 
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Several historical studies have estimated pre-Euro-American settlement amounts of older forest and other 

successional stages for the eastern Cascades of Oregon 

(Andrews and Cowlin 1940 as cited in Davis et al. 2015; 

Baker 2015b; Hagmann et al. 2013, 2014; Kennedy and 

Wimberly 2009). The estimates of the percentage of forests 

of the eastern Oregon Cascades (across all lands in the 

ponderosa pine to moist mixed-conifer potential vegetation types) with large old trees are 35 percent (Kennedy 

and Wimberly 2009); 76 percent (Baker 2015b); 42 to 76 

percent (Hagman et al. 2013); and 91 percent (Hagmann et 

al. 2014). LANDFIRE estimated that “late development” 

(both open and closed-canopy classes) covered 55 to 65 

percent of the dry ponderosa pine and mixed-conifer forest 

environments that occur in the eastern Cascades of Oregon 

and Washington. Using empirical reconstructions from 

early 20th century aerial photos from this area, Hessburg 

et al. (1999a, 2000) showed that more than 40 percent of 

the eastern Oregon Cascades area contained patches with 

medium and large-size old trees in the overstory. They also 

noted that given logging in the ponderosa zone during the 

early 20th century, which they documented via photointerpretation, that amount may have been 50 percent larger, 

i.e., 60 percent of the area with medium- and large-size 

trees in the overstory. The much lower numbers from the 

Kennedy and Wimberly (2009) modeling study may be a 

result of the assumptions about the frequency and severity 

of fire in this region, which is not well-known given the 

lack of fire history studies that were available at that time 

(Baker 2015b). The estimates of historical older forest 

structure among these studies are not strictly comparable 

because of use of different definitions, geographies, potential vegetation types, disturbance regimes, and methods 

and data sources. It is especially difficult to compare 

different studies because of the environmental heterogeneity of the region, including strong precipitation and 

topographic gradients. Also, some moist mixed-conifer 

forests in the eastern Cascades of Washington have high 

fire frequencies (<25 years), which can be similar to that 

of drier ponderosa pine forests (Wright and Agee 2004); 

that relationship would mean that the moist mixed-conifer 

potential vegetation type is not necessarily a good indicator of regimes with longer frequencies or higher fire 

severity. The frequent and very frequent fire regimes are 

spatially intermingled in many landscapes and are difficult 

to separate. 

Most estimates of older forest described above are 

from landscape simulation studies and do not take into 

account canopy cover or forest density, with the exception 

of Hessburg et al. (2007), which is limited to the early and 

mid 20th century. The historical percentage of the eastern 

Cascades in denser older forest (e.g., areas that have not 

had fire for many decades, including areas that could 

potentially support northern spotted owls) has been estimated to be 9 percent (Kennedy and Wimberly 2009) and 

as much as 22 to 39 percent by Baker (2015b). Hagmann 

et al. (2014) estimated that areas of higher density forest 

(>185 trees per acre—“group 1”) and grand fir trees were 

historically rare in dry and moist mixed-conifer forests of 

the northern eastern Oregon Cascades, which would have 

included mixed- and low-severity fire regimes. Perry et al. 

(2004) also found relatively little grand-fir in the central 

Oregon Cascades. 

The fire regimes and forest dynamics of frequent 

mixed-severity regime forests in California have been 

described by Taylor and Halpern (1991), Taylor (1993, 

2000), Taylor and Solem (2001), Bekker and Taylor (2001, 

2010), and Skinner (2003) and summarized by Skinner and 

Taylor (2006). Although no direct estimates of HRV have 

been made, these studies show that fire-return intervals 

tend to be at the low end of the range for this regime. The 

frequent mixed-severity fire regime is characteristic of the 

upper montane forests of red fir/noble fir, western white 

pine, mountain hemlock, and lodgepole pine. These forests 

are typified by precipitation being predominantly snow with 

snowpacks often lasting into early summer contributing 

to a relatively short, yet mostly dry, fire season (Skinner 

and Taylor 2006). Higher productivity (e.g., more fuels) 

and greater sensitivity of the species to fire compared 

to the very frequent, low-severity fire regime may help 

drive occurrence of moderately large patches (hundreds to 

thousands of acres) of high-severity fire despite the high 

frequency of fire. 
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Very frequent, low-severity fire regimes—The very frequent fire (<25-year interval), low-severity regime dry 

forests often occur in association with the forests of the 

infrequent, low-severity regime especially in the eastern 

Cascades and Klamath provinces in areas of topographic 

variability and strong climatic gradients (fig. 3-4). This fire 

regime would have been common in ponderosa pine, dry to 

moist mixed-conifer and oak woodlands vegetation types. 

The successional dynamics, structure, and composition of 

low-severity regime forests can be simplified into two pathways that lead to very different major types of old growth 

(Stine et al. 2014). In the first, a dominant low- or mixed-severity fire-dependent pathway maintained old-growth 

conditions (primarily old live and dead trees) in a shifting 

mosaic of open and moderately closed canopy patches (e.g., 

20 to 60 percent canopy cover) (figs. 3-18 and 3-19). 

A second, historically much less common pathway 

occurred where local climate and topoedaphic circumstances (e.g., rocky ridges) reduced wildfire frequency and 

led to development of patches of denser (60 to 90 percent 

canopy cover), multistory old-growth with shade-tolerant 

species (Agee 1993; Camp et al. 1997; Hessburg et al. 1999a, 

1999b, 2000, 2007; Merschel et al. 2014; Sensenig et al. 

2013). Levels of large standing and dead down wood would 

be much lower than in old-growth forest types in the other 

fire regimes (see Youngblood et al. 2004 for density estimates), owing to lower densities of large trees and frequent 

consumption of down wood (Safford and Stevens 2016, 

Skinner 2002). Despite the lower densities relative to denser 

old growth, large standing dead trees would have been 

present throughout though they would have been patchy 

and not found on every acre (Stephens and Fulé 2005). The 

pattern of seral stages within the forest matrix would be 

a fine-meso-scale mosaic of patches (<1 ac [<0.40 ha] to 

thousands of acres). The dominant pathway was maintained 

by high- to moderate-frequency, low- to mixed-severity 

fire (Baker 2012, Hessburg et al. 2007); scattered small- to 

medium-size patches with canopy tree mortality (individuals 

or small- to medium-size clumps) would have been present 

with medium and large fire-tolerant trees occurring in low to 

locally moderate densities (Churchill et al. 2013, Larson and 

Churchill 2012). For old-growth ponderosa pine in Oregon 

and California, canopy trees were not uniformly distributed 

and tended to occur in either clumps of up to 80 ft (24 m) 

in diameter (Youngblood et al. 2004) (figs. 3-17 and 3-18). 

These forests are sometimes characterized as being open, 

low-density forests, “park-like” stands (Agee 1993, Hessburg 

et al. 2015, Sensenig et al. 2013, Youngblood et al. 2004) (fig. 

3-1). Bark beetles, which attack trees in small groups, may 

have interacted with fire in these forests to promote patchy 

regeneration of ponderosa pine. This would occur where 

beetle-killed patches of dead trees had accumulations of 

small branches and coarse woody debris that burned with 

high severity, killing rhizomatous grasses and promoting 

patchy regeneration of ponderosa pine regeneration in ash of 

the burned logs and sterilized mineral soil (Agee 1993). 

The second successional pathway would lead to denser 

patches of pine and Douglas-fir or true fir regeneration, as 

mentioned above, often associated with variation in topography (steeper slopes and higher elevation), microclimate, 

and fire frequency that allowed trees to develop on moister 

microsites associated with north-facing lower slopes, 

concave areas, riparian areas, and wetter soils (Camp et al. 

1997, Merschel et al. 2014). However, Baker (2012) did not 

Figure 3-19—Hypothetical structural profile and typical historical 

fire behavior in a ponderosa pine forest of the eastern Cascade 

Range of Washington. From Van Pelt (2008).
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

find that concentrations of fir were associated with aspect 

or topography in an analysis of General Land Office (GLO) 

survey data from the eastern Oregon Cascades. Following 

low- to moderate-severity fire on these more moist sites, 

white fir or grand fir could establish in the understory and 

occasionally reach the canopy where bole diameters and 

bark thickness was sufficient to withstand surface fires. On 

some productive sites (e.g., benches), old-growth grand-fir 

or white-fir patches developed even while experiencing 

frequent surface fires that burned in from adjacent drier 

ponderosa pine and grassland sites (Hessburg et al. 1999, 

Taylor and Skinner 2003). The relative amount of open 

and denser older forests may have varied over time with 

climate. Many studies across the area support this characterization of forest structure and dynamics for this type in 

some portions of the region (Bisson et al. 2003; Hann et 

al. 1997; Hessburg et al. 1999a, 1999b, 1999c, 2000, 2003, 

2005; Keane et al. 2002, 2009; Lehmkuhl et al. 1994). 

With fire exclusion, the dense late-successional and oldgrowth pathway (either with ponderosa pine, Douglas-fir, 

or Abies spp.) has become dominant (fig. 3-20). White fir 

and grand fir have widely expanded out of their historical 

environments and fire refugia into sites that were historically dominated by ponderosa pine (or sugar pine (Pinus 

lambertiana) in California) or pine mixes with Douglas-fir 

(Camp et al. 1997; Hagmann et al. 2017; Merschel et al. 

2014; Taylor and Skinner 1998, 2003), or grassy woodlands 

often originally dominated by hardwoods (Skinner et al., 

in press). This expansion of shade-tolerant trees (which is 

discussed more below) has been widespread across a range 

of topographic settings and forest types, including drier 

mixed-conifer and ponderosa pine types (Hagmann et al. 

2014; Hessburg et al. 1999a, 1999b, 2000a, 2003, 2005, 

2015, 2016; Merschel et al. 2014; Stine et al. 2014).

Figure 3-20—Old-growth ponderosa pine in the eastern Cascade Range of Oregon with understory of grand fir that established in the 

early 1900s after fire exclusion.

Thomas Spies
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Woodlands, shrublands, and grasslands—A significant 

portion of some of the dry forest landscapes was occupied 

by patches of semistable, woodlands, shrublands, and 

grasslands (Hessburg et al. 2007) (figs. 3-21 and 3-22). 

These included oak, juniper, and pine woodlands that did 

not succeed to denser forest as a result of climate, soils, and 

frequent fire (Agee 1993, Franklin and Dyrness 1973, 

Hessburg and Agee 2003, Skinner et al. 2006). In many 

cases, a frequent grass- or shrub-driven fire cycle was 

responsible for maintaining low tree cover (Hessburg et al. 

2016). These areas were so dominated by grasses over a 

geologically long timeframe that mollisols can be seen 

today as the characteristic soil type. Open stands and oak 

dominance were maintained by American Indians in many 

areas using fire to promote desired resources associated 

with such habitats (Anderson 2005, Skinner et al. 2006) 

(chapter 11). Figures 3-21 and 3-22 illustrate these landscapes, and although large fires in the early 1900s would 

have affected these patterns, many of the large fires would 

have occurred in grasslands and shrublands (that were 

historically maintained by frequent fire) as evidenced by the 

lack of snags and dead trees in the large nonforest patches 

in these photos. Interestingly, the concept of old growth (in 

a general sense of a vegetation type that persisted for very 

Figure 3-21—Photographs of the Mission Peak area on the Okanogan-Wenatchee National Forest in 1934 and 2010. The 1934 image 

illustrates the mosaic of closed forests, open forests, woodlands, and grasslands that would have characterized many landscapes with 

low- and mixed-severity fire regimes. Open areas typically lack snags that would be indicative of recent high-severity fire in forests. 

Landscapes in 1934 may have been influenced by settlement fires, logging, and fire exclusion.

John F. Marshall Reino R. Sarlin
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

long periods under natural processes) has also recently been 

applied to these nonforest vegetation types (Veldman et al. 

2015) because they have distinct conservation values that 

arise as a result of being “ancient”10 ecosystems with 

characteristic biotic and soil properties that have been lost 

owing to changes in fire regimes, grazing, and other land 

use changes. 

10 Grasslands have existed for millions of years, and some 

grasslands may take 100 to as much as 1,000 years to develop; and 

clonal grasses can live for over 500 years. 

Oak woodlands dominated by California black oak 

(Quercus kelloggi) and Oregon white oak (Q. garryana) 

and other hardwoods were maintained in an open oldgrowth state by very frequent low-severity fire (Agee 

1993, Cocking et al. 2012, Franklin and Dyrness 1973). 

These species can form large, old trees with high value 

because they produce mast or berries, as well as large 

cavities for wildlife. They often support a high diversity 

of understory plants, fungi, and associated wildlife of 

particular importance to tribes (see chapter 11). However, 

a lack of fire in many of these areas has permitted conifer 

Figure 3-22—View from Eddy Gulch Lookout in the Salmon River watershed of the Klamath National Forest in 1935 (top) and 1992. The 

1935 image illustrates the mosaic of closed forests, open forests, shrub fields, woodlands, and grasslands that would have characterized 

many landscapes with low- and mixed-severity fire regimes. Open areas typically lack snags that would be indicative of high-severity 

fire in forests. Landscapes in 1935 may have been influenced by settlement fires, logging, and fire exclusion.

Klamath National Forest
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GENERAL TECHNICAL REPORT PNW-GTR-966

trees such as Douglas-fir to increase shade, accumulate 

conifer litter, and form ladder fuels, which consequently, 

render mature hardwoods more vulnerable to top-kill from 

fires. These trends are particularly evident in riparian 

forests of southwestern Oregon, where the shift in fire 

regime has led to reductions in both hardwoods and large 

trees (Messier et al. 2012). 

Role of shrubs and hardwoods in Klamath-Siskiyou 

forest dynamics—The successional dynamics of low- and 

mixed-severity regime forests in the Klamath-Siskiyou 

region of Oregon and California are distinctive for the 

prominent role of shrubs and hardwoods in the vegetation community and their interaction with both fire and 

forest succession. In the northern and western part of this 

region, mixed-severity fire can lead to patchy old growth 

with tanoak (Notholithocarpus densiflorus) understories 

(as small trees) intermixed with Douglas-fir that either 

survives the lower intensity fire as a large tree or regenerates in patches of high-severity fire that kill the tanoak 

(Agee 1993). In other areas of this region, and extending 

into the southern Cascades and northern Sierra Nevada, 

dense stands of the shrub form of tanoak (N. densiflorus var. echinoides) can be found. These stands often do 

not burn well under less-than-severe conditions but will 

strongly sprout following severe fires even though the 

acorns are killed by fire. 

Throughout the Klamath-Siskiyou region, shrub 

species resprout after fire and are also stimulated to 

germinate from seeds stored for long periods in soil seed 

banks following fires (Knapp et al. 2012b) with areas of 

higher severity fire leading to greater density of shrubs 

(Crotteau et al. 2013). Hardwoods (especially oaks, tanoak, 

and madrone (Arbutus menziesii) mixed in with the often 

more dominant conifers are often able to resprout following high-severity fires that kill the conifers (Cocking et al. 

2012, 2014; Skinner et al. 2006). This adaptation facilitates the reestablishment of trees in severely burned forest 

areas at an early-seral stage. For conifer forests to again 

occupy these areas requires sufficient time between severe 

burns to allow conifer trees to reestablish and mature. 

Where severely burned areas are reburned before such 

conditions are achieved, shrubfields and hardwoods are 

likely to be maintained and can become a more permanent 

part of the landscape (Cocking et al. 2014, Coppoletta et 

al. 2015, Lauvaux et al. 2016). Several recent studies have 

documented how severely burned areas that are reburned 

within a few decades are likely to again burn severely 

(Coppoletta et al. 2015; Odion et al. 2004; Perry et al. 

2011; Thompson and Spies 2010; Thompson et al. 2007, 

2011). In other cases, hardwoods in mixed-wood forests 

may play an important role in protecting some of the 

coniferous forest cover from severe fire effects via their 

foliar moisture content (Agee 2002, Perry 1988, Perry 

et al. 2011, Raymond and Peterson 2005, Skinner 2006, 

Skinner and Chang 1996). Likewise, depending upon the 

forest community type, hardwood trees and shrubs may 

in fact facilitate conifer succession via mycorrhizal fungi 

shared by both hardwood and coniferous species (Horton 

et al. 1999).

In complex topography, such as that found in the 

Klamath-Siskiyou area, it is unlikely that disturbance 

regimes and seral stages randomly moved about the 

landscape. Rather, particular parts of the landscape were 

more prone to severe burns. Upper thirds of slopes, and 

especially south- and west-facing slopes, were prone to 

repeated severe burning that perpetuated shrub dominance (Jimerson and Jones 2003, Taylor and Skinner 

1998, Weatherspoon and Skinner 1995). Shrubfields may 

be places where forests burned severely or places where 

fires have long maintained shrubfields (Baker 2012, 2014; 

Lauvaux et al. 2016; Nagel and Taylor 2005). In the latter 

case, these were not places that periodically contributed 

large wood and snags but reburns of shrubs, grasses, and 

occasional small conifers.

Alternative views of disturbance regimes of the dry 

forests—Some have argued that most ponderosa pine 

and mixed-conifer forests in the Western United States, 

including the area of the NWFP that we define as having 

had a very frequent, low-severity regime, have been mischaracterized. They contend that these forests are better 

characterized instead as having a more variable-severity fire regime, with significant components of mixed and 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

high-severity fire as well (Baker 2012, Odion et al. 2014, 

Williams and Baker 2012). Hessburg et al. (2007) has 

also been cited in support of this argument (Baker 2012); 

however, the results of Hessburg et al. (2007) do not fully 

support the claims of Baker (2012); there are some key 

differences. The classification of high-severity fire from 

aerial photos in Hessburg et al. (2007) included areas with 

small trees, grasslands, shrublands, and sparse woodlands. 

These nonforest areas would have typically burned with 

high-severity given the low stature of their vegetation 

driven by a predominantly grass-fire cycle. When Hessburg 

et al. (2007) restricted their analysis to forest cover types, 

they found that less than 20 percent of each cover type was 

consistently affected by high-severity fires (fig. 3-23). For 

example, the dominating ponderosa pine and Douglas-fir 

cover types exhibited 13 and 18 percent high-severity fires 

across the study area, respectively. Similarly, when they 

restricted their analyses to forest structural classes (fig. 

3-24), they found that no structural class experienced more 

than 17 percent high-severity fire across the study area. 

Furthermore, Baker (2012) uses Hessburg et al. 2007 to 

support his claim that “substantial” areas of high-severity fire occurred in ponderosa pine and dry mixed-conifer, 

but he cites Hessburg et al. (2007) data from Ecological 

Subregion 5 (ESR5), which is not a dry forest environment, but is classified as “moist and cold forest” type, with 

lesser amounts of dry forests. Hessburg et al. (2007) found 

considerable evidence of high-severity fire in their regional 

analysis of dry pine and mixed-conifer forest landscapes, 

but much of the high-severity fire was associated with 

grasslands and shrublands that where common in these 

landscapes in the past and were intermingled with forested 

patches. These vegetation types would typically burn with 

high severity. Figure 3-23 shows the proportion of forest 

structural classes affected by low-, mixed-, and high-severity fire in three ecoregions. 

Figure 3-23—The proportions of premanagement-era total forest area (hectares) by forest cover type in low-, mixed-, and high-severity 

fire (corresponding with percentage of canopy mortality values of ≤20 percent, 20.1 to 69.9 percent, and ≥70 percent, respectively) of 

Ecological Subregions (ESRs) 5, 11, and 13. Cover type abbreviations are TSHE/THPL = western hemlock/western redcedar; PIMO 

= western white pine; POTR/POTR2 = Populus and Salix spp.; LAOC = western larch; TSME = mountain hemlock; PIAL/LALY = 

whitebark pine/subalpine larch; ABAM = Pacific silver fir; ABGR = grand fir; PICO = lodgepole pine; ABLA2/PIEN = subalpine fir/

Engelmann spruce; PSME = Douglas-fir; PIPO = ponderosa pine. From Hessburg et al. (2007).
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GENERAL TECHNICAL REPORT PNW-GTR-966

Williams and Baker (2012) and Baker (2012) use GLO 

survey data from the 1880s and 1890s on live tree sizes and 

species to infer historical stand densities and fire regimes 

from central Oregon. The evidence and methods used to 

support the claims that the historical role of high-severity 

fire in low-severity regimes has been underestimated has 

been the subject of several published critiques and counter 

arguments by both sides of the debate. In one critique, 

Fulé et al. (2013) point out three problems with using GLO 

survey data to infer disturbance history (e.g., Baker 2012): 

(1) the use of tree size distributions to reconstruct past fire 

severity and extent is not supported by empirical age-size 

relationships nor by local disturbance history studies; (2) 

the fire-severity classification based on the survey data is 

qualitatively and quantitatively different from most modern 

classification schemes, limiting the validity of comparisons 

to history; (3) their finding of ”surprising” heterogeneity 

within these stands does not actually differ substantially 

from other previous studies (some from ponderosa pine 

forests outside the NWFP area but still potentially relevant 

to dry forests in the NWFP area) that found areas and 

clumps of relatively high density in ponderosa pine and 

mixed-conifer forests (e.g., Brown and Cook 2006, Youngblood et al. 2004) (fig. 3-25). For example, the lower left 

corner (66 by 66 ft [20 by 20 m]) of the old-growth plot that 

Youngblood analyzed had 16 trees (equivalent to a density 

of upper canopy trees of about 160 trees per acre), while the 

upper right corner had one tree (an acre-scale density of 10 

trees per acre). 

Williams and Baker (2014) responded to that critique 

of Fulé et al. (2013) by arguing that the concerns are 

unfounded and based on misquoting their 2012 paper. 

Figure 3-24—The proportions of the premanagement-era dry forest area (hectares) by forest structural class in low-, mixed-, and highseverity fire (corresponding with percentage of canopy mortality values of ≤20 percent, 20.1 to 69.9 percent, and ≥70 percent, respectively) of Ecological Subregions (ESRs) 5, 11, and 13. Structural class abbreviations are: SI = stand initiation, SEOC = open canopy stem 

exclusion, SECC = closed-canopy stem exclusion, UR = understory reinitiation, YFMS = young multistory forest, OFMS = old multistory 

forest, OFSS = old single-story forest. New, intermediate, and old designations are used to group structural classes into broad age groups. 

From Hessburg et al. (2007).
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Williams and Baker (2012) used tree density and relative 

proportions of small and large trees to classify GLO data 

areas as either low- or high-severity fire. According to 

Baker (2012), 26 percent of pine and dry mixed-conifer 

forests in the eastern Oregon Cascades showed evidence of 

high-severity fire based in part on tree density. The findings 

of Baker (2012) depend on many assumptions, the most 

important being that the method for calculating tree density 

from GLO survey data (Williams and Baker 2011) produces 

an unbiased estimate. However, a recent paper by Levine 

et al. (2017) indicates that the method (Williams and Baker 

2011) used by Baker (2012) overestimates tree density by a 

factor of 1.2 to 3.8. This finding could help explain why the 

estimates of historical tree densities that Baker has reported 

(mean of 100 trees per acre) are considerably higher than 

those reported from other studies, e.g., 62 trees per acre 

(Munger 1917) or 26 to 32 trees per acre (Hagmann et 

al. 2013, 2014). Other assumptions made by Baker (2012) 

could explain the higher densities relative to other studies 

including the assumption that his survey points represent 

dry environments and not wetter mixed-conifer sites that 

often occur in the eastern Cascades where topographic 

and precipitation gradients are strong, and produce high 

variability in forest structure, composition, and dynamics 

(Merschel et al. 2014). 

Odion et al. (2014) have also argued for the occurrence 

of more high-severity fire in ponderosa pine and mixedconifer forests of western North America using inferences 

from analysis of current tree-age data from unmanaged 

areas collected through the U.S. Forest Service Inventory 

and Analysis (FIA) program (Odion et al. 2014). Age data 

were analyzed and it was assumed that if stand-age diversity was low, then fire effects represented low- or mixed-severity regimes; if stand-age diversity was high, then the 

Figure 3-25—Spatial patterns of live (filled circle) and dead trees (open circle) in the 

upper canopy of an old-growth ponderosa pine forest in central Oregon. 

From Youngblood et al. (2004). 
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GENERAL TECHNICAL REPORT PNW-GTR-966

forest came from a mixed-severity regime with significant 

areas of high severity. However, a critique by FIA and 

other scientists argues that the assumptions, analysis and 

conclusions of this paper are invalid (Stevens et al. 2016). 

First, the FIA stand-age estimator underestimates the age 

range of trees in plots, and it routinely undersamples old 

trees, which would be relatively common in forests subject 

to low-severity fire regimes (see Merschel et al. 2014). 

Forests with a low-severity fire regime also continuously 

recruit new cohorts of regeneration, which would be poorly 

reflected in the stand-age estimator. Second, recruitment 

events are not necessarily related to high-severity fire 

occurrence as we have described above. Odion et al. (2016) 

responded to Stevens et al. (2016) and identified areas of 

“agreement and disagreement.” Areas of agreement include 

high-severity fire was a component of forests in low-severity fire regimes, that tree recruitment occurs in the 

absence of fire, and FIA stand data may provide evidence 

of past high-severity fire. Areas of continued disagreement according to Odion et al. (2016) include deciding 

what threshold to use for mortality from high-severity 

fire, plot sizes needed to detect high-severity fire, use of 

diameter-age relationships for reconstructing basal area, 

and historical data sources that document high-severity 

fire in patches larger than 2,500 ac (~1000 ha). We disagree 

that their historical sources present many examples that 

document the occurrence of large patches of high-severity 

fire in forests with low-severity regimes. Historical maps 

we found from the early 1900s document three patches of 

high-severity fire larger than 2,500 ac (~1000 ha) in Oregon 

and Washington that account for 1 percent of the area of 

this regime (fig. 3-6). In addition, the so called large patch 

of high-severity fire in the “eastern Cascades” of Oregon 

that is cited in Dellasala and Hanson (2015: 30–31) from 

mapping of Leiberg (1903) as evidence of a 35,000-ac (~14 

200-ha) patch of high-severity fire in ponderosa pine forests 

actually comes from a township in the western Cascades in 

an area of mixed-conifer forest, containing red fir and noble 

fir. This township and the boundaries of this fire straddle 

the infrequent high-severity regime and moderately 

frequent to somewhat infrequent mixed-severity regimes 

of our regime map (fig. 3-6). 

These concerns about interpretation of forest history data 

notwithstanding, there is essentially no disagreement that 

very frequent, low-severity regime forests (e.g., ponderosa 

pine and mixed conifer) included occasional small- to medium-size (e.g., tens to hundreds of acres) patches of high-severity fire. In addition, the broader landscapes would have 

contained grasslands or shrublands maintained by high-severity fire (relative to that life form) (e.g., see Hessburg et al. 

2007, Perry et al. 2011). Given that many larger landscapes 

(including forested areas and nonforest areas) are often a 

mosaic of environments that support both low- and highseverity fires, it would not be surprising to find landscapes 

where the amount of high-severity fire to forest and nonforest 

vegetation exceeded 20 percent (e.g., see historical landscape 

data in Hessburg et al. 1999a, 2000, 2007). However, over 

smaller areas or areas with less topographic variability and 

within environments that predominantly supported forests, 

the amount of high-severity fire in low-severity regime forests 

would be expected to be lower than 20 percent. For example, 

Hagmann et al. (2014) found that only 9 percent of forest 

survey transects in 123,500 ac (~50 000 ha) of mixed-conifer 

landscape in eastern Oregon showed potential evidence of 

high-severity fire based on absence of large trees. 

In summary, we believe the preponderance of evidence 

supports the view that large patches of high-severity fire 

were not a major component of dry forests with very high 

frequency, low-severity forest fire regimes. However, they 

were an important component of the frequent, mixedseverity regime. Remember that these regimes exist along 

a continuum of environments that differ across regions 

and landscapes. This means that landscapes often do not fit 

neatly into one regime or another. These alternative views 

of the role of high-severity fire in low-severity fire regimes 

highlights that generalizations either for or against management interventions across a wide range of forest types 

and environments should be made with caution. Different 

definitions of severity, scales of observation, and types of 

evidence (e.g., maps, surveys, aerial photos, tree age and size 

distributions, etc.) make it difficult to compare across studies 

because these factors influence the scope of inferences that 

can be made. In addition, subregional and landscape-scale 

variation in ecosystems and interactions among climate, 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

topography, soils, vegetation, and disturbance agents make it 

difficult to accurately extrapolate over to large areas. Efforts 

to infer process (e.g., disturbance history) from pattern (e.g., 

ages, sizes, or densities of trees, and patches of trees in maps 

and aerial photos), as is done in many of the fire history studies we cite, can also be fraught with some degree of uncertainty because similar patterns in biotic communities can 

arise from different processes (Cale et al. 1989). For example, 

much of the open forest reported by Baker could have been 

made up of aggrading meadows and shrublands that were 

much more common during the early 20th century (Hessburg 

et al. 2005, 2007). A lack of information on the presence of 

snags and dead wood limits any inference on fire severity 

in forests from studies based only on live trees (Reilly and 

Spies 2015, Reilly et al. 2017). Uncertainties about fire history are unlikely to be resolved given the limits of historical 

information (especially prior to Euro-American colonization) 

and the heterogeneity of ecosystems. In the end, the details 

of historical regimes (e.g., the level of high-severity fire in 

the past) may not be as important as what society wants and 

can have for their forests given changing climate, succession, 

and fire behavior (see chapter 12). 

Effects of Fire Exclusion

Forest structure and composition— 

Dry forests—There is less debate in the literature about 

the effects of fire exclusion on forest structure and composition in dry forests where fire was historically frequent. 

Nationally, over 95 to 98 percent of all wildfires are suppressed while small during initial attack (i.e., 2 to 5 percent 

escape initial attack) with suppression in the NWFP area 

especially common in dry forests (fig. 3-10, table 3-3). Many 

of these fire starts would have resulted in larger fires that 

would have altered forest structure and fuel beds and created or maintained early- and mid-successional vegetation 

over much of the region in the ensuing century. 

The recent trends in fire extent and severity in the 

NWFP area (chapter 2) suggest that fire has generally been 

less common in recent decades than would be expected 

under the historical fire regimes (Reilly et al. 2007) (table 

3-4), especially given the occurrence of the warmest decade 

(~1995–2005) since the early 1900s (Abatzoglou et al. 2014) 

and the historical link between fire and temperature, and 

drought. The amount of fire (fire rotation) in the frequent 

and very frequent regimes (117 to 182 years for federal 

lands) has been considerably less than the historical range 

for these two dry forest regime classes (5 to 50 years) (table 

3-4). For example, in the very frequent regime, most areas 

would have burned at least once (e.g., a fire rotation of less 

than 25 years), if not more, during 30 years, the length of 

the recent satellite record. 

Forests have responded to the lack of fire in the two dry 

forest fire regimes through increases in density and changes 

in composition. It is well documented that the structure and 

composition of these forests have changed across the 

Western United States since Euro-American settlement 

(Hann et al. 1997; Hessburg and Agee 2003; Hessburg et al. 

2005, 1999a, 1999b, 1999c, 2000; Lehmkuhl et al. 1994) as 

a result of fire exclusion. For example, forests are now 

typically several times denser in most locations than under 

native fire regimes (Camp 1999; Dolph et al. 1995; Hagmann et al. 2013, 2014; Merschel et al. 2014; Perry et al. 

Table 3-3—Number of lightning fire startsa between 1992 and 2013 in summer months (June–

September) on federal forest lands in the Northwest Forest Plan areab

Regime Total fire starts Number per 25,000 ac (10 117 ha)

Infrequent, high severity 4,271 12.2

Moderately frequent, mixed severity 2,350 13.4

Frequent, mixed severity 2,511 15.2

Very frequent, low severity 4,240 17.4

a

 Most of these would have been suppressed by fire crews. 

b Sources of data: Bureau of Land Management Wildland Fire Management Information system; U.S. Fish and Wildlife Wildland 

Fire Information System; U.S. Forest Service fire statistics.
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GENERAL TECHNICAL REPORT PNW-GTR-966

2004; Reilly and Spies 2015; Ritchie et al. 2008; Stephens et 

al. 2015; Youngblood et al. 2004), and composition has 

shifted toward shade-tolerant species. Baker (2012) did not 

agree with this characterization and described these forests 

of the late 1800s as historically “generally dense.” However, 

the finding that his method overestimates tree density by 20 

to 380 percent (Levine et al. 2017) suggests that forests 

were not generally dense as he claims, and data may be 

coming from a period in which shifts from a more frequent 

fire regime had already occurred as a result of various 

effects of Euro-American colonization (Fry and Stephens 

2006, Norman and Taylor 2005, Skinner et al. 2009), 

including the loss of burning11 by American Indians. Even 

if the overestimates of the Baker (2012) method are at the 

low end of the range of bias found by Levine et al. (2017), 

they are still lower than the least dense areas found in 

contemporary forests (Merschel et al. 2014, Reilly and Spies 

11 Note that American Indians were marshalled onto reservations 

by 1850, and with this came the loss of intentional burning that 

occurred near seasonal encampments and customary food production and gathering places (Stewart 2002).

Table 3-4—Comparison of historical fire frequencies and rotations (in years) with recent (1985–2010) fire 

rotation estimates from satellite remote sensing for the Northwest Forest Plan area by fire regime classa

Historical regime 

class and fire 

frequencies in 

years

Range of 

frequencies from 

historical studies, 

all fires (number 

of studies) 

Range of 

estimates of 

historical 

rotations, all 

fires (number of 

studies) 

Recent rotations 

(all severities) for 

USFS lands/all 

ownerships

Recent rotation 

(high severity) 

for USFS lands/

all ownerships

Recent frequency 

(low severity) for 

USFS lands/all 

ownerships

Infrequent, high 

severity (200 to 

1,000 years) 

No data 296–834 (5) 758/1,525 1,628/3,326 3,056/6,069

Moderately 

frequent, mixed 

severity (50 to 200 

years)

40–246 (19) 78–271 (6) 582/1,055 2,398/4,530 1,321/2,342

Frequent, mixed 

severity 

(15 to 50 years)

21–27 (2) No data 110/276 333/851 305/761

Very frequent, low 

severity (5 to 25 

years) 

3– 36 (18) 11–64 (4) 111/143 690/852 218/286

a See appendix 3 for fire history data. Recent data from Reilly et al. (2017). USFS = U.S. Forest Service.

2015). Baker (2012) estimated that the interquantile range 

(25th to 75th) for density in mixed conifer was 69 to 142 

trees per ac (170 to 352 trees per ha), whereas the interquantile range in current forests was 298 to 586 trees per acre 

(736 to 1,447 trees per hectare) an increase of 67 to 75 

percent. Consequently, the 2012 Baker paper cannot be used 

as evidence that forest density has not substantially 

increased since the 1900s—only that the increase may not 

be as large as some studies indicate. 

A consequence of succession in these forests is that 

dense understories of shade-tolerant species can shade out 

pine regeneration and eventually provide abundant seed 

sources that compete with pine regeneration in lower fire 

severity postfire environments. Restoring the dominance 

of large fire-tolerant tree species in these forests is a key 

component of restoration strategies (Hessburg et al. 2016). 

The accumulated seed source of shade-tolerant species in 

these landscapes and large-landscape inertia has probably 

altered the successional probabilities following fire disturbances toward shade-tolerant pathways as Stine et al. (2014: 

140) indicates:
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Landscapes exhibit varying degrees of inertia. 

The degree of change over the 20th century in 

forest structure, tree species composition, and 

disturbance regimes has given landscapes an 

inertia (which can be thought of also as ecological 

momentum or resistance to change) that will be 

difficult to alter through restoration-based management. For example, field observations suggest that 

after recent wildfires, instead of regenerating to 

ponderosa pine or western larch, some areas now 

quickly regenerate to Douglas-fir and white, grand, 

or subalpine fir, or lodgepole pine, despite intentional efforts (which often fail unless done well) to 

reestablish ponderosa pine or larch. The presence 

of abundant seed from shade-tolerant tree species 

(e.g., firs) provides this inertia. Likewise, high 

contagion of surface and canopy fuels creates large 

homogeneous patches that reinforce the occurrence 

of a higher than normal number of large and very 

large fires, and higher than normal fire severity.

This landscape-scale successional trend may be locally 

disrupted by large disturbances, but if the rate of disturbance is not high enough, or the disturbance does not kill 

the shade-tolerant species over large areas, the trend is likely 

to continue unless climatic changes alter the disturbance 

regime and the growth or survivorship of tree species. 

Moist forests—Fire suppression also appears to be having an 

effect on the amount of fire in the moist, west-side forest fire 

regimes (Agee 1993) (figs. 3-4 and 3-10). Over 6,600 lightning-started fires were recorded in this region over a recent 

21-year period, and most of these would have been actively 

suppressed (table 3-3). Although the vast majority of these 

fires probably would not have turned into large high-severity 

or mixed-severity fires, a few probably would have. Before the 

era of fire suppression, a few of these starts likely smoldered 

for weeks as small fires or as burning snags until a dry east 

wind event occurred, when those fires could spread rapidly 

producing large patches of high-severity fire along with 

patches of moderate- to low-severity fire. Recent fire rotations 

for high-severity fires in the two west-side fire regimes also 

appear to be at the high end of the historical range for U.S. 

Forest Service lands (table 3-4) (Reilly et al. 2017). Historical 

fire occurrence in these regimes varied at centennial scales 

with climate and human population density (e.g., Weisberg 

and Swanson 2003). Thus, given the occurrence of warm, dry 

conditions during much of the contemporary fire period, a 

rotation exceeding the upper end of the range suggests we are 

currently experiencing much less fire than would have 

occurred historically under a similar climate.12

The effects of fire suppression in the moist, west-side 

forests are quite different than in the dry forests. Fire 

suppression in relatively productive forests with longfire-return intervals has little effect on fuel accumulation 

at the stand level (Agee 1993). However, fire suppression 

would drastically reduce the amount of early- and mid-successional vegetation in the landscape and thereby, reduce 

landscape-scale heterogeneity in forest composition, 

structure, and patch sizes. Mixed-severity fires burning at 

rotations of 50 to 200 years would have created a mosaic of 

forest successional stages, including multicohort old-growth 

stands (figs. 14, 16, and 17) (Tepley et al. 2013). 

Fire severity in dry forests—

Although weather is the primary controller of fire occurrence, size, and severity, in some cases, in the NWFP area 

(Littell et al. 2009, Reilly et al. 2017), local controls (e.g., 

topography and fuels) are also important (Cansler and 

McKenzie 2014). There is significant concern that accumulation of live and dead fuels in understories as a result of 

fire exclusion and suppression has increased the threat and 

occurrence of larger areas of high-severity fire (Hessburg 

et al. 2000, 2005; Miller and Urban 1999a, 1999b, 2000; 

Parsons 1978, Parsons and DeBenedetti 1979). This threat is 

thought to arise from two processes: (1) increased accumulations of surface and ladder fuels (shrubs, small trees, 

lower canopy base heights) that increase flame length and 

fireline intensity under extreme fire weather conditions, and 

risk of mortality, even in large fire-resistant canopy trees; 

and (2) higher spatial continuity of fuel beds that can lead to 

12 Note, however, that for the infrequent and moderately frequent 

regimes, the recent 25-year record is very short and does not 

necessarily indicate deviation from historical regimes where fires 

were relatively infrequent (e.g., 505 to 1,000 years). Note also the 

relatively small sample sizes of fire history studies.
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GENERAL TECHNICAL REPORT PNW-GTR-966

more rapidly spreading and larger patches of high-severity 

fire (fig. 3-26). These changes in fire behavior as a result of 

fuel accumulation are supported by theory, simulation models of fire behavior, and empirical studies of differences in 

fire behavior between stands where fuels have been reduced 

by mechanical and prescribed fire and those that have not 

been treated (North et al. 2012, Ritchie et al. 2007, Safford 

et al. 2012b, Schmidt et al. 2008, Stephens 1998, Stephens 

and Moghaddas 2005, Stephens et al. 2009, Weatherspoon 

and Skinner 1995). Evaluation of changes in fire patch size 

distributions with those of pre-Euro-American settlement 

era fire regimes are problematic because we lack landscape-scale quantitative data on frequency-size distributions 

of fire-severity patches for most areas (Collins et al. 2006; 

Collins and Stephens 2010; cf. Perry et al. 2011; Reilly et al. 

2007; Williams and Baker 2014) (app. 3). 

Empirical evidence for increasing total area of fire, 

and increasing area of fire patch sizes in recent decades, 

exists from studies across the Western United States, which 

are relevant to the NWFP area (Cansler and McKenzie 

2014, Littell et al. 2010, Miller et al. 2008, Odion et al. 

2004, Reilly et al. 2017, Westerling et al. 2006). However, 

evidence for increased proportion of high-severity fire in 

recent decades is mixed. Lutz et al. (2009) found evidence 

for increasing proportion of high-severity fire in the Sierra 

and southern Cascades of California, but Miller et al. (2012) 

did not find evidence of increasing total proportion of high 

fire severity from northwest California between 1987 and 

2008. Miller et al. (2012) did find the sizes of high-severity 

patches to be increasing along with the overall increasing size of fires. Baker (2015a) did not find evidence for 

increasing proportion of high-severity fire in recent years 

in a study of ponderosa pine and mixed-conifer forests of 

the Western United States. Reilly et al. (2017) found no 

increases in the proportion of area burned at any level of 

severity between 1985 and 2010 in the Pacific Northwest but 

did see increasingly severe fire effects (e.g., large patches 

of high-severity fire) related to drought and annual area 

burned. Cansler and McKenzie (2014) found significant 

positive relationships in the northern Washington Cascades 

between climate and fire size, and between fire size and the 

proportion of fire events found in high-severity fire patches. 

They also found that the spatial aggregation of high-severity area within fires was greater in ecoregions with more 

contiguous subalpine forests and less complex topography.

It also appears that while recent fire frequencies for all 

severity classes are below what would have been expected 

for all the historical fire regimes in the region, the proportion of high-severity fire in fire-frequent regimes may 

be somewhat higher than it would have been historically. 

However, note that the recent rotations of high-severity 

fire in dry forests are still very low (table 3-4). Reilly et al. 

(2017) found that the amount of recent high-severity fire 

(23 to 26 percent) in the ponderosa pine, grand-fir, white fir, 

and Douglas-fir potential vegetation types was higher than 

what would be expected for these types under historical fire 

regimes. Mallek et al. (2013) reported that the percentage of 

high-severity fire in mixed-conifer forest types of the Sierra 

Nevada and southern Cascades of California was 5 to 8 percent during the pre-Euro-American period but was 22 to 42 

percent in dozens of fires between 1984 and 2009. Miller and 

Safford (2012) reported that larger recent fires in pine and 

mixed-conifer forests in the southern cascades of California 

experienced 33 percent high severity, which was probably 

higher than the historical amount of high-severity fire. However, Odion et al. (2004) found that fires in 1987 in remote 

areas of the California Klamath had relatively low percentages (12 percent) of high-severity fire (defined as 100 percent 

scorch or consumed) and the percentage of high-severity fire 

in the 2002 Biscuit Fire was only 14 percent (Azuma et al. 

2004). The relatively low percentage of high-severity fire in 

1987 may be a result of weather conditions that were not as 

extreme as those of more recent fires (Taylor and Skinner 

1998, Weatherspoon and Skinner 1995). Although the forests 

of the Klamath may have been less affected by fire suppression than more accessible forests, fire-return intervals during 

the suppression period are still nearly 50 percent longer (21.5 

vs. 14.5 years) than during the presettlement period (Taylor 

and Skinner 1998). As fire sizes increase with climate 

warming (Odion et al. 2004), patch sizes of high-severity 

fire may also increase (e.g., Miller et al. 2012, Reilly et al. 

2017). Very large patches of high-severity fire that kill older, 

dense forests would not be characteristic of the very frequent 

low-severity regime (Taylor and Skinner 1998), and efforts 
141

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Figure 3-26—Reconstructed historical (1900s) and current (1990s) maps of dry forest subwatershed of the Lower Grand 

Ronde subbasin in the Blue Mountains province displaying historical and current structural classes (A and B), fuel loading (C and D), crown fire potential under average wildfire conditions (E and F), and flame length under average wildfire 

conditions (G and H), respectively. (From Hessburg et al. 2005). Although this is from a landscape outside of the Northwest 

Forest Plan (NWFP) area, similar changes have likely occurred in dry forests in many areas within the NWFP. 
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GENERAL TECHNICAL REPORT PNW-GTR-966

to restore frequent fire and reduce fuels in older and younger 

forests would contribute to maintaining the biodiversity 

(including spotted owls in the southern part of their range) 

that was adapted to a dynamic and heterogeneous mix of 

forest ages and structures. 

Factors explaining variation in how fire-excluded 

forests burn when wildfire returns are not well understood. 

The observation that dry forests are experiencing less fire 

(excluding a direct effect of fire suppression), but more 

high-severity fire, or larger patches of high-severity fire than 

was true historically, is related to climate and fire suppression, but may also be due to shifts in vegetation-fire feedbacks. For example, it may be that with the absence of fire, 

coupled with succession to shade-tolerant and fire-intolerant 

species, is leading to forests that are less flammable under 

typical fire weather owing to a number of factors, including 

moister microclimate, denser stands that inhibit the free flow 

of wind, lower air and fuel temperatures owing to less direct 

sunlight, and more compact fuel beds (Engber et al. 2011, 

Estes et al. 2012, Kitzberger et al. 2011, Odion et al. 2004). 

For example, Weatherspoon et al. (1992) suggested that: 

...success of initial attack on wildfires evidently is 

greater in areas of owl habitat within the Sierran 

mixed-conifer type. Countryman’s (1955) description of fuel conditions within old-growth stands 

applies in large measure to fuel conditions within 

many mixed-conifer stands used by the California 

spotted owl. These stands are less flammable under 

most conditions, because the dense canopies maintain higher relative humidities within the stands and 

reduce heating and drying of surface fuels by solar 

radiation and wind. The reduction of wind velocity 

within closed stands discussed by Countryman is 

supported by wind reduction factors identified by 

Rothermel (1983) for stands with closed canopies. 

Windspeed at mid-flame height for fires burning 

in surface fuels is approximately one-tenth of the 

windspeed 20 ft (6.1 m) above the stand canopy.

However, they go on to say that: 

As fuels accumulate, however, fires that do escape 

initial attack—usually those burning under severe 

conditions—are increasingly likely to become 

large and damaging. Success in excluding fire from 

large areas that were once regulated by frequent, 

low- to moderate-severity fires has simply shifted 

the fire regime to one of long-interval, high-severity, stand-replacing fires... .

Some areas within the 2002 Biscuit Fire (which had 

relatively low total area of high-severity fire) could be an 

example of this shift in this regime, where moist multistoried older forests on north-facing slopes burned with high 

severity during the most extreme weather periods (hot dry 

east winds) of the fire (Thompson and Spies 2009).

Note that Countryman (1955) and Weatherspoon et 

al. (1992) never directly tested the hypothesis of higher 

humidity and fuel moisture in closed stands vs. more open 

stands. This was simply assumed to be so. Estes et al. (2012) 

measured an array of different sizes of fuels in closed, 

unthinned stands and open, thinned stands from spring 

snowmelt through fire season to the onset of fall rain/snow 

in the southern Cascades. They found moisture differences 

only in the early part of fire season (May–June). Moisture 

differences were gone by mid-season (July), and this carried 

through the remainder of the fire season. Further, the more 

open stands responded more quickly to the few rain events 

(thunderstorms) than did the closed stands. It appears that 

the long, dry summers of the Mediterranean climate areas 

in the southern parts of the NWFP area negate potential 

differences in moisture conditions because the closed stands 

catch up with the dry conditions of the open stands as the fire 

season progresses. Thus, the ability for crews to more readily 

catch fires in closed stands appears to be due to differences 

in exposure to sunlight creating higher air and fuel temperature and greater ease of windflow in the open stands. 

Thinning can alter fire potential and microclimate. 

Higher windspeeds in thinned stands compared to unthinned 

stands may have contributed to the former burning with 

higher fireline intensity (Raymond and Peterson 2005) than 

the latter in the 2002 Biscuit Fire. Although most of the differences in fire effects in that study were attributed to higher 

fine fuel loading and lower moisture in the stands that had 

been thinned but were not underburned to reduce fine fuels. 

Bigelow and North (2012) noted that thinning and group 

selection can change microclimates of forests but they did 

not find that such changes had a large effect on fire behavior. 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

The interaction between vegetation and fire severity is 

also determined by foliar moisture of the herbaceous, shrub, 

and hardwood fuels. For example, in open dry forests subject to frequent fire, well-developed herbaceous layers can 

reduce flammability because moisture contents can remain 

high into September (Agee et al. 2002). In the Klamath 

Mountains and western Cascades, hardwood understories 

can significantly reduce fire intensity (Agee et al. 2002, 

Perry 1988, Perry et al. 2011, Skinner 2006, Skinner and 

Chang 1996). Some species of evergreen shrubs can also 

reduce flammability of forests landscapes under most 

weather conditions but provide dense flammable fuels under 

extreme fire weather conditions (Skinner and Weatherspoon 

1996, Weatherspoon and Skinner 1995). 

Weatherspoon and Skinner (1995) suggested that 

another reason for the differences between stands of larger, 

old trees and those of smaller young trees and plantations 

experiencing different levels of fire severity in the Klamath 

could be simply the susceptibility of trees of different sizes 

to damage by fire. Large trees, especially stands dominated 

by old Douglas-fir and ponderosa pine, would be more 

likely to survive fires than younger trees, especially small 

trees in plantations (Agee and Skinner 2005, Skinner et al. 

2006). Although these multistoried stands have similar-size 

trees that succumb to the fires as do the young stands or 

plantations, the mortality is often hidden from satellite 

sensors by the surviving older, main canopy trees. Thus, 

the older stands become classified as experiencing mostly 

low-severity fire effects, while the others are classified as 

moderate- to high-severity fire effects even though fire 

intensity and sizes of trees actually killed could have been 

very similar (Weatherspoon and Skinner 1995). This is 

another example of the challenge of defining fire severity 

using single or simple metrics across variable vegetation 

types, and a potential source of confusion and debate 

(Reilly et al. 2017). 

Use of Historical Ecology in Conservation 

and Restoration

As illustrated above, knowledge of the ecology of the period 

prior to Euro-American settlement and widespread changes 

in land use can be very useful in understanding these forests 

and can serve as a starting place for developing conservation and restoration plans and management practices for 

them (Allen et al. 2002; DellaSala et al. 2003; Demeo et 

al. 2012; Hessburg et al. 1999a, 1999b, 1999c, 2000, 2005; 

Keane et al. 2002, 2009; Landres et al. 1999; Morgan et al. 

1994; Safford et al. 2012a; Swetnam et al. 1999). Knowledge 

of ecological history and knowledge of the historical range 

of variation (HRV) are not necessarily the same thing. General knowledge of ecological history may be more useful in 

management than a precise understanding of the range of 

variation in forest conditions (Hiers et al. 2016), which cannot be fully achieved for a number of ecological and social 

reasons. For example, while we may lack precise models or 

reconstructions of HRV for many landscapes in the region, 

we do have a reasonable foundation of historical knowledge 

for most areas. Ecological history reveals that forests were 

dynamic and best understood in terms of a HRV or its 

equivalent natural range of variation. The concept of HRV 

recognizes that habitats and ecosystems are dynamic in 

space and time, with historical ranges of behavior that are 

strongly constrained by the dominant climate, environment, 

and disturbances of an ecoregion. For the NWFP area, the 

HRV of forest structure among the four major fire regimes 

would have differed based on fire frequency and severity 

patterns and scale as described in the previous sections 

(fig. 3-27). Likewise, the HRV of forest structure would 

have differed across the major disturbance regimes based 

on whether small- to medium-size severity patches or 

high-severity patches were the major successional influence 

controlling patch dynamics.

Application of historical ecology HRV concepts 

and potential vegetation types in the Pacific Northwest 

and northern California must recognize the central role 

of climate variability in forest dynamics (Keane et al. 

2009, Wiens et al. 2012, Wimberly et al. 2000). Temporal 

variation in climate drove the variability of historical fire 

regimes (Hessburg et al. 200b, 2004; Trouet et al. 2010), 

which are the product of interactions between forest 

composition and structure, fire weather, and ignitions. 

Variation in climate and fire regime was the driving force of 

the “range” in the HRV in forest structure and composition. 

For example, fire occurrences in many of the moist and 

cool forests of the region are “climate limited” (Briles et 

al. 2011, Colombaroli and Gavin 2010, Littell et al. 2009) or 
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GENERAL TECHNICAL REPORT PNW-GTR-966 Live biomass/tree canopy cover Low High

A

Time (years) 

0 1,000 

Time (years) 

0 1,000 

Low High

Infrequent, high severity

Moderately frequent,

 mixed severity 

Very frequent,

 low severity 

Live biomass/tree canopy cover 

B

Frequent, mixed severity 

Figure 3-27—Hypothesized dynamics (historical range of variation) in live forest structure (biomass or 

cover) over a hypothetical 1,000-year period during the pre-Euro-American settlement period for an area 

of several thousand acres for (A) moist forest fire regimes and (B) dry forest fire regimes. Large declines 

in live biomass result from fire or wind; small declines result from fire, wind, insects, and disease. 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

“ignition limited” (sensu Agee 1993), but not fuel limited 

as the environments are typically productive enough to 

produce adequate fuels for burning within 10 to 15 years of 

a fire. If shrubs such as Ceanothus are present, they can act 

as a barrier to fire spread under less-than-extreme burning 

conditions (Briles et al. 2005, Mohr et al. 2000, Whitlock et 

al. 2004), or encourage rapid and intense fire spread under 

extreme fire weather conditions (Agee 1993, Moritz 2003, 

Schmidt et al. 2008). 

Regionally, wildfire was episodic and could be synchronous in parts of the region especially in wetter climates 

of the high- and mixed-severity regimes (Weisberg and 

Swanson 2003). Although fires were frequent in the driest 

forest regions, variability in frequency existed, and climatically driven synchrony of widespread fire still exists even in 

the fire-frequent forests of the Western United States (Falk 

et al. 2011, Heyerdahl et al. 2008). Wildfire frequency was 

variable at decadal to millennial scales, i.e., it was nonstationary. According to Whitlock et al. (2008), who examined 

paleo fire history of forests of the Northwestern United 

States, “There is no stable fire regime on millennial time 

scales, because fire-episode frequency varies continuously 

as a consequence of long-term climate variations and their 

influence on vegetation.” They go on to say, “Without 

supporting long-term paleoecologic data, short-sighted 

inferences about natural disturbance regimes and forest 

sensitivity are likely to be incorrect.”13 In other words, there 

were periods with relatively less frequent fire and other 

periods with relatively more frequent fire, creating a larger 

HRV if climate context is not taken into account. However, 

13 Although paleoecological fire histories can give us a broader 

perspective on HRV, they are subject to methodological limitations. For example, fire history studies based on charcoal 

occurrence in sediment cores are subject to bias because charcoal 

production is partially determined by the nature of the fuels (e.g., 

herbaceous vs. woody). In the Klamath Mountains, the frequency 

of fire generally exceeded the resolution of the sediment cores, 

which was usually no finer than 30 years at best (Briles et al. 2005, 

Mohr et al. 2000, Whitlock et al. 2004). Further, over most of the 

Holocene, there was rarely a time when charcoal was not entering 

the lakes in the Klamath region. Rather than being an indicator of 

fire events, the amount of charcoal at different periods appeared to 

be more indicative of biological productivity. Charcoal varied by 

amount with the periods of light, flashy fuels characteristic of pine/

oak woodlands represented by lower charcoal influx than in more 

productive periods characteristic of mixed-conifer forests (Mohr et 

al. 2000, Skinner et al. 2006, Whitlock et al. 2004).

in drier parts of the region with more frequent fire, largescale temporal variability and regional synchrony in fire 

was probably less than in regions with less frequent but 

larger fires (Hessburg et al. 2005; Heyerdahl et al. 2001, 

2008; Kitzberger et al. 2006; Mohr et al. 2000; Morgan et 

al. 2008; Skinner et al., in press; Taylor et al. 2008; Trouet et 

al. 2010). Nevertheless, regionally extensive fire events 

associated with drought did occasionally occur in the 

eastern Cascades of Washington (Hessl et al. 2004).

Going forward, several authors have argued that given 

climate change, invasive species, and widespread landscape 

change, using historical conditions or ranges of variation 

as a narrow goal or target for conservation and restoration 

can be unrealistic, impossible, or even incongruent with 

conservation goals (Millar et al. 2007, Palmer et al. 2005). 

This is especially true if the goals include threatened and 

endangered species, such as the northern spotted owl in dry 

forests, whose habitat can be the product of human land use 

activities and altered disturbance regimes. However, it is 

self-evident that knowledge of historical forest dynamics is 

essential for conservation and restoration of native (historical) vegetative communities and associated wildlife species 

even under climate change. The challenge for application of 

the concept is to be aware of limitations and apply historical 

knowledge with caution. Hessburg et al. (2016) offer four 

caveats to using historical reference conditions as management guidelines: 

• Mimicking historical conditions is not an 

end in itself, but is a means of accomplishing 

objectives (e.g., resilience to fire), and therefore 

appropriate only when it meets those objectives.

• The true value of historical information is in 

understanding how interacting fire and climate, 

and their variability through time and space, 

influenced ecological patterns of forest structure 

and successional conditions. This information 

can provide valuable direction for the complex 

process of ecological goal setting in management 

planning and implementation.

• Past conditions may not fully reflect future 

climate-vegetation-disturbance-topography 

linkages as a result of pervasive climate and 
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land-use changes. Hence, one of the challenges may be deciding the degree to which 

past lessons are relevant to future management. 

Relevance will depend on goals, reasonable 

expectations of the future climate, and resources 

required to attaining those goals.

• Because regional landscapes are highly altered, 

restoration restricted to local landscapes is insufficient to address large-scale restoration needs. 

Remember that we understand recent HRV (e.g., past 

500 years) better than we understand HRV of the more distant past or what the range of variation will be in the future. 

Consequently, planning efforts based on ecological history 

or HRV will need to be flexible, adaptive, and periodically 

revised to keep up with new knowledge and changing 

ecosystems. To deal with the challenges of restoration or 

managing for resilience, Hobbs et al. (2014) recommended 

that landscape frameworks and assessments be used to identify where it is possible to retain or restore native biodiversity and where novel or “hybrid” (seminatural) ecosystems 

might be a management goal either because of human values (e.g., areas of dense forests for wildlife created by fire 

exclusion) or because of the impracticality or impossibility 

of returning those areas to their pre-Euro-American state 

or HRV (see chapter 12 for more discussion of this issue). 

We further discuss scientific understanding of approaches 

for dealing with these and other challenges of restoration or 

creating resilient forests in sections below. 

Ecosystem Function 

The preceding sections have emphasized forest structure, 

composition, and disturbance process, but ecosystems can 

also be characterized through their functions (ecological 

processes or activities), which also differ with successional 

stage and disturbance regime. Key functions include 

primary productivity and carbon fixation, nutrient cycling, 

hydrological functions, and habitat for biota (Franklin et al. 

2018). We briefly review how these differ with succession 

here with a focus on productivity, carbon and nutrient 

cycling. For more information about hydrological functions 

and habitat, see chapters 6 and 7.

Old-growth forests are productive ecosystems, fixing 

a large amount of solar energy in what is termed gross 

primary production (Franklin and Spies 1991). Following 

major disturbances, ecosystem live biomass and net primary 

productivity (difference between carbon fixed through 

photosynthesis and lost to respiration) are relatively low 

(Bormann et al. 2015, Spies 1997), in contrast with later 

successional stages. As trees grow and canopies close, the 

rate of carbon sequestration and biomass accumulation 

becomes high. Biomass reaches its highest level in older 

forests, but net primary production declines toward zero 

because growth and mortality are roughly equal. While 

stand-level net primary productivity and carbon accumulation is low in older forests, the rate of biomass growth 

for individual trees continues to increase with tree size 

(Stephenson et al. 2014). 

Carbon, which primarily resides in the wood and soils, 

is highest in old forests (Law and Waring 2015). Douglas-fir/

western hemlock forests can continue to be a net sink for 

carbon for more than 500 years, thanks to the contribution 

of primary production of shade-tolerant understory trees 

(Harmon et al. 1990, 2004). Older moist forests of the 

NWFP area can attain higher stand (tree) carbon biomass 

than tropical or boreal forests (Law and Waring 2015). 

Young forests store less carbon but accumulate it at higher 

rates than old forests. 

Recent large wildfires in coniferous forests of the 

region release carbon, but the total emitted carbon is less 

than previously thought, partly because most fires in the 

region have burned with mixed severity. For example, 

Campbell et al. (2007) found that only 1 to 3 percent of the 

carbon in trees larger than 3 inches (7.6 cm) was combusted 

in the 2002 Biscuit Fire (Campbell et al. 2007). Total carbon 

emitted from four fires in Oregon averaged 22 percent 

of prefire carbon for all pools (Meigs et al. 2009). As the 

biomass killed in fires slowly decomposes over decades to 

centuries, carbon is emitted to the atmosphere as carbon 

dioxide and other trace hydrocarbons. About half the carbon 

remaining after a fire stays in the soil for about 90 years; 

the other half persists for more than 1,000 years as charcoal 

(Deluca and Aplet 2008, Law and Waring 2015). 
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Forest management effects on carbon differ with 

management intensity, rotation length, and forest type. It is 

often thought that managing forests on a short rotation (e.g., 

40 to 50 years) would provide the most effective long-term 

carbon sequestration, but longer rotations and selective or 

no harvest provides the most carbon sequestration (Harmon 

et al. 1990, Ryan et al. 2010). Forest management under the 

NWFP to promote older forests with a low level of timber 

harvest would provide for more carbon sequestration than 

more intensive management (Creutzburg et al. 2017, Kline 

et al. 2016). 

In forests prone to frequent fires, the carbon and forest 

management picture is more complex, with some studies 

showing a positive benefit of forest fuel reduction on carbon 

sequestration and others showing a negative effect. Some 

modeling suggests that carbon stocks over the long term 

are best protected by fuel treatments that create relatively 

low-density stands dominated by large, fire-resistant trees 

(Hurteau and North 2009). Other studies (Ager et al. 2010, 

Loudermilk et al. 2016, Spies et al. 2017) found that active 

management reduced carbon stored in the forest landscape 

by 5 to 25 percent for at least several decades. The effect 

of management on carbon depends on how frequently 

management treatments encounter fire and reduce fire 

severity. When a fire encounters a recently treated area, less 

carbon is likely to be emitted than when it encounters an 

untreated forest of the same type. However, at a landscape 

scale, many treatments will not experience a fire and the 

management actions there will reduce carbon sequestration. 

The net effect at a landscape scale may be to reduce carbon 

sequestration unless those treatments are strategically 

placed and occur where fire is most likely to happen. Further, the more active the fire regime becomes under climate 

warming scenarios, the more important strategically placed 

fuels treatments (e.g., Finney et al. 2007, Schmidt et al. 

2008) become in protecting carbon stores (Loudermilk et 

al. 2013, 2016). 

Nutrient cycling varies with successional stages and 

forest region. Old-growth forests are highly retentive of 

nutrients, and sediment outputs from old-growth watersheds 

are typically low (Franklin and Spies 1991, Swanson et al. 

1982). Many of the forests of the NWFP area are nitrogen 

limited, but several natural processes exist that capture 

nitrogen and make it available for vegetation growth. oldgrowth forests can support canopy lichens such as Lobaria 

oregana, L. pulmonaria, and others that fix nitrogen and 

then “leak” significant amounts of nitrogen to the ecosystem 

(Antoine 2004). Immediately following stand-replacement 

disturbance, rates of erosion and nutrient loss can be 

elevated until vegetation recovers (Ice et al. 2004). As plants 

establish and cover increases during early-successional 

and young forest stages, sediment losses return to predisturbance levels, and N2-fixers such as Ceanothus spp. and 

hardwoods such as red alder (Alnus rubra) begin to increase 

organic matter and nutrient availability (Borman et al. 2015, 

Compton et al. 2003). While red alder can add available 

nitrogen to forest ecosystems, the high rates of nitrification 

can accelerate cation leaching and soil acidification relative 

to conifer-dominated stands (Compton et al. 2003). Shrubs 

and sprouting hardwood trees can also help reduce nutrient 

losses after wildfire in forests of southwestern Oregon. 

While the longer term benefits of early-seral plant communities to conifer tree growth are still not well understood 

(Bormann et al. 2015), it is generally understood that early-seral herbaceous, shrub, and hardwood tree communities 

can all play an important role in supporting forest nutrient 

cycling and productivity. 

Restoration efforts in dry forests can also benefit soil 

fertility and productivity. Fire suppression can lead to 

increases in nitrogen pools in ecosystems, but the majority 

is bound in forms that are less available to plants (Ganzlin 

et al. 2016). Forest restoration treatments, including prescribed burning, can produce short-term pulses of nitrogen 

in forms that are available to plants. Thinning alone will 

not produce these nutrient benefits and is not an effective 

surrogate for fire in terms of nitrogen. Frequent prescribed 

fire that emulates historical fire frequency and severity is 

necessary to maintain rapid rates of nutrient cycling in these 

dry forest ecosystems. However, while the nutrient effects 

of fire may be ephemeral, benefits to other soil resources 

and processes such as available water and photosynthetic 

rates may be longer term (Ganzlin et al. 2016). 
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Conservation and Restoration Needs 

In this section, we summarize the major conservation (e.g., 

protection of existing vegetation) and restoration (e.g., 

promotion of desired conditions) needs for moist and dry 

forests relative to the original goals of the NWFP and of 

the 2012 planning rule under which the NWFP currently 

operates (table 3-5). 

Estimates of forest change for the NWFP region 

suggest that the need for conservation and restoration of 

the ecological integrity of old-growth forests and other 

successional stages of the region spans a wide range of the 

disturbance regimes and forest types. For example, Haugo 

et al. (2015), found that at least 40 percent of all coniferous 

forests in eastern Washington and eastern and southwestern 

Oregon are in need of management to restore wildfire, fuel, 

or forest structure conditions to be more consistent with the 

natural range of variation. After more than 125 years of land 

clearing, timber harvest, 20th century high-severity wildfire 

associated with early logging and land use, fire suppression 

and succession, the sum of mature and old-growth forest 

(OGSI 80) across all the fire regimes is 17.8 million ac 

(7.2 million ha), or ~ 39 percent of all public and private 

forest-capable lands in the Plan area (Davis et al. 2015). 

When only the oldest multilayered forests with trees >200 

years old (OGSI 200) are considered, the current amount is 

~7.6 million ac (3.1 million ha), or 17 percent of all public 

and private forest-capable lands. Of that 17 percent, more 

than 80 percent is on federal lands. It is difficult to estimate 

what percentage of the historical range of older forests this 

represents for several reasons, including lack of quantitative 

studies of HRV across the region, uncertainties in estimates 

of HRV, and the current definitions do not fully capture 

the diversity of older forest conditions, especially for older 

ponderosa pine and mixed-conifer forests of the low- and 

mixed-severity regimes. If we focus on trees older than 

200 years (OGSI 200) in moist forests zones west of the 

Cascade crest, then the total remaining may represent 17 to 

23 percent of the amount that was present on average before 

the mid-1800s. This assumes that at least 60 percent of these 

forests areas were covered by forests containing trees older 

than 200 years (FEMAT 1993, Wimberly 2002). 

Moist forests—

In the moist forests zone, losses of older forest have resulted 

mainly from clearcutting for timber management (Spies 

et al. 1994). The decline in older forest has been sharp as 

indicated above. For example, the vegetation structure of 

northern spotted owl habitat (not necessarily the same as 

Table 3-5—Summary of vegetation conservation and restoration needs for moist and dry forests of the 

Northwest Forest Plan (NWFP) region related to the ecological goals of the NWFP and the 2012 planning rule 

Forest region Conservation needs Restoration needs

Moist forests Protect existing older forests stands and 

large patches of older forests from 

logging and high-severity fire. These 

have been greatly reduced by timber 

management and other land uses.

Increase vegetation diversity in plantations and accelerate 

development of older forest structure and composition.

Reduce fragmentation and increase connectivity of older 

forest patches.

Create or promote early-seral vegetation where needed to 

provide seral stage and landscape diversity.

Restore disturbance processes (e.g., fire) where feasible.

Dry forests Protect existing large fire-tolerant trees in 

areas of dense and open forest.

Manage and protect existing dense oldgrowth forest stands as necessary to meet 

late-successional species and ecosystem 

integrity needs.

Restore low- and mixed- severity fire as key ecological 

process.

Increase areas of open old forests to promote resilience to 

fire and climate change and meet needs of species.

Develop landscape-level strategies to create desired mosaics 

of open and dense old forest and to increase resilience 

and meet simultaneous needs of wildlife species and 

ecological integrity.

Restore diversity to plantations, including tree species mixes.
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

old-growth forests) has declined by 20 to 52 percent across 

the different provinces between 1930 and 2002 (Lint et al. 

2005). Many plantations on federal lands are 30 to 60 years 

old and average about 20 to 25 ac (8.1 to 10.1 ha) with some 

as large as 60 ac (24.3 ha) (Cohen et al. 2002). They were 

often planted primarily with Douglas-fir (or at most a total 

of one or two additional species) at an even spacing. Logging and site-preparation treatments to control competing or 

unwanted vegetation resulted in uniform stand density with 

lower levels of shrub and hardwood components, and fewer 

snags and down wood structures (Bailey and Tappeiner 

1998, Spies and Cline 1988). A large percentage of federal 

forest land outside of wilderness areas is covered by such 

plantations—as much as 40 to 55 percent of some landscapes, including many late-successional reserves (LSRs) 

(fig. 3-28). In summary, management efforts to ensure high 

density and species uniformity were often so successful that 

conditions in these stands do not match the heterogeneity 

and growth trajectories of naturally regenerated postwildfire stands (Donato et al. 2011, Freund et al. 2014, Larson 

and Franklin 2005, Tappeiner et al. 1997, Tepley et al. 2014, 

Winter et al. 2002a) (fig. 3-14).

Other vegetation restoration needs for the moist forests 

zone relate to early-seral and other mid-successional stages 

that have been reduced by fire-suppression reforestation, 

timber stand improvement treatments that ensured full 

stocking, optimal sawtimber growing conditions, and 

control of unwanted vegetation (Agee 1993, Cole and 

Newton 1987, White and Newton 1989). Fire suppression 

in these infrequent-fire regimes has little impact on the risk 

of high-severity fire but it does reduce the amount of earlyseral and vegetation diversity in a landscape. Numerous 

small- to mid-size fires would likely have served as barriers to fire spread where they created patches of deciduous 

shrubs and trees. The vegetation diversity created by these 

fires probably regulated the frequency-size distributions, 

especially of the larger fires. The amount of early-seral 

condition may have been relatively high (<30 percent) in 

these regimes during the late 1800s and early 1900s when 

the legacy of aboriginal burning was still evident (Robbins 

1999) along with Euro-American-ignited fires from land 

clearing and logging (fig. 3-6). The amount and diversity 

of early-seral vegetation created by these fires would have 

been reduced where snags were cut down and large-scale 

planting efforts reduced the period of time before tree 

canopy closure. The patterns of early-seral patch size 

shapes, distribution, and structural heterogeneity created 

by logging and reforestation in the late 20th century are 

not representative patterns typically found under historical 

fire regimes (Nonaka and Spies 2005). The structure and 

composition of early-successional vegetation and young 

forests created by clearcut logging significantly differed 

from those of postwildfire conditions because intensive 

timber management removed all live and dead trees, and 

herbicides (in early years on federal lands), and planting of 

Douglas-fir seedlings reduced diversity of vegetation and 

shortened the nonforest period of succession. Moreover, 

harvest unit boundaries often followed land ownership 

boundaries on private lands, and older cutting units on 

federal lands (the most recent occurred in the early 1990s) 

represented small-size (25 to 40 ac [10.1 to 16.2 ha]), regularly shaped units with landscape patterns that differed 

from those created by fire. 

Dry forests—

We have already described many of the changes that have 

occurred in the dry forests as a result of fire exclusion and 

logging. Analysis from the Interior Columbia Basin Ecosystem Management Project (Hann et al. 1997; Hessburg 

et al. 1999a, 2000) provides a picture of how the area of 

dense multilayered older forest has changed from historical 

to current (late 1990s) in dry forests of eastern Washington 

and Oregon (fig. 3-26) (table 3-6).

In another study, Lint (2005) estimated that the amount 

of dense older forest with grand fir and Douglas-fir that is 

suitable for spotted owls (we use this as an approximation of 

multilayered old growth, but it is not necessarily the same 

as dense old-growth forest structure) has actually increased 

by 16, 6, and 11 percent in the eastern Cascades of Washington, Oregon, and California Klamath Provinces, respectively, from 193014 to 2002. These data suggest that the 

historical fire regime in these provinces did not favor large 

areas of either late-successional, multilayered old forest or 

northern spotted owl habitat. 

14 Landscapes of the 1930s would have already been altered by 

logging, grazing, fire exclusion, and occurrence of fires associated 

with land use activities. Fire exclusion would have increased the 

amount of dense forest by 1930 (McNeil and Zobel 1980, Merschel 

et al. 2014). 
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Figure 3-28—Plantations and the dates of their origin in a landscape containing late-successional reserve (in white), wilderness (striped), 

and matrix (orange) lands on the Siuslaw National Forest in coastal Oregon. From Stewart Johnston (retired), Siuslaw National Forest.
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Changes in area of medium and large old trees have 

also occurred. Hessburg et al. (1999a) documented reductions in province area of forest patches with medium and 

large trees in the overstory (>40 percent canopy cover) in 

the interior Columbia River basin. In the Northern Cascades 

and Upper Klamath provinces, area of medium- and largesize trees in the overstory declined from 30 to 24.9 and from 

28.9 to 25.3 percent, respectively. However, area of medium 

and large trees in the overstory significantly increased in the 

Southern Cascades province from 17.1 to 32.8 percent. They 

also show historical landscapes with significant areas of 

grassland, shrubland, woodland, and stand initiation forest 

conditions and young forests that had invaded meadows 

(figs. 3-21 and 3-22). These mid- to late-20th century 

increases in forest density are in addition to the substantial 

increases in stand density and shade-tolerant species that 

occurred between 1890 and 1930 as a result of fire exclusion 

(owing to grazing, logging, and eventually active fire suppression) and other factors (Merschel et al. 2014, Taylor and 

Skinner 2003). Currently, the percentage of relatively open, 

low-density (<80 trees per acre) forest with large old trees in 

mixed-conifer and Douglas-fir potential vegetation types is 

about 10 percent, while the area of dense forest (>584 trees 

per acre (1,442 trees per hectare)) with old trees covers about 

35 to 42 percent of the potential vegetation types (Reilly and 

Spies 2015). These increases in shade-tolerant densities have 

made forests less resilient to fire as described above. 

Increases in forest density are not the only conservation 

and restoration concerns in the dry forests. Loss of large, 

fire-resistant trees to logging and wildfire has also strongly 

affected forest ecosystem integrity, resilience, and wildlife 

habitat in both the very frequent low-severity and frequent 

mixed-severity fire regimes of the dry forest zone. For 

example, the density of large fire-tolerant tree species (e.g., 

ponderosa pine and Douglas-fir) has decreased substantially 

as a result of high-grade logging (selective removal of 

large mostly commercially valuable trees) (e.g., Hessburg 

et al. 1999a, 2000, 2003, 2005; Merschel et al. 2014) and 

clearcutting and plantation establishment. Hagmann et al. 

(2014) estimated that the area of forests dominated by large 

old trees has been reduced from 91 to 29 percent for dry 

and moist mixed-conifer in one landscape in the eastern 

Oregon Cascades. Increases in future development of 

large, old fire-intolerant trees may be limited as a result of 

forest densification and fire suppression. We could find no 

disagreement in the literature on the issue of restoration 

needs and concerns for large old conifers (e.g., Baker 2012, 

Stine et al. 2014). This issue is prominent in the eastern Cascades of Washington and Oregon and in California, where 

topography and proximity to settlement made these large 

valuable trees an easy target for logging (Hessburg and 

Agee 2003; Hessburg et al. 2005, 2015, 2016; Merschel et 

al. 2014; Richie 2005). Loss of large trees is less of an issue 

in more remote sites in rugged and difficult-to-access areas 

such as the less roaded areas of the Klamath Mountains. 

Timber Management and Old-Growth 

Conservation 

The NWFP strategy was based on the assumption that historical timber management approaches (e.g., removal of large 

or old early-seral and fire-tolerant trees) are not compatible 

with the full ecological functions of old-growth forests and 

other successional stages. Since FEMAT (1993), no scientific 

evidence has emerged that intensive timber production (e.g., 

clearcutting and short-rotation plantation forestry) and oldgrowth forest conservation are compatible at stand levels for 

any of these forest types and disturbance regimes. 

Table 3-6—Historical and 1990s percentages of total forest area in late-successional multistory forest in 

provinces of the Interior Columbia Basin Ecosystem Management Project 

Province

Time period Northern Cascades Southern Cascades Upper Klamath

Historical 7.0 0.7 4.8

Current 16.6 4.0 3.5

Sources: Hann et al. 1997; Hessburg et al. 1999a, 2000.
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Moist forests—

In moist forests zones, partial cutting, in the form of 

green tree retention harvest (see section below for more 

discussion of this method), patch cutting (creating gaps less 

than a few acres), or selection harvest methods may retain 

the habitats of some late-successional animal and plant 

species (Baker et al. 2016, Gustafsson et al. 2012, Halpern 

et al. 2012, Hansen et al. 1995a, Rosenvald and Lohmus 

2008). It also retains some of the ecological functions of 

old growth, but could strongly affect dead wood amounts. 

The accompanying road and harvest systems would add 

additional impacts. Very long management rotations (e.g., 

more than 150 or 200 years) could in theory produce some 

of the habitat and ecosystem service benefits of older 

forests (Kline et al. 2016), but it would take at least a 

century to quantify these effects, and no long-term studies 

are currently underway. 

One of the only operational plans to meet both older 

forest conservation goals and timber production in moist 

forests in the literature is the “structure-based management” approach proposed by the Oregon Department of 

Forestry for the state forests in the northern Oregon Coast 

Range (Bordelon et al. 2000). In this approach, management targets were sorted into five stand types, with the two 

oldest, “layered” and “older forest structure” intended to 

meet late-successional conservation goals. There are no 

reserves, and older forest conditions are met through long 

rotations. The areas in each stand type can differ over time, 

e.g., between 20 and 30 percent of older forest structure, as 

harvesting and succession shift age and structure classes 

over the landscape. Spies et al. (2007) and Johnson et al. 

(2007) used a landscape model to approximate this strategy. 

Modeling results suggest that, over time, this approach 

created a greater diversity of habitat benefits, including 

increases in older forest habitats and higher levels of wood 

compared to federal management under the NWFP. No 

formal field assessment of the ecological or economic 

implications of this approach has been attempted. At 

this stage, the Oregon Department of Forestry is under 

pressure from the counties to increase revenues and is 

in the process of modifying or abandoning the approach 

(http://www.nwtimberblog.blogspot.com/2013/11/board-offorestry-seeks-better.html; http://www.northcoastcitizen.

com/2016/12/officials-say-county-will-not-opt-out-of-classaction-lawsuit-over-timber-harvest/).

Other examples of management agency efforts to meet 

biodiversity and timber management goals exist for moist 

forests but have not been published or reviewed in the peer 

reviewed literature. The most prominent and well-developed approach for integrating timber management with 

old-growth forest conservation in moist forest zones may be 

the Washington Department of Natural Resources Habitat 

Conservation Plan for state trust lands (http://www.dnr.

wa.gov/programs-and-services/forest-resources/habitat-conservation-state-trust-lands), which has been implemented 

across more than a million acres of state and private land 

with the goal of maintaining old-growth forest species 

and providing sustainable levels of timber production. It is 

based on maintaining a mosaic and network of patches of 

old-growth and mature forest structure for terrestrial and 

aquatic species. 

Until more research is done, including field-based tests 

and monitoring, there is little debate that the best way to 

conserve and maximize old-growth values in moist forests 

is to exclude intensive timber management activities (e.g., 

clearcutting and plantation establishment) in old growth. 

This was the direction of the NWFP when it placed 80 percent of the remaining old-growth forest patches on federal 

lands into LSRs. The remaining 20 percent was placed into 

matrix lands—open to timber management, using innovative silviculture (e.g., ecological forestry) according to 

approved plans (USDA FS 1994) (fig. 3-29). The suggested 

management approach of the NWFP in the matrix lands, 

along with experiments in adaptive management areas, had 

they been implemented, would have enabled scientists and 

managers to learn about tradeoffs associated with managing 

for timber and ecosystem values at patch levels. As it stands 

now, we know relatively little about these tradeoffs because 

of a lack of implemented studies—the exceptions being 

the simulation studies of Cissel et al. (1999) and Spies et al. 

(2007) for moist forests. 
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Dry forests—

Clearcutting and plantation management are also not compatible with management for ecological integrity and resilience in dry forests (Franklin et al. 2013). However, other 

forms of management (table 3-5) may be needed to promote 

ecological integrity and resilience to climate change as 

characterized by the 2012 planning rule. Restoration 

thinning and prescribed fire in forests containing trees over 

80 years would promote resistance and resilience to fire 

and climate change both within and outside LSRs. Some 

of these restoration activities could provide economically 

valuable wood products. Areas of dense old, multilayered 

forests and owl habitat can still be provided at landscape 

scales, but they would be more dynamic, shaped by fire and 

other natural disturbance agents. A holistic landscape-restoration strategy has been proposed for the 4-million-ac 

(~1.6-million-ha) Okanogan-Wenatchee National Forests. 

The plan seeks to use a variety of vegetation and fuels management techniques to reduce wildfire vulnerability across 

the landscape, including in areas adjacent to owl habitats in 

“critical habitat” (USDI 2012), and to restore fire regimes, 

to increase resilience to climate change. More research is 

needed in these dry dynamic landscapes to develop and 

evaluate approaches for achieving both ecosystem and focal 

species goals (see chapter 12). 

Reserves in Dynamic Ecosystems 

Concepts— 

Protected areas or reserves are a well-established strategy 

for conserving biodiversity by limiting human activities 

(e.g., intensive timber management and development) that 

are incompatible with certain ecological objectives (Lindenmayer and Franklin 2002). However, the efficacy of reserves 

as the sole basis for conserving biodiversity has been 

challenged by a number of authors (e.g., Fischer et al. 2006, 

Lindenmayer and Franklin 2002). These challenges relate 

to several concerns: (1) biodiversity reserves cover only 

a small part of the Earth’s land surface (e.g., <6 percent) 

(Fischer et al. 2006); (2) globally, the majority of reserves 

tend to be small in area (tens to <25,000 ac [~10 000 ha]) 

(Bengtsson et al. 2003), making them susceptible to impacts 

from large rare events (e.g., fire and wind) and influences 

(e.g., invasive species and human activities) from outside 

the reserves; and (3) most reserves are static and climate 

change may shift environments and species distributions to 

unreserved areas (Carroll et al. 2010). 

A fundamental design recommendation for reserves 

is that they should be considerably larger than the largest 

disturbance patch size if they are to maintain habitat and 

populations of the most extinction-prone species (Pickett 

and Thompson 1978). This concept, which is known as 

“minimum dynamic area” requires knowledge of patch 

size distributions of infrequent disturbances that would 

Figure 3-29—Example of a green tree retention unit created on 

Central Cascades Adaptive Management Area on the Willamette 

National Forest. The goal was to emulate stand structure created 

by a partial stand-replacement fire and produce timber.

Gary Rost
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GENERAL TECHNICAL REPORT PNW-GTR-966

be considered incompatible with conservation goals. 

Such knowledge is lacking for most disturbance regimes, 

especially under climate change, but it can be estimated 

using historical information and power laws (e.g., see 

Moritz et al. 2005). 

The reserve design of the NWFP was a late-successional forest coarse-filter strategy that was based heavily 

on the needs of the northern spotted owl and leveraging 

existing reserves (e.g., wilderness) where appropriate. 

The reserve strategy attempted to mitigate the shortcomings of other reserve-based conservation approaches by 

increasing the proportion of reserves on federal lands 

to 80 percent (including congressional reserves, LSRs, 

riparian reserves, and administratively withdrawn areas). 

The congressional reserves and LSRs represented 28.1 

percent (15.8 million ac or 6.4 million ha) of all public 

and private forest lands in the NWFP area, which made it 

one of largest reserve systems for any temperate forested 

ecoregion in the world. The individual LSRs under the 

NWFP are also relatively large. For example, 47 percent 

of the individual LSRs are larger than 25,000 ac (~10 

000 ha), and three are larger than 250,000 ac (~100 000 

ha) (fig. 3-30). Compared to the size of recent patches of 

high-severity fire (fig. 3-30), the sizes of the reserves are 

typically larger, although many (>120) LSRs are relatively 

small (e.g., <25,000 ac) and could be completely burned in 

a single fire event with large patches of high-severity fire 

(e.g., 25,000 ac). 

The NWFP hypothesis was that a large network of 

reserves well-distributed across the region would be resilient to expected losses from wildfire over a period of 100 

years. While losses were expected, there was no estimate 

of how much loss would be too much for the goals of the 

Plan. The reserve patch size and fire-size analysis indicated 

that, for the most part, the reserves have been large enough 

and numerous enough to absorb many recent large fires 

with limited loss of OGSI 80 or OGSI 200 forests in many 

but not all provinces. However, it must be remembered that 

recent historical fire history trends will not necessarily 

continue in the future. Given current trends, it is likely that 

one to several of the LSRs, especially the small ones, will 

experience significant losses of OGSI to large patches of 

high-severity fire over the next few decades. The infrequent 

fire regimes of the area have the potential to burn with 

very large fires, and it remains to be seen if the sizes and 

numbers of LSRs are sufficient to meet the goals of the Plan 

under climate change or other threats (e.g., invasive species). 

The effectiveness of the NWFP regional reserve-matrix strategy in meeting ecological goals under current 

and future climate has received relatively little attention 

in scientific literature. The limited studies suggest that the 

existing network and standards and management guidelines 

of reserves, which spans a wide range of elevations and 10 

degrees of latitude, will provide a good (but not necessarily 

optimal) foundation for meeting conservation goals in 

moist forest zones under a changing climate (Carroll et 

al. 2010, Spies et al. 2010). However, other than Carroll et 

al. (2010) and Carrol (2010), no quantitative studies of the 

NWFP reserve network or the regional plan as a whole have 

been conducted outside of efforts focused on conservation 

planning for the northern spotted owl (USDI 2012, USFWS 

2008). In general, the science of regional conservation 

planning and assessment, including evaluation of reserve 

networks, has advanced considerably since the NWFP was 

implemented. For example, Margules and Pressey (2000) 

presented a systematic approach for evaluating reserve 

network plans and implementation and Virkkala et al. 

(2013) demonstrated a methodology to evaluate the viability 

of reserve networks for protecting biodiversity in the face 

of climate change in Finland. According to Carroll (2010), 

“Rigorous assessment of the implications of climate change 

for focal species requires development of dynamic vegetation models that incorporate effects of competitor species 

and altered disturbance regimes.” In his assessment of the 

resiliency of the NWFP reserve network for multispecies 

conservation under climate change, Carroll (2010) did not 

address how wildfire might affect the conservation goals of 

the Plan, which is a significant concern. The development 

of regional-scale vegetation and species occurrence data 

and vegetation dynamics models, including spatial fire 

landscape models (e.g., Scheller et al. 2011, Spies et al. 

2017), in recent years suggests that a more rigorous and 

comprehensive evaluation of the NWFP regional strategy 

would now be possible. 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Reserves or protected areas are not necessarily areas 

where all human activities are excluded or are inconsistent 

with ecological conservation goals (Soule 1985). There are 

many types of protected areas with different degrees of 

human activity permitted (Spies 2006), including recreation 

areas, management allocations for degree and type of vegetation manipulation, invasive species removal areas, and 

fire management (prescribed fire or fire suppression) areas 

(Pressey et al. 2007). In most cases, including the NWFP 

standards and guidelines, biodiversity reserves permit and 

encourage restoration activities that further the species 

and ecosystem goals of the reserved area. For example, the 

1

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Figure 3-30—Patch distributions of recent fire (2000–2012) and sizes of Northwest Forest Plan (NWFP) reserves: (A) frequency distribution of number patches by fire severity and area class, (B) fire-severity patch sizes by area class, (C) frequency distribution of number of 

NWFP reserves (late-successonal reserves [LSRs] alone and LSRs plus congressionally designated reserves [CRs] by area, and (D) area 

of reserves by size class. 
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GENERAL TECHNICAL REPORT PNW-GTR-966

NWFP indicated that restoration activities within reserves 

were needed for both moist and dry forests (USDA FS 

1994) in plantations in wetter and drier forests, and in older 

forests in fire-frequent regimes where forest structure and 

composition has been altered by fire exclusion and logging 

of older trees. 

Wildfire and fire exclusion both pose serious challenges and dilemmas to managers seeking to conserve 

biodiversity using reserves or any other conservation 

approach (Driscoll et al. 2010, Fischer et al. 2006, Spies 

et al. 2012). This observation may seem contradictory or 

ironic, but it is the reality when conserving fire-prone forests in the Western United States. The multifaceted nature 

of wildfire makes it difficult to find a conservation and 

management “sweet spot.” For example, fire is a vital and 

dynamic ecological process that maintains some communities, renews other communities, and increases plant growth 

and productivity (Ahlgren and Ahlgren 1960), but it also 

kills trees and destroys valued habitats, forest resources, 

and human infrastructure and lives (DellaSala and Hanson 

2015). The assumption that reserves could conserve habitat 

for the northern spotted owl and other old-growth-associated species in dynamic ecosystems subject to fire, 

succession, and climate change was a major hypothesis of 

the NWFP. We examine this hypothesis below using data 

from the monitoring program (Davis et al. 2015) and new 

scientific knowledge. 

Is the reserve system meeting the original goals of the 

Northwest Forest Plan?— 

The reserve-matrix system was intended to protect and 

recover older forests in response to threats from logging 

and natural disturbances that destroy older dense forests. 

The general goal was to increase the amount of late-successional/old-growth forest in the reserves to recover toward 

levels that were present before extensive logging began on 

federal lands in the early 1950s. No specific targets for the 

future proportion of late successional/old growth in reserves 

were made in terms of HRV at the LSR scale, but the 

expectation was the amount of late successional/old growth 

in general on federal land would approach 60 percent over 

100 years (Davis et al. 2015), including expected losses 

owing to wildfire. Dry zone forests were included in this 

rough estimate though the likelihood of achieving this goal 

was considered to be lower in dry forest zones than in moist 

forest zones (FEMAT 1993: fig. IV-3). It was expected that 

millions of acres of younger forests and plantations would 

eventually grow into an old-growth condition making up for 

any losses to wildfire or other disturbance agents. Between 

1993 and 2012, disturbances, including wildfire and planned 

timber harvest, have reduced older forest (OGSI 80) area 

by 6.0 percent and OGSI 200 by 7.6 percent (Davis et al. 

2015). Wildfire has accounted for the greatest reduction in 

older forest: annualized losses to wildfire were 0.22 percent 

and 0.28 percent for OGSI 80 and OGSI 200, respectively. 

In comparison, FEMAT (1993: IV-55) assumed that the 

annualized percentage of high-severity fire in reserves 

across all provinces would be about 0.25 percent over the 

first 50 years. At the scale of the entire NWFP, the losses 

from wildfire approximated expectations (Davis et al. 2015, 

FEMAT 1993) across the entire plan area (no projected 

losses were made by province), but losses from timber 

harvest were much less than planned. 

The rates of change in OGSI 80 were not uniform 

across the physiographic provinces. Provinces with net 

declines that were higher than the regional averages are in 

order: Oregon Klamath (-9.9 percent), Oregon Western 

Cascades (-4.9 percent), and California Klamath (-4.1 

percent).15 Net change in OGSI 80 in eastern Oregon and 

eastern Washington Cascades, where wildfires have been 

relatively common (Davis et al. 2015), (table 3-6) were at or 

less than the regional average (e.g., -2.8 and -2.2 percent). 

While losses to fire and other disturbances get much 

attention, monitoring reveals that forest dynamics are also 

about succession, which will always at least partially offset 

losses: 757,900 ac (306 842 ha) of loss to disturbance 

appears to have been partially offset by 396,100 ac (160 364 

ha) of gain from succession (Davis et al. 2015) (table 3-6). If 

losses from timber harvest are excluded (to highlight the 

role of natural disturbance agents), those losses (609,800 ac 

15 For OGSI 200, more physiographic provinces exceeded the 

regional average of -2.8 percent net change: Washington western 

lowlands = 7.0 percent; Oregon western Cascades = - 6.0 percent; 

Oregon Klamath = -10 percent; California Coast Range = -3.0 

percent; California Klamath = -7.9 percent. table 3-8. From Davis 

et al. (2015).
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

[246 882 ha]) from all disturbance agents drop to 4.7 

percent from 6.0 percent as gains from succession replaced 

about 65 percent of those over 20 years. Some provinces 

(e.g., Washington western Cascades, Oregon Coast Range, 

California Coast Range, and California Cascades) actually 

showed a net increase in OGSI 80 on federal lands (Davis et 

al. 2015) (table 3-6). 

At the scale of individual LSRs, the range in net 

changes in OGSI 200 forests ranged widely (from -52 to 

>100 percent) (fig. 3-31) as would be expected for relatively 

small land areas. Most of the LSRs with the largest net 

changes are relatively small in area, with the exception of 

those in the Klamath regions of Oregon and California, 

where large patches of high-severity fire have occurred in 

the past 20 years. Three reserves in the eastern Cascades of 

Washington show relatively high rates of net loss, but all of 

these are relatively small reserves and the total net change 

in this province is about the regional average. The majority 

of the LSRs show little or no change. In general, large 

reserves have been more stable than smaller ones (fig. 3-32), 

which was why some of the largest reserves were drawn in 

fire-prone areas during FEMAT. 

If rates of loss of dense old-growth were much higher, 

LSR function would be threatened because they were 

designed to be dominated by dense, complex older forests 

and serve as stepping stones for connectivity of old-forest 

species across the NWFP area. The loss of large areas of 

older forest in one or more of these reserves could challenge 

the connectivity design functions; however, no research has 

investigated the degree of change in the reserve network 

that might affect its overall function. At the recent rate 

of net change (-0.15 percent per year) (Davis et al. 2015) 

(table 3-6), the original matrix and reserve system appears 

sufficient to maintain areas of OGSI 80 at a regional scale, 

with greater declines (-0.23 percent per year) in the dry 

forests. This is especially so if it is assumed that the rate 

of ingrowth into denser older forest types will increase 

dramatically in coming decades as large areas of younger 

plantations and early 20th century wildfire-initiated stands 

begin to reach the age and structure where old-forests characteristics appear (Davis et al. 2015). However, the current 

trends may not hold given that fire activity is projected to 

increase across the NWFP area. With increasing drought 

fire sizes, including patches of high-severity, fire may 

increase (Reilly et al. 2017). Projections of the amount of 

increase in area or size of fires differ considerably across 

the NWFP area and among studies. For example, Stavros et 

al. (2014) found that the probability of very large fires will 

increase for Oregon and Washington, but increases would 

be minor in northern California. Littell et al. (2010) found 

that area burned is likely to increase by two to three times 

for Washington. Ager et al. (2017) modeled increases in 

fire and their effect on northern spotted owl habitat and fire 

regimes in the eastern Cascades of Oregon. They found that 

increases of two to three times in rates of wildfire would 

reduce spotted owl habitat by 25 to 40 percent within 30 

years. They also found, however, that as fire increased, 

negative feedbacks on fire area and intensity occurred, 

suggesting that as fire increases, fuel limitations would 

affect future fire behavior. Most climate projection studies 

focus on area burned and not on severity and do not include 

fire feedbacks. Studies are needed to evaluate how climate 

change and fire might affect the LSR network conservation 

goals for different network configurations and management 

guidelines (e.g., levels and types of restoration).

While understanding annual rates of change in LSRs 

during the past 23 years is important to assessing Plan 

outcomes, it is also important to acknowledge that annual 

rates of disturbance or loss over short periods of time (e.g., 

23 years) have limited value in the infrequent, high-severity 

regimes and across all regimes given climate change. Large 

fire or wind disturbances may be rare or episodic in infrequent regimes but can strongly control landscape dynamics 

and leave legacies that persist for centuries or longer (Foster 

et al. 1998, Spies and Franklin 1989). The real test of the 

reserve network can only be done over very long periods of 

time, and ultimately managers will have to be prepared for 

surprises and inevitable large events. Knowledge of trends 

and annual rates of change are useful but are of limited 

value for predicting the future in ecosystems, where fire, 

wind, volcanic eruptions, earthquakes, or invasive species 

can change forests rapidly over large areas. 

The “losses” of late-successional/old-growth structure 

in reserves to fire may be a loss from the perspective of 
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GENERAL TECHNICAL REPORT PNW-GTR-966

Figure 3-31—Map of 192 late-successional reserves (LSRs) in the Northwest Forest Plan area showing percentage of net 

change (gain or loss) in old-growth structure index (OGSI) 200 from 1993 to 2012. The LSRs are color coded by degree 

of gain (blue) or loss (red). The LSRs with little net change are shown in gray. Pie charts only show LSRs with greater 

than 20 percent net change (e.g., annualized rate of 1 percent), either gains or losses. Colored sections and numbers in pie 

charts indicate percentage of OGSI 200 in LSRs that was gained or lost. Percentages can exceed 100 percent where gains 

occur. Data based on Davis et al. 2015.
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

conservation of dense older forests, but they do not necessarily represent a loss from a broader biodiversity perspective (e.g., ecosystem integrity), especially where those fires 

burn at lower severities and thin out understories, leaving 

lower densities of fire-tolerant species. This is especially the 

case in dry forest landscapes, where open old growth and 

mosaics of old and early successional were characteristic. 

However, as mentioned above, the OGSI thresholds in 

frequent and very frequent fire regimes were based on plots 

from existing older forests that have been subject to fire 

exclusion and succession that would have increased stand 

density, layering, and amounts of shade-tolerant and 

fire-intolerant species. Hence, the reference conditions for 

older forests do not typically represent the older forest 

structure and composition types that developed under more 

frequent fire regimes. Large fires such as the 2002 Biscuit 

Fire often have less than 20 percent of their total area in 

high-severity patches and have large areas of historically 

moderate to low severity (Reilly et al. 2017, Thompson and 

Spies 2009). Lower and moderate-severity wildfire shifts 

stands from dense old forests to more open old forests (i.e., 

thins out understories but leaves many of the older firetolerant trees) that were characteristic of forest structure and 

composition under frequent fire regimes (Kane et al. 2013). 

However, monitoring and inventory definitions for these 

more open older forest types do not exist (Spies et al. 2006b, 

Taylor and Skinner 1998) and were not applied in the 

monitoring program.16 Reilly and Spies (2015) classify 

forest structure in the NWFP area using existing inventory 

plots and identify conditions that may approximate the 

historical structure of more open old-growth forests. The 

lack of focus on open types of old growth was probably the 

result of the original emphasis of the NWFP on dense 

late-successional old-growth forest habitats of the western 

Cascades of Oregon and Washington which are associated 

with northern spotted owl and other species. 

16 The OGSI for pine types was based solely on density of large 

live trees, which may approximate historical amounts, but they do 

not include canopy cover and layering. 

-100

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Loss

Gain

Figure 3-32—Relative change in old-growth strucuture index (OGSI) 200 in reserves between 1993 and 2013 in relation 

to late-successional reserve (LSR) size. Reserves smaller than 50,000 ac (20 224 ha) tended to show more change than 

larger reserves.
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GENERAL TECHNICAL REPORT PNW-GTR-966

Concerns— 

Although general trends revealed by monitoring at the 

regional scale appear consistent with NWFP goals and 

expectations, there are other less obvious trends that may 

be cause for concern in dry forests. First, in the Klamath 

Mountains and other regions, where chaparral and other 

shrub species are an important component of the vegetation, an increase in size and frequency of high-severity 

fire patches can lead to more extensive areas of early-seral 

or chaparral vegetation that can become a semipermanent 

landscape feature (Lauvaux et al. 2016, Tepley et al. 2017). 

It is not clear how much of this type of change would be 

desirable to meet ecological or social goals, and management may be needed to promote succession toward trees 

that are resistant to fire and climate change. On the other 

hand, Donato et al. (2011) suggested that low-density conifer regeneration in the presence of hardwoods and shrubs 

is an alternative successional pathway to promote early 

development of old, complex old-forest structure. 

Very large patches of high-severity fire also occur in 

other low- and mixed-severity forest types in the NWFP 

area (Hessburg et al. 2016) with the possibility that recovery 

to forest is slowed or precluded as a result of lack of conifer 

seed rain (Dodson and Root 2013). This is especially in 

large reburn patches and may require planting to mitigate 

these effects (see restoration section below). The degree 

to which large patches of high-severity fire are slowing 

forest succession after recent large fires in the NWFP area 

is not known. On the other hand, relatively large patches of 

high-severity fire can result in areas of nonforest vegetation 

(e.g., grasslands and shrub lands) that were more common 

in the past than today in many dry forest landscapes (figs. 

21, 22, and 26). 

A second concern in dry forests is that older forests and 

landscapes in reserves and outside of reserves are slowly 

transitioning to conditions characterized by denser forests, 

more shade-tolerant species, buffered microclimate (less 

wind and shaded and cooler forest), and less flammable 

fuel beds. Thus, they become less likely to burn under low 

to moderate weather conditions and more likely to burn 

under high-severity conditions. Assuming continued fire 

suppression (Calkin et al. 2015, Stephens and Ruth 2005) 

and increased warming, the forests of the reserves in 

mixed- and low-severity regimes will continue to change 

in ways that do not support the historical dynamics of these 

forest types. 

On balance, the science reveals that fire-dependent 

forests in LSRs are continuing to be squeezed into altered 

states and dynamics by two forces: (1) succession toward 

historically unprecedented structure, composition that 

affects biodiversity, landscape structure (e.g., larger 

more connected dense forest patches), and ecosystem 

function in absence of fire; and (2) a shift toward much 

less frequent but higher severity fire regimes as a result of 

fire exclusion, climate change, and changes in vegetation, 

including increased fuel loading and contagion. Losses 

of old growth and owl habitat to high-severity fire are the 

focus of the current monitoring reports and strategies, 

and succession toward dense forests with shade-tolerant 

species (e.g., owl habitat) is typically considered a positive 

outcome relative to the goals of the NWFP. However, 

within the dry forest zone and some drier parts of the 

moist forest zone, these types of forests are not a desirable outcome if the goal is ecological integrity based 

on frequent fire, open fire-resilient old growth, diverse 

successional conditions, and disturbance processes and 

landscape dynamics that maintain resilience and a full 

complement of native biodiversity. Landscape-scale 

research and strategies are needed to find options that 

provide for late-successional species while improving the 

overall resilience and functions of dry forests (Hessburg 

et al. 2016; Sollmann et al. 2016; Spies et al. 2006, 2017). 

Frameworks based on knowledge of ecological history or 

on NRVs or the HRV and departure from those references 

(Haugo et al. 2015) could be used to guide development 

and implementation of alternative approaches for dry forests to meet the goals of the NWFP and the 2012 planning 

rule. For more discussion of reserves and possible alternatives to static reserves, see chapter 12.
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Connectivity and Fragmentation

Connectivity and fragmentation of mature and old-growth 

forests were important considerations in developing the 

NWFP (FEMAT 1993). The spatial pattern, size, and 

isolation of habitat patches of older forests can affect species 

richness, population dynamics, as well as the spread of fire 

and other disturbances. Davis et al. (2015) found that older 

forests on federal lands have become slightly more fragmented by disturbance over the period of the Plan. However, this analysis only takes into account late-successional 

and old-growth conditions and does not factor in changing 

connectivity relations over the remainder of the landscape, 

which may be the larger story. Consequently, it is not clear 

what the cumulative ecological effects (e.g., species richness, microclimate) of spatial pattern changes have been 

as a result of disturbance and succession over the past 20 

years. It is now recognized that the ecological effects of spatial pattern of vegetation types and successional stages (e.g., 

edge effects, patch size effects, connectivity) differ with 

species and processes and are difficult to generalize about 

using a coarse-filter approach (Betts et al. 2014). Cushman 

et al. (2008) found that maps of existing forest cover types 

and successional stages in the Oregon Coast Range were not 

effective in estimating abundances of breeding birds and 

cautioned that maps based only on coarse vegetation classes 

may not provide a good metric of species abundance. If 

maps of vegetation types have limitations for conservation, 

then the analysis of spatial pattern is also likely to have 

limited value for predicting community or species outcomes. Fahrig (2013) has recently hypothesized that habitat 

amount is a better predictor of species richness than patch 

size and isolation for community-scale (i.e., coarse-filter) 

approaches to conservation. However, this does not mean 

that patch size, isolation, and connectivity are not important 

components of habitat at the scale of individual species 

(e.g., fine filter) or for key processes. The implication for 

the NWFP is that patch size and connectivity concerns are 

best dealt with at the individual-species scale (e.g., northern 

spotted owl, carnivores) or processes (e.g., fire spread 

through landscapes). The question of connectivity for late 

successional/old growth as a coarse-filter metric and even 

use of maps of late successional/old growth to represent 

“habitat” in general (e.g., concern of Cushman et al. 2008) 

is an area of uncertainty and needs research. See chapter 12 

for more discussion of regional-scale issues. 

Restoration Approaches

Here we address our scientific understanding of management actions that could be used to achieve goals for ecosystem restoration, especially those related to successional 

diversity and natural disturbance regime processes. We use 

a loose definition of restoration given that climate, landscape, and species changes make it from difficult to impossible or perhaps undesirable to really restore the structure, 

composition, and function of past ecosystems (Spies et al., 

chapter 12). Ecological restoration has been defined as “the 

process of assisting the recovery of an ecosystem that has 

been degraded, damaged, or destroyed” (http://www.ser.

org/resources/resources-detail-view/ser-international-primer-on-ecological-restoration). Despite the limitations of restoration, management can promote resilience of ecosystems 

to fire or climate change or increase vegetation diversity 

that has been lost as a result of management actions such as 

timber management or fire suppression. Restoration may be 

able to promote some of the features of the pre-Euro-American period (e.g., dead wood, large fire-resistant trees, or multistoried old-growth habitats), but ecosystems may not have 

the same overall structure and function (or even fall within 

their historical ranges) as those of the pre-Euro-American 

period. We address these management actions by forest zone 

and disturbance regime, acknowledging that these ecological management approaches may be similar across regimes. 

Numerous authors have addressed restoration needs specified in the NWFP (Baker 2012; Franklin and Johnson 2012; 

Franklin et al. 2008, 2013; Haugo et al. 2015; Hessburg et al. 

2016; North et al. 2009, 2012; Stephens et al. 2009; Stine et 

al. 2014). In general, these restoration needs are to restore 

disturbance processes (e.g., fire) and longer times for natural 

succession to operate without disturbance (Haugo et al. 

2015) as young forests develop following logging (table 3-5). 
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GENERAL TECHNICAL REPORT PNW-GTR-966

Moist forests— 

Stand scales—Forest plantations are the primary focal 

point of restoration in these forests. Approaches to restoring 

old-growth forest conditions in plantations include: 

• Passive management—increasing the amount of older 

forests by electing to simply allow younger postlogging forests to naturally progress, through growth 

and mortality to older life stages (Haugo et al. 2015). 

• Active management—using variable-density thinning 

(restoration thinning) (Carey 2003, Churchill et al. 

2013, Haugo et al. 2015, Muir et al. 2002) to increase 

structural and compositional diversity in unnaturally uniform plantations that reduced typical shrub 

and herb layers and accelerate development of future 

mature and old-forest structures (figs. 3-33 and 3-34). 

Currently, the most common approaches are to allow 

younger stands to age and mature on their own and to 

use variable-density thinnings (i.e., restoration thinning) 

to increase habitat diversity within uniform plantations 

(especially 30- to 80-year-old stands, where thinning is 

typically profitable) and thus accelerate the development of 

older forest structure and composition (Carey 2003) (figs. 

3-33 through 3-35). They can also be used to promote elk 

habitat, huckleberries, and other species associated with 

forest openings (chapter 11). While restoration thinning is 

a relatively new practice for ecological goals, the effects of 

standard thinning (Tappeiner et al. 2007) on tree growth 

and mortality in regular-spaced plantations are relatively 

well known. For example, growth-growing stock relationships for Douglas-fir suggest minor differences in stand 

volume growth over a range of residual densities (Marshall 

and Curtis 2002), which provides some flexibility in terms 

of thinning prescriptions (Dodson et al. 2012). However, 

extremely low residual densities and gap creation obviously 

lead to lower stand-level tree growth. However, where 

stand-level foliage biomass is concerned (which is important 

for tree growth and litter production), thinning can stimulate growth of foliage biomass on a branch and tree scale, 

which may not be a desirable outcome from a restoration 

perspective where reducing canopy fuels is a goal (Ritchie 

et al. 2013a). Decreases in stand growth owing to low tree 

numbers are partially offset by better growth of residual 

trees (Dodson et al. 2012), and by establishment and growth 

of regenerating trees.

Given the recency of restoration thinning practices 

and studies, our understanding of how this practice 

affects older forest development is based on only shortterm results (typically less than 20 years) (Poage and 

Anderson 2007). To understand possible ecological 

effects, we extrapolate from the many studies of standard 

thinning operations, which suggest that such approaches 

would not produce many of the outcomes associated with 

old-growth forests (e.g., spatial heterogeneity, large dead 

trees, compositional diversity) in the short term (up to 50 

years), other than larger diameter trees (Anderson and 

Ronnenberg 2013). 

In contrast to standard thinning operations, restoration 

thinning includes preferentially retaining minority species 

and creating a wider range of density conditions from open 

gaps to unthinned patches of various sizes (Carey 2003, 

Davis et al. 2007, Neill and Puettmann 2013). This appears 

to be key to increasing the heterogeneity in thinned stands 

and accelerating development of late-successional elements 

(Anderson and Ronnenberg 2013, Cissel et al. 2006, 

Poage and Anderson 2007). Also, the initial responses to 

variable-density thinning treatments suggest that not all 

structural components and processes react in synchrony 

(Puettmann et al. 2016). For example, one study found 

that after a brief delay, likely due to increases in crown 

size (Ruzicka et al. 2014), restoration thinning led to an 

increase in average-tree-diameter growth. However, larger 

trees, which would likely become the dominant trees 

that are the major features of an old-growth stand, barely 

responded unless they were growing in extremely low 

densities, e.g., adjacent to gaps (Davis et al. 2007, Dodson 

et al. 2012). Also, diameter growth responded rather 

quickly within the first 5 years, while changes in other 

vegetation components were slower or delayed, such as in 

crown structures (Davis et al. 2007, Seidel et al. 2016) or 

bark furrows (Sheridan et al. 2013). That study also found 

that other vegetation components followed a counterproductive trend relative to late-successional/old-growth biodiversity goals. For example, the shrub layer was knocked 

down during harvesting operations and did not recover to 
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preharvest levels within the first decade (Puettmann et al. 

2013). Also, the understory vegetation composition shifted 

toward a higher component of early-successional species. 

This trend started to reverse within a decade (Ares et al. 

2009, 2010) but was still detected 20 years after a precommerical thinning (Lindh and Muir 2004). Exotic species 

remained a minor component after restoration thinning 

and showed a similar trend of decline after a decade. With 

little postharvest mortality after thinning, snag recruitment 

was reduced 11 years after thinning (the time of the last 

measurement) (Dodson et al. 2012) and likely in the longer 

term as well (Garman et al. 2003, Pollock and Beechie 

2014). This trend can be counteracted by creating snags 

(Lewis 1998); however, if this is done during restoration 

thinning, these snags would be smaller and shorter than 

in older stands. Alternatively, leaving untreated patches 

of high tree density ensured that competition-related 

mortality continued, allthough this led to snags at the 

Figure 3-33—Aerial image from 2011 of management units and unmanaged stands in an area of late-successional and 

riparian reserves and matrix allocation on the Siuslaw National Forest, and private lands in the Oregon Coast Range: (A) 

plantations treated with variable density thinning, (B) uniform plantations that have not been thinned (these plantations are 

younger than those that have been treated), (C) recent clearcuts on private land, and (D) older naturally regenerated forests 

that have not been managed. Note areas of hardwood and shrub gaps in the older conifer forests that occur in root rot (Phellinus sulphurascens pockets). Roads are indicated by white lines. From Oregon Explorer Natural Resources Digital Library.
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GENERAL TECHNICAL REPORT PNW-GTR-966

smaller end of the size distribution (Dodson et al. 2012). 

Tree regeneration typically increased right after restoration 

treatments (Dodson et al. 2014, Kuehne and Puettmann 

2008, Urgenson et al. 2013), showing three general trends. 

First, while stand-level differences were obvious, studies 

showed very high spatial variability at small spatial scales. 

Second, seedling establishment increases after thinnings, 

but densities appeared to be similar, regardless of thinning 

intensities. Third, seedling and sapling growth differed 

by species and responded to higher degrees of overstory 

removal (e.g., Shatford et al. 2009).

The benefits of restoration thinning relate as much or 

more to increasing spatial heterogeneity as to reducing 

density per se, as high-density patches are not uncommon in 

natural stands. For example, Spies and Franklin (1991) 

reported that stand densities (trees >2 inches [5.1 cm] diameter at breast height) in young stands (40 to 79 years old) that 

regenerated naturally after wildfire in western Washington 

and Oregon averaged about 400 stems per acre (1,000 stems 

per hectare.). Some plantations 40 to 60 years old that 

regenerated naturally after logging (Curtis and Marshall 

1986) or following clearcutting and planting can have similar 

densities, though plantations with much higher densities (e.g., 

800 stems per acre [~2,000 stems per hectare) occur.17 In 

some places, natural regeneration (e.g.,, western hemlock) 

will establish itself in Douglas-fir plantations (Puettmann, 

personal observation) leading to extremely high tree densities. 

While average tree density can be high in plantations, density 

differences do not explain all potential differences between 

natural young stands and plantations. The differences are also 

17 Pabst R. Personal communication. Senior faculty research 

assistant, College of Forestry, Oregon State University, Corvallis, 

OR 97331.

Figure 3-34—Canopy (fisheye) and understory photographs of unthinned and thinned 30 to 40 year old plantations of Douglas-fir on the 

Siuslaw National Forest. Densities of four stands from left to right: unthinned; 100 trees/acre; 60 trees/acre; and 30 trees/acre.

Sam Chan
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Figure 3-35—Example of variable-density thinning from 2013, including skips and gaps (1 to 2 ac [0.40 to 0.80 ha]), in a 56-year-old 

plantation on the Willamette National Forest: (A) the pattern across the entire treatment area and the surrounding unthinned plantation, 

(B) a view from inside the thinned area, and (C) the view looking across the gap. The goal was “volume production, promotion of 

high-quality elk forage in the short term, while encouraging development of elk-optimal cover.”

Gary RostGary Rost Google EarthTM

A

B C
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GENERAL TECHNICAL REPORT PNW-GTR-966

expressed in spatial variation in density and variability of tree 

age and size (Tappeiner et al. 1997). The age ranges and 

spatial heterogeneity of trees in naturally regenerated stands 

may lead to greater variability in canopy differentiation than 

would occur in plantations where trees are the same species, 

the same age, and are planted with uniform spacing (Oliver 

and Larson 1990). A combination of tall shrubs, hardwoods, 

or other vegetation would have occupied much of the open 

growing spaces (i.e., spaces not occupied by conifer regeneration in naturally regenerated stands). The short-term effects of 

variable-density thinning aimed at improving longer term 

structural and compositional diversity may be to fragment 

canopies and root systems and temporally reduce habitat 

quality for animal, plant, or fungal species keying in on 

canopy and root structure (Davis and Puettman 2009, Pilz et. 

al. 2006). This is an important issue requiring more research. 

Alternative ways of implementing thinning prescriptions (e.g., 

leaving larger unthinned areas or thinning very young stands) 

may actually improve conditions for lichens (Root et. al. 2010) 

and may help to mitigate some of the short-term negative 

effects of discontinuous forest canopies on canopy species 

(Wilson and Forsman 2013).

Empirical studies are critical, but evaluating longterm and landscape-level effects of variable-density 

thinnings requires landscape simulation models. Traditional growth and yield models provide fairly reliable 

information about tree growth for more or less evenly 

spaced, even-aged Douglas-fir plantations (Fairweather 

2004). Most models assume the absence of disturbances, 

but ongoing efforts include a better representation of 

disturbance (e.g., insects and pathogens) on tree and stand 

growth (Crookston and Dixon 2005). Predictions for open 

or irregular-spaced conditions (Lord 2005) and growth 

of other species are less reliable or missing (Gould et al. 

2011, Kuehne et al. 2015, Weiskittel et al. 2007). Similarly, 

there is a broad understanding and agreement about 

general trends, e.g., in understory vegetation, but specific 

dynamics cannot be modeled with high precision because 

they are based on interactions of initial conditions, species 

traits, local environmental conditions, and stochastic 

events (Ares et al. 2010, Burton et al. 2014), which may 

vary over time (Thomas et al. 1999) and space (Burton et 

al. 2014, Chen et al. 1992). 

In the few modeling studies (Garman et al. 2003, Pollock 

and Beechie 2014), thinning promoted the development of 

large boles, vertical diversity, and tree-species diversity over 

100+ years, compared to controls. At the same time, less 

dead wood was produced over many decades compared to no 

thinning, highlighting that at least some of the early trends 

found in the experimental studies (e.g., Dodson et al. 2012) 

may last longer. As mentioned above, the negative effects of 

thinning on deadwood production can be countered by creating snags (Lewis 1998) or leaving cut trees on the sites where 

they can immediately contribute to terrestrial and ecological 

functions (Huff and Bailey 2009, Walter et al. 2005). 

Thinning has variable effects on wildlife and plant 

communities. In the short term, it can increase species 

diversity and abundance of some species, especially those 

associated with more open forest conditions (Ares et 

al. 2009, Berger et al. 2012). This can lead to increased 

flowering and seed productions, i.e., provision of food 

resources for selected insects, mammals, or songbirds (Neill 

and Puettmann 2013, Wender et al. 2004). The response of 

songbird populations showed similar trends (Hagar et al. 

2004), but responses appear to vary by species and over 

time (Yegorova et al. 2013). Thinning may also attract avian 

predators that prey on marbled murrelet (Brachyramphus 

marmoratus) nests (chapter 5). 

Although general stand-level trends from restoration 

thinning are mostly understood, uncertainties remain. For 

example, vegetation development for specific locations 

appears partially unpredictable for several reasons, including microclimatic conditions, initial variability in plantations, and stochastic events such as seed crops, disease, and 

windthrow (Dodson et al. 2012, Lutz and Halpern 2006). 

In addition, there are important effects of thinning on 

residual trees, such as harvesting damage to residual trees. 

Damage is typically higher the more wood is harvested 

and often concentrated near skid trails (Han and Kellogg 

2000). Through careful layout and logging (e.g., Picchio et 

al. 2012) and avoidance of early summer harvests, damage 

can be reduced to levels that are not likely to affect future 

health of Douglas-fir stands (Bettinger and Kellogg 1993, 

Kizer et al. 2011). However, other species such as western 

hemlock may be more affected (Hunt and Krueger 1962). 

With proper logging layout, techniques, and timing (e.g., 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

avoidance of wet soil conditions), the impact of thinning 

operations on soils should be limited to removal of humus 

and upper soil layers (Froehlich et al. 1981). However, 

these impacts that are concentrated near or in skid trails 

are only temporary as patches of exposed soils are reinvaded quickly.18 In this context, harvesting operations that 

removed limbs and crowns before skidding (and in some 

cases limited maximum log length that could be skidded) 

not only scattered down wood throughout the stand, but led 

to lower soil damage, as well as lower damage to residual 

trees (K.J. Puettmann, personal observation). 

In summary, ecosystem dynamics after restoration 

thinning are generally predictable, but specific responses 

can be highly variable owing to small-scale variability in 

environmental conditions and initial vegetation composition. In addition, other factors, such as weather patterns; 

seed availability; impacts of insects, diseases, and herbivores on seed or seedlings; as well as harvesting impacts as 

described above, suggest that restoration treatments are not 

likely to hit any specific target perfectly in terms of vegetation conditions and dynamics. Instead, restoration efforts 

may be better off acknowledging these inherent uncertainties by setting structural goals that allow for a range of conditions; e.g., between 10 and 30 percent of the restored area 

should have regeneration at a density from 50 to 500 trees 

per acre. Similarly, rather than locking in a spatial layout of 

prescriptions, any treatment prescription that can accommodate already existing variability within the homogenous 

stands that are to be restored will likely be more efficient 

at increasing heterogeneity in that stand (Puettmann et al. 

2016). For example, a goal to provide more broadleaf shrubs 

and trees may be achieved more easily with prescriptions 

that protect existing patches of broadleafs during harvesting than by creating open conditions that facilitate their 

development (Davis et al. 2007). Similarly, the provision of 

snags may be more efficient if it accounts for the harvesting 

damage to residual trees. Finally, flexibility in restoration 

prescriptions and adequate monitoring is key to efficient 

and successful operations.

Landscape scale—Landscape-level effects of restoration 

thinning are not well-studied, and experimental studies are 

very difficult at this scale. In a simulation study, thinning in 

plantations on federal ownerships increased habitat for olivesided flycatchers (Contopus cooperi) but had only a slight or 

no effect on total habitat for northern spotted owls and other 

associated late-successional species (Spies et al. 2007a). The 

lack of effects on habitat of owls and other late-successional species was probably due to several factors, including a 

relatively short simulation period (100 years) compared to the 

several hundred years needed for old growth to fully develop. Also, the thinning prescriptions were conservative, the 

number of thinned trees retained for dead wood recruitment 

was fairly low, and the proportion of landscape thinned in the 

first 10 years was limited to less than 8 percent of the entire 

federal landscape (Spies et al. 2007a). The scope of landscape-scale restoration benefits is also limited by the state 

and rate of succession in the population of plantations. While 

young plantations cover up to 30 percent of federal forest 

ownerships, not all of them have the structure (high density of small and relatively young conifers) that would benefit 

from restoration thinning. Also, even with increased resources, it likely will take decades to treat an area that is sufficiently large enough to have a major landscape-level impact, 

especially as some of the ecological benefits do not show up 

instantly but develop slowly over time. Lack of information 

about the structural and compositional conditions of plantations (and location amount of restoration treatments) as well 

as limited understanding of the importance of fragmentation 

and connectedness across the region limit our ability to assess restoration needs and potential at landscape scales. 

A byproduct of any large-scale restoration program 

is the need to maintain or even increase infrastructure. 

Road systems and associated travel, which are needed for 

various management objectives, have also been shown to 

negatively affect terrestrial and aquatic biological diversity 

and ecosystem processes (Forman and Alexander 1998, 

Trombulak and Frissell 2000) by serving as travel corridors 

for invasive species (Parendes and Jones 2000), for example. 

Consequently, scientific reviews note that reducing roads 

through decommissioning is important for meeting many 

biodiversity goals (chapter 7) (Franklin and Johnson 2012, 

Trombulak and Frissell 2000).

18 Unpublished data. On file with: K.J. Puettmann, Oregon State 

University, Forest Ecosystems and Society, 301L Richardson Hall, 

Corvallis, OR 97331.
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The 80-year rule—Under the NWFP, harvesting for any 

goal, including thinning for old-growth restoration, is generally restricted in moist forests in LSRs to stands less than 

80 years old (USDA and USDI 1994: c-13) (though some 

exceptions may occur). The NWFP allowed management in 

stands >80 years old in the matrix lands. This 80-year rule 

for LSRs is a one-size-fits-all approach that does not take 

into account that stand age is only a rough proxy for stand 

structure and development potential, both of which can 

differ greatly based on site conditions and history (Pabst 

et al. 2008, Reilly and Spies 2015) (fig. 3-15). That said, in 

general, treatments of stands >80 years old are not expected to result in substantial short- or medium-term shifts in 

developmental trajectories, as characterized by size and 

shape of trees and crowns, because trends established early 

in a tree’s life are not easily reversed (Wilson and Oliver 

2000). Understory vegetation would be more responsive. 

In that context, restoration thinning to promote development of complex older forest structure (e.g., large live 

and dead trees in stands >80 years old) of moist west-side 

forests is less likely to have large benefits for development 

of old-growth forests in the long term compared to younger forests, as many stands around age 80 begin to have 

some characteristics of older forests (Spies 1991, Spies and 

Franklin 1991) (fig. 3-15). 

Our scientific understanding of the ecological effects 

of restoration thinning in older forests has not changed 

much since the early 1990s, as few empirical studies 

and modeling of management in older forests have been 

conducted (see Cissel et al. 1999 for a landscape-level 

modeling study). Removing larger trees could have negative impacts on the number of large live and dead trees, 

as trees over this age are often beginning to function 

as habitat for late-successional species in middle-aged 

stands; e.g., they develop bark characteristics that may 

act as microhabitat for a variety of species (Sheridan et 

al. 2013). However, the age, or better, the set of structural 

conditions (e.g., density, spatial pattern, size distribution) 

at which such negative impacts become important will 

differ with tree, stand, site, and landscape conditions, and 

such relationships have not been quantitatively tested. 

Research and adaptive management studies are needed to 

test and evaluate the alternative approaches and assess the 

relative benefits and tradeoffs of restoration thinning in 

forests >80 years old. 

Fire and early-successional vegetation—Possible activities relative to restoring or emulating the beneficial effects 

of wildfire in moist forests include creating early-seral 

forest and creating some of the effects of partial stand-replacement fire that were common in mixed-severity regimes 

of the drier part of this region. There is relatively little 

research and management experience with either of these 

activities. Managing wildfire to promote desirable fire 

effects may be increasingly feasible in the dry forests and 

remote areas of the wetter forests. However, relatively little 

is known about public perceptions of risk in moist forests 

and their willingness to tolerate wildfire in remote areas, 

but they do understand that any fire in moist forest is likely 

to be “catastrophic” (Hall and Slothower 2009). This leaves 

mechanical treatments and prescribed fire as the primary 

way to schedule and produce fire effects. The first problem 

in creating early-seral vegetation is determining where to 

create these habitats on a landscape that has already experienced a significant decline in old forests from clearcutting. 

Creating early-seral habitat from older forests is possible 

(Cissel et al. 1999, Hansen et al.1993) and would most closely mimic natural processes that have been disrupted; however, such treatments could also reduce habitat for at-risk, 

older forest species and have encountered public resistance 

(Franklin and Johnson 2012). Consequently, Franklin and 

Johnson (2012) suggested that forest plantations (<80 years 

old) be the primary focus of any efforts to create early-seral habitat. Heavy partial harvest (i.e., retention harvest), 

leaving dead trees and islands of live trees, and prescribed 

fire would constitute an approach to creating early-seral 

vegetation in plantations and create variable within- and 

between-stand patterns for late-seral development. Such efforts would be a compromise between how wildfires would 

have created such communities—they would lack large live 

and dead trees, might not have some of the same ecological 

effects of fire on soil surfaces and vegetation, and would 

not occur in very large patches—but they would restore 

some components and values of this ecosystem. Combining 

plantations into large groups would help address the patch 
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size issue. A larger problem is how to determine how much 

of this vegetation should be created and how to schedule 

and distribute it in landscapes where wildfires could appear 

in any year and create thousands of acres of this vegetation 

type in a few days. 

Moderately frequent mixed-severity fire regimes—

Similarly, little published research exists on restoration in 

moderately frequent to somewhat infrequent, mixed-severity fire regimes, which occur in the drier parts of the moist 

forest zone (Tepley et al. 2013) (fig. 3-6). Managers have 

had some experience implementing treatments that attempt 

to emulate partial stand-replacement fire in older forests 

(fig. 3-29). Cissel et al. (1999) modeled stand and landscape 

management based on the mixed-severity fire regimes of 

the western Cascades of Oregon. They found that it produced more old-forest habitat and larger patches of older 

forests than would have occurred if the NWFP reserve-matrix strategy had been implemented as originally designed. 

However, it probably would have produced less older forest 

structure than if no timber harvests had occurred in the 

matrix and wildfire was suppressed. The broader ecological effects of mixed-severity fire in forests more than 80 

years old have not been studied. One hypothesis is that 

some late-successional conditions (e.g., spatial heterogeneity, species cohort composition, diameter diversity and 

development of large-diameter trees) in the drier parts of 

the western hemlock and Pacific silver fir zones are no longer developing at the same rate because lower severity fire 

would have thinned the older stands, creating gaps, initiating new shade-tolerant cohorts, and accelerating growth 

of surviving canopy trees (Brown et al. 2013, Tepley et al. 

2013, Weisberg 2004). In general, landscapes with more 

fire-severity diversity (“pyrodiversity”) (e.g., mixed-severity 

landscapes) are known to support more biodiversity (Kelly 

and Brotons 2017, Perry et al. 2011, Tingley et al. 2016). 

Landscapes with more vegetative diversity would likely 

affect the rate of wildfire spread and wildfires would create 

more heterogeneous vegetation. Research is needed to evaluate alternative approaches to restore successional diversity 

in this moist forest regime through mechanical treatments, 

prescribed fire, and wildfire. 

Ecological forestry—The “ecological forestry” approach 

(Franklin and Johnson 2012, Seymour and Hunter 1999), 

which seeks to use knowledge of disturbance ecology and 

retention-based management to achieve ecological and 

commodity goals simultaneously, has been promoted as 

a restoration approach for meeting goals of the NWFP. It 

can be applied to both moist and dry forests and is, to some 

degree, a branding of a collection of management actions 

(including those already identified for moist and dry forests 

[table 3-5]) that can be applied to meet ecological and social 

goals. Ecological forestry encompasses restoration thinning 

in plantations, prescribed fire, and retention silviculture 

(focusing on what to retain rather than on what to remove) 

to create early-successional patches in plantations or older 

forests (e.g., >80 years old) where appropriate (figs. 3-35 and 

3-36). The theory behind ecological forestry is supported 

by scientific understanding and rooted in established concepts in silviculture and ecology (Batavia and Nelson 2016; 

D’Amato et al. 2017; Franklin et al. 2007b, 2018; Seymour 

and Hunter 1999). 

No published empirical research studies exist that 

evaluate long-term ecological and socioeconomic effects 

of ecological forestry in the NWFP area. However, several 

of its components, including retention silviculture and 

disturbance-based forest management, have been evaluated 

in the Pacific Northwest and other places with shorter term 

studies. For example, global studies (Baker et al. 2016, Gustafson et al. 2012) and work in the Pacific Northwest (Halpern et al. 2012, Hansen et al. 1995a, Urgenson et al. 2013) 

show that retention silviculture can provide habitat and “life 

boats” (i.e., refugia) for older forest species (Rosenwald 

and Lohmus 2008) within patches of early-successional 

vegetation. Cissel et al. (2002) simulated a landscape-scale 

design for a watershed in the western Cascades that contained many elements of Franklin and Johnson’s ecological 

forestry approach. They found that their approach produced 

better ecological outcomes than implementation of the 

current NWFP standards and guides; however, relatively 

little empirical research has been published on this issue in 

the NWFP area. 

Batavia and Nelson (2016) recently criticized ecological 

forestry for its lack of a clear normative or ethical goal 
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GENERAL TECHNICAL REPORT PNW-GTR-966

(e.g., conserve all species, or maximize timber production). 

They suggested that this deficiency will limit its practical 

application and subject it to the same social pitfalls as earlier 

and current management concepts or frameworks for finding 

solutions to balancing ecological and social objectives, such 

as “new forestry” (Franklin 1989), ecosystem management 

(Christensen et al. 1996, Grumbine 1994, Franklin 1997), 

or sustainable forestry (Lindenmayer and Franklin 1997). 

Different world views and values appear to present a major 

challenge to the implementation and acceptance of any of 

these approaches that attempt to achieve multiple goals 

from the same stands or locations. For example, DellaSala 

et al. (2013) criticized ecological forestry on federal lands 

as placing too much emphasis on timber production and not 

enough on protecting habitat for the northern spotted owl, 

especially given the threat posed by the barred owl (Strix 

varia). At the same time, Oregon county commissioners 

are seeking higher levels of timber production, especially 

from Bureau of Land Management lands, and complain that 

ecological forestry does not produce enough timber for local 

lumber mills (Hubbard 2015). Clearly, the social aspects of 

active management to restore or create desired ecological 

patterns and processes (in any of the disturbance regimes) 

and producing socioeconomic values are as important to 

consider as the biophysical aspects (see chapter 12 for more 

discussion of the tradeoffs and value issues). 

Figure 3-36—Management unit designed to create a mosaic of early habitat and leave trees, and produce wood from a young Douglas-fir 

forest on Bureau of Land Management (BLM) land in western Oregon. VRH = variable-retention generation harvest.

Bureau of Land Management
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Dry forests with frequent, mixed-severity fire regimes—

Restoration approaches in both fire regimes of the dry 

forests include mechanical treatments and use of fire in 

plantations and older forests to restore or create seral stages, 

surface fuel beds, forest density conditions, and spatial 

patterns of trees that are more resistant and resilient to fire 

and better adapted to warming climate. Restoration strategies for the frequent mixed-severity regime in the area of 

the NWFP have recently been summarized in Hessburg et 

al. (2016) who provide an indepth review. Restoration challenges are large in this regime because of the complexity of 

successional pathways and variable disturbance patterns. 

The management strategies outlined include:

• Restoring pyrodiversity at landscape levels through 

prescribed fire and managed wildfire.

• Creating and maintaining successional heterogeneity 

based on local disturbance regimes and the needs of 

late-successional forest species.

• Using topography to tailor restoration treatments 

across landscapes.

• Protecting and restoring large and old, early-seral 

fire-resistant trees.

• Restoring diversity to plantations.

• Creating and maintaining early-seral vegetation, 

including grasslands and shrublands. 

• Mitigating threats from climate change, forest 

insects, and pathogens.

Prescribed fire and wildfire—The literature on restoring forest fire regimes indicates that prescribed fires and 

wildfires managed under moderate conditions are vital 

components of ecological restoration. Thinning and other 

mechanical manipulations can achieve many structural and 

composition restoration goals. However, they cannot replace 

many important ecological processes and effects of fires, 

whether prescribed or wild (McIver et al. 2013). Fire, in particular, reduces surface fuels and coarse woody debris and 

can both increase and decrease snags and large-diameter 

logs depending on severity. Fire also affects soils (Certini 

2005), insects (e.g., carabid beetle) (Niwa and Peck 2002), 

and other arthropod communities (Apigian et al. 2006). 

On the other hand, fires can also lead to increases of exotic 

plant species (Keeley 2000) and weaken high-value trees as 

well as attract bark beetles (Gibson and Negrón 2009). This 

may be viewed negatively in a narrow sense, but in a larger 

ecosystem context, such indirect impacts can feed a whole 

suite of ecosystems processes. For example, larger bark 

beetle populations can attract more woodpeckers that in 

turn spread more wood decaying fungi, thus providing more 

cavities, dead and down wood and associated habitat for a 

whole suite of species. 

Prescribed fire is often implemented at least initially 

following variable-density thinning to reduce stand density. 

Here, thinning and prescribed fire can be implemented 

in denser stands with or without large fire-resistant trees. 

Such treatments can increase the range of microclimate and 

resource conditions (e.g., soil moisture, light) (Ma et al. 2010). 

For example, Dodson et al. (2008) found a neutral to positive 

treatment effect from thinning and prescribed fire on understory vegetation, while other studies showed a short-term 

decline followed by an increase (Abella and Springer 2015). 

The high variability of responses appear to reflect (among 

others) the variability in initial conditions and the scale of 

observation (Dodson and Peterson 2010), with areas of low 

understory richness benefiting most (Dodson et al. 2008). At 

the same time, such treatments would reduce the likelihood of 

very large patches of high-severity fires that are incompatible 

with ecosystem and habitat needs for many species (Harrod, 

et al. 2009, Hessburg et al. 2016, Knapp et al. 2012a). 

Landscape-scale perspectives are needed to understand 

the potential effectiveness of fuel treatments in modifying 

fire behavior. Fuel treatments affecting a small area of landscape have a low probability of intersecting a fire, given the 

relatively low frequencies of fire in these dry forests under 

full fire suppression strategy (Rhodes and Baker (2008). To 

be effective, treatments need to be widespread enough to 

influence the current level of landscape inertia (see Stine et 

al. 2014), and then be allowed to interact more commonly 

with wildfire ignitions not influenced by suppression. Spies 

et al. (2017), using a landscape dynamics model, found 

that a doubling of rates of restoration in central Oregon, 

which is still a relatively small area compared to historical 

fire frequencies, led to only a small reduction in the mean 

occurrence of high-severity fire over a projected 50-year 
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period. That study found that treatments were more effective in reducing high-severity fire years with more fire and 

that resilience of the entire landscape and the potential for 

high-severity fire was significantly lowered by higher rates 

of fuel treatment. Similar findings about the effectiveness of 

fuel treatments in altering fire outcomes have been reported 

by Loudermilk et al. (2013, 2014) for the relatively dry forests of the Lake Tahoe basin. Treatments to reduce density 

and surface fuels will need to be repeated at intervals that 

depend on the treatment intensity and productivity of the 

site (Collins et al. 2010). Given the widespread effect of 

fire exclusion, large areas will need to be treated (Hessburg 

2016), which may be difficult for administrative and social 

reasons. Strategic spatial optimization of treatments can 

improve effectiveness per unit area treated (Finney et al. 

2007), where prior commitments of land area to reserves 

or unique management allocations are minimal. Where 

major parts of the landscape are already committed to any 

management allocation that prevents optimal treatment allocation, spatial optimization efforts are essentially equivalent 

to random treatments (Finney et al. 2007)

Use of naturally ignited wildfires to achieve resource 

objectives is very important because, in most areas, current 

amounts of prescribed fire are too little to affect a sufficient 

area (North et al. 2012, 2015). Managing wildfire to promote 

ecological benefits is especially well suited for remote areas, 

with steep, complex topography, although it can become 

a more viable option in other landscapes when used in 

conjunction with prescribed fires, fuel reduction treatments, 

and footprints from past fires to create a patchwork that 

helps to contain the spread of natural ignitions to achieve 

desirable outcomes. Such fires will promote a high diversity 

of fire effects under moderate weather, including patches 

of low-, mixed-, and high-severity fires (Miller et al. 2012; 

Skinner et al., in press). Fire suppression and exclusion 

would also still be an important management tool, especially 

where dense older forest habitat conditions are desired, 

where landscapes may not yet be adapted for wildfire (e.g., 

contain many younger unthinned forests), or where human 

values are at risk from fire or smoke. Effectively managing 

wildfire depends on having moderate weather conditions 

that reduce the risk of high-severity fire effects (e.g., Estes 

et al. 2017). There are few published studies about restoring 

fire processes and structural diversity in older forests within 

the mixed-severity fire regimes in the NWFP area. However, 

examples exist from forests of the Sierra Nevada that are 

quite relevant to the dry forests of the NWFP area (Collins et 

al. 2006, 2008, 2010; North et al. 2009; North and Sherlock 

2012; van Wagtendonk et al. 2012; Webster and Halpern 

2010) and the Rocky Mountains (Holden et al. 2010; Larson 

et al. 2013; Parks et al. 2013, 2016). Among other things, 

these studies point out the importance of patch heterogeneity 

and topography as a driver in dry forest restoration. 

Landscapes and resilience to climate change—

Successional heterogeneity is a product of pyrodiversity and 

is fundamental to biodiversity and resilience of forests to 

climate change (Hessburg et al. 2016). This heterogeneity 

occurs across a range of spatial scales from tree clumps, 

patches and patch neighborhoods, to landscapes (Hessburg 

et al. 2015). Using variable-density thinning or varying prescribed fire treatments can promote heterogeneity at these 

fine scales (Churchill et al. 2013, Fry et al. 2014, Lyderson 

and North 2012). Developing landscape-scale prescriptions 

for use of thinning, prescribed fire, and managing wildfire 

can help promote landscape-scale heterogeneity. Landscape 

strategies are also important to maintaining and providing habitat for species that used dense, late-successional 

forests (Hessburg et al. 2015, 2016) or a mosaic of lateand early-successional forests (e.g., Franklin et al. 2000). 

Landscape-scale models and scenario analysis are needed 

to better understand tradeoffs associated with managing 

mixed-severity landscapes for a diversity of seral stages and 

biodiversity objectives (Lehmkuhl, et al. 2007, Roloff et al. 

2005, Spies et al. 2017). Topography can provide a valuable 

template for implementing landscape strategies in mixed-severity regimes (Hessburg et al. 2016). Topography, whose 

patterns and effects differ regionally can be used to help 

set goals for seral stages and prioritize treatment locations 

(Lyderson and North 2012, Taylor and Skinner 2003). 

Increasing resilience of forests to insects, pathogens, 

and drought can be accomplished through efforts described 

above related to managing for pyrodiversity, and successional diversity in a landscape context. Altering species 

composition can address a number of insect and disease 
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concerns, including spruce beetle (Dendroctonus rufipennis), 

laminated root rot (Phellinus sulphurescens), and western 

spruce budworm (Cholristoneura freemaii) (Hessburg el. 

2016). Thinning forests can lower the likelihood of mortality 

associated with mountain pine beetle (D. ponderosae) 

and western pine beetle (D. brevomis) (Fettig et al. 2007). 

Thinning can reduce dwarf mistletoe infestations. Strategies 

to increase resilience to climate include reducing surface and 

ladder fuels, reducing and maintaining lower tree densities, 

and restoring horizontal spatial heterogeneity in forest 

structure, including openings where early-seral species can 

establish (Churchill et al. 2013). Baker and Williams (2015) 

argued that efforts to remove most small trees may compromise resilience, because the presence of small trees can 

increase resilience to insect outbreaks, which can disproportionately affect large trees. They further argued that reducing 

stand density is not consistent with restoration of forests, 

because most dry forests were historically dense (based on 

their GLO survey, which overestimates tree densities as we 

discussed above). Allen et al. (2010) in a global review of 

drought-induced mortality found situations where mortality 

in forests increases with tree density as a result of increased 

competition, and situations where mortality was not related 

to density. Bradford and Bell (2017) examined thousands of 

forest inventory plots from the Southwestern United States 

and found that mortality during warm and dry conditions 

was related to basal area. Similarly, Guarin and Taylor 

(2005) found mortality associated with basal area and tree 

density in mixed-conifer forests of Yosemite. Both Allen 

et al. (2010) and Bradford and Bell (2017) suggested that 

thinning is one option for increasing resilience of forests 

to drought. Baker and Williams (2015) argued that forest 

resilience is a function of diverse sizes of trees and species, 

which is consistent with the literature that supports the idea 

that efforts to increase resilience should focus less on stand 

or landscape averages but focus on increasing heterogeneity 

and forest structure and composition at multiple scales 

(Hessburg 2016). 

Large, old, fire-resistant trees—The number of large, old, 

early-seral, and fire-resistant trees have been reduced in 

many areas as mentioned above. These keystone forest structures promote forest resilience to fire and climate change 

(Agee and Skinner 2005, Hessburg et al. 2016). Management 

actions for maintaining and promoting these trees include (1) 

identifying environments that support them; (2) protecting 

them from logging, crown fires, and drought stress; and (3) 

developing future cohorts through stand management practices (e.g., reducing stand densities and prescribed fire) that 

promote their regeneration, growth, and crown development. 

Plantations—Although plantations are a strong focus of 

restoration in the wetter forests, many thousands of acres 

of plantations also exist in dry forests landscapes that are 

in need of attention to promote resilience to fire and other 

threats. For example, precommercial thinning and prescribed burning can be used to reduce the near-term risk of 

loss of young, dense plantations to high-severity fire, while 

variable-density thinning can promote development of early-seral fire-resistant species where they are lacking in commercial-aged plantations (Stephens and Moghaddas 2005, 

Weatherspoon and Skinner 1995). Where desired species 

are lacking, planting may be needed (Hessburg et al. 2016). 

Where thinning is done, it will be important to treat surface 

fuels because logging slash will typically increase severe fire 

behavior in the residual stand (Huff et al. 1995, Raymond 

and Peterson 2005, Weatherspoon and Skinner 1995) unless 

trees are whole-tree yarded and slash piles are burned. 

Early-successional vegetation—To cover the full suite 

of landscape conditions found under natural conditions, 

restoration efforts in the mixed-severity regimes may also 

consider providing early-successional habitats (Haugo et al. 

2015), as mentioned above (Hessburg et al. 2016). Collins et 

al. (2010) suggested that silviculture could be used to mimic 

stand-replacing fire patches in a portion of the mixed-severity fire regime landscape. Other restoration treatments in 

older forests would not be stand replacing but may be targeted to remove at least part of the vegetation that established 

after fire exclusion, thus improving growing conditions and 

vigor for dominant residual trees (Latham et al. 2002). We 

lack research that provides guidance on how to implement 

restoration for early-seral conditions at landscape scales given that wildfires will continue to create this vegetation type, 

but early-seral conditions may highly differ from those of 

historical conditions depending on the successional stage of 
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GENERAL TECHNICAL REPORT PNW-GTR-966

the predisturbance forest. Collins et al. (2010) cautioned that 

silvicultural prescriptions may never achieve the complexity that freely burning fire can. Instead, allowing for more 

freely burning wildland fires would increase patch heterogeneity across landscapes and decrease potential for spread 

of very large high-intensity fires. However, cautions apply. 

Fires freely burning through dense layered stands produce 

very different fire effects in comparison to those where 

stands are open canopied and surface fuels are more limited 

(Miller and Urban 2000b). 

Dry forests with very frequent, low-severity regimes— 

Management approaches—The restoration needs and 

approaches (e.g., variable-density thinning, prescribed 

fire, and promotion of large fire-tolerant trees) in the very 

frequent, low-severity regime have many similarities to 

the frequent mixed-severity regime, but targets in terms of 

density, tree sizes and species, spatial patterns, and disturbance processes (e.g., frequent fire) are quite different. We 

emphasize some of the approaches that are unique to this 

fire regime. The overall needs for restoration in the very 

frequent, low-severity fire regime forests are larger given 

that fire suppression and widespread logging of large trees 

in many ecoregions has had a greater overall effect on forest 

structure and composition than in other dry zone forests; 

e.g., the larger number of fire cycles that have been missed 

owing to fire suppression. 

Guidance for restoration of forests of this disturbance 

regime can be found in Franklin et al. (2008), North et al. 

(2009, 2012), Stephens et al. (2009), Franklin and Johnson 

(2012), Franklin et al. (2013), Stine et al. (2014), Haugo et 

al. (2015), and Hessburg et al. (2015, 2016). Strategies to 

restore old hardwood components of forests and woodlands 

are described for California black oak in Long et al. (2016), 

for Oregon white oak in Devine and Harrington (2006), 

and for riparian areas in southwestern Oregon in Messier 

et al. (2012). We summarize some of the recommendations 

from these publications below. A combination of harvesting 

and fire management is important to foster regeneration 

and development of large shade- and fire-tolerant canopy 

trees, associate understory and midstory vegetation, and to 

increase structural heterogeneity (e.g., areas of relatively 

open patches with large canopy trees). In forests that have 

become denser as a result of fire exclusion, the old-tree 

component is often diminished or absent. This is especially 

prominent in drier forest areas, likely owing competition 

from the higher number of younger, competing trees (Dolph 

et al. 1995, Ritchie et al. 2008). Restoration thinning that is 

aimed at improving growing conditions for the larger trees 

appears to reverse this process (Latham et al. 2002). Thinning stands for resilience to drought and fire will require 

very low densities, especially of small trees and shifting 

composition to fire- and drought-tolerant species (Churchill 

et al. 2013). Studies by Hagmann et al. (2013, 2014, 2017) 

provide snapshots of the structure of low-density pine forests 

in central Oregon. Where large trees are lacking, sufficient 

numbers of intermediate-size trees will be needed to produce 

future large trees (Ritchie 2005). Flexible tree size criteria 

for thinning are needed to remove relatively large shade- and 

fire-intolerant trees that have developed in the past century 

of fire exclusion. It will be important to treat fuels created 

by mechanical treatments to reduce the risk of high-severity 

fire. Thinning and fuel treatments and prescribed fire should 

seek to reintroduce spatial heterogeneity into stands and 

landscapes (Haugo et al. 2015, 2016). Prescribed fire should 

aim for low levels of canopy mortality (e.g., 5 to 10 percent) 

to promote snag recruitment and spatial heterogeneity. 

In some cases, it may be necessary to plant drought- and 

fire-tolerant tree species. Topographic and soil patterns 

can provide a template for distributing treatments across 

landscapes (Hessburg et al. 2016, North et al. 2009). It will 

be important to consider understory plant communities in 

restoration plans (Franklin et al. 2013) as they have been 

severely degraded by grazing and are important for wildlife 

habitat, productivity, and providing fine fuels to promote the 

movement of low-severity surface fire through the landscape. 

For example, introducing prescribed fire after a long period 

of fire exclusion and accumulation of litter can lead to locally 

intense fires that still kill trees and rhizomatous grasses 

that are important for browse and form surface fuels that 

are needed to sustain relatively frequent surface fires. Other 

important considerations in restoration planning include 

developing efficient and effective marking guides that promote spatial heterogeneity (e.g., the individuals, clumps, and 

openings method) (Churchill et al. 2013, Franklin et al. 2013). 
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Landscapes—Landscape-scale considerations are important for altering successional patterns, general resilience to 

drought and wildfire, and for providing habitat for wildlife 

species that depend on different types of habitat, including 

dense conditions that may not be resilient to fire. Where 

restoration actions such as thinning and prescribed fire are 

done, it will be important to treat large patches to reduce 

the likelihood that treated areas will be rapidly recolonized 

by shade-tolerant tree species and certain shade-intolerant 

trees (e.g., lodgepole pine) that seed-in from nearby untreated areas. The landscape inertia (e.g., mass effects) (Stine 

et al. 2014) created by large areas dominated by shade-tolerant tree species will be a major influence on the rate 

and potential for restoring successional dynamics in these 

landscapes. Patch types and sizes differ in their susceptibility to high-severity fires and considering their patterns 

and relative abundances in landscapes is critical for restoration planning in low-severity forests and in other fire 

regimes. The following patch types are listed from highest 

to lowest susceptibility to high-severity fire (Odion et al. 

2004, Thompson and Spies 2009). Note that order is not 

necessarily the same as management priorities, which take 

multiple factors into account. Landscape context (e.g., edge 

effects, also can play a large role in determining fire severity (Weatherspoon and Skinner 1995): 

• Young homogenous plantation vegetation without 

slash treatment greater than 10 years after logging 

or fire. 

• Young naturally regenerated and shrubby vegetation 

greater than 10 years after fire. 

• Dense uniform stands of young conifers with low 

crown base heights. 

• Dense young to mature forests without large trees. 

• Dense forests containing large fire-tolerant trees and 

fuel ladders.

• Relatively open forests with large fire-resistant trees 

and low fuel ladders.

This list does not account for deciduous and evergreen 

hardwoods that can make patches less flammable, under less 

than extreme burn conditions. The appropriate mix of these 

types and management actions can only be determined 

using multiscale (patch, landscape, ecoregion) approaches 

that integrate fire protection, fire restoration, and wildlife 

habitat goals (Hessburg et al. 2016, North et al. 2009). An 

overarching aim of restoration efforts could be to introduce 

more heterogeneity in fuel conditions at landscape levels 

with the goal to reduce the likelihood of rapidly spreading 

large fires that include large patches of high-severity fire. 

Such landscapes would have lower threats to large overstory 

fire-resistant trees that were once common and widely distributed across a large percentage of these forest landscapes 

(Baker 2015; Hagmann et al. 2013, 2014; Sensenig et al. 

2013). A special concern with large fires that may burn as 

large high-severity patches is that they can remove habitat 

for the northern spotted owl and other late-successional 

species (Camp 1999, Camp et al. 1997). However, the effect 

on spotted owl habitat at landscape scales is a subject of 

uncertainty and active research (chapter 4). 

Williams and Baker (2012) argued that restoration programs for ponderosa pine and dry mixed-conifer forests are 

“misdirected in that they are seeking to reduce all high-severity fire.” Eliminating all high-severity fire patches from 

forests with predominantly low-severity or mixed-severity 

regimes would not be supported by our understanding of 

fire history and ecology in these systems. Instead, efforts to 

reduce the size of high-severity patches or the homogeneity 

of current fuel loads that lead to large high-intensity fires 

can be justified where knowledge of local landscape conditions and fire regimes indicates that such patches would not 

be characteristic of the landscape or would pose a threat to 

important social and ecological values. 

Consideration should also be given in these regimes for 

promoting open woodlands (e.g., oaks), open shrublands, 

and meadows and grasslands that have been lost as a result 

of overgrazing, fire exclusion, succession to forest, and 

other land use changes (Hessburg and Agee 2003, Hessburg 

et al. 2005). However, because reintroduction of fire to 

these systems may increase exotic species or have other 

unintended effects, restoration actions need to be done 

thoughtfully (Perchemlides et al. 2008). 

Invasive Plant Species and Pathogens

Nonnative invasive plants, insects, and disease can have 

major economic and ecological effects on forests (Lovett et 
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GENERAL TECHNICAL REPORT PNW-GTR-966

al. 2016, Moser et al. 2009). While the problem of invasive 

plants and pathogens is most severe in the forests of the 

Northeastern United States, there are several species of 

plants and pathogens that are having or could have significant impacts on forests within the NWFP area (Brooks et al. 

2016, Gray 2005, Lovett et al. 2016, Moser et al. 2009). 

Invasive plant species often have early-successional life 

histories and are well adapted to colonizing disturbed areas. 

Examples of this type of invasive plant in this region include 

Scotch broom (Cytisus scoparius) and Himalayan blackberry (Rubus armeniacus), which can invade disturbed areas 

and oak savannas, altering soil nutrient conditions, limiting 

tree regeneration, and promoting growth of other nonnative 

species (Gray 2005, Shaben and Myers 2009). Management 

of these species requires an understanding of their ecology 

and does not lend itself to a one-size-fits all solution (D’Antonio and Meyerson 2002). Once tree canopy closure is 

attained, these species typically drop out of the ecosystem. 

Although many invasive species invade disturbed, 

early-successional and open-canopy forests, closed-canopy 

forests, including old-growth forests, are not immune to 

invasive species (Martin et al. 2009). Invasion of forests by 

shade-tolerant species may just be slower but not necessarily 

less impactful in the long run than invasion of disturbed 

nonforest vegetation. Shade-tolerant invasive species of 

concern in this region include the perennial false brome 

(Brachypodium sylvaticum) and English holly (Ilex aquifolium). These species can outcompete native species, alter 

fire regimes, and possibly alter soil conditions where they 

occur within forests (Berger and Fischer 2016, Stokes et al. 

2014, Taylor and Cruzan 2015). Management strategies for 

reducing spread of false brome, which is most likely to be 

found in lower elevation forests, include limiting disturbance 

within stands, cleaning clothes and equipment to reduce seed 

dispersal, and possibly promoting hardwoods, whose litter is 

less suitable for germination (Taylor and Cruzan 2015). False 

brome may increase flammability of forests, and short-interval fire may promote it; as climate warms, invasion of forests 

by false brome is expected to increase (Brooks et al. 2016). 

Invasive pathogens with significant effects on forests of 

the NWFP area include white pine blister rust (Cronartium 

ribicola), Port Orford cedar root disease (Phytophthora lateralis), and sudden oak death (SOD) (P. ramorum) (see also 

chapter 11). Whitebark pine (Pinus albicaulis), a high-elevation species of the Cascades, is in decline throughout 

its range as a result of the combined effects of white pine 

blister rust and native bark beetles (Ellison et al. 2005). The 

loss of this species is having cascading effects on hydrology 

and other species. 

Sudden oak death is of particular concern because it 

has caused extensive mortality of tanoak (Notholithocarpus densiflorus), coastal live oak (Quercus agrifolia var. 

oxyadenia), California black oak (Q. kelloggii), and several 

other oaks in coastal forests of northern California and 

southern Oregon. The pathogen also infects a number of 

other tree and shrub species, many of which have special 

cultural significance to tribes (see chapter 11). Management 

strategies for SOD have focused on preventing or reducing 

transmission through quarantines that limit commercial 

movement of wood and host plants, and stand-level treatments, including killing and removal of infected trees and 

host plants, especially California bay laurel (Umbellularia 

californica), through cutting, burning, or herbicide application (Rizzo et al. 2005, Swiecki and Bernhardt 2013). 

Moritz and Odion (2005) reported that infections in stands 

that had experienced fire since 1950 were extremely rare; 

they suggested that a lack of fire could contribute to infestation by increasing shading, stand density, and abundance of 

hosts. Meentemeyer et al. (2008) concluded that reductions 

in fire frequency have likely facilitated SOD by increasing 

woodland cover and continuity at the expense of grasslands 

and chaparral, and by increasing bay laurel and creating 

more shaded, cooler microclimates.

The loss of mature tanoaks and various oaks has 

significant impacts on forest ecosystems in the infested 

areas. In heavily infested areas in conducive environments, 

stands formerly dominated by tanoak have been converted 

to shrubfields (Cobb et al. 2017, Klein et al. 2013). Additionally, infested stands could form stands with multiaged 

structures, a higher proportion of redwood and a lack of 

tanoak, and large canopy gaps (Waring and O’Hara 2008). 

While such changes could enhance stand structural heterogeneity, they could also jeopardize valuable ecological 

services such as nut production and abundance of large tree 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

cavities in hardwoods, which are important for fisher, owls, 

and other animals (Long et al. 2016). Other likely effects of 

the dieback include increased fuel loads, risk of high-severity burns, hazardous conditions for firefighters, increased 

soil erosion, and spread of invasive plants (Forrestel et al. 

2015, Swiecki and Bernhardt 2013). Research in one burned 

landscape indicated that stands with recent SOD establishment may experience higher vegetation burn severity, 

while stands where dead trees have fallen may experience 

increased soil burn severity (Metz et al. 2011). Although 

high-severity fire in particular can reduce pathogen load, 

infected bay laurel plants that survive within such burns 

may infect the resprouting vegetation (Beh et al. 2012). The 

combination of severe fires and SOD infection may increase 

the likelihood of extirpating tanoak in redwood-dominated 

areas, because redwood generally outcompetes tanoak after 

fire (Ramage et al. 2010). Consequently, it is important for 

managers to consider landscape-scale strategies that could 

promote resilience to both the disease and other disturbance 

agents such as severe wildfire and drought. Evaluating 

restoration strategies through an adaptive management 

framework seems particularly important given the complex 

dynamics among vegetation, SOD and other diseases, and 

fire (Odion et al. 2010, Rizzo et al. 2005). Use of managed 

wildland fire, especially in stands that are not already 

heavily infested with SOD, may be particularly important 

as a means of promoting forest resilience. Meanwhile, 

infected stands may be a priority for silvicultural treatments 

to reduce the potential for severe crown fires (Kuljian and 

Varner 2010).

Postfire Salvage and Management

Ecological effects—

Postfire salvage logging is typically proposed as a means 

of recovering some of the lost economic value in dead or 

damaged trees. The ecological consequences of salvage 

logging are often considered negative from the perspective 

of soils, hydrology, postfire seedling establishment, and 

wildlife habitat resources, although species responses differ. 

Early scientific understanding of salvage logging after 

wildfire was hindered by a lack of studies with sufficient 

replication and controls (McIver and Starr 2001), but recent 

research offers a more complete understanding of some ecological effects of salvage logging (Long et al. 2014). Table 

3-7 summarizes key findings from several reviews to help 

inform management decisions surrounding postfire salvage; 

research on this topic is developing as more large and severe 

fires occur in fire-excluded landscapes. We focus on effects 

of salvage logging (i.e., the removal of dead trees and those 

that are likely to die following wildfire) rather than a broad 

range of other postfire management activities. However, it is 

important to recognize that managers often avoid replanting 

in areas that have not been salvage logged for crew safety 

and for silvicultural reasons. 

Immediate stand-level effects of fire are primarily 

related to intensity, duration, and corresponding severity, 

most commonly interpreted through some measure of tree 

mortality and combustion of surface fuels, including dead 

and down wood and organic matter stored in duff, litter, 

and soils. Fire can reduce live tree density and canopy 

cover and increases the density of standing dead trees 

(snags) and the future abundance of dead and down wood. 

Although enormous amounts of carbon stored in live and 

dead biomass may be lost to the atmospheric carbon pool 

in a large fire (Campbell et al. 2007), most is retained 

in biological legacies, including snags, dead and down 

wood, charcoal, and live remnant trees (Acker et al. 2013, 

Baird et al. 1999, Donato et al. 2013). This carbon pool is 

then slowly lost from the forest as the retained deadwood 

decomposes or is consumed in subsequent fires (Campbell 

et al. 2016b, Donato et al. 2016). These biological legacies 

play important ecological roles that differ from the enrichment of recovering vegetation to providing microhabitats, 

stabilizing soils, and moderating harsh environmental 

conditions on burned sites (Lindenmayer and Noss 2006, 

Lindenmayer 2004). 

Salvage logging alters postfire vegetation structure by 

reducing the basal area and density of live and dead trees 

(McIver and Otmar 2007) and decreasing the persistence 

of remaining snags (Russell et al. 2006) and altering the 

microclimate of a site (Marañón-Jiménez et al. 2013). 

What’s more, once a tree dies, it functions as a snag, down 

log(s), mulch, and charcoal in soils for a period that can 

far exceed the period spent as a live tree (DeLuca and 
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Aplet 2008), although those dynamics should vary widely 

based upon moisture and fire regimes. Cumulatively, these 

reductions result in decreases in live and dead biomass 

(Donato et al. 2013) and reduced soil carbon. However, 

the down dead wood would not likely have been able to 

decompose in frequent fire regimes before the onset of 

fire suppression (Skinner 2002). Studies have shown that 

as wood becomes more decayed, it is more likely to be 

consumed in subsequent fires (Knapp et al. 2005, Uzoh 

and Skinner 2009). Numerous studies document initial 

short-term decreases in natural regeneration following 

salvage (McIver and Starr 2001) for various reasons, 

including direct mortality from mechanical damage 

(Donato et al. 2006) as well as indirect effects of altered 

competitive interactions with shrubs and harsher microclimate (Marañón-Jiménez et al. 2013, Ritchie and Knapp 

2014, Stuart et al. 1993). However, one study 10 years 

after salvage showed no difference in natural regeneration 

following a severe fire with different levels of salvage 

ranging from leaving everything to taking everything 

(Ritchie and Knapp 2014). Planting following salvage may 

be needed to mitigate any effects on regeneration or to 

establish tree species and genotypes that are better suited 

to climate warming or diseases. The effects of salvage 

logging versus no intervention on loading of fine fuels and 

coarse fuels and the effects of reburn are expected to differ 

considerably over time. If not followed by fuel treatment 

or accomplished through whole tree harvesting (Ritchie et 

al. 2013b), salvage logging can increase fine fuels to levels 

that support high-severity fire, which kills regeneration 

(Donato et al. 2006). There are few studies of the effects of 

salvage on subsequent wildfire, but Thompson et al. (2007) 

found higher reburn severity in stands that were salvaged 

and planted than in unmanaged stands. The Thompson et 

al. (2007) study hypothesized that salvage logging without 

sufficient treatment of the slash after logging and uniform 

Table 3-7—Suggestions for ecologically based postfire management in terrestrial ecosystems from three 

major reviews

Recommendations

Karr et al. 

2004

Beschta et al. 

2004

Lindenmayer 

and Noss 2006

Promote natural recovery  

Retention of old, large trees and snags   

Protect soils against compaction and erosion   

Protect ecologically sensitive areas (e.g., reserves, roadless areas, steep 

slopes, fragile soils)

  

Rehabilitation of roads and fire lines, avoid creation of new roads  

Limit reseeding and replanting  

Protect and restore watershed before fire  

Continue research, monitoring, and assessment of the effects of 

salvage treatments



Educate public on the natural role of wildfires, allow natural regimes  

Ban introduction of exotic species 

Curtail livestock grazing 

Low-intensity or no harvesting in unburned or partially burned patches  

Limit removal of biological legacies from particular areas (e.g., burned 

old-growth stands)



Ensure maintenance and creation of essential habitat elements for 

species of concern


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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

conifer plantations likely contributed to higher surface fuel 

loads after salvage and consequently to the higher reburn 

severity. More work is needed to evaluate the effects of 

salvage logging and adequate slash disposal on risk of 

high-severity fire. One study found that fine fuel loading 

following salvage returned to untreated levels after about 

25 years (McIver and Ottmar 2007). 

Salvage logging also reduces large fuel loads over time 

through removal of snags that would otherwise begin to fall 

and increase large dead wood on the ground as early as the 

first 10 years following fire, but typically over much longer 

periods (Dunn and Bailey 2015, McIver and Ottmar 2007, 

Peterson et al. 2015). One study showed that regardless of 

intensity of salvage logging, more than 80 percent of tree 

biomass left standing had transitioned to become surface 

fuel after 8 years (Ritchie et al. 2013b) with pines falling 

more rapidly than either white fir or incense cedar (Calocedrus decurrens) (Ritchie and Knapp 2014). Greater log 

biomass in unsalvaged stands resulted in higher surface 

temperatures during prescribed fire 20 to 30 years following 

wildfire (Monsanto and Agee 2008). Large areas of the 

Western United States have been burned by high-severity 

fire or killed by bark beetle outbreaks. The resulting dead 

fuels will become future surface fuels. Long-term research 

is needed to better understand the tradeoffs among postfire 

salvage logging and future surface fuels, and the ecological 

benefits of dead and down wood and future fire severity and 

community succession.

Salvage logging can affect ecosystem processes by 

altering microclimate and hydrology, increasing sediment production, and reducing soil nutrients and carbon 

sequestration in the forest. Removal of snags can affect 

microclimate by reducing shade (sometimes referred to as 

dead shade) and consequently reducing temperatures at 

night and increasing temperatures during the warming part 

of the day (Fontaine et al. 2010). Risk of accelerated erosion 

comes with ground disturbance during salvage logging 

(Wondzell 2001); however, there is a noticeable lack of 

studies from the Northwest on this issue. In one Western 

United States study, Wagenbrenner et al. (2015) found that 

salvage logging increased soil compaction, decreased soil 

water repellency, and slowed recovery of vegetation, but the 

degree of impact depended on the method of logging, local 

climate, and soils. Where a winter snowpack is typical, the 

potential for hydrological impacts is greatest where harvest 

operations occur outside of the winter months. Logging 

over snow and frozen ground could reduce the effects on 

soil and sediment (Poff 1989). Indeed, Peterson and Dodson 

(2016) found that postfire commercial logging on dry or 

frozen soils in northeastern Oregon displaced or compacted 

an average of 15 percent of the soil surface in commercial 

logging units and 19 percent of the soil surface in the 

fuel reduction logging units, yet they found no persistent 

impacts on understory vegetation 15 years following 

treatment. In a study from central Oregon, compaction 

following salvage logging decreased soil respiration and 

available nitrogen, while later subsoiling to alleviate 

compaction decreased available phosphorus (Jennings et al. 

2011). In several studies of boreal forests, postfire removal 

of snags reduced soil carbon for several years (Bradford et 

al. 2012, Kishchuk et al. 2015, Poirier et al. 2014). In two 

studies from relatively dry Sierra Nevada forests, Johnson 

et al. (2005) and Powers et al. (2013) found that postfire 

salvage resulted in a substantial reduction in onsite carbon 

compared to fire alone, although the authors of both studies 

noted that their studies, as with many other studies, did not 

account for sequestration in the resulting wood products. 

Moreover, it is important to consider long-term carbon 

dynamics given future fires (Carlson et al. 2012), because 

planting treatments can potentially accelerate carbon storage in trees, and fuel reduction treatments can potentially 

reduce future tree mortality. 

The impacts of salvage logging on biota are mostly 

associated with the removal of snags and deadwood, 

which are important habitat components for a variety of 

terrestrial and aquatic organisms. Salvaging has been 

reported to have negative effects for several species of 

cavity-nesting birds, such as black-backed woodpeckers 

(Picoides borealis), three-toed woodpeckers (P. tridactylus), and mountain bluebirds (Siala currucoides), (Hutto 

2006, Hutto and Gallo 2006, Saab et al. 2007), but neutral 

or positive effects have been documented on a few species 
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(Peterson et al. 2009). In a recent study from the Sierra 

Nevada, White et al. (2015) suggested that it was important 

to retain some relatively dense stands of dead or dying 

trees (40 to 60 per acre) at the landscape scale, to promote 

snag-associated species such as black-backed woodpecker, 

mountain bluebird, and olive-sided flycatcher, rather than 

evenly thinning all stands and retaining smaller numbers of 

snags; they suggested further research would be needed to 

guide the extent and configuration of such treatments. Soil 

bacteria and fungi appear resilient to salvage (Jennings et 

al. 2011). Removal of snags and large coarse woody debris 

could adversely affect habitat for carnivores such as fisher 

(Pekania pennanti) and Pacific marten (Martes caurina), 

if the large dead wood would have otherwise persisted into 

closed-forest stages where the animals use large structures 

for den and rest sites (Bull et al. 2001). 

Fire may have positive effects by contributing wood 

and coarse sediment for aquatic habitats (Benda et al. 2003, 

Reeves et al. 1995) that may be partially negated by removal 

of wood during salvage logging, especially when the large 

wood is removed from key source areas to streams. Many 

aquatic and riparian organisms are adapted to fire (Flitcroft 

et al. 2016, Reeves et al. 2006) so postfire management 

is typically not needed to support aquatic ecosystems. 

Hillslope processes and subsequent erosion after periodic 

fires are critical to aquatic habitat succession, and native fish 

populations can often rebound within a decade after a wildfire, especially when they can recolonize altered reaches 

from connected refugia (Bisson et al. 2003, Dunham et al. 

2003, Rieman and Clayton 1997, Rieman et al. 1997). 

Management of postfire environments—

The ecological effects of postfire salvage logging can 

differ depending on treatment, fire severity, and biophysical setting (Peterson et al. 2009). In general, research 

supports the conclusion that salvage logging does not 

benefit native species and terrestrial or aquatic ecosystems 

(Beschta et al. 2004, Karr et al. 2004); an exception might 

include, e.g., fire-suppressed forests with high densities of 

trees. Further long-term research on contemporary salvage 

practices would greatly enhance understanding of the 

circumstances under which salvage might be beneficial. 

Peterson et al. (2015) and Hessburg et al. (2016) identified 

situations, including elevated long-term woody fuel 

loads, lack of seed sources, and potential for reburns that 

maintain undesirable shrubfields, in which postfire management might be used to meet ecological goals. These 

include (1) fuel reduction treatments that reduce long-term 

levels of large woody fuels (which may be elevated as 

shade-tolerant species increased under fire suppression 

and that may pose a risk to soil fertility were the area to 

reburn), (2) fuel treatments or planting trees to reduce 

potential for high-severity reburns and forest succession 

where potential for large semistable patches of shrubs 

is high and regeneration is lacking (Dodson and Root 

2013), and (3) removing surface fuels that may impede 

establishment of trees. The effects of particular strategies 

may differ considerably with ecological conditions across 

the NWFP area. In some cases, shrub removal may be 

important for promoting native plant species richness 

(Bohlman et al. 2016) in subsequent decades. However, 

shrubs may also have important roles in increasing soil 

carbon and nutrients, especially nitrogen. For example, in 

a dry ponderosa pine site in central Oregon, Busse et al. 

(1996) found that shrub removal aided tree growth in the 

first two decades, but the effect then leveled off and shrub 

removal was associated with decreases in soil carbon and 

nitrogen 35 years later.

Tree replanting, which as mentioned above is often 

practically tied to postfire snag removal, may be an 

important strategy to consider in areas where natural 

regeneration is too low to meet objectives for a landscape 

in the time desired. One example of such low regeneration 

was reported for several fires in the northern Sierra Nevada 

(Collins and Roller 2013) bordering the NWFP area but that 

has similar species to the Klamath region. The authors of 

that study noted that several studies from mixed-conifer 

forests in the mixed-severity regime of the Klamath-Siskiyou Mountains (Donato et al. 2009, Shatford et al. 2007) 

had found generally abundant conifer regeneration in 

stand-replacing patches. Where sites reburn and high-severity patches are large, regeneration can be low (Tepley 

et al. 2017). Lower and less consistent moisture may also 
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contribute to incidents of sparse conifer regeneration in 

regions predisposed to a frequent fire regime. Because 

promoting vegetation heterogeneity may reduce fire spread 

and burn severity (Thompson et al. 2007) and promote 

biodiversity, managers have experimented with more 

variable planting patterns (e.g., spacing and clustering) than 

have traditionally been used, but more research is needed to 

evaluate outcomes from such strategies. 

Accumulation of large dead fuels can lead to severely 

burned soils if forests reburn. A study from the eastern 

Cascades of Oregon found that severely burned soils can 

have lower fertility and depleted microbial communities 

(Hebel et al. 2009). However, this study also found that 

several native plants appeared highly competitive in 

severely burned, low-resource soils; based upon a laboratory study component, they suggested that those native 

plants might be more competitive in those burned soils 

than invasive nonnative species. Relationships between 

plant diversity and fire severity are complex because they 

reflect variation in environment (especially precipitation 

and fire regime) and species composition (such as presence 

of invasive species). For example, DeSiervo et al. (2015) 

hypothesized that diversity would be promoted in fires that 

matched the reference fire regime, and they indeed found 

that native species richness was greater in areas of low to 

moderate vegetation burn severity of northern California 

(in a region of frequent fire), while areas burned at higher 

severity experienced more incursion by cheatgrass and 

other nonnative species. Similarly, Stevens et al. (2015) 

found that high burn severity shifted composition toward 

nonnative species and native species with southern-xeric 

affinity and away from native species with northern-temperate affinity.

Application of salvage logging in these contexts would 

need to consider overall effects of a wildfire on the larger 

affected landscape, and tradeoffs with other ecological and 

economic objectives. More research is needed to better 

understand the ecological effects of low to moderate levels 

of salvaging that may be done to recover economic value 

(Campbell et al. 2016a) from fire-killed trees. 

Research Needs, Uncertainties, 

Information Gaps, and Limitations

While much as been learned about the ecology, conservation, and restoration of these forests, many knowledge gaps 

and uncertainties remain. We mention them throughout the 

document and summarize the major ones here:

1. While the range- and regional-scale patterns of disturbance regimes are known, much less is known 

about them at subregional and landscape scales. 

Our knowledge of the region is based on extrapolation from relatively few fire and forest history 

studies. Research is needed to help fill in the gaps 

in our knowledge especially as they relate to fire 

sizes, frequencies, and function in mixedseverity regimes of both the moist and dry forests. 

2. We know much about the structure of old-growth 

forests from studies of contemporary older forests 

across all forest types but lack stand-structure 

definitions for use in monitoring and inventory 

related to old-growth forests that developed in the 

mixed- and low-severity fire regimes of moist and 

dry forests. Our current monitoring efforts (e.g., 

definitions and indices) use reference conditions 

for old growth that are based on forests that have 

been altered by fire exclusion and do not take into 

account structures associated with historical disturbance regimes. Research is needed to develop 

old-forest definitions and landscape-scale targets 

based on HRV, desired levels of resilience given 

fire, and future climate change or other considerations such as species habitat needs. 

3. We lack information about the biodiversity and 

ecosystem functions of early-seral vegetation as 

well as frameworks for developing landscape-scale 

goals for these conditions given fire suppression. 

Mechanical treatments and prescribed fire can be 

used to approximate some of the ecological functions of diverse early-successional habitats. We 

also lack knowledge of what restoration actions 

(e.g., planting in post-wildfire environments) might 

be beneficial for longer term successional goals 

(e.g., recovery of conifer forest canopies). 
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4. The effects of fire suppression on forest biodiversity and ecosystem function in older forests are 

not well studied in much of the NWFP area. This 

is apart from knowledge of how succession has 

altered fire regimes and fire risk. Lack of fire in 

high-fire-frequency forests is altering plant community diversity, but more research is needed on 

the long-term ecosystem effects of increased stand 

density and shade-tolerant species in forests that 

were burned frequently to moderately frequently 

by low- to moderate-severity fire. 

5. We lack a solid understanding of how drought, beetles, and disease are likely to affect forests given 

climate change and interactions with fire. 

6. The ecological tradeoffs associated with variable-density thinning (i.e., restoration thinning) 

to restore or create ecological diversity in forest 

plantations are not well understood at stand or 

landscape scales and are known only from relatively 

short-term studies. Long-term research is needed to 

understand how ecosystems and the biota respond to 

these management actions and to learn more about 

the possible ecological costs and benefits of these 

actions in stands older than 80 years that might 

have undesirable densities or uniformity of trees. 

Similarly, long-term effects of postfire management 

warrant further study at large and long-term scales.

7. Given tradeoffs associated with restoration actions 

or inactions for different types of habitats and 

successional stages, research is needed to explore 

options for managing for a dynamic mosaic of 

vegetation and habitats at landscape scales under 

climate change. For example, how much do the 

pace, scale, and pattern of restoration activities at 

landscape scales affect fire severity and patterns of 

successional stages under a changing climate? 

8. It will also be important to better understand the 

tradeoffs associated with use of both coarse- and 

fine-filter approaches to conservation. Dynamic 

landscape modeling is needed, and where feasible, 

landscape-scale experiments and demonstration 

areas will be important to advancing our understanding of this issue. 

Conclusions and Management 

Considerations

Timber harvest, fire exclusion, fire suppression, and the loss 

of burning by American Indians have profoundly changed 

the moist and dry forests of the NWFP area. Although the 

motivation for the NWFP arose from clearcutting of old 

growth and loss of spotted owl habitat in moist forests, the 

dry zone forests, which occupy about 43 percent of the Plan 

area, have actually experienced more pervasive ecological 

changes as a result of human activity. Key changes in dry 

forests are loss of large, fire-resistant trees to logging, large 

departures in amounts and patterns of surface and canopy 

fuels, widespread shifts in proportions of seral stages, 

and changes in the patch sizes of those seral stages. These 

changes have affected all species and all processes; some in 

favorable ways (e.g., more habitat for dense forest species) 

and others in unfavorable ways (e.g., loss of open oldgrowth forests and ecological resilience to fire and drought). 

Changes in the moist forests are also significant, but they 

have been affected to a lesser and different degree by fire 

exclusion. Here, intensive timber harvest has been the 

primary impact on biodiversity by dramatically reducing 

the amount of dense old-growth forests and fragmenting 

habitats for species associated with these older forests. Fire 

exclusion in moist forests has had an important but different 

and less visible effect: the loss of diverse early-seral vegetation and associated reduction in landscape diversity. 

The 2012 planning rule adds a new context for NWFP 

national forests that will undergo plan revision in the coming 

years: management for ecological integrity (ecosystem 

characteristics) and species conservation using coarse-filter 

approaches; fine-filter approaches are to be used for a limited 

number of species where coarse-filter approaches may not be 

sufficient. Coarse-filter approaches based on managing for 

ecological integrity (as opposed to coarse-filter approaches 

based on one vegetation type, i.e., dense old growth) 

would promote basic ecological processes, including major 

disturbances that regulate successional and fuel patterns 

(i.e., “habitat” for fire). Ecosystem-dynamics approaches 

are needed to rebuild more functional ecosystems, reduce 

threats to and possible listing of additional species, and 

provide a more ecologically viable approach to maintaining 

existing listed or sensitive species within the context of 

meeting other ecological and socioeconomic goals. 
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Management Considerations Summarized

• The 2012 planning rule sets a new context for ecosystem management under the NWFP: it focuses 

on ecological integrity based on maintaining and 

restoring disturbance and other ecological processes. 

Natural range of variation is a guide but not necessarily a target. This is a broader focus than the 

original coarse-filter approach of the NWFP, which 

focused primarily on one type of forest condition: 

dense, multilayered older forest. 

• The goals and standards and guides for LSRs of the 

moist forests with infrequent fire are a relatively 

good match for managing for ecological integrity and 

resilience, especially in the face of climate change 

and invasive species. 

• Focusing restoration (e.g., variable-density thinning) 

in LSRs in moist forests on plantations makes sense 

from a conservation perspective, and can provide 

jobs and economic returns. However, there will be 

tradeoffs with some ecological goals (e.g., amounts 

of dead wood) that may need mitigation. 

• Fire suppression has had an effect on vegetation 

conditions in moist forests, especially in the drier 

part of the zone where fire was historically more 

frequent and mixed-severity effects more common. 

The effect is not the same as in dry forests. Fire 

exclusion in moist forests has reduced the amount 

of early-successional vegetation in the landscape, 

reduced diversity of structure in old-growth forests that were subject to partial stand-replacement 

fire, and thus reduced landscape-scale diversity. 

Managers may want to consider restoring fire or 

using fire surrogates to promote early-successional 

forests and landscape-scale diversity in plantations 

and forests more than 80 years old in the matrix. 

Managing for diverse early-seral stages would 

require a landscape-scale approach to ensure that 

old-growth goals are not compromised. 

• The goals, standards, and guides for LSRs in dry 

forests are inconsistent with management for ecological integrity and resilience to climate change 

and fire. Dense late-successional older forests would 

have been historically uncommon in dry forests, 

and their current higher abundance is a function of 

fire exclusion and suppression. Fires have become 

much less frequent than historically, but, when they 

burn, they are more likely to include large patches 

of high-severity fires that kill fire-resistant older 

trees and alter landscape-scale patch patterns. In the 

absence of fire, the forest structure and composition 

are shifting toward denser forests and shade-tolerant 

species that are less resistant to fire and drought. 

• Management actions that promote resilience in dry 

forest landscapes include reducing the continuity of 

surface and canopy fuels to reduce patch sizes and 

thus the extent of high-severity fires and using prescribed fire or managing wildfire for ecological benefits where appropriate. Landscape-level strategies 

are needed to provide for dense forest conditions 

as indicated by the NWFP in environments where 

they are more likely to persist in the face of fire and 

climate change. 

• Alternative approaches to the LSR network 

and standards and guides may better meet both 

coarse- and fine-filter goals by incorporating the 

evolving understanding of the ecological dynamics of dry forests and threats from climate change 

and invasive species that apply to both moist and 

dry forests. 

Our main findings and conclusions are listed below 

by general topic. We also indicate which of the following 

questions the conclusion applies to: 

Guiding Questions

1. What are the structures, dynamics, and ecological 

histories of mature and old-growth forests in the 

NWFP area, and how do these features differ from 

those of other successional stages (e.g., early and 

mid successional)? 

2. How do these characteristics differ by vegetation 

type, environment, physiographic province, and 

disturbance regime? 

3. What is the scientific understanding about using 

historical ecology (e.g., historical disturbance 

regimes and natural range of variation) to inform 

management, including restoration?
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4. What are the principal threats to conserving 

and restoring the diversity of old-growth types 

and to other important successional stages (e.g., 

diverse early seral), and to processes leading to 

old growth? 

5. What does the competing science say about needs 

for management, including restoration, especially 

in dry forests, where fire was historically frequent? 

6. How do the ecological effects of treatments to 

restore old-growth composition and structure 

differ by stand condition, forest age, forest type, 

disturbance regime, physiographic province, and 

spatial scale? 

7. What are the roles of successional diversity and 

dynamics, including early- and mid-seral vegetation, in forest conservation and restoration in the 

short and long term? 

8. What is the current scientific understanding 

concerning application of reserves in dynamic 

landscapes?

9. How do recent trends of forests in the NWFP 

reserve network relate to both original NWFP 

goals, those of the 2012 planning rule, and climate 

change adaptation needs? 

10. What is the current understanding of postwildfire 

management options and their effects?

Ecology of Old-Growth and Other Vegetation 

Types (Questions 1 and 2) 

1. Knowledge of historical disturbance regimes and 

successional dynamics is essential for conserving, 

restoring, and promoting resilience of old-growth 

forests and other successional stages to climate 

change, fire, and other disturbances. 

a. All seral stages contribute to maintaining 

native forest biodiversity, ecosystem function, 

and other ecosystem services. Moist forests 

and dry forests have fundamentally different 

disturbance regimes, developmental pathways, 

and ecological potentials. 

b. We developed a fire regime map (fig. 3-6) to 

provide a framework for planning and managing these diverse forests. Four major fire 

regimes are recognized, two in the moist forests and two for the dry forests. 

c. The major regimes of the moist forests are: 

i. Infrequent (greater than 200 years), high 

severity 

ii. Moderately frequent to somewhat infrequent to (50 to 200 years) mixed severity. 

d. The major regimes of the dry forests are:

i. Frequent (15 to 50 years) mixed severity 

ii. Very frequent (5 to 25 years) low severity 

e. Of these four regimes, the two mixed-severity 

regimes are the most variable and complex. 

All severities of fire occur in all regimes, but 

the regimes differ in proportion and spatial 

pattern of high-severity fire. 

2. Old-growth forest structural elements common 

to all forests of the region include relatively large 

and old live, decadent and dead trees, and spatial 

heterogeneity of forest structure and composition. 

Other characteristics such as multiple canopy layers, shade-tolerant associates, and large amounts of 

dead and down wood are not necessarily characteristic of all old-growth forest types under the historical disturbance regimes of the region. Large-tree 

elements can also be found in younger forests, and 

patches of early-seral vegetation that developed following high-severity disturbance in older forests. 

3. Definitions of old growth that recognize oldgrowth structural features as a continuum across 

stands of various ages and disturbance histories are 

more ecologically realistic and useful for restoration planning than a definition that has only one 

threshold with the result that forests are either old 

growth or not. 

4. Current definitions of old growth used in monitoring are based on current forest inventory plots. This 

means that definitions for dry forests, which have 
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been heavily influenced by fire exclusion, are not 

reflective of historical forest structure and composition that were typical of this environment. Better 

definitions or reference conditions that reflect the 

variety of old growth are needed for conservation and restoration of old-growth and landscape 

dynamics for dry forest types, as well as communities with significant hardwood components. 

5. Older forests differ in tree density, spatial heterogeneity, and species composition between moist 

and dry forest zones and across their associated 

disturbance regimes. Dense, multilayered old forests were typical of infrequent/high-severity fire 

regimes in moist forests parts of the region, while 

relatively open forest of pine, Douglas-fir, and 

other conifers were typical of very frequent/lowand mixed-severity regimes in dry zone forests. 

Dense multilayered older forest in dry forest landscapes occurred in fire refugia such as topographic 

settings where fire was infrequent. Old-growth forest structure and composition were most diverse in 

the mixed-severity regime of the moist forests and 

the mixed-severity regime of the dry forests. 

6. Early-seral and “pre-forest” vegetation was 

an important component of many landscapes. 

Early-seral vegetation that results from high- and 

mixed-severity disturbance provides distinctive 

biodiversity and ecosystem function. Grasses, 

herbs, shrubs, hardwoods, and legacy live and dead 

trees that develop during these stages can influence 

forest development, biotic communities, and ecosystem function for decades to centuries. 

7. Landscape diversity also varied across the disturbance regimes. In the infrequent/high-severity 

regime of the moist forests, the dominant landscape pattern was medium to coarse grained with 

very small to very large patches of older forests of 

complex structure, patches of younger more homogeneous forests, and rare to common (depending 

on climate period) very large patches of early-successional vegetation. Patches of hardwoods and 

shrubs would have occurred along many streams. 

In the mixed-severity regime of the moist forests, 

the landscape would have been a relatively dynamic 

mosaic of well-connected and dispersed mature and 

older forests and differently aged and sized patches 

of younger forests and preforest vegetation forests, 

often containing remnant live and dead large trees. 

8. The forest landscape of the frequent/mixed-severity regime of the dry forests would have been 

a complex mosaic of forest structural types that 

was very strongly controlled by frequent fire. In 

the very frequent fire regimes, the forested part of 

the landscape would have been a fine- to medium-grained mosaic of older trees and very small 

to small patches of early-successional conditions. 

The open nature of the forest combined with the 

fine grain of patches often led to blending of areas 

of old trees with understory vegetation (forbs, 

grasses, shrubs) otherwise typical of early-seral 

conditions. In steep, dissected topography (e.g., 

northwest California), the mosaic of forest conditions would have been more strongly expressed as 

a function of topography and fine-scale variability 

in disturbance regimes and successional pathways. 

Value of Ecological History (Question 3)

1. Knowledge of ecological history is essential for 

conducting and guiding conservation and restoration. Using HRV in forest structure, composition, and landscape patterns can be a useful guide 

for conservation and restoration efforts. However, 

returning forests and landscapes to a narrowly 

defined state of historical conditions and dynamics 

will not be possible nor desirable in many landscapes given anthropogenic forest change (e.g., 

land ownership patterns and forest management) 

and climate change. Approximations of historical 

regimes and forest conditions or management for 

resilience to fire as a recurring ecological process 

and climate change will be a more realistic and 

sustainable goal for many areas. 
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Conservation and Restoration Needs 

(Questions 4 and 5)

1. While the restoration needs differ between the 

moist and dry forests, logging and plantation silviculture have affected forests in all of the regimes. 

In the moist forests, clearcutting and plantation 

establishment for timber production reduced 

the area of old-growth forests and fragmented 

the landscape across millions of acres of forest 

lands. Intensive timber management has reduced 

stand-level diversity, reduced dead wood and snag 

abundance, increased the amount of sharp edges, 

and increased road densities. Clearcutting and 

plantation establishment affected the drier forests 

as well, but a more pervasive effect may have been 

the partial harvest of old pines that significantly 

reduced the abundance of large, fire-resistant 

trees, leaving existing older forests with far fewer 

large live and dead trees than they would have 

had under natural disturbance regimes. Moreover, 

often the larger overstory trees are species (e.g., 

Douglas-fir, grand fir, or white fir) that are not as 

resistant to fire. 

2. Fire exclusion effects are also present in all 

regimes but are significantly different between 

the dry and moist forest zones. In the dry forests, 

lack of fire has greatly increased tree density and 

reduced resilience to fire, drought, insects, and 

disease. Specifically, the area of multilayered, 

closed-canopy older forest has increased outside the historical range over the past 100 years 

despite logging and recent fires. Fire suppression has also had an effect in the moist forests, 

but there has generally been little impact on fuel 

accumulation (except where logging has occurred 

and slash has not been treated) and fire risk as 

these productive forests naturally have high 

fuel loads. Instead, the effects of fire suppression in moist forests have been to reduce the 

area of high-severity fire (relative to historical 

dynamics), and, consequently, the area of diverse 

early-successional vegetation. Thus, lack of fire 

in the moist, mixed-severity-regime forests has 

likely reduced landscape diversity. 

3. Fire exclusion and succession toward shade-tolerant, fire-sensitive species may be leading to more 

fire-resistant older forest vegetation in some dry 

forests under a wider range of fire weather conditions. Forests in these areas are more shaded, 

dry out more slowly, have lower windspeeds, and 

have more compact fuel beds that are less able to 

carry fire than more open pine-dominated older 

forests. However, under extreme weather, these 

forests are less resistant and resilient because 

they are more likely to burn with high severity 

than historically, when forests were more open 

and contained less fuel. As climate changes, 

such extremes (e.g., drought and high winds) are 

expected to increase. 

Competing Science Related to Need for 

Restoration (Question 5)

1. Some have argued that restoration is not needed 

because most ponderosa pine and dry mixed-conifer forests have been mischaracterized as simply 

having a low-severity fire regime. Instead, they 

contend that these forests were historically denser 

than most other studies indicate and are better 

characterized as having a more variable-severity fire regime, with significant components of 

mixed- and high-severity fire. Baker (2012) and 

others cite Hessburg et al. (2007) in support of 

their arguments; however, the results of Hessburg 

have been misinterpreted in these papers and do 

not fully support claims about the importance 

of high-severity fire in dry forests. In addition, 

recent research (Levine et al. 2017) indicates that 

the method used by Baker (2012) overestimates 

tree densities. We believe the preponderance of 

evidence supports the view that prior to EuroAmerican settlement, pine and dry mixed- and 

some moist mixed-conifer forests had relatively 

low tree densities and that large patches of 

high-severity fire were not common in dry forests 
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with very high frequency (<25 years) and low 

severities. However, larger patches of high-severity fire were an important component of dry 

forests (e.g., mixed conifer) with frequent (15 to 

50 years) mixed-severity regimes. 

Trends in Forests in the NWFP Reserve Network 

(Question 9)

1. At the scale of the NWFP area, losses of older 

forest owing to logging and wildfire over the 20 

years of the NWFP have been relatively small and 

compensated for by significant gains from succession offsetting almost two-thirds of the losses 

from high-severity disturbance. However, dynamics differ geographically and with scale, and some 

areas, especially the Klamath region in Oregon and 

California, have had much higher net losses as a 

result of very large high-severity patches (mainly 

from a single large fire [Biscuit]). The NWFP 

reserve strategy, which focused on closed-canopy older forests is currently meeting many of the 

expectations of the NWFP, but it appears unlikely 

that this network will support the original conservation goals and new goals of the 2012 planning 

rule in dry forests under climate change. Threats 

include more frequent and larger patches of 

high-severity fire, which are promoted by high canopy fuel continuity and elevated surface fuel loads. 

Reserve Approaches in Dynamic Landscapes 

(Questions 8 and 9) 

1. Reserves are a valuable strategy for conserving 

biological diversity in the face of development 

and many extractive land uses. The literature 

indicates that goals and management guidelines for reserves need to be clearly defined. 

Management within reserves also may be needed 

in many cases to address past management 

effects or restore ecological processes and ecosystems that have been altered by past land use, 

including timber management, fire exclusion, and 

invasive species. 

2. The options that were developed in FEMAT (1993) 

and set the foundation for the NWFP were based on 

the best available science at the time, but that science 

emphasized moist zone forest ecology and did not 

adequately deal with the substantially different ecology of forests and landscapes of the dry forest zone 

(Spies et al. 2006b). Although the LSRs are currently 

providing for late-successional/old-growth forest 

conservation, new science and increased understanding of fire regimes and climate change indicate that 

focusing only on dense older forest as the primary 

conservation goal across the entire NWFP area will 

likely have unintended negative consequences in 

terms of diversity of successional stages, resilience 

to fire and climate change, and biotic disturbance. 

3. The current LSR standards, guidelines, and spatial 

patterns for dry forests do not appear to be consistent with emphasis on ecological integrity and 

other approaches for conserving biodiversity under 

the 2012 planning rule. In addition, threats from 

climate change and invasive species including the 

barred owl would appear to justify a reassessment 

of the reserve network in both dry and moist forests (see chapter 12). Development and evaluation 

and testing of new, highly integrated conservation 

approaches is encouraged to deal with changing 

knowledge, new perspectives on fire regimes, 

climate change, invasive species, and recognition 

of tradeoffs among biodiversity goals (e.g., coarse 

filter and fine filter) and between the ecological and 

social dimensions of forest ecosystem management 

(see chapter 12 for more information). 

Restoration Approaches (Questions 6 and 7)

1. Restoration is more about creating landscapes 

for the future that are resilient to future fires and 

changes in climate and support native species than 

it is about recreating past conditions. We can use 

historical ecology at the community and landscape 

scales to understand how various patch- and landscape-level patterns will respond under these new 

conditions. Restoration strategies include: 
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a. Variable-density thinning in plantations to 

increase ecological heterogeneity and accelerate growth of large trees and tree crowns. 

b. Variable-density thinning from below and 

prescribed fire in dense older forests in very 

frequent/low-severity and frequent/mixed-severity regimes to increase resilience of those 

forests to fire and climate change through 

restoring more diverse structures and compositions of older forests. 

c. Careful use of prescribed fire and managing 

wildfires away from the wildland-urban interface in dry forests and mixed-severity regimes 

of moist forests to restore key ecological processes while protecting critical areas of dense, 

older forest conditions. 

d. Creating diverse early-successional habitat 

where feasible given other ecological goals and 

social constraints. This could include partial 

cutting (retention silviculture) and prescribed 

fire (e.g., “ecological forestry”) in plantations 

and perhaps in forests over 80 years old (which 

is allowed in the NWFP in the matrix of moist 

forests and within LSRs in dry forests) where 

this practice would be consistent with other 

landscape goals (e.g., resilience to fire and climate change, habitat for spotted owls, creating 

landscape-scale successional diversity).

e. Using landscape-level strategies based on disturbance regimes, topography, spatial pattern, and 

departure from desired historical conditions. 

2. The scientific understanding of using 80 years as 

a threshold for restoration of stands within LSRs 

in moist forests has not improved much since the 

NWFP was established. The 80-year rule from 

the NWFP was based on expert opinion of stand 

development from data collected in natural forests 

of different ages. Eighty years is a one-size-fits-all 

threshold that does not recognize that stand age is 

only a rough proxy for stand structure and development potential, both of which can vary greatly 

based on site conditions and disturbance history. 

Depending on the structure and composition of 

stands, and landscape context and objectives, restoration treatments in forests over 80 years could promote old-growth characteristics or reduce them (e.g., 

reduce number of large dead trees). However, in general, and given a lack of new information, treatments 

of stands over 80 years in moist forests would still be 

expected to have less benefit for reaching old-growth 

structure than restoration in stands under 80.

3. There is no new ecological science that undercuts 

the guideline of using alternative silviculture to 

meet both wood production and ecological goals 

in stands over 80 years in the NWFP matrix of 

the moist forests. Studies of retention silviculture 

suggest that some biodiversity elements of older 

forests can be retained in stands managed for a 

combination of timber and structural and compositional diversity. 

4. All management (including restoration activities 

and lack of activities) involve ecological tradeoffs: 

a. Commercial thinning can provide short-term 

early-seral habitat and accelerate the development of large live trees and habitat diversity 

for some species but may have a short-term 

impact on habitat quality for other late-successional species and can reduce amounts of 

deadwood in the future (although deadwood 

may be higher than the historical range owing 

to fire exclusion). 

b. Thinning and restoring fire to forests with 

a history of very frequent fire can increase 

resilience to wildfire and increase habitat for 

species that use more open older forests and 

are dependent on fire, but these actions can 

degrade habitat quality for species that use 

dense older forests, which may have developed 

owing to fire exclusion. 

c. Excluding fire from dry forests will increase 

surface and canopy fuel continuity and 

increase size of patches of high-severity fire 

when fires escape suppression and burn under 

extreme conditions. 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

d. Excluding fire and disturbance from dry 

forests will typically increase stand density 

and shift species composition toward late-successional species and species that use dense 

forests and lower the resilience of these forests 

to fire and drought. 

e. Excluding fire from moist forests (especially in 

the drier parts of the moist forests) likely reduces 

landscape-scale vegetation diversity and the area 

of diverse early-successional forest and may 

increase the sizes of high-severity fire patches.

f. The effects of stand-level management actions 

may be different when examined at different 

spatial scales and time periods. Multiscale and 

multitemporal analysis can help reveal how 

management effects differ with spatial and 

temporal scale. 

g. Tradeoffs among goals are particularly 

strong in managing road networks, because 

existing road networks can negatively affect 

some native species and ecosystem processes, but they also can support landscape 

restoration, fire management, and active 

management to support other ecological and 

socioeconomic goals. 

Post-Wildfire Management (Question 10) 

1. Salvage logging after wildfire does not typically 

generate ecological benefits for species and processes associated with patches of high-severity 

wildfire. However, in some cases (e.g., where fire 

exclusion has led to dense forests), post-wildfire 

management may be justified, including: 

a. Planting key tree species after wildfires in 

uncharacteristically large patches of high-severity fire that may otherwise be slow to 

regenerate where seed sources are lacking

b. Thinning high-density post-wildfire regeneration as appropriate to increase heterogeneity 

and resilience to drought and wildfire.

c. Salvaging postfire pole and small-sized 

trees that have grown in during the period 

of fire exclusion in dry zone forests, where 

these may constitute a significant fuel bed 

for reburns in the future, while retaining the 

medium, large, and very large trees as dead 

snags and down logs.

2. Actions can be taken to mitigate many of the 

potentially undesirable effects of salvage logging, 

particularly by retaining many areas that are not 

salvaged to ensure heterogeneity and availability of 

those distinctive postfire communities. 

Acknowledgments

We acknowledge Ramona Butz, Tom Demeo, Malcolm 

North, Robyn Darbyshire, Kim Mellen-Mclean, Emily Platt, 

Hugh Safford, Joe Sherlock, Max Wahlberg, and managers 

from the Pacific Northwest and Southwest Regions of the 

U.S. Forest Service for their reviews on earlier versions of 

this chapter. Seven anonymous reviewers provided many 

valuable suggestions. Keith Olsen, Rob Pabst, and Kathryn 

Ronnenberg helped prepare figures. 

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Appendix 1: Crosswalk of Simpson (2013) Potential Vegetation Zones With Existing 

Vegetation From the Classification and Assessment With Landsat of Visible Ecological 

Grouping (CALVEG)

Table 3-8—Potential vegetation zones with existing vegetation from CALVEGa

Potential vegetation zone CALVEG Regional Dominance 1

Western hemlock Douglas-fir (40.3%), white fir (18.5%), Jeffrey pine (15.5%), tanoak (madrone) (9%), black oak 

(3.9%), ultra mafic mixed conifer (3.7%), California bay (2.9%), red fir (2.4%) 

Tanoak Douglas-fir (40.3%), tanoak (madrone) (11.3%), Oregon white oak (6.2%), California bay (5%)

Shasta red fir Red fir (33.2%), white fir (10.1%), Jeffrey pine (10.1%), barren (10%), mixed conifer–fir (8.1%), 

alpine grasses and forbs (5.1%), pinemat manzanita (5%), subalpine conifers (4.9%), upper 

montane mixed chaparral (2.9%), perennial grasses and forbs (2.1%)

Port Orford cedar Douglas-fir (46.6%), ultramafic mixed conifer (24.8%), Douglas-fir–white fir (7.9%), tanoak 

(madrone) (2.9%), Douglas-fir–ponderosa pine (2.9%), mixed conifer–pine (2.2%), Oregon 

white oak (2%)

Other pine Lower montane mixed chaparral (16.5%), gray pine (10.1%), chamise (8%), Oregon white oak 

(7.1%), interior mixed hardwood (6.6%), canyon live oak (5.6%), blue oak (5.6%), annual 

grasses and forbs (4.8%), Douglas-fir–ponderosa pine (4.4%), scrub oak (3.6%), Douglas-fir 

(3.5%), mixed conifer–pine (3.3%), Sargent cypress (3.2%), black oak (2.5%), knobcone pine 

(2.2%), ponderosa pine (2%) 

Grand fir/white fir Mixed pine conifer (27.1%), white fir (19%), Douglas-fir–white fir (14%), Douglas-fir (10.6%), 

Douglas-fir–ponderosa pine (6.3%), red fir (5.9%), mixed conifer–fir (2.5%), upper montane 

mixed chaparral (2%)

Douglas-fir Douglas-fir (29.3%), Douglas-fir–ponderosa pine (13.3%), Oregon white oak (12.7%), mixed 

conifer–pine (7.8%), lower montane mixed chaparral (5.3%), canyon live oak (4.6%), black 

oak (4%), interior mixed hardwood (3.8%), ponderosa pine (3.2%), annual grasses and 

forbs (2%).

Juniper Annual grasses and forbs (45.3%), mixed conifer–pine (17.2%), barren (8.3%), Douglas-fir–

ponderosa pine (7%), upper montane mixed chaparral (4.3%), perennial grasses and forbs 

(2.9%), manzanita chaparral (2.8%), ponderosa pine–white fir (2.3%), Jeffrey pine (2%) 

a

 Percentages indicate the percentage of the potential vegetation zone that falls into the CALVEG class. Existing vegetation comes from 

the Regional Dominance Type 1 field in the CALVEG database and indicates the primary, dominant vegetation alliance. The listed existing 

vegetation alliances comprise 95 percent of each potential vegetation zone in northern California. Current vegetation types with less than 2 

percent cover in a potential vegetation zone are not shown. For information on CALVEG, see: http://www.fs.usda.gov/detail/r5/landmanagement/

resourcemanagement/?cid=stelprdb5347192 .
232

GENERAL TECHNICAL REPORT PNW-GTR-966

Appendix 2: Fire Regime Mapping Method

Wildfire studies in Pacific Northwest forests have shown 

strong correlations between fire occurrence and area 

burned with summer temperature and precipitation (Dalton et al. 2013, Littell et al. 2009, McKenzie et al. 2004). 

Accordingly, we used climate variables for temperature and 

precipitation that coincided with the regional fire season as 

covariates in this mapping method. Our climate data source 

was the parameter-elevation regressions on independent 

slopes model (PRISM) climate normal data (PRISM 2015) 

for the period 1971–2000. We included a third variable for 

density of lightning-ignited wildfires data from 1970 to 

2002 (Brown et al. 2002) because fires in some regions may 

be limited by lack of ignitions during dry periods. Each 

mapping variable was classified into categories based on 

the equal divisions of the distributions in the forested areas. 

Thus, each class covered a relatively equal proportion of 

the forested landscape. Temperature was divided into five 

classes, and the other two variables were divided into three 

classes (table 3-9).

Potential vegetation zones (potential vegetation types) 

were summarized across all combinations of variable 

classes. Review of these data (e.g., temperature, precipitations, lightning ignition, density, and vegetation types) and 

expert opinion were used to assign each variable combination to one of four fire regimes: (1) infrequent (>200-year 

return interval) stand replacing; (2) somewhat infrequent 

to moderately frequent (50- to 200-year return interval), 

mixed severity; (3) frequent (15- to 50-year return interval), 

mixed severity; and (4) very frequent (5- to 25-year return 

interval), low severity (table 3-10). The final map product 

was filtered to remove pixel noise using a 3 by 3 majority 

filtering process.

Table 3-9—Variable map classification scheme based on quantile (by forested area) breaks

Rank

July–August mean monthly 

maximum temperature 

May–September 

mean monthly precipitation 

Lightning ignition density

1970–2002 

°C Millimeters Ignitions/km2

Very low 15–23 NA NA

Low 23–25 6–32 <0.05

Moderate 25–27 32–54 0.05–1.2

High 27–30 54–189 >1.2

Very high 30–37 NA NA

NA = not applicable.
233

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

1.1. Infrequent (>200-year return intervals) stand 

replacing (Landfire group V)

a. Potential vegetation type (PVT): wetter/colder parts of western hemlock, 

Pacific silver fir, mountain hemlock. 

Cover types: Douglas-fir, western hemlock, Pacific silver fir, noble fir, mountain hemlock

b. Area dominated by large to very large 

patches (103 to 106 ac) of high-severity fire, low and moderate severity also 

occur. Small- to medium-size patches 

were most frequent. 

1.2. Moderately frequent to somewhat infrequent 

(50- to 200-year return intervals) mixed 

severity (Landfire regime group III)

a. PVT: drier/warmer parts of western 

hemlock, Pacific silver fir and others. 

Cover types: Douglas-fir, western hemlock, Pacific silver fir, noble fir.

b. Mixed severity in space and time, 

typically including large (103 to 104 ac) 

patches of high-severity fire and areas 

of low- and moderate-severity fire. 

Small patches of high severity would 

be frequent. 

1. Dry forests, primarily east side of Washington and 

Oregon, southwest Oregon, northwest California

1.1. Frequent (15- to 50-year return intervals), 

mixed severity (Landfire regime group I 

and III) 

a. PVT: Douglas-fir, grand fir, white fir, 

tanoak. Cover type: Douglas-fir, white 

fir, red/noble fir, western white pine

b. Mixed-severity fire with medium to 

large (102 to 104 ac) patches of high-severity fire

1.2. Very frequent (5- to 25-year return intervals) 

low severity (Landfire regime group I)

a. PVT: ponderosa pine, dry to moist grand 

fir, white fir. Cover types: ponderosa 

pine, Douglas-fir, mixed pine, oak

b. Dominated by low-severity fire with 

fine-grained pattern (<10° to 102 ac) 

of high-severity fire effects, large 

patches of high-severity fire rare in 

forests except in earlier seral stage 

(e.g., shrub fields). 
234

GENERAL TECHNICAL REPORT PNW-GTR-966 Table 3-10—Temperature, precipitation, and lightning class levels (see table 3-9) of fire regimes and percentage of vegetation zones in that set of environmental classes (continued) Regime Temp. Precip. Lightning PISI THPL TSHE CHLA LIDE SESE ABAM TSME ABLA ABMAS PSME ABGRC PIPO PINUS OAK Infrequent—high severity Very low Low Low 0 0 0 0 0 12 0 0 0 0 0 0 0 0 0 Infrequent—high severity Very low Low High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 Infrequent—high severity Very low Moderate Low 14 1 0 0 0 2 0 0 0 0 0 0 0 0 0 Infrequent—high severity Very low Moderate Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 Infrequent—high severity Very low Moderate High 0 0 0 0 0 0 3 7 31 1 0 1 0 0 0 Infrequent—high severity Very low High Low 32 0 2 0 0 0 2 0 0 0 0 0 0 0 0 Infrequent—high severity Very low High Moderate 2 0 1 0 0 0 32 15 13 3 0 0 0 0 0 Infrequent—high severity Very low High High 0 0 0 0 0 0 5 10 9 0 0 0 0 0 0 Infrequent—high severity Low Low Low 0 3 0 0 0 8 0 0 0 0 0 0 0 0 0 Infrequent—high severity Low Low High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Infrequent—high severity Low Moderate Low 4 41 4 0 0 2 0 0 0 0 1 0 0 0 0

Infrequent—high severity Low Moderate Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Infrequent—high severity Low Moderate High 0 0 0 0 0 0 2 18 13 6 1 6 0 0 0

Infrequent—high severity Low High Low 37 18 23 0 1 0 2 0 0 0 1 0 0 0 0

Infrequent—high severity Low High Moderate 9 0 7 1 2 0 30 4 0 0 0 0 0 0 0

Infrequent—high severity Low High High 0 0 1 0 0 0 7 21 0 0 0 0 0 0 0

Infrequent—high severity Moderate Low Low 0 0 0 0 0 1 0 0 0 0 0 0 0 0 0

Infrequent—high severity Moderate Low High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Infrequent—high severity Moderate Moderate Low 0 1 0 0 0 0 0 0 0 0 0 0 0 0 0

Infrequent—high severity Moderate Moderate Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Infrequent—high severity Moderate Moderate High 0 0 0 0 0 0 0 1 0 0 0 0 0 0 0

Infrequent—high severity Moderate High Low 1 1 11 0 0 0 0 0 0 0 4 0 0 0 0

Infrequent—high severity Moderate High Moderate 0 0 0 0 0 0 1 0 0 0 0 0 0 0 0

Infrequent—high severity Moderate High High 0 0 0 0 0 0 1 1 0 0 0 0 0 0 0

Infrequent—high severity High Low Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Infrequent—high severity High Moderate Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Infrequent—high severity High High Low 0 0 2 0 0 0 0 0 0 0 0 0 0 0 0

Infrequent—high severity High High Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Infrequent—high severity Very high High Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Infrequent—high severity Very high High Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Very low Low Moderate 0 0 0 0 0 0 0 1 0 0 0 0 0 0 0
235

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Table 3-10—Temperature, precipitation, and lightning class levels (see table 3-9) of fire regimes and percentage of vegetation zones in that set of environmental classes (continued)Regime Temp. Precip. Lightning PISI THPL TSHE CHLA LIDE SESE ABAM TSME ABLA ABMAS PSME ABGRC PIPO PINUS OAKModerately frequent—

mixed severity

Very low Low High 0 0 0 0 0 0 0 1 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Very low Moderate Moderate 0 0 0 0 0 0 3 4 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Very low Moderate High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Very low High Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Very low High High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Low Low Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Low Low High 0 0 0 0 0 0 0 1 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Low Moderate Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Low Moderate Moderate 0 0 1 0 0 0 2 1 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Low Moderate High 0 0 0 0 0 0 0 1 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Low High Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Low High Moderate 0 0 1 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Low High High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Moderate Low Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Moderate Low High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Moderate Moderate Low 0 14 7 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Moderate Moderate Moderate 0 0 4 0 0 0 1 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Moderate Moderate High 0 0 0 0 0 0 0 6 0 0 0 0 0 0 0
236

GENERAL TECHNICAL REPORT PNW-GTR-966 Table 3-10—Temperature, precipitation, and lightning class levels (see table 3-9) of fire regimes and percentage of vegetation zones in that set of environmental classes (continued) Regime Temp. Precip. Lightning PISI THPL TSHE CHLA LIDE SESE ABAM TSME ABLA ABMAS PSME ABGRC PIPO PINUS OAK Moderately frequent— mixed severity Moderate High Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 Moderately frequent— mixed severity Moderate High Moderate 1 0 7 0 0 0 2 0 0 0 0 0 0 0 0 Moderately frequent— mixed severity Moderate High High 0 0 3 0 0 0 4 3 0 0 0 0 0 0 0 Moderately frequent— mixed severity High Low Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 Moderately frequent— mixed severity High Low Moderate 0 0 1 0 0 0 0 0 0 0 0 0 0 0 0 Moderately frequent—

mixed severity

High Low High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

High Moderate Low 0 21 5 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

High Moderate Moderate 0 0 8 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

High Moderate High 0 0 3 0 0 0 0 2 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

High High Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

High High Moderate 0 0 3 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

High High High 0 0 3 0 0 0 0 1 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Very high Low Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Very high Moderate Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Very high Moderate High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Very high High Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Moderately frequent—

mixed severity

Very high High High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Frequent—mixed severity Very low Low Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Frequent—mixed severity Very low Low Moderate 0 0 0 0 0 1 0 0 1 0 0 1 0 0 0
237

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Table 3-10—Temperature, precipitation, and lightning class levels (see table 3-9) of fire regimes and percentage of vegetation zones in that set of environmental classes (continued)Regime Temp. Precip. Lightning PISI THPL TSHE CHLA LIDE SESE ABAM TSME ABLA ABMAS PSME ABGRC PIPO PINUS OAK

Frequent—mixed severity Very low Low High 0 0 0 0 0 0 0 0 4 0 0 1 0 0 0

Frequent—mixed severity Very low Moderate Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Frequent—mixed severity Very low Moderate Moderate 0 0 0 0 0 3 0 0 9 0 0 0 0 0 0

Frequent—mixed severity Very low Moderate High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Frequent—mixed severity Very low High Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Frequent—mixed severity Low Low Low 0 0 0 0 0 1 0 0 0 0 0 0 0 0 0

Frequent—mixed severity Low Low Moderate 0 0 0 0 0 0 0 0 0 0 1 2 0 0 0

Frequent—mixed severity Low Low High 0 0 0 0 0 0 0 0 10 67 1 4 0 0 0

Frequent—mixed severity Low Moderate Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Frequent—mixed severity Low Moderate Moderate 0 0 0 0 0 4 0 0 2 0 0 2 0 0 0

Frequent—mixed severity Low Moderate High 0 0 0 0 0 0 0 0 0 0 0 1 0 0 0

Frequent—mixed severity Low High Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Frequent—mixed severity Low High Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Frequent—mixed severity Low High High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Frequent—mixed severity Moderate Low Low 0 0 0 0 0 9 0 0 0 0 0 0 0 0 0

Frequent—mixed severity Moderate Low Moderate 0 0 0 0 0 1 0 0 0 0 1 3 4 0 0

Frequent—mixed severity Moderate Low High 0 0 0 0 0 0 0 1 1 0 0 0 0 0 0

Frequent—mixed severity Moderate Moderate Low 0 0 0 0 0 2 0 0 0 0 12 0 0 0 0

Frequent—mixed severity Moderate Moderate Moderate 0 0 0 0 2 4 0 0 0 0 0 1 0 0 0

Frequent—mixed severity Moderate Moderate High 0 0 0 0 1 0 0 0 0 0 0 8 0 0 0

Frequent—mixed severity Moderate High Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Frequent—mixed severity Moderate High Moderate 0 0 0 13 3 1 0 0 0 0 0 0 0 0 0

Frequent—mixed severity Moderate High High 0 0 0 0 0 0 0 0 0 0 0 1 0 0 0

Frequent—mixed severity High Low Low 0 0 0 0 1 28 0 0 0 0 3 2 1 0 8

Frequent—mixed severity High Low Moderate 0 0 0 0 1 1 0 0 0 0 3 7 19 0 1

Frequent—mixed severity High Low High 0 1 0 0 0 0 0 0 0 0 0 0 0 0 0

Frequent—mixed severity High Moderate Low 0 0 0 0 0 2 0 0 0 0 4 0 0 0 0

Frequent—mixed severity High Moderate Moderate 0 0 0 0 8 4 0 0 0 0 2 2 0 0 0

Frequent—mixed severity High Moderate High 0 0 0 0 3 0 0 0 0 0 0 10 0 0 0

Frequent—mixed severity High High Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Frequent—mixed severity High High Moderate 0 0 0 75 11 0 0 0 0 0 1 0 0 0 0

Frequent—mixed severity High High High 0 0 0 0 1 0 0 0 0 0 0 1 0 0 0
238

GENERAL TECHNICAL REPORT PNW-GTR-966 Table 3-10—Temperature, precipitation, and lightning class levels (see table 3-9) of fire regimes and percentage of vegetation zones in that set of environmental classes (continued) Regime Temp. Precip. Lightning PISI THPL TSHE CHLA LIDE SESE ABAM TSME ABLA ABMAS PSME ABGRC PIPO PINUS OAK Frequent—mixed severity Very high Low Low 0 0 0 0 1 12 0 0 0 0 12 0 0 21 59 Frequent—mixed severity Very high Low Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 Frequent—mixed severity Very high Low High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 Frequent—mixed severity Very high Moderate Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 Frequent—mixed severity Very high Moderate Moderate 0 0 0 0 13 0 0 0 0 0 2 1 0 5 0 Frequent—mixed severity Very high Moderate High 0 0 0 0 0 0 0 0 0 0 0 1 0 0 0 Frequent—mixed severity Very high High Moderate 0 0 0 11 5 0 0 0 0 0 0 0 0 0 0 Frequent—mixed severity Very high High High 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0 Very frequent—low severity Very low Low High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 Very frequent—low severity Low Low High 0 0 0 0 0 0 0 0 0 1 0 1 0 0 0

Very frequent—low severity Low Moderate High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Very frequent—low severity Moderate Low Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Very frequent—low severity Moderate Low High 0 0 0 0 0 0 0 0 3 22 2 10 9 0 0

Very frequent—low severity Moderate Moderate High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Very frequent—low severity High Low Moderate 0 0 0 0 0 0 0 0 0 0 1 0 1 0 0

Very frequent—low severity High Low High 0 0 0 0 3 0 0 0 1 0 7 17 48 0 0

Very frequent—low severity High Moderate Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Very frequent—low severity High Moderate High 0 0 0 0 0 0 0 0 0 0 0 1 0 0 0

Very frequent—low severity High High High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Very frequent—low severity Very high Low Low 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Very frequent—low severity Very high Low Moderate 0 0 0 0 10 0 0 0 0 0 14 2 14 55 26

Very frequent—low severity Very high Low High 0 0 0 0 22 0 0 0 0 0 20 8 3 13 4

Very frequent—low severity Very high Moderate Moderate 0 0 0 0 0 0 0 0 0 0 0 0 0 1 0

Very frequent—low severity Very high Moderate High 0 0 0 0 5 0 0 0 0 0 5 3 0 5 1

Very frequent—low severity Very high High High 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Temp. = temperature; Precip. = precipitation; ABAM = Abies amabilis; ABLA = Abies lasiocarpa; ABMAS = Abies magnifica var. shastensis; ABGRC = Abies grandis/concolor; CHLA = Chamaecyparis lawsoniana; LIDE = Libocedrus decurrens; OAK = Quercus spp.; PIPO = Pinus ponderosa; PINUS = Pinus spp.; PISI = Picea sitchensis; PSME = Pseudotsuga menziesii; 

SESE = Sequoia sempervirens; THPL = Thuja plicata; TSME = Tsuga mertensiana; TSHE = Tsuga heterophylla. 
239

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Appendix 3: Summary of Fire History Studies in the Northwest Forest Plan Table 3-11—Fire history studies in the Northwest Forest Plan area by potential vegetation zone (continued) Vegetation zone Study Extent/time period Method Frequency/ return interval Rotation Low/ moderate/ high High- severity patch size Interpreted regime Mapped regime Hectares Years Percent Hectares Redwood: Stuart 1987 300 ha 1898–1940 Scars 7.8

—

—

— Low F-MS

Finney and Martin 

1989

~600 ha

1300–1860

Scars 10.1

—

—

— Low I-HS

Brown and 

Swetnam 1994

<1000 ha

1714–1962

Scars 9.9

—

—

— Low F-MS

Brown et al. 1999 Unknown Age, scars 7–13 

—

—

— Low F-MS

Brown and Baxter 

2003

20 316 ha

1550–1930

Scars 6–20

—

—

—

Low F-MS

Western hemlock:

Means 1982 Unknown Scars 100

—

—

Mixed MF-MS

Fahnestock and 

Agee 1983

Western 

Washington

pre-1934

Age class from 

historical survey 

records

— 598

—

— High I-HA

Stewart 1986 <1 ha

~1200–1982

Age, scars 50a —

—

— Mixed MF-MS

Yamaguchi 1986 Unknown

Post-1480

Age, scars 40–150 

—

—

— Mixed I-HS

Teensma 1987 11 000 ha

1482–1952

Age, Scars 114 78

—

— Mixed MF-MS

Agee et al. 1990a 

or b?

3500 ha

1573–1985

Age, Scars 137

—

—

—

Mixed I-HS

Morrison and 

Swanson 1990

1940 ha

1150–1985

Age, Scars 96 95 0–86/

0–60/

0–100

<110 ha Mixed/high MF-MS

Garza 1995 3540 ha

Pre-1910

Age, scars 93-158 134 24–41/9–23/ 

25–54

Mixed MF-MS
240

GENERAL TECHNICAL REPORT PNW-GTR-966 Table 3-11—Fire history studies in the Northwest Forest Plan area by potential vegetation zone (continued) Vegetation zone Study Extent/time period Method Frequency/ return interval Rotation Low/ moderate/ high High- severity patch size Interpreted regime Mapped regime Hectares Years Percent Hectares Impara 1997 ~140 000 ha 1478–1909 Age, scars 85 271 — — Mixed MF-MS/ I-HS

Wetzel and Fonda 

2000

2500 ha

1400–1985

Age, growth 

release

21.3b — — — Low I-HS

Agee and 

Krusemark 2001

26 000 ha

Pre-1900

Age, live residual 

structure from 

air photos

— 296 7–9/

18–31/

62–90

— High MF-MS

Robbins 1999 ~1562 km2

1700–1990

Age, scars 49 (2–191) — — — Low/mixed MF-MS

Olsen and Agee 

2005

~7000 ha

1650–1900

Age, scars 2-167 — — — Mixed MF-MS

Weisberg 2009 14 504 ha

1550–1849

Age, scars — 162 — — Mixed MF-MS

Wendel and 

Zabowski 2010

1873 ha

1568–2007

Age, scars 127 140 — — Mixed/high I-HS

Silver fir:

Hemstrom and 

Franklin 1982

~53 000 ha

1200–1850

Age — 465 — — High I-HS

Fahnestock and 

Agee 1983

Western 

Washington

pre-1934

Age class from 

historical survey 

records

— 834 — — High I-HS

Agee et al. 1990a 3500 ha

1573–1985

Age, scars 108–137 — — — Mixed I-HS

Morrison and 

Swanson 1990

1940 ha

1150–1985

Age, scars 239 149 0–80/

0–78/

0–100

<50 Mixed/high MF-MS
241

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Table 3-11—Fire history studies in the Northwest Forest Plan area by potential vegetation zone (continued)Vegetation zone Study

Extent/time 

period Method

Frequency/

return interval Rotation

Low/

moderate/

high

Highseverity 

patch size 

Interpreted 

regime

Mapped 

regime

Hectares Years Percent Hectares

Garza 1995 3540 ha

Pre-1910

Age, scars 154–246 — 24–57/20–22/ 

45–50

— Mixed MF-MS

Mountain hemlock:

Dickman and Cook 

1989

18 000 ha

Post-1400

Age — — — >3200 High I-HS

Fahnestock and 

Agee 1983

Western 

Washington

pre-1934

Age class from 

historical survey 

records

— 598 — — Mixed I-HS

Agee et al. 1990a 3500 ha

1573–1985

Age, scars 137 — — — Mixed I-HS

Subalpine:

Fahenstock and 

Agee 1983

Western 

Washington

pre-1934

Age class from 

historical survey 

records

— 800 — — High I-HS

Agee et al. 1990a 3500 ha

1573–1985

Age, scars 109 — — — Mixed I-HS

Douglas-fir and grand fir/white fir:

Leiberg 1903 Southern Oregon

~1900

Historical land 

survey

— — ~14 000 High MF-MS/ 

F-MS

Weaver 1959 Unknown Scars 47 — — — Mixed VF-LS

Agee et al. 1990a 3500 ha

1573–1985

Age, scars 52-93 — — — Mixed I-HS

Agee 1991 197 ha

1760–1930

Age, scars 16 37-64 — — Low/mixed VF-LS

Bork 1985 ~100 ha

Pre-1900

Scars 8 — — ~400 Low VF-LS

Wills and Stuart 

1994

~20 ha

1745–1849

Age scars 10.3–17.3 — — — Low VF-LS
242

GENERAL TECHNICAL REPORT PNW-GTR-966 Table 3-11—Fire history studies in the Northwest Forest Plan area by potential vegetation zone (continued) Vegetation zone Study Extent/time period Method Frequency/ return interval Rotation Low/ moderate/ high High- severity patch size Interpreted regime Mapped regime Hectares Years Percent Hectares Taylor and Skinner 1998 1570 ha 1627–1849 Age, scars 14.5 19 59/ 27/ 14 — Low/ mixed VF-LS

Van Norman 1998 45 000 ha

1480–1996

Age, scars 123 — — — Mixed MF-MS

Brown et al. 1999 2000 ha

1820–1945

Age, scars 7.7–13 — — — Low F-MS

Everett et al. 2000 3240–12 757 ha

~1700–1860

Scars 6.6–7 11–12.2 — 2.4–40 Low F-MS

Stuart and Salazar 

2000

~120 ha

1614–1944

Age, scars 27 (12–161) — — — Low VF-LS

Taylor and Skinner 

2003

2325 ha

Pre-1905

Age, scars 11.5–16.5 19 – — Low/mixed VF-LS

Wright and Agee 

2004

~30 000 ha

1562–1995

Scars 19–24 — — 10-100 Low/mixed MF-MS

Hessburg et al. 

2007

~72 000 ha

~1930

Historical aerial 

photos

— — 18/

58/

24

~10 000 Mixed MF-MS/ 

F-MS

Baker 2012 140 400 ha

~1770–1880

Live structure 

from historical 

inventory

— 496c 18/

59/

23

— Mixed F-MS/VF-LS 

Ponderosa pine:

Weaver 1959 Unknown Scars 11–16 — — — Low VF-LS

Soeriaatmadja 

1966

1500–5000 ha

Unknown

Scars 3–36 — — — Low VF-LS

West 1969 Unknown Age — — — <0.26 Low VF-LS
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Table 3-11—Fire history studies in the Northwest Forest Plan area by potential vegetation zone (continued)Vegetation zone Study

Extent/time 

period Method

Frequency/

return interval Rotation

Low/

moderate/

high

Highseverity 

patch size 

Interpreted 

regime

Mapped 

regime

Hectares Years Percent Hectares

Bork 1985 ~100 ha

Pre-1900

Scars 4–7 — — — Low VF-LS

Morrow 1985 2 ha

Pre-1900

Age — — — <0.35 Low VF-LS

Hessburg et al. 

2007

~106 000 ha

1930–1940

Live structure 

from historical 

aerial photos

— — 30/

58/

12

— Low/mixed VF-LS

Baker 2012 123 500 ha

~1770–1880 Live structure 

from historical 

inventory

— 705c 40/

44/

16

— Low/mixed VF-LS

— = No value in cella Interpreted regimes are classified on fire frequency classes: low <35 years, mixed 35 to 200 years, high >200 years). In cases in which fire frequency was not available, we considered fire rotation and the percentage of high-severity fire. Mapped regimes are predicted from combinations of summer precipitation, summer temperature, and lighting frequency and follow the four class regimes used in this chapter: F-MS is frequent–mixed severity, I-HS is infrequent–high severity, MF-MS is moderately frequent–mixed severity, and VF-LS is very frequent–low severity. 1 ha = 2.47 ac.b Stewart noted 15 fires over a 750-year period.c Estimated at a 200-ha (494 ac) scale.d Rotation for high-severity only.
244

GENERAL TECHNICAL REPORT PNW-GTR-966

A northern spotted owl in the McKenzie River Basin in Oregon.

Photo by John and Karen Hollingsworth, U.S. Fish and Wildlife Service.
245

Chapter 4: Northern Spotted Owl Habitat and 

Populations: Status and Threats 

Damon B. Lesmeister, Raymond J. Davis, 

Peter H. Singleton, and J. David Wiens1

Introduction 

The northern spotted owl (Strix occidentalis caurina) was 

listed as threatened under the Endangered Species Act 

in 1990 (USFWS 1990). Providing adequate amounts of 

suitable forest cover to sustain the subspecies was a major 

component of the first recovery plan for northern spotted 

owls (USFWS 1992) and a driver in the basic reserve 

design and old-forest restoration under the Northwest 

Forest Plan (NWFP, or Plan) (USDA and USDI 1994). 

The reserve design included large contiguous blocks 

of late-successional forest, which was expected to be 

sufficient to provide habitat for many interacting pairs of 

northern spotted owls. As such, the selection of reserves 

generally favored areas with the highest quality oldgrowth forests, but some areas of younger forest were also 

included with the expectation that they would eventually 

develop suitable forest structure characteristics and 

contribute to spatial patterns that would sustain spotted 

owl populations. 

Northern spotted owls are now one of the most studied birds in the world. Much of the research and interest 

in spotted owls stem from the economic and ecological 

implications surrounding management for the subspecies. 

Courtney et al. (2004) and the U.S. Fish and Wildlife 

Service (USFWS 2011b) completed comprehensive reviews 

and syntheses of scientific information regarding the status, 

ecology, and threats to the northern spotted owl. In the 

10-year science synthesis of the NWFP, Raphael (2006) 

detailed the expectations and observations for northern 

spotted owl populations and suitable forest types under 

the Plan. Here we provide a 20-year synthesis of northern 

spotted owl science and review key information concerning 

the ecology and expectations for conservation of northern 

spotted owls under the NWFP. We build upon previous 

syntheses and address guiding questions by focusing on the 

scientific understanding accumulated from 2005 to 2016 

on the ecology, conservation, and management of northern 

spotted owls. We also provide an overview of the main 

scientific debates surrounding conservation and management of northern spotted owls. We discuss the distinction 

between associated forest cover types and the relative 

value of habitat in different forest types for the subspecies. 

Where needed, we review and draw inference from research 

related to Mexican spotted owls (S. o. lucida) and California 

spotted owls (S. o. occidentalis), but keep the focus of this 

synthesis on published literature specific to northern spotted 

owls (spotted owl hereafter). 

Major threats to spotted owls identified at the time of 

design and initial implementation of the NWFP and species 

recovery plan included the effects of past and current timber 

harvest, loss of old forest to wildfire, and competition with 

rapidly encroaching barred owls (Strix varia) (USDA and 

USDI 1994, USFWS 1992). Studies of associations between 

spotted owls and forest cover published since 2005 have 

reinforced previous work indicating a strong association of 

nest and roost sites with older forest conditions and a wider 

range of forest cover types used for foraging and dispersal 

(Anthony et al. 2006; Carroll and Johnson 2008; Dugger 

et al. 2005, 2016; Forsman et al. 2011, 2015; Hamer et al. 

2007; Irwin et al. 2012, 2013; McDonald et al. 2006; Olson 

et al. 2005; Sovern et al. 2015). In the southern portions of 

the range, abiotic environmental factors begin to play larger 

roles in territorial owl use (Glenn et al. 2017), and at the 

very southern end of the range (Marin County, California), 

spotted owls occur at higher densities and tend to nest in a 

wider variety of forest cover types and ages (Stralberg et 

al. 2009). The difference in localized spotted owl densities 

and generalist vegetation associations appear to be driven 

by the diversity of forest conditions and high prey density 

prevalent in that landscape. 

1

 Damon B. Lesmeister is a research wildlife biologist, U.S. 

Department of Agriculture, Forest Service, Pacific Northwest 

Research Station, 3200 SW Jefferson Way, Corvallis, OR 97331; 

Raymond J. Davis is the northern spotted owl and old growth monitoring lead, U.S. Department of Agriculture, Forest Service, Pacific 

Northwest Region, 3200 SW Jefferson Way, Corvallis, OR 97331; 

Peter H. Singleton is a research wildlife biologist, U.S. Department 

of Agriculture, Forest Service, Pacific Northwest Research Station, 

1133 N Western Avenue, Wenatchee, WA 98801; and J. David 

Wiens is a research wildlife biologist, U.S. Department of the Interior, Geological Survey, Forest and Rangeland Ecosystem Science 

Center, 3200 SW Jefferson Way, Corvallis, OR 97331.
246

GENERAL TECHNICAL REPORT PNW-GTR-966

Every study that has assessed rangewide population 

trends of spotted owls found steady declines since standardized monitoring efforts started in 1985 (Anthony et al. 

2006, Dugger et al. 2016, Forsman et al. 2011, Franklin et al. 

1996). Loss of suitable forest and competitive interactions 

with barred owls are the primary threats that have contributed to those declines. In the following sections, we review 

recent information on the status and trends of spotted owl 

populations and suitable forest, effects of interactions with 

barred owls, prey ecology, disturbance impacts, climate 

change, and other threats. We also review population trends 

and range expansion of barred owls, their habitat and 

prey, and identify other sensitive wildlife and ecological 

processes that may ultimately be affected by the invasion of 

barred owls. We conclude by outlining considerations for 

management and research needs for spotted owls and forest 

types most critical to their persistence.

Guiding Questions

We used the following questions received from forest 

managers to guide our synthesis and focus on relevant 

spotted owl literature. Following each question, we provide 

the section that most effectively addresses the question, or 

if a question could not be adequately addressed because of a 

lack of published literature on the subject.

1. What is the current understanding about spotted 

owl population status? Will continuing to implement the NWFP reverse the downward trend in 

spotted owl populations?

• Information can be found in the “Population 

Status and Trends” and “Conclusions and 

Management Considerations” sections.

2. Is the NWFP maintaining or restoring forest conditions necessary to support viable populations of 

spotted owls?

• Despite old-forest loss to wildfire and timber 

harvest, implementation of the NWFP has been 

successful for putting federal lands on a trajectory for restoring forest capable of supporting spotted owls on federal lands. Information 

can be found in the “Habitat Status and 

Trends,” “Disturbance,” and “Conclusions and 

Management Considerations” sections. 

3. What are the effects of various timber management practices and wildfire on forests used by 

spotted owls?

• Information can be found in the “Habitat Status 

and Trends,” “Disturbance,” and “Research 

Needs” sections.

4. How is space use by spotted owls affected by timber management? Are there ways to modify management activities (i.e., silvicultural treatments) 

to benefit spotted owls? How do managed stands 

compare to untreated forests in terms of use by 

spotted owls?

• Information can be found in the “Habitat 

Status and Trends,” “Disturbance,” “Research 

Needs,” and “Conclusions and Management 

Considerations” sections. 

5. Do spotted owls use forests following wildfire? If 

so, how? Do the impacts of treatments that reduce 

risk of wildfire outweigh the risks of suitable forest 

loss resulting from wildfire?

• The short- and long-term response by spotted 

owls to wildfire remains largely unknown, and 

scientific debate remains. We were unable to 

fully address this question, but do provide a 

synthesis of available literature in the “Habitat 

Status and Trends,” “Disturbance,” “Research 

Needs,” and “Scientific Uncertainty” sections. 

6. How effective are protections for buffered areas 

around nest sites in retaining spotted owls across 

treated landscapes? Are site buffers equally 

effective as landscape-scale forest management 

in ensuring species persistence, dispersal, and 

habitat connectivity?

• We were unable to address this question fully 

owing to the lack of published literature, but 

some information about the effectiveness of 

buffered management areas can be found in the 

“Habitat Status and Trends,” and “Forest protection effectiveness” sections.
247

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

7. Which provides a higher level of spotted owl persistence: the current spotted owl critical habitat or 

the NWFP late-successional reserve network? 

• Information can be found in the “Habitat Status 

and Trends,” and “Forest protection effectiveness” sections.

8. Does treating late-successional stands improve 

spotted owl persistence when wildfire, insects, 

disease, and climate change threaten the ability of 

these forests to provide habitat for spotted owls? 

• Information can be found throughout the 

chapter in the “Habitat Status and Trends,” 

“Barred Owls,” “Disturbance,” “Climate 

Change,” “Other Threats,” “Research Needs,” 

“Scientific Uncertainty,” and “Conclusions and 

Management Considerations” sections. 

9. What are the effects of barred owls on spotted 

owls? What is the relationship of wildfires to 

barred owl encroachment? Can a barred owl management program be effectively implemented at a 

scale that will have meaningful conservation value 

for spotted owls? 

• Information about the effects of barred owls 

is found in “Barred Owls.” We were unable to 

adequately address questions about the relationship between barred owls and wildfire, and 

barred owl management, because of a paucity 

of literature. In addition, some of this research 

was ongoing at the time this synthesis was 

being prepared. We provide further details in 

“Research Needs, Uncertainties, Information 

Gaps, and Limitations.”

10. What are the management considerations and 

research needs for spotted owls?

• Based on our synthesis of available literature 

within the context of the guiding management 

questions we received, we specifically address 

high-priority information needs in “Research 

Needs, Uncertainties, Information Gaps, and 

Limitations.” We conclude the chapter with 

“Conclusions and Management Considerations.” 

Key Findings

Population Status and Trends

Understanding vital rates (e.g., birth, death) and the 

factors affecting those parameters over time and space can 

provide crucial information for management and conservation. Since the listing of the spotted owl, demographic 

rates have been monitored in up to 14 demographic study 

areas distributed across the spotted owl’s geographic 

range. Franklin et al. (1996) developed a general framework to estimate demographic parameters and population 

trends of spotted owls that has been used in subsequent 

spotted owl population analyses. In the past 10 years, 

three meta-analyses (Anthony et al. 2006, Dugger et 

al. 2016, Forsman et al. 2011) documented a continued 

decline in spotted owl populations throughout their range. 

Those meta-analyses built upon the Franklin et al. (1996) 

methods to analyze survival, reproduction, and territory 

occupancy data that has been collected consistently for 

nearly three decades. 

The number of study areas in which spotted owls 

have been monitored has changed through time owing to 

changes in funding and institutional support. Anthony et 

al. (2006) used data from 14 study areas (1985 to 2003), 

Forsman et al. (2011) used data from 11 study areas (1985 

to 2008), and Dugger et al. (2016) used data from 11 

study areas (1985 to 2013) to evaluate survival, fecundity, 

recruitment, and rate of population change of spotted owls 

throughout the subspecies’ geographic range (fig. 4-1). 

Dugger et al. (2016) also investigated territory occupancy 

dynamics (gains and losses of occupied territories; 

i.e., local colonization and extinction rates). All three 

meta-analyses investigated relationships between population demography of spotted owls and the distribution of 

suitable forest cover types, local and regional variation in 

climatic conditions, and presence of barred owls. Study 

areas included in these meta-analyses comprised about 

9 percent of the spotted owl’s range, were distributed 

throughout the geographic range, and were selected to 

encompass the broad range of forest conditions used by 

the subspecies. 
248

GENERAL TECHNICAL REPORT PNW-GTR-966

Olympic

Rainier

Cle Elum

Coast Ranges

Tyee

Klamath

H.J. Andrews

South Cascades

Green Diamond

Hoopa

Northwest California

0 100 200 50

Kilometers

o

Figure 4-1—Locations of 11 study areas used in the analysis of vital rates and population 

trends of northern spotted owls, 1985 to 2013 (Dugger et al. 2016).
249

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

When the NWFP was developed, populations of 

spotted owls were estimated to be declining at about 4.5 

percent (confidence interval [CI] 1.1 to 7.9) per year (Burnham et al. 1996, USDA and USDI 1994). The population 

was expected to continue declining for up to 50 years until 

younger second-growth forest in reserves matured to a 

point at which it would provide suitable structural conditions for nesting and roosting (Lint 2005, USDA and USDI 

1994). During the first 10 years of the NWFP, the overall 

rate of population decline in Washington was much greater 

than in Oregon and California (Anthony et al. 2006, Lint 

2005). Three study areas in southern Oregon had stable 

populations during the first decade. Anthony et al. (2006) 

estimated an annual decline of 3.7 percent (CI = 1.9 to 5.5) 

across the range, but that analysis included lands outside 

of the NWFP monitoring area. The eight federal study 

areas within the boundaries of the NWFP area (i.e., lands 

under federal management) used for effectiveness monitoring of the NWFP had a decline of 2.4 percent (CI = 1.0 to 

3.8) compared to a 5.8 percent (CI = 2.6 to 9.0) decline for 

study areas composed primarily of nonfederal lands, suggesting that implementation of the NWFP had a positive 

effect on the demography of spotted owls (Anthony et al. 

2006, Raphael 2006). Forsman et al. (2011) estimated an 

annual decline of 2.9 percent (CI = 1.7 to 4.0) throughout 

the northern spotted owl’s range, and Davis et al. (2011) 

estimated an annual decline of 2.8 percent (CI = 1.5 to 

4.2) within the eight federal study areas. The most recent 

meta-analysis indicated that spotted owl populations were 

continuing to decline throughout the range of the subspecies, and that annual rates of decline were accelerating 

in many areas (Dugger et al. 2016). The population was 

declining by about 3.8 percent (CI = 0.1 to 7.5) per year 

and declines ranged from 1.2 percent to 8.4 percent per 

year depending on the study area (fig. 4-2) (Dugger et al. 

2016). For monitored populations, population change was 

more sensitive to adult survival than to recruitment (Glenn 

et al. 2010). Other studies have also documented declines 

in populations throughout the range of the spotted owl 

(Farber and Kroll 2012, Funk et al. 2010, Kroll et al. 2010). 

1.2

1.1

1.0

0.9

0.8

0.7

Mean lambda

Washington Oregon California

Study area

Cle Elum

Rainier

Olympic

Coast Ranges

H.J. Andrews

Tyee

Klamath

South Cascades, Oregon

Northwest California

Hoopa

Green Diamond — CB 

Green Diamond — TB 

Green Diamond — CA 

Green Diamond — TA 

Figure 4-2—Estimated mean rates of 

population change (mean lambda) and 

95 percent confidence limits for spotted 

owls from 1985 to 2013 at 11 sites: Cle 

Elum, Rainier, and Olympic, Washington; Coast Range, H.J. Andrews, and 

Tyee, Klamath, and South Cascades, 

Oregon; and northwest California, 

Hoopa, and Green Diamond, California 

(from Dugger et al. 2016). Estimates for 

Green Diamond are presented separately for control and treatment areas 

before (1990 to 2008) and after (2009 

to 2013) barred owls were removed 

on the treatment area (CB = control 

before removal; TB = treatment before 

removal; CA = control after removal; 

TA = treatment after removal) (Diller et 

al. 2016, Dugger et al. 2016).
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GENERAL TECHNICAL REPORT PNW-GTR-966

Habitat Status and Trends

Background and definitions—

Habitat for a species is an area that encompasses the necessary combination of resources and environmental conditions 

that promotes occupancy, survival, and reproduction of 

that species (Morrison et al. 2006). Typical wildlife habitat 

components include food, water, shelter (including nesting 

or denning sites), security from predators and competitors, 

and proper spatial arrangement of those features (Morrison 

et al. 2006). Although this concept of habitat may seem 

simple, the ways in which these individual components 

and animal needs interact in space and time result in very 

complex relationships (Mathewson and Morrison 2015). 

Spotted owl habitat has often been characterized as 

older forest with large trees and moderate to closed canopy 

(Courtney et al. 2004, Forsman et al. 1984). Spotted owl site 

occupancy has repeatedly been shown to be influenced by 

the presence of these forest conditions (e.g., Dugger et al. 

2016), likely because they often provide important habitat 

components that are suitable for nesting (e.g., cavities or 

platforms) (Sovern et al. 2011), abundant prey populations 

(Carey et al. 1992, Forsman et al. 2004, Wilson and Forsman 2013), and security from predators, including other 

raptors (Forsman et al. 1984, Sovern et al. 2014). An advantage of characterizing spotted owl habitat based on forest 

structure is that these forest types can be mapped for the 

entire subspecies’ range using remotely sensed data (Davis 

et al. 2016). Other habitat components like prey abundance, 

predation risk, and presence of competitors are much more 

difficult, if not impossible, to map independently. For example, the recent colonization of the range of northern spotted 

owls by barred owls has confounded efforts to quantify the 

amount of habitat available for spotted owls because barred 

owls use similar forest types and can displace spotted owls 

from those areas (see “Barred Owl” section below). 

In addition to availability, the arrangement of habitat 

components at a variety of scales is also important for understanding spotted owl habitat. Typically, spotted owl habitat 

is discussed in terms of forest cover types (stand-level 

forest structure and composition) most suitable for nesting, 

roosting, foraging, or dispersal (Davis et al. 2016, Lint 2005, 

Thomas et al. 1990). However, the spatial and temporal 

dynamics of suitable forest cover types, and how environmental conditions including climate and topography interact 

with vegetation patterns, are also important for producing 

and sustaining habitat for spotted owls (USFWS 2012a, 

2012b). For example, Glenn et al. (2017) constructed habitat 

models using forest cover types and abiotic environmental 

conditions, and estimated the density of spotted owl territories on a landscape before and after barred owl invasion. 

In this chapter, we define spotted owl habitat as those 

areas with the full suite of resources (e.g., abundant prey, 

available nest structures) and environmental conditions 

(e.g., appropriate climate, suitable forest structure, and 

infrequent presence of barred owls) suitable for occupancy, 

reproduction, and survival of the subspecies. As such, 

habitat is more analogous to a species’ realized niche 

rather than the fundamental niche because habitat is more 

constrained than the availability of a vegetation type and a 

subset of environmental conditions. All published models 

of spotted owl habitat fall short of this definition because 

the distribution of spotted owls in relation to abundant 

prey is not known, and the distribution of an important 

competitor—barred owls—is not fully known. Throughout 

this chapter we distinguish between spotted owl habitat and 

components of that habitat (e.g., forest cover types used 

for nesting and roosting) regardless of the terms used in 

published literature. 

Differing concepts regarding habitat definitions have 

long caused confusion and uncertainty in the interpretation 

of scientific literature (Bamford and Calver 2014, Hall et 

al. 1997, Morrison et al. 2006). The differences in how 

spotted owl habitat is defined and modeled has also caused 

confusion. The NWFP monitoring program estimates trends 

in forest types used by spotted owls (Davis and Lint 2005). 

The Fish and Wildlife Service (USFWS 2012a) modeled 

suitable forest and considered the amount and spatial 

arrangement of forests associated with specific life history 

requirements (e.g., forest types used for foraging in relation 

to forests used for nesting and roosting), as well as abiotic 

factors (e.g., slope, climate). The resulting models were 

used for delineation and designation of what was considered 

critical habitat (USFWS 2011b). The models of potential 

spotted owl habitat developed by the NWFP monitoring 
251

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

program and the Fish and Wildlife Service have important 

differences that result in different amounts of what is 

considered suitable forest for spotted owls. Estimates of the 

amount of suitable forest for spotted owls are highly scrutinized because of the conflict caused by the importance of 

that forest type for the reproduction and survival of spotted 

owls and because merchantable large timber is important 

economically for many of the rural areas where old forest 

occurs. The different estimates of suitable forest cover for 

spotted owls resulted in litigation filed in relationship to 

critical habitat designation. Carpenters Industrial Council et 

al. vs. Ashe and Salazar (District of Columbia District Court 

case number 1:2012cv00111 filed January 24, 2012) claimed 

that the USFWS (2012a) estimate of approximately 18 

million ac (7.3 million ha) of suitable forest conditions (they 

used the term habitat) for spotted owls was an overestimate 

of 5.9 million ac (2.4 million ha) because previous documents produced by the agency had used estimates of approximately 12.1 million ac (4.9 million ha) as found in (Davis et 

al. 2011). The 2.4 million ha difference can be explained by 

an examination of how habitat was defined and modeled in 

the different efforts. For example, estimates from Davis et 

al. (2011) were based on a stand-level designation of forest 

cover suitable for nesting and roosting (fig. 4-3A), whereas 

USFWS (2011b) and USFWS (2012a) delineated critical 

habitat based on a model that included suitable forest stands 

(Davis et al. 2011) and other landscape components essential 

for spotted owls at the core-area scale (200 ha) (fig. 4-3B). 

The NWFP defined suitable forest for spotted owls as 

an area with the species of trees, structure associated most 

commonly with late-successional forest, sufficient area, and 

adequate food source to meet some or all of the subspecies’ 

life needs, including nesting, roosting, and foraging (USDA 

and USDI 1994). This definition relied heavily on the work 

in the Interagency Scientific Committee report (Thomas 

et al. 1990), which acknowledged the difficulty in defining 

habitat and chose to characterize the concept based on 

relative value or suitability of forest stands for spotted owls. 

Forest cover can be viewed as supporting different spotted 

owl life functions (e.g., nesting, roosting, foraging) and a 

suitability gradient in terms of its influence on individual 

fitness (Thomas et al. 1990). Partitioning of forest cover 

0 500 1000 2000 Meters

(A)

Unsuitable Suitable

(B)

Figure 4-3—Examples of the suitable forest cover at (A) the stand scale developed by the Northwest Forest Plan monitoring program 

(Davis et al. 2011), and (B) the 200-ha (~250-foot radius) core-area scale used for modeling and delineating critical habitat (USFWS 2012a). 
252

GENERAL TECHNICAL REPORT PNW-GTR-966

into discrete categories based on established measures of 

suitability for particular life functions facilitates a common 

frame of communication and standardization. A monitoring 

framework to measure relative suitability of forest cover 

types used by spotted owls was developed as part of a 

rangewide monitoring program for the subspecies (Davis et 

al. 2011, 2016; Lint 2005). Monitoring divided a continuous 

gradient of cover-type suitability into four discrete classes 

(table 4-1), based on use-versus-availability analyses using 

documented territorial pair locations. The unsuitable class 

was used for nesting and roosting by spotted owls less than 

expected by chance based on availability, the marginal 

class was used in proportion to its availability, the suitable 

class was used more often than expected by chance, and the 

highly suitable class was used much higher than one would 

expect from chance based on its availability. For monitoring 

purposes that dates to the life of the NWFP, the suitable and 

highly suitable classes were combined into a single class 

to identify forests that were most strongly associated with 

nesting and roosting locations. Thomas et al. (1990) characterized highly suitable forest cover as forests that include 

a multilayered, multispecies canopy dominated by large 

(>30 inch diameter at breast height [d.b.h.]) conifer trees; 

an understory of shade-tolerant conifers or hardwoods; 

moderate to high (60 to 80 percent) canopy cover (they used 

the term closure, but by definition they had described cover) 

(Jennings et al. 1999); substantial decadence in the form of 

large, live coniferous trees with deformities (e.g., cavities, 

broken tops, and dwarf mistletoe infections); numerous 

large snags; large accumulations of logs; and other woody 

debris. The unsuitable or marginal classes do not imply 

unimportance to spotted owls because the classification 

was restricted to describe only suitability for nesting and 

roosting activities by spotted owls. The marginal class is 

likely important for supporting dispersal, foraging, and 

nonbreeding (i.e., floater) individuals that can replace adult 

mortality and dispersal at nesting territories. Likewise, 

unsuitable and marginal classes may be important forest 

types for many prey species used by spotted owls. Forests 

that are suitable for nesting and roosting have similar characteristics throughout the range of spotted owls, but the path 

of development to those conditions typically differ based on 

the fire regime within the area (chapter 3; table 4-2, fig. 4-4).

Thomas et al. (1990) defined forest suitable for dispersal as having ≥11 inch (28 cm) d.b.h. trees and ≥40 

percent canopy cover occurring on ≥50 percent of a 36 mi2

township; this definition became known as the 50/11/40 

rule. Analyses of movement data of spotted owls suggest 

that most (90 percent) dispersal occurred through landscapes meeting these criteria and are generally considered 

Table 4-1—General descriptions of forest cover type classes used to estimate the amount of suitable forest 

available for nesting and roosting by spotted owls.

Cover type class General description

Unsuitable Younger forests or older forests with higher basal area of pine or high-elevation tree species or more 

open canopies. Usually smaller than average tree diameters, and lacking the presence of residual large 

trees and multiple canopy layers. 

Marginal Usually mid-seral forests, but can also be older forests lacking large-diameter trees, having simpler 

stand structure, or primarily composed of pine or high-elevation tree species.

Suitable Forest stands older than 125 years of age, except in the California redwoods, where younger stands are 

used. Average tree diameters are usually above 20 inches (50 cm) d.b.h., with the presence of at least 

a few large trees exceeding 30 inches (75 cm) d.b.h. Canopy cover is usually greater than 60 percent, 

and the stand has multiple canopy layers.

Highly suitable Typically forests 150 and 200 years of age or older. Average tree diameters often in excess of 30 inches 

(75 cm) d.b.h. except in drier portions of the range, where tree ages and sizes are typically smaller 

(e.g., 120 years and 24 inches). Canopy cover is usually in excess of 70 percent, and the stand has 

multiple canopy layers with high diversity of tree sizes.

d.b.h. = diameter at breast height.

Source: Davis et al. 2016.
253

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

capable of supporting dispersal (Davis et al. 2011, 2016; 

Forsman et al. 2002; Lint 2005). However, the Thomas et 

al. (1990) 50/11/40 hypothesis was not based on juvenile 

resource selection data and remains largely untested. Only 

two studies (Miller et al. 1997, Sovern et al. 2015) have 

empirically studied forest-type selection during juvenile 

dispersal. Both studies found that juveniles strongly select 

for old forest with closed canopy (>70 percent canopy cover) 

and large-diameter trees (>20 inch d.b.h.), which are similar 

forest conditions selected by adult spotted owls for nesting 

and roosting (Miller et al. 1997, Sovern et al. 2015). Given 

the importance of forest cover classified as suitable for 

nesting and roosting to juvenile dispersal, the canopy cover 

recommendations of Thomas et al. (1990) are unlikely to be 

sufficient to facilitate juvenile movements on the landscape. 

Sovern et al. (2015) suggested that stands managed for 

dispersing spotted owls should be at least 80 percent canopy 

cover and have large average tree diameter. 

Both the nesting/roosting and dispersal maps of 

suitable cover types produced by the NWFP monitoring 

program were designed to match the conceptual descriptions of forest vegetation components defined by Thomas et 

al. (1990) and used at the time of the NWFP development. 

Mapping of forests used by spotted owls is continuing to 

Table 4-2—General descriptions of how forest cover types suitable for nesting and roosting by spotted owls 

typically develop within four general fire regimes within the Northwest Forest Plan (NWFP) area

Fire regime Typical development of suitable nesting/roosting forest

Infrequent—high severity

(Coast Range, fig. 4-4)

Large contiguous patches that form following infrequent, yet very large, highseverity wildfires. Once established, these large patches persist for long periods 

until the next large high-severity wildfire. Immediately following a large 

wildfire, large areas of the landscape are unsuitable for nesting and roosting for 

decades until closed canopies redevelop in areas that had remnant tree structures 

that could serve as nest trees. During this period, fine-scale gaps created by rootrot pockets, windstorms, landslides, and other small-scale processes produce 

complex stand structure. Complex structure sometimes does not develop 

over large areas for several decades following a wildfire. Produces the largest 

diameter and tallest nest trees; nests are usually in cavities or broken tops. 

Moderately frequent—mixed severity

(West Cascades, fig. 4-4)

Abundant to moderately abundant on the landscape, but very well connected 

across the landscape owing to the lack of extremely large high-severity wildfire 

patches. High-severity wildfire created smaller patches of complex early-seral 

forest cover type within an otherwise older forest matrix. Through time, these 

wildfire-created patches produced complex forest structure at the stand scale and 

a diverse mosaic of seral stages at the landscape scale. 

Frequent—mixed severity

(Klamath Mountains, fig. 4-4)

Moderately abundant on the landscape but more confined to topographic positions 

that functioned as wildfire refugia (e.g., lower slopes, north aspects, etc.). These 

areas allowed for the development and persistence of large trees required for 

nesting structures. In the Klamath Mountains and California Coast Range 

physiographic provinces, evergreen hardwoods (e.g., tanoak) are an important 

component that increase the suitability of use in these stands. In addition to forest 

stand structure and species composition, climate, and topography are important 

predictors of use by spotted owls. 

Very frequent—low severity

(East Cascades, fig. 4-4)

Not naturally abundant within the NWFP area; primarily restricted to the east side 

of the Cascade Mountains and eastern parts of northern California. Occurred 

historically in areas where the topography or soil conditions created a productive 

environment suitable for the development of large Douglas-fir and grand fir. 

Once established, these closed-canopy, structurally complex forest cover 

conditions can be relatively resistant to most fires, but burn with high severity 

under extreme weather conditions (chapter 3).
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GENERAL TECHNICAL REPORT PNW-GTR-966

evolve (Ackers et al. 2015). For example, recent maps of 

suitable forest types (e.g., Glenn et al. 2017; USFWS 2011b, 

2012a) differed from the original monitoring maps in that 

they factored in the spatial arrangement of discrete forest 

cover types (e.g., nesting, roosting, foraging) as well as 

abiotic factors (e.g., slope, topographic position, etc.) to 

produce maps describing a more comprehensive view of 

suitable forest (i.e., potential habitat). However, even the 

most recent efforts are not complete models of spotted owl 

habitat because they lack the impact of prey and barred owls 

on restricting distribution by limiting access to otherwise 

suitable forest for spotted owls. An important need is 

a better understanding and mapping of the differences 

between the potential and realized habitat for spotted owls. 

This is discussed in the “Research Needs, Uncertainties, 

Information Gaps, and Limitations” section below.

Patterns of change— 

Federal vs. nonfederal lands—Davis et al. (2016) estimated 

that there were about 12.6 million ac (5.1 million ha) of suitable nesting and roosting cover type distributed across the 

spotted owl’s geographic range at the time of NWFP development (1993), the majority (73 percent) of which occurred 

on federal lands. By 2012, suitable nesting/roosting forest 

BEND

EUGENE

MEDFORD

COOS BAY

ROSEBURG

Coast Range

Klamath Mountain

West Cascades

East Cascades

±

0 10 20 5 Miles

Figure 4-4—Differing historical patterns of old-growth Douglas-fir and mixed-conifer forest (green shaded areas) in west-central Oregon (Andrews and Cowlin 1940) within four areas with different fire regimes (Coast Range, infrequent—high severity; West Cascades, 

moderately frequent—mixed severity; Klamath Mountains, frequent—mixed severity; East Cascades, very frequent—low severity). 
255

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

cover decreased to 12.1 million ac (4.9 million ha) (74 percent occurring on federal lands), resulting in an overall net 

change of -3.4 percent. Net decreases were -1.5 percent on 

federal lands (primarily caused by wildfire) and -8.3 percent 

on nonfederal lands (primarily caused by timber harvest). 

During those two decades, forest cover suitable for dispersal 

decreased from 26.2 to 25.7 million ac (10.6 to 10.4 million 

ha) (-2.3 percent net change) on all lands. On federal lands, 

forest cover suitable for dispersal increased by 2.2 percent, 

and it decreased by 8.6 percent on nonfederal lands (Davis 

et al. 2016). Gains occurred because of forest succession, 

whereas losses were primarily a result of wildfire, disease, 

and timber harvest (Kennedy et al. 2012).

Timber harvest accounted for the majority (63 percent) 

of the losses across all lands. The vast majority of losses on 

nonfederal lands was caused by timber harvest (94 percent), 

whereas timber harvests accounted for 18 percent of total 

losses on federal lands (Davis et al. 2016). In Washington 

alone from 1996 to 2004, most (85 percent) of the timber 

harvest that resulted in lost forest cover suitable for nesting 

and roosting of spotted owls occurred on private lands (Kennedy et al. 2012, Pierce et al. 2005). Following timber harvest, wildfire was the next largest cause of loss (31 percent 

of total losses), which was 73 percent of the losses on federal 

land and only 3 percent of the losses on nonfederal land. 

Moist vs. dry forests—Primary causes of loss differed by 

ecoregion and forest type. The loss of nesting and roosting 

forest cover from wildfire occurred primarily in drier, fireprone portions of the spotted owl’s geographic range (i.e., 

northern California, southern Oregon, and eastern Cascade 

Range). Losses owing to insects and disease (and other natural disturbances) was the next most significant disturbance 

and mainly occurred in the eastern Cascades of Washington 

and Oregon (Davis et al. 2016, Kennedy et al. 2012). 

Recruitment of forest cover suitable for nesting and 

roosting by spotted owls was estimated at 257,591 ac (104 

288 ha) from 1993 to 2012 (Davis et al. 2016). Most of the 

gain occurred on nonfederal lands within the redwood 

(Sequoia sempervirens) zone of coastal California (fig. 4-5). 

On federal lands, the largest net gain (40,385 ac [16 350 ha]) 

occurred in the eastern Cascades of Oregon, where fire suppression allowed forest succession of Douglas-fir (Pseudotsuga menziesii) and grand fir (Abies grandis) to develop in 

areas that historically had frequent low-severity fires and 

were formerly dominated by open ponderosa pine-dominated forests (Pinus ponderosa) (Davis et al. 2016). 

Effects of forest change—

Loss of suitable forest cover for nesting and roosting, 

especially on nonfederal lands, has been an important 

contributor to declining populations of spotted owls 

(Dugger et al. 2016). Those spotted owls that had territories 

with more forest cover associated with nesting and roosting 

conditions typically had better survival, fecundity, occupancy dynamics, recruitment, and rate of population change 

(Dugger 2016; Dugger et al. 2005, 2011; Forsman et al. 2011; 

Seamans and Gutiérrez 2007). For example, Dugger et al. 

(2005) found that owl territories with the greatest fitness 

potential were characterized by >50 percent old-forest 

habitat within a 412-ac (167-ha) circle centered on used nest 

locations. Relationships among population parameters of 

spotted owls and older forests vary over different spatial 

scales (e.g., individual territory vs. study area), and can 

be independent of, or interact with, the presence of barred 

owls. Concentrated areas of older forest suitable for nesting 

and roosting, or increased amounts of heterogeneity (i.e., 

mixture of conditions used for foraging), have positive 

effects on the vital rates of spotted owls (Dugger et al. 2016, 

Forsman et al. 2011, Franklin et al. 2000, Olson et al. 2004). 

In some landscapes, fragmentation of older forest can 

have negative or positive effects on spotted owl occupancy 

depending on the scale of fragmentation and edge characteristics. Schilling et al. (2013) found that spotted owls had 

decreased survival and increased home-range size with 

increased forest fragmentation in southwestern Oregon. 

In Washington, territory-level extinction rates decreased 

with increased amount of late-seral edge, and colonization 

decreased with more late-seral patches within a territory 

(Sovern et al. 2014). It is also important to consider spatial 

scale, and level of contrast between edge, when assessing 

the influence of forest edges on foraging and space use by 

spotted owls. Comfort et al. (2016) found that spotted owls 

radio-marked in southern Oregon were negatively associated with hard edges (high contrast in forest structure and 

height) at a fine scale (telemetry location), but showed a lack 
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GENERAL TECHNICAL REPORT PNW-GTR-966

of negative response to hard edges at broader scales (territory or home-range scales). At least at the territory scale, 

heterogeneity can contribute to accessibility to different 

forest types. Regardless of spatial scale, spotted owls were 

positively associated with softer, more diffuse edge types 

created by disturbances such as low- and mixed-severity fire 

(Comfort et al. 2016). Collectively, these and other studies 

suggest that spotted owls select for abundant, structurally 

diverse closed-canopy forest with diffuse late-seral forest 

edge at the territory scale, and relatively lower fragmentation in nesting areas (Franklin et al. 2000, Olson et al. 2004, 

Sovern et al. 2014).

Forest protection effectiveness—

The NWFP included a network of large late-successional 

reserves (LSRs) that were designed to conserve forest for 

species dependent on older forests (FEMAT 1993). The LSR 

network was intended to meet the resource needs of many 

species, but a substantial focus was placed on creating and 

maintaining forest cover features from a draft recovery 

plan for the spotted owl (USFWS 1992). LSRs contained 

enough suitable forest cover to support multiple pairs of 

spotted owls and were distributed to facilitate movement 

of spotted owls across their geographic range. Although 

many of the LSRs contained large areas of older forest, a 

Percentage of area

0 to 10

10 to 20

20 to 30

30 to 40

40 to 50

50 to 60

60 to 70

70 to 80

80 to 90

90 to 100

7

3 2

5

8

10

1

6

12

11

9 7

3 2

5

8

10

1

6

12

11

9 7

3 2

5

8

10

1

6

12

11

9

1993 2012 Net change

 1. Olympic Peninsula (Washington)

 2. Western Lowlands (Washington)

 3. Western Cascades (Washington)

 4. Eastern Cascades (Washington)

 5. Coast Range (Oregon)

 6. Willamette Valley (Oregon)

 7. Western Cascades (Oregon)

 8. Klamath (Oregon)

 9. Eastern Cascades (Oregon)

10. Coast Range (California)

11. Klamath (California)

12. Cascades (California)

Physiographic provinces

Percentage of change

>5% loss

2.5 to 5% loss

1 to 2.5% loss

<1% change

1 to 2.5% gain

2.5 to 5% gain

>5% gain

4 4 4

Figure 4-5—Loss and recruitment of forest types suitable for nesting and roosting by spotted owls in 1993 versus 2012, based on change 

in percentage of cover estimated within 18.6 mi (30 km) (center to center) hexagons across the Northwest Forest Plan area. 
257

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

significant portion of them were delineated in fragmented 

landscapes that contained stands of younger forest. Dispersal between LSRs is important for spotted owl conservation, 

and the NWFP was expected to facilitate that dispersal by 

designated riparian reserves, retention of green trees in 

timber harvest units in the matrix, protection of 100 ac (40 

ha) areas at known owl sites (managed as LSRs within the 

matrix), and other administratively withdrawn areas (USDA 

and USDI 1994). However, these assumptions are largely 

untested, so it remains unknown if the NWFP is sufficient 

to facilitate adequate dispersal, which may be a limiting 

factor of spotted owl populations. 

In addition to broad-scale LSRs, forest protections 

for spotted owls include circles of varying radii centered 

on used nest locations, within which various amounts of 

suitable nesting, roosting, and foraging forest cover types 

are protected. For example, the Fish and Wildlife Service 

developed guidelines for consultation under section 7 of 

the Endangered Species Act that included a 2.9-km-radius 

circle (6,424 ac [2600 ha]) around spotted owl nest locations for evaluating ‘‘incidental take’’ for projects affecting suitable habitats (USDA and USDI 1994). The rationale 

for this circle size was developed based on preliminary 

analysis of the median home-range size of radio-marked 

spotted owls. States also developed rules for state and 

private forestry practices to protect spotted owl nest sites. 

For example, the 2006 Washington State Forest Practices 

Board Rules called for protection of 40 percent cover of 

suitable nesting and roosting forest within a 6,422 ac (2600 

ha) circle around nest sites (WAC 222-10-041). Forsman 

et al. (2015) suggested that level of protection would not 

be sufficient because spotted owl home ranges contained 

more suitable forest cover than would be protected under 

the Washington forest practices rules. Furthermore, new 

methods for delineating owl territories (e.g., Thiessen 

polygons) used by Dugger et al. (2016) provide better 

representations of the territory.

At the time LSRs were delineated, it was estimated 

that they contained on average 43 percent older forest 

(USDA and USDI 1994). The expectation was that all 

LSRs would eventually fill in and achieve the 60-percent-or-greater area threshold needed to support multiple 

breeding pairs and collectively would facilitate spotted owl 

population recovery. The success of meeting that threshold 

depends on the frequency, severity, and spatial extent of 

disturbance (e.g., wildfire, timber harvests), as well as the 

rate of forest succession, and interactions among these 

processes on forest recruitment (chapter 3). As of the most 

recent monitoring report (Davis et al. 2016), the rangewide 

estimate for suitable nesting/roosting forest cover in LSRs 

was an average of 42.4 percent in 1993. As of 2012, this 

average decreased to 42.0 percent. Larger LSRs (≥10,000 

ac) averaged 45.0 percent, decreasing to 44.5 percent 

by 2012. These losses were due mainly to wildfire and 

exceeded the regional-scale expected rate of loss (2.5 

percent per decade) (FEMAT 1993). Most of the losses of 

nesting and roosting forest cover have been in the more 

fire-prone portions of the spotted owl’s range (Davis et al. 

2011, 2016). For example, within LSRs and other reserves 

(e.g., administratively withdrawn, wilderness areas, etc.) 

in the Klamath Mountains physiographic province, losses 

were as high as 18.9 percent between 1993 and 2012 (fig. 

4-5), and largely the result of the 2002 Biscuit Fire, which 

burned 494,000 ac (>200 000 ha). 

Forest cover trends on federal lands during the next 

two to three decades are expected to benefit spotted owls 

because significant recruitment of suitable nesting/roosting 

forest cover is expected to offset many pre-NWFP losses 

(chapter 3) (Davis et al. 2016). However, this expectation is 

based on current rates of harvest and wildfire occurrence on 

federal lands, which may change depending on future forest 

plan revisions and the predicted increased spatial extent, 

frequency, and severity of wildfires due to climate change 

(chapter 2) (Jones et al. 2016, Westerling et al. 2006). In 

addition, competitive pressure from established barred owls 

(see below) has raised uncertainties about whether recruitment of suitable forest cover will be enough to conserve 

spotted owls over the long term. If spotted owls are to 

persist in LSRs under competitive pressure from barred 

owls, it will likely be only in localized areas that support 

few barred owls. However, it remains doubtful if there are 

any areas where spotted owls hold a competitive advantage 

over barred owls (Pearson and Livezey 2007, Singleton 

2013, Wiens et al. 2014). 
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GENERAL TECHNICAL REPORT PNW-GTR-966

The potential effects of climate change add to the 

uncertainty of how competitive dynamics with barred 

owls and availability of suitable habitat will affect spotted 

owls in the future. Carroll et al. (2010) used a climatic 

niche modeling approach to evaluate the regional system 

of LSRs for resiliency to climate change for providing 

necessary resources of species associated with old forest. 

They developed distribution models integrating climate 

data with vegetation variables for a large suite of species, 

including the spotted owl. The LSRs functioned better 

than expected by chance for capturing all of the species, 

but community composition and interspecific interactions 

were also important to consider in evaluating effectiveness of the reserves. A network of fixed reserves with 

a high level of climatic and topographic heterogeneity 

(i.e., designed for resilience) has an increased likelihood 

of retaining the biological diversity of old-forest ecosystems under climate change. Under this scenario, even 

those species with limited dispersal capability are able 

to colonize future habitat. Carroll et al. (2010) projected 

a northward and higher elevation movement of suitable 

forest for spotted owls; therefore, the current fixed system 

of LSRs may not have enough climatic and topographic 

heterogeneity to be adequate for spotted owls into the 

future. Other reserves designated before the NWFP, such 

as parks and wilderness areas, may become increasingly 

important for the subspecies’ persistence. LSRs successfully protected areas with greater biological importance 

for spotted owls when the NWFP was developed, but 

in the face of climate change, it may be necessary to 

have another evaluation and planning phase that results 

in a reserve system designed for more robust resilience 

(Carroll et al. 2010) (see chapter 3 for more discussion of 

alternative reserve designs), especially in the dry forest 

zone where management for ecosystem and spotted owls 

may not be compatible at stand and small landscape scales 

(chapter 12). Even with relatively little modification in 

response to climate change, suitable forest conditions on 

the east side and southern portions of the range are at risk 

of losses. Dense, multilayered forests in the dry forest 

zone are vulnerable to a host of mortality forces, especially wildfire (see chapters 3 and 12).

Barred Owls 

Barred owl range expansion and population trends—

Competition with established populations of barred owls 

has emerged as a much more prominent and complex threat 

to the long-term persistence of the spotted owl than was 

anticipated during the development of the NWFP. Once 

confined to forests of eastern North America, the barred 

owl is a medium-size, ecologically similar species whose 

newly extended geographic range now completely overlaps 

that of the northern spotted owl (Gutiérrez et al. 2007, 

Livezey 2009). Newly colonizing barred owls in the Pacific 

Northwest have been classified as native invaders—species 

that, under the influence of events such as climate change or 

human modifications to the landscape, have become invasive 

by expanding their populations into new areas (Carey et 

al. 2012, Valéry et al. 2009, Wiens et al. 2014). The range 

expansion of barred owls in western North America is well 

documented (Dark et al. 1998, Dunbar et al. 1991, Kelly et al. 

2003, Livezey 2009, Taylor and Forsman 1976). Initial colonization of different regions by barred owls was variable, but 

barred owls now appear to co-occupy and outnumber spotted 

owls throughout the entire range of the threatened subspecies 

(Dugger et al. 2016, Pearson and Livezey 2003, Singleton et 

al. 2010, Wiens et al. 2011, Yackulic et al. 2012). Barred owls 

have also invaded the range of the California spotted owl in 

the Sierra Nevada (Seamans et al. 2004). The cause of this 

range expansion is unknown, but landscape changes facilitated by European settlement or historical changes in climate 

are factors that may have enabled barred owls to expand 

their range from eastern to western North America (Livezey 

2009, Monahan and Hijmans 2007). 

With few exceptions, barred owls have not been 

systematically surveyed in the Pacific Northwest, and the 

majority of information on their distribution and population 

trends is limited to incidental observations during surveys 

of spotted owls (Dugger et al. 1991, 2016; Gutiérrez et al. 

2007; Wiens et al. 2011). Despite this shortcoming, incidental field data show a rapid increase in barred owls as they 

expanded their populations westward and southward into 

the range of the spotted owl (fig. 4-6) (Dugger et al. 2016). 

Studies focused on barred owls found much higher densities than estimates based on incidental field observations 
259

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

1.0

0.8

0.6

0.4

0.2

0

Barred owl covariate

1989

199119931995

19971999

2001

2003

200520072009

2011

2013

Washington

Cle Elum

Rainier

Olympia

1.0

0.8

0.6

0.4

0.2

0

Barred owl covariate

1988199019921994199619982000200220042006200820102012

Oregon

Coast Range

H.J. Andrews

Tyee

Klamath

South Cascades

1.0

0.8

0.6

0.4

0.2

0

Barred owl covariate

198519881991199419972000200320062009

2012

California

Northwest California

Hoopa

Green Diamond

Green Diamond—treatment

Green Diamond—control

A

B

C

Year

Figure 4-6—Annual increase in the 

proportion of spotted owl territories 

with detections of barred owls at (A) 

three sites in Washington, (B) five 

sites in Oregon, and (C) three sites in 

California from 1985 to 2013 (Dugger et 

al. 2016). Estimates for Green Diamond 

are presented separately for control and 

treatment areas and after (2009 to 2013) 

barred owls were removed from the 

treatment area from 2009 to 2013. 
260

GENERAL TECHNICAL REPORT PNW-GTR-966

(Hamer et al. 2007; Singleton et al. 2010; Wiens et al. 2011, 

2014; Yackulic et al. 2012, 2014). For example, Wiens et 

al. (2011) conducted surveys of barred owls during 2009 in 

the Oregon Coast Range and identified approximately 11 

territorial pairs of barred owls per 100 km2 (39 mi2; 3 to 8 

times higher density than spotted owls) with 89 percent of 

the landscape occupied, which peaked on publicly owned 

lands with greater amounts of mature and old coniferous 

forest. More recent (2015–2016) surveys of barred owls 

indicate an even greater probability of landscape occupancy 

in the Oregon Coast Range (~0.94) (Wiens et al. 2017). 

The degree to which the colonizing population of barred 

owls has reached carrying capacity within the geographic 

range of the spotted owl is currently unknown, but studies 

are underway that can help address this uncertainty (e.g., 

Wiens et al. 2017). Barred owl populations may continue 

to increase depending on the capacity of available habitat 

and food resources, which varies regionally with forest 

composition and latitudinal changes in prey communities 

and climate. 

Barred owl effects on spotted owls—

Compared to spotted owls, barred owls are slightly larger 

(Gutiérrez et al. 2007), have more diverse diets (Hamer 

et al. 2001, Wiens et al. 2014), and use a broader range 

of forest conditions for nesting (Herter and Hicks 2000, 

Livezey 2007, Pearson and Livezey 2003) and foraging 

(Hamer et al. 2007, Singleton 2015, Singleton et al. 2010, 

Weisel 2015, Wiens et al. 2014). Barred owls also have 

higher annual survival (fig. 4-7), higher reproductive 

output, and, in most areas, use much smaller home ranges 

than spotted owls (Hamer et al. 2007, Singleton et al. 2010, 

Wiens et al. 2014). The exception is in northern California, 

1.0

0.9

0.8

0.7

0.6

0.5

0.4

0.3

Predicted survival probability

Proportion of home range with old forest

0 0.10 0.20 0.30 0.40 0.50

Spotted owl

Barred owl

Figure 4-7—Survival (with 95-percent confidence limits) of individual adult barred owls and spotted owls increased with increasing 

amount of old forest (≥120 years) conifer forest within their home ranges in the Oregon Coast Range (Wiens et al. 2014).
261

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

where the two species used relatively small home ranges of 

similar size (Weisel 2015). Barred owls also defend their 

territories more aggressively than spotted owls (Van Lanen 

et al. 2011), which can result in increased mortality of 

spotted owls from agonistic interactions and direct killing 

of spotted owls by barred owls (Leskiwand Gutiérrez 1998, 

Wiens et al. 2014).

The dramatic increase in populations of barred owls 

since implementation of the NWFP has significant implications for management of forests inhabited by spotted 

owls. Several lines of evidence indicate that increases in 

the abundance of barred owls has had a strong and negative 

impact on spotted owls. Increasing abundance of barred 

owls has been documented to have the following effects on 

spotted owl populations:

1. Occupancy of historical spotted owl territories 

is lower (fig. 4-8) (Bailey et al. 2009, Dugger 

2016, Dugger et al. 2011, Kelly et al. 2003, Kroll 

et al. 2010, Olson et al. 2005, Sovern et al. 2014; 

Yackulic et al. 2014). 

2. Apparent survival is lower (Anthony et al. 2006, 

Diller et al. 2016, Dugger et al. 2016, Forsman et al. 

2011, Glenn et al. 2011a). 

3. Reproduction is lower (Dugger et al. 2016, 

Forsman et al. 2011, Olson et al. 2004). 

4. Population size declines more rapidly (Anthony et 

al. 2006, Dugger et al. 2016, Forsman et al. 2011).

5. Hybridization between the species is increased 

(Barrowclough et al. 2005, Dark et al. 1998, 

Gutiérrez et al. 2007, Haig et al. 2004, Hamer et al. 

1994, Kelly and Forsman 2004).

6. Detection rates during surveys are lower (Bailey 

et al. 2009, Crozier et al. 2006, Dugger et al. 2011, 

Dugger et al. 2016, Kroll et al. 2010, Olson et al. 

2005, Sovern et al. 2014, Yackulic et al. 2014). 

Moreover, studies of competitive interactions and 

resource partitioning showed that barred owls can directly 

alter the movements, resource use, and reproduction of 

spotted owls (Wiens et al. 2014). Barred owls also display 

demographic superiority over spotted owls; annual rate of 

0.70

0.60

0.50

0.40

0.30

0.20

0.10

0

Local extinction rate (ε)

Without barred owl

With barred owl

Study area

Cle Elum

Rainier

Olympic

Coast Ranges

H.J. Andrews

Tyee

Klamath

South Cascades, Oregon

Northwest California

Hoopa

Green Diamond

Figure 4-8—Mean annual local 

extinction rates (with 95 percent 

confidence limits) for northern 

spotted owls on 11 study areas 

in Washington, Oregon, and 

California relative to with (gray 

triangles) and without (black 

circles) the presence of barred 

owls (Dugger et al. 2016).
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GENERAL TECHNICAL REPORT PNW-GTR-966

survival was greater for barred owls (0.92 ± 0.04) than for 

spotted owls (0.81 ± 0.05), and mean reproductive output of 

barred owl pairs was 4.4 times greater than that observed 

for spotted owls over 3 years in western Oregon (Wiens et 

al. 2014). More recently, studies in California have demonstrated a positive association between removal of barred 

owls and population trends of spotted owls (fig. 4-2) (Diller 

et al. 2016, Dugger et al. 2016). Collectively, these studies 

provide strong evidence that interspecific competition with 

an increasing number of barred owls, combined with continued loss of potentially suitable forest cover, is contributing 

to population declines of spotted owls despite widespread 

conservation of old forest under the NWFP. 

Barred owl densities are now thought to be high enough 

across the range of the spotted owl that, despite the continued management and conservation of suitable forest cover 

on federal lands, the long-term persistence of spotted owls 

is in question without additional management intervention 

(Buchanan et al. 2007, Diller et al. 2016, Dugger et al. 2016, 

USFWS 2013). In a few cases, populations of spotted owls 

have responded positively to the removal of barred owls 

during pilot removal experiments; supporting the hypothesis that along with forest conservation and management, 

removal of barred owls might slow or reverse local declines 

in spotted owl populations in some areas (Diller et al. 

2016, Dugger et al. 2016). However, the effectiveness and 

moreover the feasibility of large-scale barred owl removal 

for conservation of spotted owls remain to be demonstrated, 

and barred owl removal activities would likely need to be 

continued for the foreseeable future to maintain low barred 

owl densities in control areas.

Barred owl habitat and prey—

Barred owls occupy a broader range of forest types and 

consume a wider variety of prey than northern spotted 

owls (Livezey 2007), and use a variety of different forest 

types in the Pacific Northwest, including fragmented 

mixed-deciduous forest in rural and urban landscapes 

(Rullman and Marzluff 2014). Hamer et al. (2007) reported 

that, in the northern Cascade Range of Washington, barred 

owls tended to use old forest more than expected, but used 

most cover types in proportion to availability. Compared to 

spotted owls, barred owls occupied areas at lower elevations 

(Hamer et al. 2007). In the eastern Cascades of Washington, 

Singleton et al. (2010) reported that barred owls typically 

established their home ranges in areas that had canopy 

cover more than 72 percent, medium to large trees (tree 

crown diameter >21 ft [>6.5 m]), low topographic position 

(<25 percent), and gentle slopes (<11 degrees). Within those 

home ranges, barred owls used structurally diverse mixed 

grand fir forest more intensively than open ponderosa 

pine or Douglas-fir (Singleton 2015). In the Oregon Coast 

Range, foraging barred owls most often used patches of old 

(>120 years) conifer forest in addition to riparian-hardwood 

forests in relatively flat areas (Wiens et al. 2014). In the 

redwood region of coastal California, barred owls most 

often used sites with greater understory vegetation height 

and more hardwood trees, perhaps in response to greater 

densities of woodrats (Neotoma spp.) in these conditions 

(Weisel 2015). Collectively, these studies showed that barred 

owls, in areas where they were sympatric with spotted 

owls, were most commonly associated with relatively gentle 

slopes in structurally diverse, mature and old-conifer forests 

or lowland riparian areas containing large hardwood trees. 

Use of older forest in combination with moist, valley-bottom 

forest was also consistent with forest associations described 

for barred owl nesting areas (Buchanan et al. 2004, Herter 

and Hicks 2000, Pearson and Livezey 2003). Barred owls 

use the full range of forest types used by spotted owls, and a 

broader range of forest cover types outside of areas historically occupied by spotted owls. However, systematic studies 

have yet to quantify the full range of forest conditions that 

support barred owls in the Pacific Northwest. There are 

currently no known forest management actions that would 

benefit spotted owls more than barred owls. 

Dietary studies are lacking for barred owls in California, but their diets in Washington and Oregon included a 

broad variety of small- to medium-size mammals, birds, 

frogs, salamanders, lizards, snakes, crayfish, snails, fish, and 

insects (Graham 2012, Hamer et al. 2001, Wiens et al. 2014). 

Mammalian prey of barred owls primarily included northern 

flying squirrels (Glaucomys sabrinus), woodrats, brush 

rabbits (Sylvilagus bachmani), snowshoe hares (Lepus americanus), moles (Scapanus spp.), Douglas squirrels (Tamiasciurus douglasii), red tree voles (Arborimus longicaudus), 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

red-backed voles (Myodes californicus), shrews (Sorex spp.), 

and deer mice (Peromyscus maniculatus) (Hamer et al. 2001, 

Wiens et al. 2014). Although there is substantial geographic 

variation in diets of barred owls corresponding with differences in prey distributions, northern flying squirrels appear 

to be a primary contributor to diets in Oregon and Washington (Graham 2012, Hamer et al. 2001, Wiens et al. 2014). 

Although there is some evidence that barred owls were 

more strongly associated with riparian areas than spotted 

owls, studies clearly indicate a high degree of ecological 

overlap between the two species, especially in their use of 

old-growth forests and associated prey species (Hamer et al. 

2001, 2007; Singleton et al. 2010; Weisel 2015; Wiens et al. 

2014). In the eastern Cascades of Washington, spotted owls 

used drier midslope areas less likely to be occupied by barred 

owls, possibly as a mechanism to minimize interactions with 

barred owls, at least in the near term (Singleton 2013). This 

pattern reflects displacement of spotted owls by barred owls 

from highly suitable forest into conditions less favorable to 

long-term reproduction and survival of spotted owls, a finding consistent with long-term demographic studies of spotted 

owls throughout the range of the subspecies (Dugger et al. 

2016, Forsman et al. 2011, Singleton 2013, Wiens et al. 2014).

In addition to impacts on spotted owls, changes in the 

abundance and distribution of an apex predator like the 

barred owl can have cascading effects on prey populations 

and food web dynamics (Holm et al. 2016, Wiens et al. 

2014), as well as populations of other small sympatric 

owls (Acker 2012, Elliot 2006). Differences in space use, 

abundance, demography, suitable forest, diets, and behavior 

collectively suggest that the barred owl is not a direct 

functional replacement of the spotted owl in old-growth 

forest ecosystems (Holm et al. 2016, Wiens et al. 2014). As 

a consequence, additional changes in community structure 

and ecosystem processes are anticipated as a result of barred 

owl encroachment into areas managed under the NWFP.

Spotted Owl Prey

Like all predators, spotted owls are dependent on abundant 

and vulnerable prey. Much is known about the ecology 

and population demography of spotted owls, but little 

information exists on how fluctuations in populations of 

prey species influence behavior, space use, reproduction, or 

population growth of spotted owls. Spotted owls in some 

areas during some periods have had a strong 2-year cycle 

of high reproduction one year followed by a year of low 

reproduction (Anthony et al. 2006). One hypothesis for the 

cycle in reproductive output is variation in prey abundance. 

However, simple prey relationship models do not explain 

the highly synchronous and temporally dynamic patterns 

of spotted owl reproductive performance (Rosenberg et 

al. 2003). Northern flying squirrels, woodrats, red-backed 

voles, and red tree voles are the primary prey of spotted 

owls throughout different regions of the spotted owl’s 

geographic range (Barrows 1980; Bevis et al. 1997; Forsman 

et al. 1984, 2004, 2005; Hamer et al. 2001; Rosenberg et al. 

2003; Wiens et al. 2014; Zabel et al. 1995). None of these 

studies had data that could be used to examine relationships 

between annual variation in prey abundance and annual 

variation in survival or fecundity of spotted owls. Although 

deer mice are not a primary prey species (<2 percent biomass consumed), one study (Rosenberg et al. 2003) found a 

positive correlation (r2

 = 0.68) between abundance of deer 

mice and reproductive success of spotted owls.

Abundance and distribution of primary prey species 

can influence space use by spotted owls. For example, 

spotted owls more frequently use riparian areas within their 

home ranges (Wiens et al. 2014), perhaps because the cool 

microclimates associated with stream drainages may be 

favorable for thermoregulatory purposes during summer 

months (Barrows 1981), or more importantly, riparian areas 

are likely to support a rich diversity of prey (primarily 

small mammals) used by spotted owls (Anthony et al. 2003, 

Carey et al. 1999, Forsman et al. 2004). Home ranges of 

spotted owls tend to be smaller in the southern portion 

of the subspecies range, where woodrats are the primary 

prey, as compared to the northern portion of the geographic 

range, where woodrats are uncommon and northern flying 

squirrels are the primary prey (Forsman et al. 2005, Zabel et 

al. 1995). In northern California, southwestern Oregon, and 

the eastern Cascades, woodrats occur in fairly open forests 

and at much greater densities compared to northern flying 

squirrels (Carey et al. 1992; Lehmkuhl et al. 2006a, 2006b; 

Wilson and Forsman 2013; Zabel et al. 1995). Differences 
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GENERAL TECHNICAL REPORT PNW-GTR-966

in space use by spotted owls in different portions of their 

range also relate to regional differences in the availability 

of prey species. Northern flying squirrels and red tree voles, 

for example, occur at highest densities in the complex 

structure of mature Douglas-fir stands with old-growth 

characteristics, whereas woodrats have greater densities 

in young stands, along edges, or in brushy areas (Carey et 

al. 1992, Price et al. 2015, Sakai and Noon 1993, Swingle 

and Forsman 2009, Walters and Zabel 1995, Zabel et al. 

1995). Spotted owls used forest edges to a greater degree 

when forage consisted primarily of woodrats (Diller et al. 

2012), but preferred forest interiors, where they foraged on 

red tree voles and northern flying squirrels. Timber harvest 

activities, including thinning of dense plantations, reduce 

the abundance of northern flying squirrels and red tree voles 

for several decades, contributing to a reduction in use by 

spotted owls (Carey 2000, Dunk and Hawley 2009, Gomez 

and Anthony 1998, Manning et al. 2012, Price et al. 2015, 

Waters and Zabel 1995, Wilson and Forsman 2013). 

Disturbance

In this section, we define disturbances as modifiers of 

the structural characteristics, species composition, and 

landscape patterns of forest cover types used by spotted 

owls. The range of the northern spotted owl encompasses 

a variety of historical disturbance regimes that are fundamental to the health and diversity of these ecosystems 

(chapter 3). Important forest disturbances result from 

wildfire, forest management (e.g., thinning), timber 

harvests, extreme weather events, or forest insect and 

disease processes (Davis et al. 2016). Effects that forest 

disturbances have on spotted owls depends on spatial scale, 

severity, and season (McKelvey 2015). Biogeographic 

variation across the large range of spotted owls also results 

in very different levels of disturbance type, frequency, 

and severity (see “Wildfire” below). Major disturbance 

events influence forest cover types that have been used by 

spotted owls for many decades, and have different effects 

depending on the magnitude of change and the time since 

disturbance. For example, in the short term, a disturbance 

that creates open canopy conditions could reduce value 

for spotted owl roosting, but have long-term benefits by 

enhancing understory vegetation diversity and conditions 

for spotted owl prey. Further, disturbances can stimulate 

the development of large-tree, complex-structure stand 

conditions over time (Lehmkuhl et al. 2015). An important 

secondary effect of forest disturbances for spotted owls 

are changes in prey abundance or vulnerability. These 

effects can be positive by creating conditions that increase 

abundance or vulnerability for some prey species, or 

negative by removing critical forest structure required by 

primary prey populations (e.g., northern flying squirrel, red 

tree vole) (Manning et al. 2012, Wilson and Forsman 2013). 

Some disturbances have a neutral affect, particularly when 

limited in severity or spatial extent, and ample suitable 

forest remains available at core and home-range scales. 

Spotted owls were listed as a threatened species under 

the Endangered Species Act largely because of concerns 

regarding loss of old forest resulting from commercial 

timber harvest (Thomas et al. 2006, USFWS 2011b). Subsequent to reductions in harvest of old forest, high-severity 

wildfire has become the leading cause of suitable forest loss 

for spotted owls on federal lands, especially in fire-prone 

landscapes. However, commercial timber harvest still 

contributes substantially to the loss of suitable forest cover 

in some areas, especially on nonfederal lands (Davis et al. 

2016, Pierce et al. 2005). Recent research on disturbance 

effects on spotted owls indicates that disturbances such as 

mixed-severity fires that generate heterogeneity at landscape and stand scales are not necessarily adverse, provided 

that adequate nesting and roosting structural conditions 

remain after the disturbance (Clark et al. 2013, Comfort 

et al. 2016). High-severity disturbances that broadly alter 

stands and landscapes within nesting territories can remove 

critical components of forest structure (e.g., high canopy 

cover and density of large live trees) required for spotted 

owl survival and reproduction (Dugger et al. 2005, Franklin 

et al. 2000, Olson et al. 2004). Timber harvesting and 

wildfire can both reduce the living tree components of a 

stand and reduce the overall suitability for spotted owls (see 

sidebars on pages 265 and 266). An important difference 

between timber harvest and wildfire is the removal of trees 

and ground disturbance in a timber harvest. For most wildfires, there is limited physical soil disturbance (although fire 
265

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

 

 



       

!!! ! ! !!!! !!! ! !!!! !!!! !!! !!!! ! !  !!!! !!!!!!  ! !!

! !   !   ! !  ! !! ! !! ! ! !  ! !  !!! !! !!!!!!!!  !!!!!

!!

!!!  ! ! ! ! ! !    ! !!  !      !  !  !!  !  ! ! ! !  !!   !  ! ! ! ! !  !         !     ! !!! !! ! !! !! !!!! !!!!!!!

 Timber

! Wildfire

Effects of Forest Disturbances on

Nesting/Roosting Forest Cover

Map data from the most recent northern spotted owl habitat 

monitoring report (Davis et al. 2016) and Forest Inventory 

Analysis and Current Vegetation Survey plots were 

used to assess changes resulting from forest disturbances on 

stand structure elements used in the Davis et al. (2011, 2016) 

nesting/roosting cover type modeling and mapping procedure.

Plots used in this analysis occurred in mapped suitable nesting/

roosting cover type in 1993 that experienced a disturbance between 1994 and 2012 from either timber harvesting or wildfire, 

which occurred between the initial plot measurement and remeasurement dates.

Changes in the mapped nesting/roosting relative suitability 

index were also analyzed by differencing the 2012 and 1993 

relative suitability maps.

LandTrendr (LT) data (Kennedy et al. 2012) of forest disturbance 

magnitude are satellite-based measurements of loss of vegetation 

cover. We divided them into three classes:

• Low (<33 percent cover loss)

• Moderate (33 to 66 percent cover loss)

• High (>66 percent cover loss

The graph to the right is from Davis et al. 

(2015) and shows the relationship between 

these classes and monitoring trends in

burn severity classes.

100

90

80

70

60

50

40

30

20

10

0

LT low 

magnitude

LT moderate

magnitude

LT high

magnitude

4

16

59

9

39

32

8 10

2

86

45

90

Percent

High

Moderate

Low

Unburned 

to low
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GENERAL TECHNICAL REPORT PNW-GTR-966

80

70

60

50

40

30

20

10

0

Cover percentage

Low Mod. High

Effects on modeled relative suitabilityb of forest 

cover types used for nesting and roosting

Conifer cover Stand height

Large tree 

density Diameter diversitya Diameter at 

breast height

Effects on stand structure elements

a

 The diameter diversity index is a measure of the stuctural diversity of a forest stand based on live tree 

densities in diameter at breast height classes. An index value of 1 represents a single-story stand. 

Higher index values represent more complex multistoried stand structure.

b

 Modeled relative suitability from Davis et al. (2016). The index for suitable nesting/roosting forest cover 

type ranges from 31 to 40, depending on the modeling region. 

Timber harvesting

Wildfire

Disturbance type

Before disturbance

After disturbance

Plot measurement

80

70

60

50

40

30

20

10

0

Cover percentage

Low Mod. High

Trees per hectare

Low Mod. High

30

25

20

15

10

5

0

30

25

20

15

10

5

0

Trees per hectare

Low Mod. High

90

80

70

60

50

40

30

20

10

0

Centimeters

Low Mod. High

40

35

30

25

20

15

10

5

0

Meters

Low Mod. High

40

35

30

25

20

15

10

5

0

Meters

Low Mod. High

7

6

5

4

3

2

1

0

Index value

Low Mod. High

7

6

5

4

3

2

1

0

Index value

Low Mod. High

90

80

70

60

50

40

30

20

10

0

Centimeters

Low Mod. High

15

5

-5

-15

-25

-35

-45

-55

-65

Change in suitability index

LOW MOD

15

5

-5

-15

-25

-35

-45

-55

-65

Change in suitability index

HIGH LOW MOD HIGH
267

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

can have substantial impacts on soil chemistry and organic 

matter composition), and patches of live trees, snags, and 

logs remain in situ, which contributes to enhanced biodiversity, future quality of complex forest, and forest succession 

(Swanson et al. 2011).

Wildfire—

Wildfires occur throughout the entire range of the spotted 

owl. Some physiographic provinces are more environmentally suitable for wildfire occurrence at a decadal 

scale, while other provinces have wildfire-return intervals 

of several centuries (see chapter 3) (Agee 1993). Beyond 

frequency, the severity and spatial extent of wildfires 

differ across the NWFP area (Davis et al. 2011). The 

physiographic provinces of the eastern Cascades, southern 

portions of the western Cascades, and the Klamath Mountains are characterized by frequent low- and mixed-severity 

fire regimes (Baker 2015, Hessburg et al. 2007, Perry et 

al. 2011). Owing to more than a century of fire exclusion 

(e.g., from grazing, fire suppression, and historical forest 

management practices), many of these fire-prone landscapes 

have experienced significant increases in stand density and 

loss of large trees, threatening forest health and biodiversity 

(Hagmann et al. 2013, 2014; Hessburg et al. 2007; Perry et 

al. 2011). The historical extent of forest cover types suitable 

for nesting and roosting by spotted owls in dry and mesic 

mixed-conifer forests in the eastern Cascades and other 

fire-frequent forests was likely historically limited but has 

increased substantially in recent decades (Hagmann et al. 

2013, 2014; Hessburg et al. 2007; Merschel et al. 2014). 

Moreover, in this fire-prone landscape, forest structure 

conditions that are more resilient to low- and mixed-severity fires (i.e., single-story old forests with large ponderosa 

pines) are not suitable for nesting and roosting by spotted 

owls. Areas occupied by spotted owls in the fire-prone 

landscapes of the eastern Cascade Range are often dense, 

closed-canopy, medium-size tree forests with a substantial 

true fir (Abies spp.) component and structural diversity 

enhanced by a variety of insect and disease processes, 

including dwarf mistletoe (Stine et al. 2014). These are 

the conditions that have been promoted through fire 

suppression and removal of fire-resilient large Douglas-fir 

and ponderosa pine trees. Compared to forest structure 

conditions that are more resilient to wildfire, areas occupied 

by spotted owls in these fire-prone landscapes are at higher 

risk to high-severity wildfire (Dennison et al. 2014, Stine et 

al. 2014). All forest types in these landscapes are vulnerable 

to substantial impacts from high-severity wildfire under 

extreme weather conditions, which are likely to be more 

common with climate change (Kennedy and Wimberly 

2009, Reilly et al. 2017). 

West of these fire-prone areas to the Pacific coastline, 

the forests become progressively moister and less prone to 

frequent large wildfire. In these moist forests, large wildfires 

tend to be infrequent to moderately frequent, and fire severity trends from mixed to high severity (see chapter 3). In 

less fire-prone landscapes, old and complex forest with large 

trees—compared to other forest types—has higher moisture 

retention and cooler microclimates compared to other forest 

types, and may enhance biodiversity under a changing climate (Frey et al. 2016). In these mixed- and low-frequency 

fire regime landscapes, old forest may be more resistant to 

wildfire than young forest with closed canopy under normal 

fire weather conditions (Thompson and Spies 2009). 

Throughout the NWFP area, the fundamental association between spotted owls and multilayer forests with 

large trees and closed canopies is well established (Dugger 

et al. 2016, Forsman et al. 1984, Franklin et al. 2000, Olson 

et al. 2005, Wiens et al. 2014). The severity of the wildfire 

has a strong influence on the degree to which these forest 

cover types are altered by wildfire (see sidebar on page 

265). Low-severity wildfire can have very little effect on 

the suitability of nesting and roosting cover types, and can 

even increase it. Moderate-severity wildfire can change 

stand structure and species composition, resulting in 

moderate decreases in cover-type suitability. High-severity 

wildfire can alter forest cover to the point at which the area 

is no longer be suitable for nesting, roosting, or dispersal. 

Multiple lines of research have confirmed the effects of 

wildfire on stand structure and composition, but much less 

is understood about the short- and long-term response of 

spotted owls to wildfire. 

Most studies focused on wildfire effects evaluated 

the short-term response of spotted owls to wildfire, but in 

one of the few studies of the long-term effects of wildfire 
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GENERAL TECHNICAL REPORT PNW-GTR-966

on spotted owls Rockweit et al. (2017) used 26 years of 

demographic data in a landscape with several wildfires and 

found that moderate and high burn severities negatively 

affected spotted owl apparent survival. They also found 

that burned territories functioned as ecological sinks where 

recruitment was high, but survival was lower than in nearby 

unburned territories. Several shorter post-wildfire studies 

have seemingly contradictory results regarding spotted owls 

and wildfire. For example, in an occupancy analysis, Jones 

et al. (2016) found high site extirpation rates of California 

spotted owls following a large, high-severity wildfire, but in 

a telemetry study, Bond et al. (2016) observed that burned 

forests were generally used in proportion to their availability. Other studies of California spotted owls and Mexican 

spotted owls have shown that wildfire does not necessarily 

decrease short-term occupancy in low- or moderate-severity 

burned areas (Bond et al. 2009, Ganey et al. 2011, Lee and 

Bond 2015, Lee et al. 2013, Roberts et al. 2011). Spotted 

owls can persist, at least for short periods, in landscapes 

that have experienced recent wildfires, as long as adequate 

moderate to closed-canopy nesting/roosting forest cover is 

retained at nesting core and home range scales. Even with 

high-severity wildfire, the effects can be insignificant or 

positive (e.g., increase vulnerability of prey) at larger spatial 

scales, especially if the forest cover changes caused by 

high-severity fire comprise only a small portion of a spotted 

owl’s territory (Comfort et al. 2016). 

Effects of wildfire interact in complex ways with 

other historic and current disturbances. Clark et al. (2013) 

found that local spotted owl site extinction probability 

was higher for sites with more combined area of past 

timber harvest, high-severity fire, and salvage logging. 

They also found evidence that colonization and occupancy 

rates were higher for sites with older forest burned at low 

severity (Clark et al. 2013). Coupling wildfire and salvage 

logging results in a high probability that a site becomes 

unoccupied after the first year postfire, especially if the 

core area burns at high severity and is subsequently logged 

(Bond 2016, Ganey et al. 2011, Lee et al. 2013). Beyond the 

effects on spotted owls, a human disturbance that directly 

follows a high-severity natural disturbance can have 

significant negative consequences to a forest ecosystem 

by disrupting abiotic and biotic processes, reducing or 

eliminating biological legacies, simplifying post-disturbance structural complexity, altering vegetation recovery, 

diminishing natural patterns of landscape heterogeneity, 

facilitating invasion of nonnative species, decreasing 

native biodiversity, increasing susceptibility to erosion 

and repeated high-severity disturbances, and eliminating 

restorative benefits of disturbance events (Lindenmayerand Noss 2006, Thorn et al. 2017).

Overall, studies suggest that spotted owls are adapted 

to a forest landscape with a mosaic of successional stages 

shaped by historical disturbance regimes, accompanied by 

abundant prey resources, few barred owls, and structurally 

diverse closed-canopy forest with diffuse late-seral edge 

at the territory scale, and limited fragmentation occur 

within nesting areas (Dugger et al. 2011, Forsman et al. 

1984, Franklin et al. 2000, Olson et al. 2004, Sovern et 

al. 2014). Research supports the premise that some spatial 

heterogeneity in forest conditions can have a positive effect 

on demography of spotted owls. At the territory scale (~500 

to 1500 ha), a mosaic of older forest interspersed with other 

vegetation types, including early-seral and riparian forests, 

can promote high survival and reproduction of spotted owls 

(Comfort et al. 2016, Franklin and Gutiérrez 2002, Franklin 

et al. 2000). In terms of the effects of wildfire on spotted 

owls, we emphasize that most available research on impacts 

to spotted owls has been based to some degree on shortterm responses and primarily focused on the other two 

spotted owl subspecies. The long-term (>5 years) effects of 

wildfire on spotted owl survival, reproduction, recruitment, 

and interactions with barred owls are not well documented. 

Forest restoration and silvicultural treatments—

To meet management objectives of the NWFP, the spotted 

owl recovery plan, and critical habitat requirements, 

researchers and federal land managers have focused on 

ecosystem function (e.g., fire as an ecological process) in 

developing silvicultural practices that provide ecologically 

sustainable alternatives to clearcutting and old-growth 

harvest while still providing for timber production (chapter 

3). As a result, alternative thinning methods, including 

variable-density thinning, have replaced clearcutting as 

the predominant form of active management on federal 
269

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

lands, whether for restoration or timber production goals 

or both (Anderson and Ronnenberg 2013, Lehmkuhl et al. 

2015). Ecological objectives for forest management differ by 

region, forest type, and historic disturbance regime (Franklin and Johnson 2012) (chapter 3). 

Moist forest—The focus of silvicultural treatments in moist 

forests of the western Cascades and Coast Ranges (historically infrequent, high-severity fire regimes) has been an 

attempt to accelerate development of old-forest conditions 

in plantations or younger closed-canopy stands (Anderson 

and Ronnenberg 2013). Typical thinning treatments that 

create canopy gaps in moist forests west of the Cascade crest 

can create relatively rapid increases in understory vegetation diversity and productivity (Johnson and Franklin 2013) 

(chapter 3). The intensity and pattern of retained trees in 

forest thinning can have dramatic influence on microclimate 

and ecological response in the short term (Aubry et al. 2009, 

Heithecker and Halpern 2006). Stand conditions can be either too open or too dense for foraging because spotted owls 

are adapted to old forest with closed canopies, and the understory must be open enough to fly and access prey (Irwin et 

al. 2015). In areas where dusky-footed woodrats are primary 

prey (e.g., southern Oregon, northern California), thinning of 

young dense stands may increase spotted owl use for foraging, but still not create preferred forest conditions for other 

life history needs such as nesting and roosting (Irwin et al. 

2015). Wilson and Forsman (2013) found that the abundance 

of mice, terrestrial voles, and shrews increased immediately 

following thinning, but that northern flying squirrels and red 

tree voles—important prey species for spotted owls—decreased dramatically in abundance in treated areas for up to 

11 years after treatment (Wilson 2010). Thus, spotted owls 

respond to silvicultural treatments differently where the 

primary prey are northern flying squirrels, which includes 

most of the northern and western portions of their range in 

Oregon, Washington, and British Columbia. 

When assessing the potential effects of thinning on 

prey species, the landscape context should be considered. 

For example, the effects of thinning within heterogeneous 

landscapes with well-connected, intact old-forest cover 

may be less detrimental to northern flying squirrels than if 

thinning occurs within a highly fragmented forest landscape (Sollmann et al. 2016). Some degree of landscape 

heterogeneity resulting from forest restoration activities in 

west-side forests does not adversely impact spotted owls, 

provided that sufficient large-tree, closed-canopy forest for 

nesting and roosting is available at core and home range 

scales (Andrews et al. 2005). For example, in northern 

California, Franklin et al. (2000) found that territories with 

the highest fitness (survival and reproduction) were those 

with a mixture of old forest and about 40 percent of other 

vegetation types. Diller et al. (2012) reported that forest 

cover heterogeneity (i.e., juxtaposition of young and older 

stands) had positive effects on survival and reproduction 

of spotted owls on commercial timberlands in northern 

California, where disturbance regimes were historically of 

mixed severity. Highly productive growing conditions and 

abundant hardwoods contribute to structural complexity in 

these managed forests. However, survival of spotted owls 

decreased in southern Oregon when the amount of nesting/

roosting forest cover within the territory center was less 

than 50 percent (Dugger et al. 2005), and a similar relationship was found in other studies (Franklin et al. 2000, Olson 

et al. 2004, Wiens et al. 2014).

Dry forest—In the drier forests of the eastern Cascades, 

southern Oregon, and northern California, wildfire was historically more frequent and burned with mixed- and low-severity effects. In these areas, forest management treatments 

have focused on accelerating the development of old-forest 

conditions, but also have focused more on restoring or promoting fire-resilient forest structure, species composition, 

and landscape pattern (Hessburg et al. 2016, Lehmkuhl 

et al. 2015, Stine et al. 2014). Landscape managers implementing forest restoration treatments in drier, mixed- and 

low-severity fire regime forests face substantial challenges 

in balancing the tradeoffs between known short-term forest 

cover impacts on spotted owls from restoration and fuel 

reduction treatments versus potential benefits of reducing 

losses of forests with larger trees from high-severity, largescale wildfire (Hessburg et al. 2015, 2016; Lehmkuhl et al. 

2015; Stine et al. 2014). Management emphasis on wildfire 

suppression combined with historical harvest of large trees 

in these landscapes over the past 100 years has contributed to the recruitment of small-tree, closed-canopy forest 
270

GENERAL TECHNICAL REPORT PNW-GTR-966

(Hessburg et al. 2016). In these regions, the moderate- to 

closed-canopy forest with multilayer canopy structure enhanced by dwarf mistletoe infestations are used by spotted 

owls for nesting and roosting areas, and appear to have 

increased over the latter part of the 20th century into the 21st

century (Davis et al. 2016, Lint 2005). Large tree, multistory canopy typical of forest cover types used for nesting 

and roosting by spotted owls across their range make them 

less flammable under most fire conditions, but, like most 

cover types, these are susceptible to burning intensely in 

extreme weather. Standard treatments focused on increasing stand-level resilience to wildfire by using prescribed 

fire and removing ladder fuels (e.g., Cochrane et al. 2012, 

Safford et al. 2012, Stephens et al. 2009), and reducing canopy connectivity (Agee and Skinner 2005) can reduce the 

risk of stand-replacement high-severity wildfires, but the 

practices also remove important forest cover elements for 

spotted owls and their prey (Lehmkuhl et al. 2006a, 2006b, 

2015). Prescribed fire treatments as part of fuel reduction 

projects can further reduce under- and mid-story canopy 

complexity, and burn up logs and snags, potentially causing 

additional negative impacts to suitable forest for spotted 

owls and their prey (Lehmkuhl et al. 2015). Silvicultural 

practices that promote spatial and structural complexity 

have been proposed for retaining suitable foraging conditions for spotted owls while also reducing fuel loads 

(Churchill et al. 2013, Gaines et al. 2010, Hessburg et al. 

2016, Johnson and Franklin 2013, Lehmkuhl et al. 2015). 

However, the effectiveness of these management practices 

to restore ecological resilience and reduce risk of loss to 

high-severity wildfire, while maintaining components of 

suitable forest for spotted owls, remains to be tested in dry 

forest landscapes (see chapters 3 and 12 for more discussion 

of this issue). 

Several simulation studies have used coupled wildfire 

and forest growth models to investigate the relative effects 

of wildfire and forest restoration treatments on recruitment 

and retention of forest cover types used by spotted owls in 

fire-prone landscapes. Some of these studies suggest that 

certain fuel treatment scenarios (i.e., active management) 

can reduce wildfire-caused losses of forest cover types used 

by spotted owls (Ager et al. 2007, Roloff et al. 2012). Other 

modeling efforts found that active management reduced 

forest cover used by spotted owls more than simulations 

with no management, (Roloff et al. 2005, Spies et al. 2017). 

As with any modeling exercise, outcomes of these studies 

reflect the assumptions incorporated into the simulations. 

Assumptions regarding wildfire severity, return intervals, 

and effects of treatments are particularly influential. One 

general theme from these simulations is that benefits of 

fuel treatments to forest types used by spotted owls depend 

on what probability of occurrence is assumed for future 

high-severity wildfires. If the likelihood and impacts of 

high-severity wildfire are assumed to be high, thinning 

treatments are more likely to have a positive outcome for 

spotted owls (e.g., Roloff et al. 2012). If the likelihood 

of high-severity wildfire is assumed to be low, however, 

then thinning treatments are more likely to produce only 

declines in the amount of suitable forest cover types used by 

spotted owls. 

Climate Change

Climate change will affect spotted owl populations through 

changes in weather, forest cover, disturbance processes, 

prey availability, and other ecological interactions. Population growth of spotted owls appears to be positively associated with wetter than normal conditions during the growing 

season (May–October), which likely increases prey populations and thus availability (Glenn et al. 2010). Population 

growth and reproduction were also negatively associated 

with cold, wet winters (pre-nesting) and the number of hot 

summer days (July–August) (Diller et al. 2012, Glenn et al. 

2011b). Annual survival was more closely related to regional 

climate conditions (Southern Oscillation Index [SOI] and 

Pacific Decadal Oscillation [PDO]), whereas recruitment 

was often associated with local weather. Projected future 

climate conditions have the potential to negatively affect 

annual survival, recruitment, and, consequently, population 

growth rates for spotted owls (Glenn et al. 2010). Climatic 

factors affecting vegetation and prey abundance likely have 

a greater effect on reproduction and population growth than 

direct effects of weather on nestlings or adult spotted owls 

(Glenn et al. 2011a, 2011b). Climate change models for the 

first half of the 21st century predict warmer, wetter winters 
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and hotter, drier summers for the Pacific Northwest (Mote 

et al. 2003) (chapter 2). These conditions are expected to 

decrease survival of spotted owls in some areas (Glenn et 

al. 2011a). Climate change can affect development of forest 

structure by altering temperature and precipitation regimes, 

and disturbance frequency and intensity (Dale et al. 2001). 

Altered understory vegetation can reduce prey availability 

and thus spotted owl fitness (Carey and Johnson 1995, 

Franklin et al. 2000). Carroll (2010) found that vegetation 

rather than climate variables best explained distributions of 

spotted owls. Potential climate-related forest cover losses 

resulting from large-scale, high-severity wildfires and 

increased mortality of old-growth trees (Van Mantgem et al. 

2009) may be particularly important for future viability of 

spotted owl populations (chapter 2). 

Franklin et al. (2000) found that forest cover patterns 

explained a high amount of spatial variation in fitness 

potential among territories occupied by spotted owls in 

northern California, but climate explained most of the 

temporal, year-to-year variation in fitness-related traits. Survival and reproduction, for example, were lower when the 

early nesting period (February–March) was cold and wet. 

Fecundity, recruitment, and survival decreased across the 

range of the spotted owl when winters or early springs were 

colder and wetter than average (Diller et al. 2012; Dugger 

et al. 2005, 2016; Forsman et al. 1984, 2011). Spotted owl 

populations in drier forests may be especially vulnerable to 

climate change because hot, dry summers can reduce prey 

abundance or availability, and subsequently reduce spotted 

owl survival (Glenn et al. 2011a). Regional climate patterns, 

including the SOI and PDO, have also been correlated with 

demographic rates of spotted owls (Dugger et al. 2016; Forsman et al. 2011; Glenn et al. 2010, 2011a, 2011b). Survival of 

spotted owls was greater when the PDO was in a warming 

phase and lower when the SOI was negative (i.e., El Niño 

events resulting in higher than average temperatures and 

below normal precipitation) (Dugger et al. 2016).

Extrapolation of the best combination of vegetation-climate models to predicted future climates suggests 

northward expansion of high-suitability forest cover for 

spotted owls (Carroll 2010). Increased winter temperature 

under future climates might be expected to increase winter 

survival and nesting success, and allow range expansion 

of prey species such as woodrats, which currently occur 

at high densities only in the southern portions of the range 

(Noon and Blakesley 2006). However, it is uncertain how 

barred owls will respond to changing prey populations, 

and model results suggest that an initial expansion in the 

suitable climatic niche may be followed by a contraction 

as climate change intensifies (Carroll 2010). An important 

qualifier is that these models did not account for losses of 

multilayered forests to wildfire and the potential for competition with barred owls to become even more prevalent as 

climatic change causes shifts in forest communities that in 

turn further constrain both owl species to a common set of 

increasingly limited resources.

Other Threats

Genetic diversity and hybridization—

Loss of genetic diversity within a population can contribute 

to inbreeding depression and decrease adaptive potential. 

Increased rates of hybridization with barred owls may 

further compromise the genetic integrity of the spotted owl 

population (Funk et al. 2010, Gutiérrez et al. 2007). Genetic 

studies have reinforced other studies that showed spotted 

owl population declines. Specifically, genetic evidence indicates a loss of genetic variation and increased potential for 

inbreeding depression in small populations. This suggests 

a vulnerability of spotted owls to extinction (Funk et al. 

2010). Genetic data from spotted owls have indicated population bottlenecks for the Washington eastern Cascades, 

northern Oregon Coast Range, and Klamath Mountains 

(Funk et al. 2010), which corresponded temporally with 

population declines in most of those regions (Anthony et al. 

2006, Dugger et al. 2016, Forsman et al. 2011). There was, 

however, no definitive evidence that suitable forest cover 

associated with dispersal was limited, or that gene flow was 

restricted in those regions (Barrowclough et al. 2005, Davis 

et al. 2011)

Hybridization with barred owls is another potential 

threat to spotted owl persistence, especially as the spotted 

owl becomes increasingly rare and the invading species 

becomes more abundant (Gutiérrez et al. 2007, Haig et 

al. 2004). Spotted owls occasionally mate with barred 
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owls (male spotted owl–female barred owl mating is most 

common) and produce fertile hybrids (Hamer et al. 1994, 

Kelly and Forsman 2004). In the southern portion of the 

spotted owl range, 3 percent of spotted owl genetic samples 

collected prior to 2004 (barred owls were still relatively 

rare on the landscape) contained barred owl mitochondrial 

DNA (Barrowclough et al. 2005). There are typical markings of hybrids that can be helpful in field identification 

(Hamer et al. 1994), but genotyping potential hybrids across 

generations has shown that field identifications were often 

wrong (Funk et al. 2007). Hybridization rates may also have 

changed substantially in recent years as barred owl populations have increased and spotted owls have decreased. 

Hybridization with other spotted owl subspecies does 

not appear to be a concern for spotted owl conservation. 

The northern spotted owl and California spotted owl are 

two well-differentiated subspecies connected by a narrow 

hybrid zone in a region of low population density for both 

subspecies in north-central California (Barrowclough et al. 

2005, 2011; Funk et al. 2008; Gutiérrez and Barrowclough 

2005). Spotted owls in the contact zone are highly differentiated and may be a distinct population from other northern 

spotted owl and California spotted owl populations (Miller 

et al. 2017). 

Diseases and pathogens—

Disease exposure could be a secondary consequence of 

climate change, blood parasites, and effects of barred 

owl interactions. Lewicki et al. (2015) found that spotted 

owls had a higher Haemoproteus spp. parasite diversity 

and probability of infection than sympatric barred owls. 

Further, avian malaria (Plasmodium spp.) is common in 

barred owls, and only recently was documented in spotted 

owls; therefore, barred owls likely have an additional 

competitive advantage because spotted owls are potentially 

immune-compromised owing to recent exposure to avian 

malaria (Ishak et al. 2008). Spotted owls are susceptible to 

West Nile virus and experience high rates of mortality when 

exposed (Courtney et al. 2004); however, it is unknown 

what, if any, population-level impacts the disease has 

caused. Wiens et al. (2014) reported that the leading cause 

of death in a sample of radio-marked barred owls was 

bacterial infection associated with endoparasitism.

Environmental contaminants—

Environmental contaminants, especially anticoagulant 

rodenticides, have recently emerged as a potential threat 

to spotted owls and their prey. In particular, anticoagulant 

rodenticides used in illegal marijuana cultivation and urban 

settings can have significant indirect impacts by the poisoning of nontarget forest predators, including owls (Albert et 

al. 2010, Gabriel et al. 2012, Riley et al. 2007, Stone et al. 

1999). To our knowledge, no studies have addressed potential effects of anticoagulant rodenticides on spotted owls. 

Research Needs, Uncertainties, 

Information Gaps, and Limitations

Research Needs

Effects of barred owls—

It has become increasingly clear that barred owls are a 

primary driver of spotted owl population declines, but 

many questions remain about the full impact of barred owls 

directly on spotted owls, and indirectly through alterations 

of forest communities. Research is needed to build on the 

work of Wiens et al. (2014) and others to identify potential 

processes by which spotted owls and barred owls use 

resources differently. More research is needed to establish 

the full suite of cause-and-effect relations of barred owl 

impacts on spotted owls, and how barred owls interact with 

other threats to spotted owls. Unfortunately, these types of 

studies are becoming increasingly difficult because spotted 

owl numbers are declining so rapidly on most study areas. 

In a pilot study, Diller et al. (2016) found that spotted owls 

responded positively to experimental removal of barred 

owls, but additional removal studies in other physiographic 

provinces, where owl populations and suitable forests are 

different, are needed. To determine the feasibility and effectiveness of barred owl removals as a tool for spotted owl 

recovery, the Fish and Wildlife Service and U.S. Geological 

Survey initiated a barred owl removal experiment on four 

study areas in Washington, Oregon, and northern California (USFWS 2013). Continued monitoring of spotted owl 

populations in those areas will be required to fully assess 

the short- and perhaps, in particular, long-term response 

of spotted owls to the removal of an important competitor. 

More genetic studies are needed to address the frequency 
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and impact of hybridization between spotted owls and 

barred owls, and how hybridization rates may have changed 

with changes in abundance of the two species.

It remains uncertain how climate change will affect 

interactions between spotted owls and barred owls, or 

even where barred owl populations are in terms of the 

invasion process. For example, little research has been 

conducted to investigate if populations of barred owls are 

continuing to increase or if carrying capacity has been met 

in some regions. Fundamental information on barred owl 

distribution and population trends is needed to address 

this important issue. Further, little is known about barred 

owl distribution and populations beyond forest cover types 

occupied by spotted owls. Ecologists are being challenged 

to predict how spotted owls will change in abundance and 

distribution under current climate, availability of suitable 

forest, and competitive interactions with barred owls. It is 

well documented that climate change influences species’ 

abundances and distributions, and can have indirect 

effects on interspecific interactions (Angert et al. 2013). 

An important area of needed research related to barred 

owl-spotted owl interactions and climate change will be to 

better understand how the combined effects of barred owl 

competition and future changes in the amount and distribution of forests used by spotted owls might contribute to 

spotted owl population persistence and range shifts under a 

changing climate. 

In addition to impacts on spotted owls, changes in the 

abundance and distribution of a generalist apex predator 

like the barred owl can have cascading effects on prey populations and food-web dynamics (Gutiérrez et al. 2007, Holm 

et al. 2016, Wiens et al. 2014). Barred owls have reached 

densities in the Pacific Northwest that are far greater than 

historical populations of northern spotted owls (Wiens et 

al. 2011, 2014). Moreover, as generalist predators, barred 

owls capture a greater proportion of diurnal, terrestrial, 

and aquatic prey than northern spotted owls (Forsman 

et al. 2004, Hamer et al. 2001, Wiens et al. 2014). These 

life-history traits indicate that barred owls are not direct 

functional replacements of northern spotted owls in forested 

ecosystems of the Pacific Northwest (Holm et al. 2016), and 

that a wide range of prey species may be affected if they 

replace northern spotted owls. Further research is needed 

to determine the potential effects of barred owls on other 

sensitive wildlife beyond spotted owls.

Finally, critical needs for managers are detailed assessments of those locations where spotted owls persist and a 

better understanding of the effects of forest management 

activities on interactions between spotted owls and barred 

owls, and the species individually. Many spotted owl sites 

with apparently suitable forest structure for nesting and 

roosting have been abandoned as a result of displacement 

by barred owls. Those sites that spotted owls have persisted 

in the face of barred owls may be a result of the behavioral 

characteristics of the territorial spotted owl, or perhaps 

those sites have unique forest characteristics that enhance 

coexistence between the two species. Thinning treatments 

could potentially affect competitive interactions either by 

displacing barred owls into areas occupied by spotted owls, 

or potentially increasing foraging opportunities for barred 

owls over spotted owls. These and many other responses 

are plausible, but it remains unknown how either species 

responds to many forest management techniques. Recent 

advances in lightweight geographic positioning system 

telemetry devices and high-resolution forest structure 

mapping technologies can provide new opportunities for 

advancing our understanding of these issues. 

Prey populations and population performance—

Previous studies have characterized the diet of spotted owls 

in different portions of the subspecies’ range (Barrows 

1980; Bevis et al. 1997; Cutler and Hays 1991; Forsman et 

al. 1984, 2001, 2004), investigated the relationship between 

forest cover selection, home-range size, and prey availability (Carey et al. 1992; Forsman et al. 1984, 2005; Irwin 

et al. 2000; Zabel et al. 1995), and evaluated diet overlap 

with barred owls (Hamer et al. 2001, Wiens et al. 2014). 

The importance of understanding relationships between 

spotted owl populations and their prey has repeatedly been 

acknowledged (Clark et al. 2011, Courtney et al. 2004, 

Forsman et al. 2004, Glenn et al. 2010, Olson et al. 2004, 

Rosenberg et al. 2003, Thomas et al. 1990, Wilson and Forsman 2013, Zabel et al. 1995). However, to our knowledge, 

no efforts have been undertaken to quantify the relationship 

between interannual fluctuations in prey abundance and 
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long-term demography of spotted owls. Research is needed 

to understand how spotted owl reproduction, stress levels, 

and survival are influenced by prey species composition 

and abundance, and how prey populations are influenced 

by disturbance or fluctuations in weather and climate. 

Population fluctuations in small mammals have been linked 

with variation in precipitation (Avery et al. 2005, Crespin et 

al. 2002). However, identifying the mechanisms by which 

climate influences population processes of spotted owls and 

their prey remains a challenge (Glenn et al. 2011a). 

A better understanding of the effects of thinning 

treatments and the impacts that anticoagulant rodenticides 

have on spotted owl prey populations will be critical for 

managers. Research and an effect analysis is needed to 

address thinning impacts on spotted owl prey, both within 

treated stands and at broader landscape scales. This information would contribute to thinning prescription development throughout the range of the spotted owl. The use of 

anticoagulant rodenticides in natural systems is increasing, 

especially in areas where illegal marijuana cultivation is 

prevalent. Studies are also needed to better understand the 

individual- and population-level impact of rodenticides on 

spotted owls, and development of management options to 

reduce the ecological impacts. 

Landscape restoration, silvicultural treatments, prescribed fire, and wildfire in moist and dry forests—

Research is needed in both dry and moist forest landscapes 

to evaluate the short- and long-term effects of silvicultural 

treatments and wildfire on spotted owl occupancy, forest 

dynamics, and prey, but research questions differ between 

forest types. For example, the optimization of forest restoration and conservation of spotted owls will require more 

knowledge about the conditions under which restoration 

activities can benefit spotted owls in the long term without 

significant detrimental impact in the short term. Restoration 

activities and objectives are different between moist and 

dry forest landscapes. Current conditions in dry forests are 

generally not sustainable, and some measure of treatment is 

needed to increase fire resiliency of forest stands in at least 

some locations (USFWS 2012b). In these fire-prone landscapes, a common objective is to modify and reduce fuels to 

alter wildfire behavior and to manage for ecological integrity 

based on the natural range of variability (USDA 2012). Additional information is needed to evaluate the consequences 

of fuels reduction and restoration treatments relative to the 

long-term benefits of forest restoration, particularly as large, 

high-severity fires are expected to become more frequent 

because of climate change. This is especially true in the 

frequent low-severity fire regime of the eastern Cascades, 

where environmental conditions favor open pine-dominated 

forests. Studies are needed to identify resilient sites for 

spotted owls in the face of changing forests (e.g., species 

composition changes) caused by climate change, active 

forest management, and increased wildfire occurrence. 

In moist forest landscapes, research is needed to 

determine how or if spotted owls use forest stands where 

thinning has been conducted to accelerate the development 

of late-successional forest characteristics. If spotted owls 

avoid these areas in the short term, work is needed to understand the time before they begin using the areas again. To 

fully understand restoration effects, long-term before/after 

control-impact studies are needed to elucidate spotted owl 

and prey responses to forest restoration treatment effects in 

different ecotypes. 

Research to address restoration and silvicultural 

treatment on spotted owl space use and forest structure 

development will also need to account for the potential 

confounding impact that barred owls are likely to have on 

spotted owl response to restoration efforts. Beyond a better 

understanding of spotted owl response to silvicultural 

treatments, managers need information regarding how 

sympatric populations of barred owls respond to treatments. 

Additionally, research is needed to understand the effectiveness of ecosystem-scale conservation versus conservation 

that targets one particular stage of succession (e.g., late-successional forest characteristics for spotted owls). Finally, 

much more information is needed to evaluate the short- to 

long-term effects that wildfire has on spotted owls in all 

landscapes, with a focus on the relative susceptibility of 

old forest and young forest to high-severity wildfire under 

a range of weather conditions. Finally, it is important to 

note that these research topics become increasingly difficult 

to address as spotted owl populations decline and fewer 

individual owls are available to study in some landscapes. 
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Physiological consequences of stress—

An animal’s ability to cope with stressors is an important 

determinant of its physiological conditions, and therefore, 

health and survival. Environmental perturbations and an 

individual’s response can affect the body’s production 

of hormones, such as glucocorticoids, with negative 

physiological consequences (Carrete et al. 2013, Strong et 

al. 2015). For many species, the level of stress hormone 

corticosterone can be an effective predictor of survival 

probabilities, reproduction, dispersal, and can have 

population-level impacts (Carrete et al. 2013, Romero 

and Wikelski 2001, Romero et al. 2000). Quantification 

of corticosterone in feathers, which is stable over time, 

represents an integrated measure of stress levels (Bortolotti et al. 2009, Sheriff et al. 2011). Stress hormones are 

accumulated in feathers during growth, so can provide 

a measure of stress levels during that time, and can be a 

strong predictor for future survival of individuals (Koren 

et al. 2012). Variation in feather corticosterone can also be 

quantified among individuals of a population, as well as 

through time to track stress over space and time to address 

questions about the health and ecology of a population 

(Bortolotti et al. 2009). 

Hayward et al. (2011) found that spotted owls had a 

glucocorticoid response to acute noise disturbance and 

that spotted owls with nests near noisy roads fledged fewer 

young than those near quiet roads. Corticosterone analyses 

are needed to determine the physiological response to acute 

and prolonged exposure to environmental stressors (e.g., 

barred owls, prey abundance, weather, and human-caused 

disturbance) and response activity for both juvenile and 

adult spotted owls. Our understanding of spotted owl ecology will be improved with studies to evaluate the associations between stress levels and survival, reproduction, and 

dispersal of spotted owls. From a management perspective, 

it is important to understand the stress response of spotted 

owls related to management activities like prescribed fire, 

road construction, various logging systems, and the timing 

of these activities. Additional research will be important 

to understand key stressors for spotted owls and inform 

seasonal restrictions on human activities that can increase 

stress levels.

Dispersal and suitable forest connectivity—

Dispersal behavior for both juveniles and adults may 

increase survival and reproductive success, but also 

increase risks to establishing a home range in an unfamiliar landscape. Juvenile spotted owls disperse within their 

first year and the condition of matrix forest types between 

natal and breeding sites can facilitate or hamper survival 

and movement processes (Forsman et al. 2002). Available 

information for spotted owls suggests that stands used 

for roosting during natal dispersal movements have very 

similar structure as those stands used for nesting and 

roosting activities of adults (>70 percent canopy cover 

and large trees >50 cm d.b.h.), but this finding is based 

on only two studies with no data throughout most of the 

geographic range (Miller et al. 1997, Sovern et al. 2015). 

Further research is needed to understand the contemporary dynamics of juvenile dispersal because many 

assumptions are made about what constitutes forest cover 

suitable to facilitate dispersal by spotted owls. A better 

understanding of the forest structure and configuration 

characteristics of forest conditions that facilitate juvenile 

dispersal is needed to ensure demographic connectivity 

among isolated patches of remaining old forests. Further, 

it remains unknown how barred owls influence juvenile 

spotted owl survival or dispersal. It is possible that some 

of these questions could be addressed with a thorough 

analysis of existing dispersal data from demographic 

study areas.

Historically, adult spotted owls exhibited strong 

nesting-site and mate fidelity, with fewer than 8 percent 

of individuals dispersing to a different territory between 

years (Forsman et al. 1984, 2002). In recent years, however, 

field observations suggest that interterritory movements by 

resident spotted owls are increasing, and that such movements appear to coincide with the colonization of barred 

owls (Dugger et al. 2011, Olson et al. 2005). Research that 

addresses how forest alteration and the presence of barred 

owls interact with social conditions on territories to affect 

movement decisions and survival of individual spotted owls 

will improve our ability to implement forest management 

practices that benefit spotted owls. In addition to helping 

land managers identify the range of conditions within 
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individual owl territories that promote high site fidelity and 

survival, such data can also provide a powerful framework 

for testing broad ecological theories about the causes and 

consequences of breeding dispersal in a long-lived predatory bird with declining populations. 

Testing alternative monitoring protocols—

When the NWFP was developed, mark-recapture and 

random census (i.e., occupancy framework; the proportion 

of sites occupied by spotted owls) population monitoring 

methods were both considered. The decision was made 

to use the mark-recapture method, which was already in 

use. Precise estimates from mark-recapture studies require 

large samples of marked spotted owls; therefore, Lint et al. 

(1999) recommended the use of an independent estimate of 

population trend for comparison with the results from spotted owl demographic studies. Monitoring in an occupancy 

framework (i.e., MacKenzie et al. 2006) could provide an 

independent, empirical assessment of population trends to 

compare with estimates of the annual rate of population 

change. Because of uncertainty about the precision of the 

occupancy-based approach, Lint et al. (1999) recommended 

that statistical power and cost effectiveness of the method 

be explored. 

The low number of spotted owls in some study 

areas suggests that passive acoustic monitoring may be 

an effective solution for future monitoring of spotted 

owl populations. Traditional call-back surveys at night 

(playing spotted owl calls and listening for a spotted owl 

response) are labor intensive, more risky compared to 

daytime work, and only generate reliable data for spotted 

owls. Further, detection probabilities for spotted owls—

using call-back surveys—are negatively influenced by 

the presence of barred owls, and barred owls often do not 

respond to spotted owl calls (Bailey et al. 2009). Callback surveys could also have unintended consequences 

by exposing spotted owls to predation or harassment by 

barred owls or great-horned owls. Primary advantages of 

passive acoustic monitoring are as follows: (1) surveys do 

not require an elicited response from target species; (2) 

surveys are able to detect and do not bias against many 

other species (e.g., barred owl, marbled murrelet, western 

screech-owl, northern pygmy-owl, northern saw-whet 

owl, and many others); (3) increased crew safety because 

all work would be conducted during daylight hours; (4) 

biological training and expertise needed for crew members 

will be much less than is needed for call-back surveys and 

demographic studies; and (5) sound recordings provide 

a permanent record of the detection. A limitation of this 

approach is the time required to process recordings and 

data storage. Automated call detection technology has 

been developed, but improvements are needed, especially 

for call recognizers for rare birds in areas with excessive 

background environmental noise (e.g., rain, streams). 

Research is needed to test alternative methods that take 

advantage of technological advancements in noninvasive 

detection equipment to monitor trends in rare populations. 

The transition to alternative methods to monitor spotted 

owl populations will be most effective if new methods 

have spatial and temporal overlap with traditional methods 

so that robust comparisons can be made between historical 

and contemporary data. 

Population simulation modeling—

The program HexSim (Schumaker 2015) provides a simulation framework for systematically investigating factors that 

influence population function, including forest conservation 

scenarios and emergent competitors. The implementation 

of HexSim by the USFWS (2011b) did not include spatially 

explicit representation of spotted owl interactions with 

barred owls. Modeling exercises that incorporate a more 

sophisticated representation of population interactions with 

barred owls are needed to simulate and predict responses 

of spotted owls to experimental removal of barred owls. 

Two-species models implemented in HexSim could also be 

used to simulate potential efficacy of long-term management programs for barred owls and spotted owls relative to 

critical habitat designations. Current modeling efforts are 

female-only models. A two-sex HexSim implementation for 

the spotted owl population is needed to get at small population processes (e.g., Alee effects and stochasticity in sex 

ratios) that can drive extinction. 
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Scientific Uncertainty

Survival estimates— 

Adult survival is typically the most important factor 

influencing population performance in long-lived raptors, 

and survival estimates for spotted owls have been the focus 

of extensive research and monitoring. As in other meta-analyses of spotted owl demographic data (e.g., Burnham et al. 

1996, Dugger et al. 2016, Forsman et al. 2011), Anthony et al. 

(2006) used capture-recapture methods to estimate apparent 

survival rates of spotted owls. Apparent survival is the product of probabilities that an animal survives and remains in 

the population. If a marked animal permanently emigrates, 

then it is, for purposes of the estimate, presumed dead, 

because emigration and mortality are confounded. Further, 

fates are not known for all individuals because recapture 

probabilities are less than one even when animals remain in 

the population. Therefore, models based on capture-recapture data account for imperfect encounter rates in estimates 

of survival (i.e., apparent survival). Apparent survival rates 

on individual study areas ranged from 0.75 (± 0.03) to 0.89 (± 

0.01) for adults, 0.63 (± 0.07) to 0.89 (± 0.01) for 2-year-olds, 

and 0.42 (± 0.11) to 0.86 (± 0.02) for 1-year-olds. They found 

negative effects of reproduction and barred owls in survival 

rates on several study areas (Anthony et al. 2006). 

Elsewhere, Loehle et al. (2005) used telemetry to study 

annual survival of spotted owls and obtained a known-fate 

estimate of 0.93 (± 0.07), which was considerably higher 

than the apparent survival estimates reported by Anthony et 

al. (2006). Known-fate models estimate survival rate when 

fates (i.e., alive or dead) of individuals can be determined 

with certainty. Loehle et al. (2005) used their results to cast 

doubt on apparent survival estimates from mark-recapture 

studies of spotted owls. They suggested that survival 

estimates from mark-recapture studies were too low 

because some marked individuals left the study areas and 

were assumed to be dead. Anthony et al. (2006) estimated 

a declining spotted owl population; Loehle et al. (2005) 

suggested that the true population change for spotted owls 

was likely stable and not declining. In response, Franklin 

et al. (2006) argued that Loehle et al. (2005) had inappropriately compared their study with the work of Anthony et 

al. (2006) in a number of ways, including (1) the manner 

in which missing radio-marked individuals were removed 

from analyses may have overestimated survival; (2) telemetry-based estimates of survival were not valid for estimating 

bias; and (3) results from the telemetry-based study should 

not be compared to the capture-recapture study because 

study areas differed dramatically in size and distribution. 

Both apparent survival estimates from mark-recapture data 

and known-fate estimates from telemetry studies are valid 

estimates of annual survival. However, in this circumstance 

it was inappropriate to compare telemetry-based survival 

estimates with results from capture-recapture studies, 

which was acknowledged by both sides of the disagreement 

(Franklin et al. 2006, Loehle and Irwin 2006). 

Wildfire risk—

The 2008 recovery plan (now withdrawn) for spotted owls 

(USFWS 2008) suggested a change in the LSR network 

as the foundation of conservation strategies established 

in the NWFP. Because of concern about wildfire, the 

plan recommended a switch from a reserve to a no-reserve strategy in up to 52 percent of the spotted owl’s 

range. For dry forests, the plan recommended thinning 

stands at regular intervals to reduce fuel loading, and 

thus wildfire risk. Hanson et al. (2009) suggested that 

the estimates of wildfire risk used by the USFWS (2008) 

were overestimated and that there was not a strong basis 

for major changes to the NWFP conservation strategy for 

the spotted owl. Spies et al. (2010) defended the estimates 

of wildfire risk and suggested that Hanson et al. (2009) 

had underestimated wildfire risk and were biased against 

active management. Hanson et al. (2010) then responded 

by calling for less focus on fuel treatments in the recovery 

plan for the spotted owl. Because of uncertainty about 

future wildfire occurrence, spatial extent, and severity, we 

cannot know with complete confidence whether wildfire 

risk has been over- or underestimated in these efforts. 

Both the 2008 critical habitat designation and the 2008 

recovery plan were challenged in court, and the inspector 

general of the Department of the Interior issued a report 

concluding that the decisionmaking process for the 
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recovery plan was potentially jeopardized by improper 

political influence (Devaney 2008, USFWS 2011a). The 

court ordered the Fish and Wildlife Service to withdraw 

the 2008 recovery plan and issue a revised recovery plan 

and critical habitat designation. 

Spies et al. (2017) projected that the extent of forest 

cover suitable for spotted owls in the eastern Oregon 

Cascades is expected to increase in coming decades under 

recent historical frequencies and severities of wildfire 

(and current levels of wildfire suppression). Treating the 

landscape to reduce potential loss of suitable forest cover 

for spotted owls with high-severity wildfire still resulted in 

increases in that forest cover type, but not as much as would 

occur without management. The results suggest that managing for resilience to fire and climate change could occur 

without necessarily reducing forest cover from its current 

levels (younger forest is growing into older closed-canopy 

forests to replace dense forests lost thinning or wildfire). 

However, these outcomes are likely to be different under 

climate change or if an alternative landscape-scale treatment design is used (Spies et al. 2017). 

Despite the potential negative effects on spotted owl 

habitat, the overwhelming consensus in the scientific literature is that active management in dry forests is appropriate to reduce wildfire risk and improve ecosystem function. 

Therefore, the 2011 revised recovery plan (USFWS 2011b) 

and 2012 critical habitat designation (USFWS 2012a) for 

spotted owls contained proposals for active management in 

dry forests. In some regions, project planning has moved 

forward, and federal land managers are consulting with 

the Fish and Wildlife Service on a case-by-case basis. 

The debate about active management related to wildfire 

risk for forests used by spotted owls remains unresolved 

and reflects different goals (e.g., ecosystem versus single species) and assumptions about wildfire risk with a 

changing climate. These differences of opinion highlight 

legitimate concerns about where to place the burden of 

proof regarding ecosystem versus species management, 

but the fundamentals of this controversy lie in the diversity 

of philosophical views about ecological goals and the role 

that active management should play on public lands (see 

chapter 12). 

Restoration framework—

Franklin and Johnson (2012) outlined a series of recommendations for an “ecological forestry” framework and a 

forest restoration strategy within the Plan area that reflect 

many of the elements of the revised spotted owl recovery 

plan (USFWS 2011b). They called for reserving older forest 

stands, thinning plantations to accelerate development of 

structural complexity, and implementing variable-retention 

harvests in younger forests to help provide diverse earlyseral ecosystems on moist forest sites. On dry forest sites, 

their strategy called for silvicultural treatments that retain 

and release older trees, reduce stand densities, shift composition toward fire- and drought-tolerant tree species, and 

incorporate spatial heterogeneity at multiple spatial scales 

(Franklin and Johnson 2012). The framework included an 

extensive set of large patches of dense forests on approximately 30 percent of the forested landscape to retain some 

suitable forest for spotted owls while reducing the potential 

for landscape-level high-severity wildfires.

DellaSala et al. (2013) identified seven areas in which 

the ecological forestry framework may fall short of the 

stated goals of the NWFP, and offered 14 recommendations 

to improve the framework and its implementation. They also 

criticized decisions to incorporate some of the elements of 

ecological forestry in the revised recovery plan and revised 

critical habitat designation. Henson et al. (2013) agreed with 

many of the recommendations made by DellaSala et al. 

(2013), but differed on two key perspectives. Henson et al. 

(2013) regarded the potential impacts of wildfire to spotted 

owls as higher risk to species persistence, and suggested that 

in many circumstances, the adverse effects associated with 

active management may be preferable to adverse effects of 

passive management. As with wildfire risk, the fundamentals of this debate reside in philosophical disagreements 

about ecological goals and what role active management 

should play in managing public lands. Most research in 

dry or frequent-fire forest landscapes suggests that active 

management is needed to achieve or accelerate restoration 

objectives, but more study is needed to advance our understanding of disturbance effects on wildlife dependent on old 

forest, especially interactions between wildfire and a range 

of prefire and postfire active management actions. 
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Modeling to inform critical habitat designation—

The Fish and Wildlife Service (USFWS 2012a) produced 

maps of distribution of potentially suitable habitat for 

spotted owls that did not include the effects of barred owls 

on spotted owl distribution, but the effort did incorporate 

the spatial arrangement of forest structure associated with 

nesting/roosting and foraging, and abiotic factors such as 

slope and topographic position, to determine the extent 

of critical habitat. In an alternate analysis, Loehle et al. 

(2015) conducted an accuracy assessment of vegetation data 

used as input to develop the USFWS (2012a) models, used 

independent locations to validate model prediction, correlated model output with spotted owl reproductive success 

in two study areas, and developed alternate models. Their 

independent locations and vegetation evaluations suggested 

a high rate of classification errors, and productivity did 

not correlate well with predictions in their study areas 

(Loehle et al. 2015). Dunk et al. (2015) defended the critical 

habitat model as scientifically rigorous and as meeting the 

goals established by the Fish and Wildlife Service. They 

suggested that Loehle et al. (2015) mischaracterized the 

literature and the Fish and Wildlife Service species distribution model, failed to demonstrate the locations used by the 

agency were biased, and failed to show significant flaws in 

analytical methods. 

Bell et al. (2015) argued that Loehle et al. (2015) 

underestimated the predictive performance of critical 

habitat maps because the field plots they used potentially 

biased the accuracy assessment toward older forests, and 

that they examined accuracy at finer scales than the model 

was intended to predict. Loehle and Irwin (2015) responded 

to Bell et al. (2015) and Dunk et al. (2015) by arguing that, 

although the habitat models average out at large spatial 

scales, errors at smaller scales may limit their utility for 

conservation. This debate underscores the importance of 

acknowledging the appropriate scale at which predictive 

distribution models can be used for conservation purposes. 

The debate also serves as another example highlighting 

the need to recognize and carefully evaluate how habitat is 

defined. The definition of habitat for spotted owls must now 

consider that forests that were once suitable for spotted owls 

are less suitable habitat if occupied by barred owls.

Conclusions and Management 

Considerations

Spotted owls are a resilient subspecies but are faced with 

significant challenges. Research and monitoring efforts 

over the past several decades have documented the population declines and risks to spotted owls despite measures 

to address their long-term sustainability. The framework, 

standards, and guidelines of the NWFP have been both 

critical and necessary for spotted owl conservation, and 

underlie species recovery plans. However, because of 

barred owls and continued forest perturbations outside of 

federal lands, the NWFP alone is not sufficient for spotted 

owl recovery. Additional measures beyond the Plan will 

be needed for long-term persistence of spotted owls. 

Suitable habitat continues to decline because of current 

and lingering effects of extensive forest disturbance, and 

the recent invasion of a formidable congeneric competitor 

has reduced the space available for spotted owl recovery. 

The need to provide habitat for spotted owls has been a 

critical component of conservation plans and was a major 

catalyst for developing the NWFP. It is now clear that 

barred owl presence reduces habitat suitabiliy for spotted 

owls, so species recovery will require protections for old 

forest and management actions focused on reducing the 

threat from barred owls. After only two decades, it is too 

early to evaluate if the Plan has been effective at improving the conservation status of spotted owls; however, the 

framework, standards, and guidelines of the NWFP have 

aided spotted owl conservation; if logging had continued at 

pre-NWFP levels, spotted owl populations certainly would 

have declined more rapidly over the past 20 years. Further, 

the NWFP has put federal lands on a trajectory for providing enough suitable forest for recovery of spotted owl 

populations over the next several decades. The effectiveness of LSRs established under the NWFP is linked to the 

frequency, severity, spatial extent, and type of disturbance, 

as well as how those disturbances are offset by recruitment 

of suitable forest, primarily through succession. Disturbance events can reduce the suitability of forests used by 

spotted owls for several decades by creating open canopy 

conditions and reducing structural complexity. Although 

disturbance rates have exceeded suitable forest-cover 
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recruitment rates during the first 20 years of the NWFP, 

recruitment will likely outpace losses if current timber 

harvests and wildfire occurrence remain constant. However, climate models suggest that wildfire occurrence may 

increase, causing significant reductions in cover for spotted 

owls, and that suitable forest cover for spotted owls will 

move northward and occur at higher elevations. Therefore, 

other reserves designated before development of the 

NWFP, such as parks and wilderness areas, may become 

increasingly important for spotted owl conservation. 

Several lines of compelling evidence indicate that interspecific competition between spotted owls and barred owls 

is causing accelerated population declines of spotted owls, 

despite widespread conservation of old forests under the 

NWFP. Competitive pressure from barred owls may negate 

the benefits of recruitment of suitable forest cover, because 

barred owls exclude spotted owls from sites that otherwise 

are suitable for spotted owls. It remains uncertain how, or if, 

spotted owls can coexist with barred owls. Although much 

research has been done on spotted owls, we identified many 

uncertainties in available information and have identified 

future research needs important for management of the 

subspecies. The long-term effects of barred owls and finescale partitioning of resources remain unknown, and studies 

are needed to identify resilient sites for spotted owls in the 

face of competitive interactions with barred owls, if they 

exist. Additionally, it remains unknown how, or if, spotted 

owls will respond to removals of barred owls from historical 

spotted owl territories. 

Abundance and distribution of primary prey species 

can influence home range size and forest selection by spotted owls. But it remains unknown how spatially and temporally fluctuating prey populations influence the survival and 

reproduction of spotted owls. Studies are needed to quantify 

relationships between interannual fluctuations in prey 

abundance and long-term demography of spotted owls. The 

short- and long-term effects of silvicultural treatments and 

wildfire on spotted owl occupancy, forest dynamics, and 

prey remain unclear. The optimization of forest restoration 

and conservation of spotted owls will require more knowledge about the conditions under which restoration activities 

can benefit spotted owls in the long term without significant 

detrimental impact in the short term.

Management Considerations

Forest management and barred owls—

Wiens et al. (2014) found that adult survival of spotted owls 

and barred owls was higher in home ranges with greater 

amounts of conifer forest dominated by trees age 120 years 

or older. Dietary studies also showed that barred owl diet 

is broader than spotted owls, but both owl species relied 

on similar prey associated with older forest types (e.g., 

northern flying squirrels and red tree voles). These findings 

have important implications for land managers because 

they suggest that (1) conservation of old forest under the 

NWFP not only promotes survival of spotted owls, but also 

survival of barred owls; and (2) availability of old forests 

(and associated food resources) is a key limiting factor in 

the competitive relationship between the two owl species 

(Wiens et al. 2014). As barred owls continue to increase 

in number, it has become clear that conservation of the 

spotted owl and its forest cover types need to be extended 

from ameliorating the effects of old-forest loss and fragmentation to accounting for the impacts of a widespread 

invasive competitor as well. Although spotted owls are 

known to use recently thinned stands (e.g., Irwin et al. 

2015), it remains unclear how such silvicultural treatments 

can affect the fitness of spotted owls in the long term 

or how barred owls may respond to those management 

actions. Those silvicultural treatments with high disturbance likely increase long-term extinction rates of spotted 

owls by reducing forest complexity and thus suitability for 

spotted owls but not necessarily for barred owls (Dugger et 

al. 2016, Singleton 2015, Singleton et al. 2010, Sovern et al. 

2014, Wiens et al. 2014). 

Barred owl densities may now be high enough across 

the range of the spotted owl that, despite the continued 

management and conservation of suitable forest cover types 

under the NWFP, the spotted owl population will continue 

to decline without intervention to reduce barred owl populations (Dugger et al. 2016). Recommendations to conduct 

experimental removal of barred owls to benefit spotted owls 

have been criticized as being too difficult to accomplish 

owing to the effort and cost required to maintain sufficiently low numbers of invasive barred owls (Livezey 2010, 

Rosenberg et al. 2012). Nonetheless, experimental removal 

of barred owls on one study area in California suggests that 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

removal of barred owls may have positive, short-term effects 

on population trends of spotted owls (Diller et al. 2016, 

Dugger et al. 2016). In 2013, the Fish and Wildlife Service 

decided to expand removal experiments to additional sites in 

California, Oregon, and Washington to determine if similar 

results can be obtained in areas with different forest conditions and densities of barred owls (USDI 2013, USFWS 

2013). Those experiments will yield information about how 

spotted owls respond, and will convey the economic and 

logistic feasibility of removal efforts as potential management actions. Such information will be useful in projecting 

possible long-term consequences and benefits of an active 

management program for barred owls in the future. 

Current evidence suggests that a combination of habitat 

protection and active management of barred owls are the 

two highest priorities for stabilizing declining trends in populations of spotted owls. A recent analysis casts doubt on 

the likely effectiveness of barred owl removals for spotted 

owl conservation (Bodine and Capaldi 2017). Experimental 

culling of barred owls will provide information to validate 

those models and about how, or if, their populations can be 

controlled at scales sufficient to promote recovery of spotted 

owls. However, detailed studies of habitat associations and 

resource use by barred owls have been conducted in only 

a few limited areas within the range of the spotted owl. 

More detailed studies in other areas will better enable an 

understanding of how specific tree species, stand densities, 

or physiographic conditions are negatively associated with 

barred owls but not spotted owls.

Wildfire and active management—

Disturbance processes that increase forest or landscape 

heterogeneity (e.g., wildfire, management activities) can 

benefit spotted owls as long as the required forest structural 

conditions are available for foraging, nesting, and roosting 

activities. Processes that substantially simplify stand structure or landscapes often have negative impacts on the suitability of forest for spotted owls. Our basic understanding 

of forest structural conditions used by spotted owls has not 

substantially changed over the past 20 years, but there has 

been a growing recognition of the contribution of diverse 

forest conditions to broader ecosystem function and species 

diversity in conifer forests of the Pacific Northwest. This is 

especially true in historically moderate- and high-frequency 

fire regime landscapes where fire suppression and forest 

management have greatly reduced fire and altered forest 

structure and composition at stand and landscape scales 

(chapter 3). For example, nonconiferous vegetation, including shrubs and broad-leaved trees, makes an important 

contribution to the diversity of forest landscapes. Therefore, 

allowing shrubs and hardwood trees to develop and persist 

in early-seral stands, and curtailing vegetation control, will 

benefit many wildlife species associated with nonconiferous 

vegetation (Hagar 2007), including some spotted owl prey 

species (Diller et al. 2012). Additionally, diversity and configuration of different forest types are important for spotted 

owls at stand, home range, and landscape scales (Franklin 

et al. 2000). The function and diversity of an ecosystem 

is enhanced by the presence of high-quality early-seral 

patches (i.e., a mix of nonforest and forest) because they 

have high species and structural diversity (Swanson et al. 

2011). These early-seral ecosystems can be created using 

low-intensity approaches for regeneration, combined with 

retention of biological legacies to promote the development 

of structurally diverse closed-canopy forest over time 

(Franklin and Johnson 2012). Indeed, under normal conditions, natural disturbances frequently result in patches of 

high-quality early-seral ecosystems, provided that intensive 

salvage and replanting does not occur after the disturbance 

(Swanson et al. 2011). 

Disturbances have different impacts on spotted owls 

depending on the scale under consideration. A hypothesis 

that has emerged from recent research is that disturbance 

processes (e.g., low- and mixed-severity wildfire, light to 

moderate thinning) that increase stand or landscape heterogeneity can have long-term benefits for spotted owls, as 

long as enough suitable forest cover for nesting and roosting 

remain within the territory. Conversely, disturbances that 

substantially simplify stands or landscapes often have 

long-lasting negative impacts on spotted owls and their 

habitat. Finally, we emphasize the importance of conserving 

sites currently occupied by spotted owls as well as those that 

are known to have been historically occupied by the subspecies. Many sites, for example, have been abandoned as a 

result of disturbance to suitable forest cover or displacement 

by barred owls, but maintain structure suitable for nesting 

and roosting. Those remaining spotted owls and sites likely 
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represent unique behavioral or forest characteristics that 

may not yet be fully recognized, thus they are an important 

research need. Conserving the unique forest structural 

conditions of those few sites that remain, particularly in the 

northern portion of the geographic range, will likely have a 

positive benefit for the long-term persistence of spotted owls. 

Prognosis for the future—

In the 2011 revised recovery plan for spotted owls, the Fish 

and Wildlife Service’s modeling team used the HexSim 

modeling program (Schumaker 2008) to simulate population-level responses to various conservation strategies 

and other threats (USFWS 2011b). They developed models 

based on demographic data (Forsman et al. 2011), dispersal 

information (Forsman et al. 2002, Thomas et al. 1990), and 

home range size (Carey et al. 1990; Forsman et al. 1984, 

2005; Glenn et al. 2004; Hamer et al. 2007). Objectives 

of the modeling effort were to (1) evaluate if future viable 

populations of spotted owls were likely given conditions 

at the time (demographic rates, LSR network, amount of 

suitable forest cover, barred owls); (2) estimate population 

viability under different conservation networks of suitable 

forest cover; and (3) quantify the effect of forest cover 

and barred owl management on recovery goals for spotted 

owls (USFWS 2011b). The modeling results suggested that 

availability of suitable forest cover was critical for territory 

acquisition and sustained occupancy by spotted owls. 

Population viability models suggest that barred owls reduce 

spotted owl survival and act to depress populations to about 

half of potential population size without barred owls (fig. 

4-9). Simulations did not include the barred owl impact on 

spotted owl reproduction, forest selection, site fidelity, or 

detection probability, and were based upon early rates of 

population growth. More recent population change estimates 

(Dugger et al. 2016) indicate a further declining growth 

Figure 4-9—HexSim model runs with five replicates each for without barred owl impacts and with barred owl 

impacts for the spotted owl’s geographic range in the United States. The apparent within-year variation that 

appears in the figure is a function of an “even-odd” year effect on reproduction (USFWS 2011b). The first 30 years 

of the simulation was a “burn-in” period, which provided for the simulated population to distribute according to 

available resources and develop an age structure determined by demographic processes. Barred owl effects were 

not included during the “burn-in” period and were introduced starting at year 30 (USFWS 2011b).

7,000

6,000

5,000

4,000

3,000

2,000

1,000

0

Number of owls

25 50 75 100 125 150 175 200 225 250

Time step (year)

Without barred owls

With barred owls
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

rate, suggesting that USFWS (2011b) projected estimates are 

more optimistic than what is likely to be observed in spotted 

owl populations. These studies provide further evidence that 

the framework, standards, and guidelines of the NWFP are 

critical components to spotted owl recovery plans, but the 

impacts of barred owls will likely need to be controlled if 

spotted owl species recovery is to be successful. 

Schumaker et al. (2014) used the HexSim model originally developed by the Fish and Wildlife Service (USFWS 

2011b) to simulate and quantify source-sink dynamics and 

landscape connectivity throughout the range of the spotted 

owl. Their results indicated that populations are likely to 

decline in most regions, but that southern Oregon and northern California may serve as source populations. Marcot 

et al. (2013) also used the HexSim model to evaluate how 

size and spacing of suitable forest cover types for spotted 

owls affected simulated population size and persistence. 

Their results indicated that long-term occupancy rates 

were significantly higher with suitable forest patches large 

enough to support 25 spotted owl pairs or more, with less 

than 9.3 mi (15 km) spacing between patches, and with 

overall landscapes of at least 35 to 40 percent suitable 

forest cover types for nesting and roosting. In a sensitivity 

analysis, Marcot et al. (2015) determined that spotted owl 

response variables in the HexSim model were most sensitive 

to the availability of highly suitable forest cover for nesting 

and roosting. All these studies used static habitat maps that 

did not incorporate climate change or wildfire impacts on 

spotted owls. Only the USFWS (2011b) model incorporated 

effects of barred owls.

Spotted owl populations have continued to decline 

under the NWFP, but because of slowed timber harvest 

on federal lands since the late 1980s, forests throughout 

most of the range of the spotted owl are on a trajectory—

through succession—to develop suitable forest characteristics for spotted owls in coming decades. When the NWFP 

was adopted, spotted owl populations were expected to 

continue declining for up to 50 years because of lingering 

impacts of previous losses of suitable forest cover, yet the 

magnitude and characteristics of barred owl impacts were 

unknown and unexpected at that time. Per assumptions of 

the NWFP, we are unable, after only two decades, to use 

stable or increasing populations (i.e., improved conservation status) of spotted owls as the success criterion for the 

NWFP. However, if the success criterion is forests capable 

of supporting interconnected populations of spotted owls 

in the absence of barred owls, then the implementation of 

the framework, standards, and guidelines of the NWFP 

has put federal lands on a trajectory for success, despite 

recent losses of suitable forest cover to wildfire. In the 

Pacific Northwest, forest succession from early-seral 

to climax forest is a slow process, which is in part the 

reasoning for the NWFP to be a 100-year plan intended to 

span several human generations (USDA and USDI 1994). 

Further, conservation and management of spotted owls 

rests critically on continued implementation of the protections afforded by the NWFP and the Endangered Species 

Act (Noon and Blakesley 2006). It also rests on improving 

our understanding of how to minimize impacts of barred 

owls, and on fine-tuning our ability to retain needed 

forest structure while also increasing resiliency of forests 

through strategic management.

U.S. and Metric Equivalents

When you have: Multiply by: To get:

Inches 2.54 Centimeters

Meters (m) 3.28 Feet

Hectares (ha) 2.47 Acres

Acknowledgments

We are indebted to the many biologists and funding sources 

dedicated to studying northern spotted owl populations 

and factors affecting this threatened subspecies. Heather 

Roberts (Oregon State University, Department of Forest 

Ecosystems and Society) assisted with the plot analysis 

presented in the inset of this chapter. Author salaries were 

supported by the USDA Forest Service Pacific Northwest 

Research Station (D. Lesmeister and P. Singleton) and 

Pacific Northwest Region (R. Davis), and by the USDI 

Geological Survey, Forest and Rangeland Ecosystem 

Science Center (J.D. Wiens). We thank E. Forsman, E. 

Glenn, members of the public, and anonymous reviewers 

for providing comments and suggested edits that greatly 

improved this chapter. 
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299
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Marbled murrelet.

Photo by Kim Nelson, Oregon State University.
301

Chapter 5: Marbled Murrelet

Martin G. Raphael, Gary A. Falxa, and Alan E. Burger1

Introduction 

In this chapter, we describe expectations of the Northwest 

Forest Plan (NWFP, or Plan) and review recent science on 

the ecology and status of the marbled murrelet (Brachyramphus marmoratus), with an emphasis on the portion 

of the species’ range that falls within the Plan area. The 

conservation strategy embodied in the NWFP evolved from 

designation and protection of a large number of relatively 

small management areas to an approach based primarily 

on the designation of fewer large areas, each designed to 

conserve functioning late-successional and old-growth 

ecosystems. These were intended to support multiple 

pairs of northern spotted owls (Strix occidentalis caurina) 

and murrelets, and to conserve habitat for other species 

associated with older forests. 

The marbled murrelet is a small seabird of the family 

Alcidae (fig. 5-1) whose summer distribution along the 

Pacific Coast of North America extends from the Aleutian 

Islands of Alaska to Santa Cruz, California (fig. 5-2). It 

forages primarily on small fish and krill in the nearshore 

(0 to 2 mi [0 to 3 km]) marine environment. Unlike other 

alcids, which nest in dense colonies on the ground or in 

burrows at the marine-terrestrial interface, murrelets nest 

in more dispersed locations up to 55 mi (89 km) inland. In 

the southern portion of the range, including the Plan area 

1 Martin G. Raphael is a research wildlife biologist (retired), 

U.S. Department of Agriculture, Forest Service, Pacific Northwest 

Research Station, 3625 93rd Ave. SW, Olympia, WA 98512; Gary 

A. Falxa is a wildlife biologist (retired), U.S. Fish and Wildlife Service, 1615 Swanson Lane, Eureka, CA 95503; Alan E. Burger is an 

adjunct professor, University of Victoria, Department of Biology, 

and a wildlife consultant, P.O. Box 2539, Merritt, BC V1K 1B8.

Figure 5-1—The marbled murrelet is a small seabird of the family Alcidae.

Martin Raphael
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GENERAL TECHNICAL REPORT PNW-GTR-966

and the area emphasized in this chapter, murrelets typically 

nest in large coniferous trees in forested areas containing 

characteristics of older forests. Throughout the forested portion of the species’ range, murrelets typically nest in areas 

containing characteristics of older forests (Baker et al. 2006; 

Binford et al. 1975; Hamer and Cummins 1991; Hamer and 

Nelson 1995; Hamer et al. 1994; Hébert and Golightly 2006; 

Quinlan and Hughes 1990; Ralph et al. 1995a; Singer et al. 

1991, 1992; Wilk et al. 2016). The marbled murrelet population in Washington, Oregon, and California nests in most 

of the major types of coniferous forests (Hamer and Nelson 

1995) in the western portions of these states, wherever older 

forests remain inland of the coast at elevations primarily 

below the extent of the true fir zone, generally <4,000 ft 

(1220 m) (table 5-1). Although murrelet nesting habitat 

characteristics may differ throughout the range of the 

species, some general habitat attributes are characteristic 

throughout its listed range, including the presence of nesting 

platforms, adequate canopy cover over the nest, larger patch 

size of mature forest, and being within commuting distance 

to the marine environment (Binford et al. 1975, Hamer and 

Nelson 1995, Nelson 1997, McShane et al. 2004, Ralph et 

al. 1995b). Because murrelets do not construct nests, they 

depend on the availability of platforms, typically tree limbs 

with a moss or other thick substrate, such as piles of needles 

collected on limbs near a tree bole, sufficiently large for 

laying their single egg and raising a nestling (Nelson 1997, 

Ralph et al. 1995). 

Figure 5-2—Range of the marbled murrelet in North America. 

Map by Terry Sohl from NatureServe data.

Table 5-1—Known inland limits of marbled murrelet nests and occupied sites

Inland distance

State/province Nesta Occupied 

site Sources

- - - Miles - - - 

Alaska 33 Nelson et al. 2010, Whitworth et al. 2000 

British Columbia 39 41 Jones et al. 2006, Lougheed 1999, Nelson et al. 2010, Ryder et al. 2012 

Washington 55 55 D. Lynch, personal communicationb

; Ritchie and Rodrick 2002

Oregon 32 47 Alegria et al. 2002; Dillingham et al. 1995; E. Gaynor, personal communicationc

; 

Witt 1998a, 1998b 

California 24 24 S. Chinnici, personal communicationd

; A. Transou, personal communicatione

Note: see table on page 338 for metric equivalents.

a

 Includes grounded fledglings and eggshell fragments.

b

 D. Lynch. Personal communication. Fish and wildlife biologist, Washington Department of Fish and Wildlife, 111 Washington Street SE, Olympia, WA 98501.

c

 E. Gainer. Personal communication. Wildlife biologist, U.S. Department of the Interior, Bureau of Land Management, 777 NW Garden Valley Blvd., 

Roseburg, OR 97471.

d

 S. Chinnici. Personal communication. Forest science manager, P.O. Box 712, Humboldt Redwood Company, Scotia, CA 95565.

e

 A. Transou. Personal communication. Environmental scientist, California Department of Parks and Recreation, North Coast Redwoods District, P.O. 

Box 2006, Eureka, CA 95502; 707-445-6547; atransou@parks.ca.gov. 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Individual tree attributes that provide conditions suitable 

for nesting (i.e., provide a nesting platform) include large 

branches (ranging from 4 to 32 inches (10 to 81 cm) diameter, 

with an average of 13 inches (33 cm) in Washington, Oregon, 

and California) or forked branches; deformities (e.g., broken 

tops); dwarf mistletoe infections; witches’ brooms; and 

growth of moss or other structures large enough to provide a 

platform for a nesting adult murrelet (Hamer and Cummins 

1991; Hamer and Nelson 1995; Singer et al. 1991, 1992). 

These nesting platforms (fig. 5-3) are generally 

located ≥33 ft (10 m) above ground (reviewed in Burger 

2002 and McShane et al. 2004). These structures are 

typically found in old-growth and mature forests, but may 

be found in a variety of forest types, including younger 

forests containing remnant large trees. Since 1996, 

research has confirmed that the presence of platforms is 

considered the most important characteristic of murrelet 

nesting habitat (Burger 2002, Huff et al. 2006, McShane 

et al. 2004). Platform presence is more important than the 

size of the nest tree because tree size alone may not be a 

good indicator of the presence and abundance of platforms 

(Evans Mack et al. 2003). Tree diameter and height can be 

positively correlated with the size and abundance of platforms, but the relationship may change depending on the 

variety of tree species and forest types that murrelets use 

for nesting (Burger et al. 2010, Huff et al. 2006, Raphael 

et al. 2011). Overall, nest trees in Washington, Oregon, 

and northern California have been greater than 19 inches 

(48 cm) diameter at breast height (d.b.h.) and greater than 

98 ft (30 m) tall (Hamer and Meekins 1999, Hamer and 

Nelson 1995, Nelson and Wilson 2002). Northwestern 

forests and trees typically require 200 to 250 years to 

attain the attributes necessary to support murrelet nesting, 

although characteristics of nesting habitat sometimes 

develop in younger western hemlock (Tsuga heterophylla) 

forests with dwarf mistletoe.

Marbled murrelets are reported to nest disproportionately on lower slopes and near streams. The recovery 

plan for the murrelet (USFWS 1997) states, “With respect 

to slope, eighty percent of nests in the Pacific Northwest 

were located on the lower one-third or middle one-third 

of the slope.” Hamer and Nelson (1995) showed the mean 

distance to streams from murrelet nests in the Pacific 

Northwest to be 159 m (509 ft). In southern California, 

Baker et al. (2006) found that murrelet nest sites were 

located closer to streams, and were located lower on 

slopes than random sites, based on analysis of variance 

models. Baker et al. (2006) found that nest sites were 

much closer to streams than would be expected based on 

randomly available sites within old-growth forests. Nest 

sites may have been located near streams because these 

Figure 5-3—Nesting platforms usually include large branches and sites afforded murrelets better access from at-sea flyways. other structures large enough to provide a platform for a nesting 

adult murrelet.

Nick Hatch
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GENERAL TECHNICAL REPORT PNW-GTR-966

Other studies have also found proximity to streams or 

other openings to be important for murrelet nesting in 

other regions as well (Hamer and Nelson 1995, Meyer 

et al. 2004, Zharikov et al. 2006). In British Columbia, 

Rodway and Regehr (2002) found that forests bordering 

major stream channels provided high-quality nest habitat 

for murrelets, with large trees, high epiphyte cover, and 

many potential nest platforms. 

Murrelets travel up to 55 mi (89 km) inland to reach 

suitable habitat in the northern part of their range in the 

Pacific Northwest; inland distances narrow in the southern 

portions of the range (table 5-1). Because murrelets depend 

on marine conditions for foraging and resting, and on 

forests for nesting, both marine and forest conditions could 

limit murrelet numbers. Population declines attributed to 

loss of mature and old-growth forest from harvesting, low 

recruitment of young, and mortality at sea, led this species 

to be federally listed as threatened in Washington, Oregon, 

and California in 1992 (USFWS 1997), and listed as threatened in British Columbia (Rodway 1990). The murrelet’s 

association with late-successional and old-growth forests 

and its listed status made conservation of the murrelet an 

explicit goal in the design of the NWFP. 

The NWFP included several elements of protection for 

murrelet nesting habitat. The Plan’s system of reserves was 

not designed, as it was for the northern spotted owl, with 

specific goals for the number and spacing of clusters of murrelets. Rather, the system of congressionally reserved lands 

and late-successional reserves was designed to encompass 

a high proportion of murrelet nesting habitat thought to 

exist on federal lands. In addition to the reserve system, the 

NWFP requires murrelet surveys to be conducted before 

harvest on any other federal lands in the murrelet’s range. If 

a survey shows likely nesting, then all contiguous existing and recruitment habitat (defined as stands that could 

become nesting habitat within 25 years) within a 0.5-mi (0.8 

km) radius is protected. These occupied sites become small 

reserves, denoted as LSR3, and are managed to retain and 

restore nesting habitat.

Guiding Questions

The mission statement for the Forest Ecosystem Management Assessment Team (FEMAT) directed the team to take 

an ecosystem approach to forest management and particularly to address maintaining and restoring biodiversity on 

federal forests within the range of the northern spotted owl. 

In addressing biological diversity, the team was directed 

to develop alternatives that met, among other things, the 

objective of maintaining or restoring habitat conditions for 

the murrelet that would provide for viability of the species 

(FEMAT 1993: iv). Now, 22 years after the NWFP was 

initiated, national forests in the Plan area are preparing to 

revise their forest plans. Accordingly, U.S. Forest Service 

managers have asked how the NWFP has been functioning 

to support the murrelet and what new science is relevant 

to murrelet conservation and management. Managers were 

polled to develop questions relating to the murrelet (as well 

as other NWFP issues), and this chapter aims to synthesize 

relevant science related to these questions:

• Are murrelets maintaining viable populations under 

current NWFP management? 

• Is forest management under the NWFP providing 

nesting habitat for murrelets as planned?

• What is the latest science surrounding the effects 

of various treatments (silvicultural and fuels) and 

wildfire on late-successional, old-growth forests and 

plantations, and what are the effects on murrelets? 

• Does the murrelet use these treated forests after harvest? If so, how? Are there ways to modify harvest 

to benefit murrelets? 

• How do these treated habitats compare to 

untreated habitat in terms of habitat use and reproductive success? 

• How have at-sea conditions affected nearby forest 

use by the murrelet? 

To address these questions, we conducted a thorough 

literature review, guided by keywords included in the 

questions, and we emphasized references pertaining to 

murrelets in the Plan area. We excluded gray literature and 

other unpublished work. We considered additional literature 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

suggested by public comments. As will be apparent in the 

text, we found little literature bearing on questions 3, 4, and 

5, as they pertain to responses of murrelets to silviculture. 

We direct readers to Spies et al. (this volume) for a summary 

of how younger forests respond to silvicultural treatments 

that might influence murrelet nesting habitat. 

Key Findings 

NWFP Expectations

The stated objective of the NWFP is to maintain and restore 

nesting habitat conditions that would provide for viability 

of murrelet populations, well-distributed along their current 

range on federal lands (FEMAT 1993: iv). The expectation 

was that the Plan “...would eventually provide substantially 

more suitable nesting habitat for murrelets than currently 

(in 1994) exists on federal lands” (USDA and USDI 1994a). 

FEMAT used an expert panel to assess the likelihood that 

nesting habitat on federal lands would support stationary and 

well-distributed populations of the murrelets. Following the 

methods described in FEMAT (1993), the murrelet expert 

panel assigned an 80 percent likelihood that nesting habitat 

would be of sufficient quality, distribution, and abundance 

to allow the murrelet population to stabilize, well distributed 

across federal lands over the next 100 years (Outcome A) 

under Option 9, the preferred alternative that was eventually adopted (with modifications) as the NWFP. The panel 

assigned a 20 percent likelihood for Outcome B, under which 

nesting habitat would be sufficient to allow the murrelet population to stabilize but with significant gaps in the historical 

distribution that could cause some limitation in interactions 

among local populations. The panel assigned no likelihood 

of Outcomes C or D. Thus, the panel’s assessment was that 

the likelihood was high that nesting habitat conditions on 

federal lands would allow the murrelet population to stabilize 

and be well distributed throughout its range (FEMAT 1993). 

In recognition of the major influence of marine conditions 

on population viability, however, including mortality from 

oil spills and gill netting, and considering the potentially 

important role of nonfederal lands, the murrelet panel 

assigned a second set of ratings that considered the cumulative effects of all major factors. The murrelet panel concluded 

that the likelihood that the murrelet population on federal 

lands would be stationary and well-distributed was between 

50 and 75 percent. The higher rating was meant to indicate 

the degree of protection conferred by nesting habitat conditions on federal lands, assuming that all other factors were 

not limiting; the lower rating from the cumulative effects 

analysis was an attempt to indicate the greater uncertainty in 

murrelet persistence, given the importance of other factors 

beyond federal nesting habitat. 

Neither the assessment team nor final supplemental 

environmental impact statement nor subsequent monitoring 

plan for the murrelet (Madsen et al. 1999) provided quantitative descriptions of expected murrelet population trends 

or nesting habitat trends over time that now could be used to 

assess NWFP performance since its implementation. There 

are, however, some more qualitative descriptions or assumptions from the period around the start of the assessment 

team and the record of decision:

• The amount of murrelet nesting habitat had declined 

over the previous 50 years, primarily because of timber harvesting (Perry 1995, USFWS 1997).

• Murrelet populations are likely to have declined as 

well, largely in response to loss of nesting habitat 

(Ralph et al. 1995a).

• Demographic projection models estimated at the 

time the NWFP was initiated suggested a population 

decline of 4 to 7 percent per year from 1990 to 1995 

(Beissinger 1995).

• Because murrelets have naturally low reproductive 

rates, population recovery will be slow, on the order 

of a maximum of 3 percent per year (USFWS 1997).

• No destruction of nesting habitat surrounding active 

murrelet nesting sites will be knowingly done on 

federal lands.

• Catastrophic and stochastic events that decrease the 

quality or quantity of nesting habitat would affect 

nesting habitat at unknown rates. 

• Over the long term, the amount of nesting habitat 

will increase in reserves as unsuitable forest matures.

• Late-successional reserves will provide large contiguous blocks of nesting habitat with increased interior 

(180 ft [55 m] or more from edge) nesting habitat.
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GENERAL TECHNICAL REPORT PNW-GTR-966

• Rates of nest depredation would decrease as the 

amount of interior nesting habitat increases in reserves.

• In the short term (less than 50 years), the availability 

of nesting habitat may remain stable or decline from 

losses from fire and other natural disturbances.

• The rate of increase in the amount of nesting habitat 

will be slow because trees do not develop structures 

suitable to support nests until they are large and old, 

often 150 or more years (USDA and USDI 1994a; 

USFWS 1997).

• Nesting habitat management on nonfederal lands 

will affect viability of murrelets on federal lands.

• Physical and biological processes in the marine environment, which operate at multiple temporal and 

spatial scales, also affect short- and long-term population trends of murrelets, independent of nesting 

habitat quantity or quality.

McShane et al. (2004) developed a population model to 

predict population change in each of five conservation zones 

comprising the Plan area (fig. 5-4). Their model, which used 

annual adult survival estimates obtained from detailed 

mark-recapture studies in British Columbia (the only such 

data then available) and fecundity estimates from ratios of 

juveniles to adults at sea or from mark-recapture studies, 

predicted annual rates of decline varying from 3 to 5 

percent per year over the first 20 years of their simulations 

in murrelet conservation zones 1 through 5.2 Rates of 

decline were generally greater going from north (zones 1 

and 2) to south (zone 5). These predictions are in line with 

those of Beissinger (1995), using models based mostly on 

comparative demographic data from other alcid species. 

These models do not directly account for the amount of 

nesting habitat, thus model projections do not respond to 

expected habitat trends.

2

 These zones are defined in the marbled murrelet recovery plan 

(USFWS 1997): Conservation zone 1 is Puget Sound and the Strait 

of Juan de Fuca in Washington; zone 2 is the outer coast of Washington to the Columbia River; zone 3 is Oregon from the Columbia 

south to North Bend (Coos Bay); zone 4 is North Bend south to 

Shelter Cove, California; zone 5 is Shelter Cove south to the mouth 

of San Francisco Bay (see fig. 5-2). Zone 6, from the mouth of 

San Francisco Bay south to Point Sur, California, is outside of the 

Northwest Forest Plan area.

NWFP Monitoring Results for Marbled Murrelets

Population size and trends—

A specific conservation goal of the plan is to stabilize 

and increase murrelet populations by maintaining and 

increasing nesting habitat. As described below, population 

monitoring results to date indicate that the plan goal of 

stabilizing and increasing murrelet populations has not yet 

been achieved throughout the Plan area, because while in 

some areas the population may have stabilized, they have 

not increased substantially. Murrelet populations were 

thought to be declining at the start of the Plan, with loss of 

more than 80 percent of nesting habitat being the central 

cause for declines and for murrelets being listed as federally 

threatened (USFWS 1997). Declines were expected to 

continue for a period (e.g., Raphael 2006), until nesting 

habitat sufficiently recovers from previous losses to lead to 

increased fecundity, and populations stabilize and increase 

(USFWS 1997). The Plan goal of increasing populations 

recognizes the large historical population declines (Peery et 

al. 2010, USFWS 1997), and the conservation value of larger 

populations than were present in 1994.

To evaluate murrelet population status and trends under 

the Plan, the murrelet effectiveness monitoring program 

designed a coordinated sampling protocol (Madsen et al. 

1999, Raphael et al. 2007) and obtained annual population 

estimates starting in 2000 by monitoring murrelet populations 

in nearshore marine waters associated with the Plan area, in 

Washington, Oregon, and northern California (fig. 5-4). The 

population monitoring uses boat-based transects and distance 

estimation methods in those coastal waters, which are divided 

into five geographic subareas corresponding to conservation 

zones established in the U.S. Fish and Wildlife Service’s 

recovery plan for the murrelet (fig. 5-4). The monitoring program estimated population size and trend for each conservation zone, for each state, and for all zones combined. Through 

2013, the entire Plan area was surveyed annually; starting in 

2014 a reduced-sampling design was instituted because of 

funding constraints, in which conservation zones 1 through 

4 are sampled every other year, and zone 5 every fourth year. 

Details about the sampling and data analysis methods used by 

the population monitoring program are described elsewhere 

(Falxa et al. 2016, Raphael et al. 2007).
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

California

Oregon

Washington

Oregon

Pacific

Oce

a

n

Zone 1

Zone 2

Zone 5

Zone 4

Zone 3

Federal lands

Marbled murrelet conservation zones

Reserved

Nonreserved

Northwest Forest Plan inland zones

Zones 1–5

0 50 100 150 200 Miles

0 80 160 240 320 Kilometers

o

Figure 5-4—Range of the marbled murrelet with boundaries of conservation zones and locations of federal reserves, 

other federal lands, and nonfederal lands. Also shown are boundaries of the “inland zones” designated by the Northwest 

Forest Plan; see “Research Needs, Uncertainties, Information Gaps, and Limitations” for a description of these zones.
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GENERAL TECHNICAL REPORT PNW-GTR-966

The 20-year murrelet status and trends report provided 

estimates through 2013 (Falxa et al. 2016); population monitoring results from 2014 and 2015 have since become available in annual reports (Falxa et al. 2015, Lynch et al. 2016). 

At the conservation-zone scale, the most recent population 

estimate shows few murrelets remaining in conservation 

zone 5 (San Francisco Bay north to Shelter Cove, California; 

estimate: 71 murrelets, 95 percent confidence interval: 5 to 

118) (Lynch et al. 2016); this is consistent with estimates 

since 2000. Considerably more murrelets remain in the other 

four conservation zones within the NWFP area, with murrelet numbers, expressed as an average of annual estimates 

over the the past 4 years with sampling (Lynch et al. 2016) 

as follows: about 7,600 murrelets in conservation zone 1 (the 

Strait of Juan de Fuca, San Juan Islands, and Puget Sound 

in Washington; for 2012–2015); about 2,000 birds in conservation zone 2 (the outer coast of Washington; 2012–2015); 

about 7,600 murrelets in conservation zone 3 (from Coos 

Bay north to the Columbia River, Oregon; 2011–2014); and 

about 6,600 birds in conservation zone 4 (from Shelter Cove, 

California, north to Coos Bay, Oregon; 2012–2015). The 

use of averages accounts for some of the annual variation 

in population estimates. Single-year estimates vary among 

years and tend to have relatively large confidence intervals. 

For example, the most recent estimate for conservation zone 

2 (3,204 murrelets in 2015) is higher than the 4-year average, 

but with a 95 percent confidence interval (1,883 to 5,609) 

(Lynch et al. 2016) that includes that average. All annual 

estimates at the conservation zone and other scales are found 

in recent reports from the NWFP’s murrelet effectiveness 

monitoring program (Falxa et al. 2016, Lynch et al. 2016).

Estimated density of murrelets on the surveyed waters 

(generally within 2 to 3 mi [3 to 5 km] of shore, depending on conservation zone) (Raphael et al. 2007) ranged 

from approximately 0.1 murrelets per square kilometer in 

conservation zone 5 to 7.5 murrelets per square kilometer in 

conservation zone 4 in 2015. Annual population estimates 

for the entire Plan area ranged from about 16,600 to 22,800 

murrelets during the 15-year period (fig. 5-5), and averaged 

about 21,000 birds over the past 4 years (2011–2014); the 

most recent estimate for the Plan area is 21,300 birds for 

2014 (95 percent confidence interval: 17,500 to 25,100) 

(Lynch et al. 2016). The confidence intervals associated 

with population estimates reflect the difficulties in sampling 

such a mobile, patchily distributed, and relatively rare 

species over a large area of ocean waters. Although this 

sampling error decreases the power to detect population 

trends, the trend estimation accounts for sampling error.

The estimates from population monitoring form the 

basis for evaluating population trends since 2000. The 

monitoring program evaluated linear trends from 2000 

to 2015 at multiple scales (Lynch et al. 2016), and found 

evidence for a declining trend in Washington, no clear 

trend in Oregon, and evidence for an increasing trend in the 

California portion of the Plan area (fig. 5-6). In Washington 

(fig. 5-7), there was strong evidence of a population decline 

in conservation zone 1 (a 5.3 percent annual decline, 95 

percent confidence interval: -8.4 to -2.0) (Lynch et al. 2016), 

and a 4.4 percent decline per year for Washington state 

(conservation zones 1 and 2 combined; 95 percent confidence interval: -6.8 to -1.9) (Lynch et al. 2016). In conservation zone 2, where past analyses found a declining trend 

(Falxa et al. 2016), the most recent trend analysis, with 2014 

and 2015 data included, indicates that a negative trend may 

continue in conservation zone 2, but the upper confidence 

interval now overlaps zero (fig. 5-7), thus the trend for this 

zone is uncertain (95 percent confidence interval: -7.6 to 

2.3) (Lynch et al. 2016). In conservation zones 3 and 5, the 

most recent data provide no evidence of a trend (confidence 

intervals broadly overlap zero) (Falxa et al. 2016, Lynch et 

al. 2016); for an earlier period, Strong (2003) described a 

decline for central Oregon, which includes part of zone 3. 

In zone 4, the trend estimate was positive (3.0 percent per 

year), and with the addition of 2015 survey data the trend 

estimate’s 95 percent confidence interval does not include 

zero (0.4 to 5.6; fig. 5-7), evidence for a positive trend on 

average for the 2000 to 2015 period for this zone (Lynch et 

al. 2016). At the state scale for Oregon and California, which 

combines conservation zones and portions of conservation 

zones, there was no evidence of a trend in Oregon (fig. 5-6). 

For California, as for zone 4, the trend estimate was positive 

for 2000 to 2015 (3.8 percent per year) and the 95 percent 

confidence interval for that estimate (0.9 to 6.8) lies entirely 

above zero, suggesting an increasing population (fig. 5-6).
309

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

For the entire Plan area, the estimated rate of population change for the 2001 to 2014 period was negative (-0.7 

percent per year), but the confidence interval for the estimate (-2.3 to 0.8) broadly overlapped zero and there was 

no clear evidence for a trend (fig. 5-7). Additional years of 

monitoring should increase the power to detect an ongoing 

trend, such as where the trend is slight and power to detect 

low, but population trajectories can also change with time, 

which adds variability and difficulty in describing trends. 

For example, the magnitude and strength of evidence for a 

NWFP-wide population decline have decreased relative to 

a previous assessment for the 2001 to 2010 period (Miller 

et al. 2012). This difference may be driven by a variety of 

factors, most notable being the higher population estimates for 2011 through 2014 compared to the previous several years (fig. 5-5), which reduced the slope of the trend 

and increased variability (Falxa et al. 2016, Lynch et al. 

2016). In 2011 and 2012, estimates of murrelet population 

size increased in all conservation zones except conservation zone 2, compared to estimates from previous years. 

Falxa et al. (2016) discuss and evaluate potential causes 

for the pattern observed, which include (1) change in the 

distribution of murrelets relative to shore that affects 

the proportion of the population sampled, (2) change in 

the model parameters used to estimate density, (3) shift 

of murrelets from nonsampled units to sampled units in 

conservation zone 1, (4) movement of birds into conservation zone 1 from the north or south during 2011 to 2013, 

and (5) potential effects of atypical timing of breeding or 

proportion of the population nesting. The cause(s) remain 

unknown, and continued monitoring and research should 

help managers better understand population trends and 

assess underlying factors that might explain trends and 

variability in annual estimates. 

10,000

15,000

20,000

25,000

30,000

2001 2002 2003 2005 2004 2006 2008 2007 2009 2010 2011 2012 2013 2014

Number of murrelets

Year

Figure 5-5—Annual marbled murrelet population estimates and 95 percent confidence intervals for the Northwest Forest Plan area 

(conservation zones 1 through 5 combined) based on 2000–2014 data (Falxa et al. 2016, Lynch et al. 2016).
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GENERAL TECHNICAL REPORT PNW-GTR-966

The population monitoring results to date indicate that, 

as expected, the NWFP goal of stabilizing and increasing 

murrelet populations has not yet been achieved throughout 

the Plan area. Although the population monitoring data 

for 2000 through 2015 are not consistent with declining 

populations in Oregon and California during this period, 

murrelets are declining in Washington. The Washington 

trend results are consistent with demographic models for 

the murrelet (McShane et al. 2004, USFWS 1997), which 

predicted declining populations based on the available data 

on rates of murrelet survival and reproductive output. The 

population monitoring data suggest a north-to-south trend 

pattern, in which population trends appear to improve 

from north to south within the Plan area based 

on the last 15 years. The observed Oregon and 

California trend results are not consistent with 

model predictions. However, major sources of 

uncertainty include (1) uncertainty in estimating 

survivorship and fecundity (reproductive output) 

in the demographic models, (2) uncertainty 

about whether the murrelet populations being 

monitored are closed or open to immigration, 

and (3) the relatively large confidence intervals 

around population estimates. Murrelets occur 

immediately to the north of the Plan area, and 

monitored populations may be subsidized by 

immigrants from British Columbia or Alaska, 

where birds are more abundant (Falxa and 

Raphael 2016, Raphael 2006). Peery et al. (2007) 

found that immigration of murrelets from north 

of the zone 6 (Santa Cruz Mountains) population 

may have been sufficient to mask an intrinsic 

decline in the zone 6 population; this could 

occur elsewhere.

Status and trend of nesting habitat—

Whereas the focus of the murrelet effectiveness 

monitoring program is on the status and trends 

of murrelet populations and nesting habitat on 

federal lands within the Plan area, the populations monitored at sea respond to nesting habitat 

conditions on both federal and nonfederal lands. 

To better understand the murrelet’s conservation 

status, and the relationship between population 

conditions and nesting habitat conditions, monitoring 

considered nesting habitat conditions across both federal 

and nonfederal lands (Raphael et al. 2016a). Also, in some 

areas, such as southwest Washington and northwest California, few federal lands occur within the murrelet’s nesting 

range, and thus nonfederal lands are likely important to 

murrelet conservation.

Baseline nesting habitat—When the NWFP was developed, no consistent map of murrelet nesting habitat was 

available. For purposes of the Plan, murrelet nesting habitat 

was then assumed to be late-successional forest with much 

-8

-4

0

4

8

Annual change (percent)

Washington Oregon California

State

Figure 5-6—Trend results: average rate of annual change by state, 2000 to 

2013, with 95 percent confidence intervals. Washington trend is based on 

2001–2015 data, Oregon on 2000–2014 data, and California on 2000–2015 

data (Falxa et al. 2016, Lynch et al. 2016).
311

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

the same characteristics as northern spotted owl habitat. 

Therefore, the existing map of spotted owl habitat, which 

was itself a mosaic derived from compilations of local maps 

based on agency judgment, classified satellite imagery, and 

existing inventory maps, was constrained to the range of the 

murrelet and used as a proxy for murrelet nesting habitat. No 

estimate or map of nesting habitat on nonfederal land was 

available. The murrelet effectiveness monitoring group has 

since developed a series of maps, using a consistent vegetation base across all ownerships throughout the Plan area 

(Raphael et al. 2016a); the maps were based first on vegetation data from CALVEG and the Interagency Vegetation 

Mapping Project (Moeur et al. 2005), and then later based on 

Gradient Nearest Neighbor (GNN) vegetation data (Davis et 

al. 2015, Ohmann and Gregory 2002, Moeur et al. 2011).

The primary objectives of the effectiveness monitoring 

plan for the murrelet included mapping baseline nesting 

habitat (at the start of the NWFP in 1993) and estimating 

changes in that forest over time. For the NWFP 20-year 

analysis and report, Raphael et al. (2016a) used maximum 

entropy (Maxent) models to estimate nesting habitat suitability over all habitat-capable lands in the murrelet’s range in 

Washington, Oregon, and California. “Habitat-capable” 

lands were defined as lands capable of supporting or 

-15

-10

-5

0

5

10

15

All 1 2 3 4 5

Annual change (percent)

Zone

Figure 5-7—Trend results: average rate of annual change by conservation zone (see fig. 5-2 for zone locations) 

and for all conservation zones combined, with 95 percent confidence intervals. All zones based on 2001–2014 

data, zones 1 and 2 on 2001–2015 data, zone 3 on 2000–2014 data, zone 4 on 2000–2015 data, and zone 5 on 

2000–2013 data (Falxa et al. 2016, Lynch et al. 2016).
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GENERAL TECHNICAL REPORT PNW-GTR-966

developing into murrelet nesting habitat (fig. 5-8). The area 

of habitat-capable lands evaluated by the 20-year analysis 

included about 20.7 million ac (8.5 million ha) of federal plus 

nonfederal lands within the murrelet range portion of the 

Plan area (Raphael et al. 2016a).3

The portion of the murrelet range included in this analysis excluded inland zone 2 of Oregon and California, where 

no murrelet nests have been observed (see Raphael et al. 

2016a for details). The models used vegetation and climate 

attributes, and a sample of 368 murrelet nest sites (184 confirmed murrelet nest sites and 184 occupied sites) for model 

training. Occupied sites are sites where murrelet behaviors 

associated with nesting have been observed during carefully 

prescribed surveys (Evans Mack et al. 2003); such sites 

do not have confirmed nests but are places deemed likely 

to have nests. Attributes used to build the model included 

estimates of canopy cover, mean tree diameter, diameter 

diversity, canopy layers, number of nesting platforms, stand 

age and stand height, an index of old-growth structure, percentage of a 124-ac (50-ha) area composed of older forest, 

and several climate variables. All of these attributes were 

derived from a regional vegetation database and a climate 

database that covered the entire Plan area as described 

in Raphael et al. (2016a). The model classified each 30-m 

pixel in the Plan area with a nesting habitat suitability score 

ranging from 0 (unsuitable) to 1.0 (most suitable); higher 

scores indicate that a pixel has vegetation and climate characteristics more similar to those in the sample of murrelet 

nest sites, compared to a random sample of available forest. 

Model validation was accomplished by withholding 25 percent of the training data, testing the model on the withheld 

data, and replicating the process 25 times.

Thresholds were defined that summarized land area 

into four classes of nesting habitat suitability; classes 1 

and 2 were deemed lower suitability, and classes 3 and 4 

were deemed higher suitability (see Raphael et al. [2016a] 

for a detailed explanation of these suitability classes and 

the cutoff values used to define them). The model was run 

25 times for each state and then summarized to provide an 

3

 Does not include conservation zone 6, which is south of San 

Francisco and outside of the NWFP area.

estimate of model error, owing to variation in model runs 

themselves and variation in underlying GNN data. Raphael 

et al. (2016a) estimated that there were 2.53 million ac 

(1.02 million ha) of higher suitability nesting habitat over 

all lands in the murrelet’s range in Washington, Oregon, 

and California at the start of the NWFP; this included 1.50 

million ac (0.61 million ha) on federal lands. Of the 2.53 

million ac of higher suitability nesting habitat, 0.46 million 

ac (0.18 million ha) were identified as highest suitability 

(class 4), matching or exceeding the average conditions for 

the training sites; of this, 0.25 (0.10 million ha) million ac 

were on federal lands. A substantial amount (41 percent) of 

baseline nesting habitat occurred on nonfederal land (fig. 

5-9). The estimate of nesting habitat on federal land from 

the 1993 final supplemental environmental impact statement was 2.6 million ac. Differences between the 1993 and 

current nesting habitat estimates were to be expected, as 

the new map was derived from a nesting habitat suitability 

model specific to the murrelet, and was built from forestand satellite-derived data that had not been available at 

the time the NWFP was written. As noted earlier, the final 

1993 supplemental environmental impact statement used 

habitat for the northern spotted owl as a proxy for murrelet 

nesting habitat.

Although a substantial amount of higher suitability 

nesting habitat occurred on nonfederal lands, federal lands 

contributed proportionately more suitable nesting habitat. 

Of the about 20.7 million ac (8.4 million ha) of forest land 

capable of supporting or developing into murrelet nesting 

habitat, federal lands comprise only about 28 percent of 

the area, but provided 59 percent of the suitable nesting 

habitat at the start of the NWFP (Raphael et al. 2016a). The 

contribution of suitable nesting habitat from nonfederal 

land varies: in Washington, 42 percent; in Oregon, 33 

percent; and in California, 80 percent (fig. 5-9). On the 1.0 

million ac (0.4 million ha) of suitable nesting habitat on 

nonfederal lands in 1993, about 39 percent was managed by 

states. In Washington, the proportion of the nesting habitat 

on federal lands that is within reserves is 93 percent; in 

Oregon, 88 percent; and in California, 93 percent. The final 

supplemental environmental impact statement estimated 

that 86 percent of murrelet nesting habitat on federal lands 
313

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

Marbled Murrelet Nesting Habitat

(2012)

Murrelet habitat suitability

Physiographic province

Below threshhold

Above threshhold

Habitat capable

Class 3 (moderately 

 high suitability)

Class 4 (highest suitability)

Not habitat capable

Northwest Forest Plan 

 inland zones

Washington Olympic Peninsula

Washington Western Lowlands

Washington Western Cascades

Washington Eastern Cascades

Oregon Western Cascades

Oregon Coast Range

Oregon Willamette Valley

Oregon Klamath

California Klamath

California Coast Range

1.

2.

3.

4.

5.

7.

8.

9.

10.

11.

0 50 100 150 Miles

0 80 160 240

200

320 Kilometers

o

Figure 5-8—Map of suitability for marbled murrelet nesting habitat, 2012 (Raphael et al. 2016a).
314

GENERAL TECHNICAL REPORT PNW-GTR-966

0

400

800

1,200

1,600

2,000

Washington Oregon California

Nesting habitat (thousands of acres)

0

400

800

1,200

1,600

2,000

Nesting habitat (thousands of acres)

1993

2012

Nonfederal Federal: reserve Federal: non-reserve

A

B

1993

2012

Figure 5-9—Estimated amounts of higher suitability nesting habitat of the marbled murrelet in 1993 and 

2012, by (A) state, and (B) land allocation (Raphael et al. 2016a). Error bars are 95-percent confidence 

intervals from 25 replicated model runs. See table on page 338 for metric equivelents.
315

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

would be in reserves. The 20-year analysis found that, in 

1993, 90 percent of potential nesting habitat on federally 

administered lands occurred within reserved-land allocations (Raphael et al. 2016a). Thus, the NWFP seems to have 

successfully captured most of the existing higher suitability 

nesting habitat on federal lands within its reserve system. 

We conclude that the NWFP had successfully encompassed 

a majority of murrelet nesting habitat within its reserve 

system but that a substantial amount of additional suitable 

nesting habitat occurs on nonfederal lands over which the 

NWFP has little or no control.

Nesting habitat losses—The intent of the NWFP is to 

conserve most of the remaining murrelet nesting habitat 

and to prevent the subsequent loss of any nesting habitat 

occupied by nesting birds, wherever that nesting habitat 

occurred on federal lands. The amount of nesting habitat was expected to increase over time, but the rate of 

increase would be very slow, and changes might not be 

observed for many decades. In the meantime, some unoccupied nesting habitat would be lost to timber harvest on 

federal land, and some losses might be caused by wildfire 

and other disturbances.

The observed trends are in line with these expectations. Raphael et al. (2016a) used satellite imagery 

and change detection methods (see Davis et al. 2015) to 

estimate a net loss of 307,957 ac (124,692 ha) of higher 

suitability nesting habitat over all lands (including nonfederal) from 1993 to 2012, or a total loss of about 12 

percent. Net loss was about 27 percent from the baseline 

on nonfederal lands, and 2.2 percent on federal lands (table 

5-2). Of those losses on nonfederal lands, the highest rate 

of loss was on private lands (37 percent); losses on state 

lands were just under 10 percent (table 5-2). Of those losses 

on federal lands, 62 percent was due to fire (most of that 

in one event, the 2002 Biscuit Fire); 23 percent to timber 

harvest; and 16 percent to insects, disease, or other natural 

disturbances (table 5-3). On nonfederal lands, 98 percent of 

losses were due to timber harvest, and 2 percent to insects, 

disease, and other causes (table 5-3). 

Table 5-2—Change in acres (thousands) of suitable nesting habitat from 1993 to 2012 by land ownership in 

the Northwest Forest Plan area (updated from Raphael et al. 2016a)

State Owner 1993 2012 Change

- - - - - - Acres (thousands) - - - - - - Percent

Washington Federal 899.7 887.1 -1.4

State 243.7 209.7 -29.8

Other nonfederal 405.6 246.3 -39.3

Oregon Federal 573.1 553.7 -3.4

State 123.3 119.6 -3.0

Other nonfederal 157.0 101.5 -35.4

California Federal 26.5 26.0 -1.9

State 32.3 31.9 -1.2

Other nonfederal 73.7 51.3 -30.4

Plan area total Federal 1,499.3 1,466.8 -2.2

State 399.2 361.2 -9.5

Other nonfederal 636.4 398.8 -37.3

Note: see table on page 338 for metric equivelents.
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GENERAL TECHNICAL REPORT PNW-GTR-966

Nesting habitat increases—One NWFP expectation was a 

gradual increase in the amount of suitable nesting habitat as 

forests mature. Previous evidence showed that the amount 

of forest with large (>20 [>51 cm] inches in diameter) trees 

had increased by about 15 percent over the first 10 years of 

the NWFP, based on analyses of inventory plots on national 

forest lands (Moeur et al. 2005). More recent work, however, 

showed a decrease of about 2.8 percent in the amount of older forest on federal lands and about 6 percent over all lands 

within the entire NWFP area; the discrepancy may be due to 

the newer definitions of older forest used in the more recent 

estimates (Davis et al. 2015); this analysis included large 

areas outside (inland to the east) of the murrelet nesting 

range. As noted above, net losses of murrelet nesting habitat 

totaled about 12 percent over all lands and 2.2 percent on 

federal lands. At some point in the future, the extent of current young forest within the reserve system on federal land 

will be such that we could see a net increase in amount of 

suitable nesting habitat. For example, trends in the Oregon 

Coast Range on federal lands show that nesting habitat can 

increase when stand-replacement rates of disturbance are 

low and forest age classes are available to grow into murrelet 

nesting habitat in a few decades. Unfortunately, however, 

we are unaware of any estimates of exactly when that point 

will be reached. There is a need to develop models to project forest conditions forward in time and to then estimate 

future nesting habitat suitability. We do know, as pointed 

out in Raphael et al. (2016a), that there is sufficient young 

and mature forest within the reserve system (fig. 5-8) to 

eventually make up for losses since the start of the NWFP, 

if future nesting habitat losses on federal 

lands remain similar to the first 20 years of 

the NWFP, and the NWFP reserve system 

remains intact and continues to be managed 

for the development of old-forest conditions. 

While at broader scales the amount of murrelet nesting habitat declined, some gains in 

nesting habitat may already be occurring locally, notably on Forest Service lands in the 

Oregon Coast Range province, where small 

net gains (about 1 percent) were observed by 

the 20-year analysis (Raphael et al. 2016a).

Status of Marbled Murrelets Elsewhere in the 

Species’ Range

The NWFP effectiveness monitoring program provides data 

on murrelet status and trends that is unparalleled elsewhere 

in geographic and temporal extent. Nonetheless, other 

monitoring programs exist elsewhere within the species’ 

range (see fig. 5-3 for range map); these provide information 

on the status and trends for some areas outside of the 

NWFP area. The most comprehensive of these in geographic scope is conducted by the Canadian government to 

assess temporal trends of the murrelet in British Columbia. 

That program recently reported on murrelet population 

trends from 1996 through 2013, based on a radar-based 

monitoring program; they found evidence for a coastwide 

decline of about 1.6 percent per year in British Columbia 

(Bertram et al. 2015a). Trends varied strongly among the six 

sampling regions within British Columbia: negative trends 

were detected in their east Vancouver Island (-9 percent per 

year) and south mainland coast (-3 percent per year) regions, 

and a weak negative trend in Haida Gwaii. A separate 

program has monitored at-sea murrelet numbers from about 

62 mi (100 km) of transects on the southwest coast of 

Vancouver Island during May to July since 1995. Results 

from this effort suggest an initial decline through 2006, 

followed by stable or increasing numbers since 2006 

(Bertram et al. 2015a; Zharikov et al. in Irvine and Crawford 2012; Y. Zharikov, pers. comm.4

). The most recent 

4

 Zharikov, Y. 2016. Personal communication. Monitoring 

ecologist, Parks Canada, Ucluelet, BC V0R 3A0.

Table 5-3—Attribution of loss, in thousands of acres, of marbled 

murrelet higher suitability habitat from the Northwest Forest Plan 

baseline (1993) to 2012 by land allocation

Lossesa

Land allocation Fire Harvest Other Total

Acres (thousands)

Federal reserved 19.1 4.6 5.3 34.8

Federal nonreserved 2.4 3.3 0.2 5.3

Nonfederal 0.6 308.7 6.9 316.3

Total 22.1 316.7 12.4 351.7

a

 Losses as verified by LandTrendR (see Raphael et al. 2016a for details).

Source: Raphael et al. 2016a.
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

population estimate in British Columbia, using extrapolations from at-sea surveys and radar counts, gives the range 

as 72,600 to 125,600 birds of all ages (mid-point 99,100 

birds) (COSEWIC 2012). 

In central California, the small murrelet population of 

conservation zone 6 (from the mouth of San Francisco Bay 

to Point Sur in Monterey County) has been monitored with 

at-sea surveys almost annually since 1999. Those surveys 

estimated population sizes of about 400 to 600 birds 

between 2009 and 2014 (Henry and Tyler 2014), with no 

clear trend during that period, but an apparent decline compared to numbers from 1999 to 2003 (Henry et al. 2012).

Data are more limited on the murrelet’s status in 

Alaska, where its range extends from the southeast corner 

of the state through the Aleutian Islands. Within that area, 

monitoring surveys have been conducted annually in 

Glacier Bay since 2009; murrelet numbers there have been 

variable, with the highest annual estimates in 2013 and 

2015 (Sergeant et al. 2015). Monitoring surveys throughout 

Prince William Sound in 11 years between 1972 and 2007 

suggest that murrelet abundance there declined by an 

annual average rate of about 4 to 5 percent per year for that 

period (Kuletz et al. 2011).

Less recent information is available from a 2007 

evaluation of the status of the murrelet in Alaska and British 

Columbia (Piatt et al. 2007). That review evaluated trends 

for Alaska using at-sea survey data from eight different and 

widely distributed sample sites. Although the sites differed 

in methods, sampling effort, and time period sampled, the 

evaluation found evidence for significant declines at five 

of eight sites, at annual rates of -5.4 to -12.7 percent since 

the early 1990s (Piatt et al. 2007). While acknowledging 

uncertainty resulting from a lack of recent survey data from 

key areas, they projected the 2007 murrelet population in 

Alaska to be roughly 270,000 birds, representing a decline 

of about 70 percent over a 25-year period (Piatt et al. 2007). 

They concluded that the declines were likely real, and 

attributed them to combined and cumulative effects from 

climate-related changes in the marine ecosystem affecting 

prey resources (including a regime shift in the Gulf of 

Alaska that reduced the abundance of important murrelet 

prey), and human activities (logging, gill net bycatch, and 

oil pollution).

As noted below, Raphael et al. (2016b) reported a 

correlation between numbers of murrelets counted at sea 

and amounts of adjacent suitable nesting habitat within the 

three-state region of the NWFP. This relationship, however, 

seems to vary considerably in different portions of the 

murrelet range, as illustrated in table 5-4. Certainly, part of 

the reason for this variation is due to differences in methods and definitions of nesting habitat, but the magnitude 

of difference (e.g., 207 ac [84 ha] per bird in Washington 

versus 15 ac per bird in Alaska) suggests that there are real 

differences in relationships between offshore numbers of 

birds and inland nesting habitat in the various regions. We 

note that there is likely a higher proportion of murrelets in 

Alaska nesting in small patches of forest, which are likely 

to be excluded in forest inventories, and on cliffs or on 

the ground (Barbaree et al. 2014). It is also possible that 

foraging prey density is much greater in Alaska, supporting 

a larger number of birds relative to available nesting habitat 

compared with other parts of the range. 

Table 5-4—Estimated amounts of potential nesting habitat (rounded to nearest 100 ac), murrelet population 

size, and ratio of habitat to population in portions of the murrelet range (as depicted in fig. 5-7) 

Region

Nesting 

habitat

Estimated 

murrelet 

population

Habitat area 

per bird Source

Acres Acres

Southeast Alaska 2,034,700 144,200 15 Piatt et al. 2007

British Columbia 3,439,100 99,100 35 Environment Canada 2014

Washington 1,549,000 7,494 207 Lynch et al. 2016, Raphael et al. 2016a 

Oregon 853,400 11,384 75 Lynch et al. 2016, Raphael et al. 2016a

California 132,600 5,666 23 Lynch et al. 2016, Raphael et al. 2016a
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GENERAL TECHNICAL REPORT PNW-GTR-966

Nesting Habitat Relationships

Patches and edges—

Although the behavior and habitat cues used by murrelets to 

locate nest sites are not known, their nests tend to be widely 

spaced across the landscape, especially if there is extensive 

suitable nesting habitat (Nelson 1997). In areas where there 

is a wide choice of suitable trees, nest trees tend not to be 

re-used in successive seasons (Burger et al. 2009). Nests 

located using radiotelemetry in Desolation and Clayoquot 

Sounds, British Columbia, had mean inter-nest distances of 

2.9 ± 2.5 (standard deviation [SD]) mi (4.6 ± 4.0 km) and 4.1 

± 2.6 mi (6.6 ± 4.2 km), respectively, although there were, 

almost certainly, undiscovered nests in between. Other 

telemetry studies showed similar wide spacing (Barbaree 

et al. 2014, Bloxton and Raphael 2009, Wilk et al. 2016), 

although in northern California where nesting habitat is very 

limited, nests were closer together and more often reused 

(Hébert and Golightly 2006). In some circumstances, nests 

might be more closely aggregated. For example, on the 

southern mainland coast of British Columbia, Manley (1999) 

found that 52 percent of nests located with tree climbing were 

within 300 ft of another nest, and on Naked Island, Alaska, 

Naslund et al. (1995) found three nests within a 43-ac (19 ha) 

stand. Additional evidence of co-location within stands and 

watersheds is reviewed by Plissner et al. (2015). 

Analyzing the distribution of marbled murrelet nests 

relative to patch size and forest edges is limited, because 

many studies lacked a statistical comparison of habitat use 

in patches or edges versus the availability of these and alternative habitats (Jones 2001), and proximity to edges was 

not considered in relation to the degree of fragmentation of 

the landscape. Marbled murrelets are known to nest within 

150 ft (46 m) of forest edges and in small, often isolated 

patches of suitable trees. The data summarized by McShane 

et al. (2004) showed that 75 percent of all nests were within 

164 ft (50 m) of forest edges. Most of these edges were 

natural edges (streams, wetlands, natural forest gaps, and 

avalanche chutes) but almost a third of all nests were close 

to edges created by human activities. These data include 

nests located from ground searches and tree climbing linked 

to audiovisual surveys, and these nests are likely to be 

biased toward being found near edges (Burger 2002). When 

considering only the nests found by climbing randomly 

selected trees and radiotelemetry to remove possible bias, 

the results were similar: most nests were located near edges 

(76 percent of 152 nests), and the most common type of 

edge was natural (69 percent of 115 edge nests) (McShane et 

al. 2004). In this unbiased sample, which covered a range of 

modified and relatively pristine nesting habitats in Oregon, 

Washington, and British Columbia, 24 percent of all nests 

were near manmade edges, even though interior forest 

existed near many of these nests. Distances of nests to all 

edges in these samples ranged from 20 to 2,100 ft (6 to 640 

m), and proximity to anthropogenic edges ranged from 9 to 

1,000 ft (3 to 305 m) (McShane et al. 2004). 

Studies using telemetry in British Columbia and Alaska 

found some murrelets nesting in small, often isolated patches 

of suitable forest; these patches were usually in higher elevation sites, where suitable trees are sparse and small patches 

of larger trees provide suitable platforms (e.g., Barbaree et al. 

2014, Bradley 2002). When small patches are used in lower 

elevation sites, this often occurred where logging had removed 

most of the low-elevation suitable forest (e.g., Zharikov et al. 

2006, 2007a). It is possible that murrelets persisted in such 

small patches because of site fidelity. Murrelets have shown 

a strong fidelity to sites where they have previously nested 

(e.g., Hébert and Golightly 2006). It is important to note that 

nest success may be lower in these smaller patches, probably 

because of higher risk of nest depredation (Barbaree et al. 

2014). Fine-scale spatial analysis of the nests found with 

telemetry in Desolation Sound, on the southern mainland of 

British Columbia, showed that murrelets were more likely to 

nest close to natural edges, but there were insufficient data to 

test whether this was true for manmade edges (Burger 2002). 

Two studies of nest placement did consider the use 

versus availability of edge habitat and patch size within the 

landscape. Raphael et al. (2016a) found that more than 60 

percent of 162 nests in Washington, Oregon, and California 

were found in interior forest (defined as further than 180 

ft (55 m) from any edge) (table 5-5). In that study, only 23 

percent of potential nesting habitat occurred as interior 

forest on all lands in the study area, indicating a greater-than-expected occurrence of nests in interior forest. Wilk 

et al. (2016) analyzed nesting habitat at nests used by birds 
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tagged with radios in the waters close to the Olympic Peninsula, Washington. Murrelet nests in Washington (n = 18) 

had greater core areas of older forest than random sites (235 

ac [95 ha] at nest sites versus 25 ac [10 ha] in random sites). 

Core area is the interior area of the forest patch after buffering edge effects (180-ft buffers); this measure integrates 

patch size, shape, and edge-effect distance into a single measure. Raphael et al. (2016a) also found that patch cohesion, 

the physical connectedness of the corresponding patch type 

(index range 0 to 100), was greater at nests than random 

sites (93 at nests, 66 at random sites). They concluded that 

stands with nests were less fragmented than available forest 

across the murrelet’s range that they sampled. 

Edge effects on forest nesting habitat: windthrow, 

microclimate, and epiphytes—

A general rule of thumb used in Pacific Northwest forests 

has been that microclimatic effects penetrate two tree 

heights (240 to 300 ft [73 to 93 m]) and sometimes farther 

(450 ft [137 m] or more) into old-growth forests bordering 

clearcuts or similar sharp-gradient boundaries (Franklin 

and Forman 1987, Kremsater and Bunnell 1999). This 

is supported by some field studies, but local variables 

like topography, wind exposure, type of forest, and the 

surrounding matrix strongly influence the magnitude and 

influence distance of these edge effects (reviewed below). 

Several studies reviewed below found differences based 

on edge type, in which “hard” edges are those with recent 

clearcuts (e.g., 0 to 20 years old) and “soft” edges are with 

regenerating forest (such as 21 to 100+ years old).

Windthrow refers to the uprooting or breakage of 

trees by wind, which can affect murrelets owing to loss of 

potential nest trees and nest limbs. Windthrow is increased 

when clearcuts, and to a lesser extent roads, increase the 

exposure of residual trees to wind (Sinton et al. 2000). 

Windthrow and physical damage to canopy branches are 

common problems at hard edges within the murrelet’s 

range. In the Pacific Northwest, factors affecting the risk 

and degree of windthrow include orientation relative to winter winds; topography; the age, height, and density of trees; 

soil type; exposure to wind prior to logging (trees exposed 

to winds are more likely to develop stronger root systems); 

and the shape and size of the clearcuts and residual stands 

(Franklin and Forman 1987, Gratowski 1956, Mitchell et 

al. 2001). Although local factors have a strong influence, 

these impacts are generally found within 150 to 240 ft (46 

to 73 m) of edges, are most prevalent in patches less than 3 

ha (7.4 ac), and are most likely within 25 years of clearcut 

logging creating the edges. In a review of data from the 

Pacific Northwest, Franklin and Forman (1987) suggested 

that wind-driven edge effects were likely to penetrate 

into remnant forests about two tree heights (240 ft [73 m]) 

from clearcut edges, but they did not distinguish between 

windthrow, canopy damage, and changes to microclimate.

Canopy epiphytes (mostly mosses) provide nest platforms for murrelets in much of the NWFP area. Exposure 

to increased wind and solar radiation at newly created edges 

could be detrimental (through wind-removal, thermal stress, 

and desiccation) or beneficial (through increased light for 

photosynthesis) to these epiphytes. Studies in the Pacific 

Northwest found variable effects of edges on bryophytes, 

although moss cover tended to be lower near hard edges. 

Local features, especially topography, time since edge 

creation, edge orientation, aspect, the nature of the surrounding harvested matrix, and even soil conditions have a strong 

effect on physical damage and changes in edge microclimates 

(Franklin and Forman 1987, Gratowski 1956, Mitchell et al. 

2001, Muth and Bazzaz 2002, Sherich et al. 2013). These 

Table 5-5—Number of marbled murrelet nestsa

located in core areas (interior forest) and near (within 

180 ft [55 m]) edges

Location Core

Core 

edgeb Edgec Total 

Washington 24 15 8 47

Oregon 29 23 4 56

California 45 8 6 59

Total 98 46 18 162

Percent 61 28 11

Available (percent)d 22 28 49

a Numbers of nests as sampled in Raphael et al (2016a), not the total 

number of known nests in this region.

b Edge of interior forest (core) patch.

c Isolated edge or stringer.

d Percentage of each type throughout range.

Source: Raphael et al. 2016a.
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GENERAL TECHNICAL REPORT PNW-GTR-966

studies of edge effects on epiphytes and microclimate, 

although not focused on murrelet nesting, indicate that in 

many cases forests within 150 ft (46 m) of hard edges are 

likely to provide adverse conditions for nesting murrelets, 

and in situations with greater wind exposure, these adverse 

conditions could extend well beyond 300 ft (91 m). These 

adverse conditions are likely to diminish as the adjacent 

regenerating forest reduces the edge gradient (i.e., creates 

“soft” edges). One study, by Van Rooyen et al. (2011) at four 

locations in British Columbia, has specifically investigated 

edge effects on factors relevant to nesting murrelets. Compared to adjacent interior forest, epiphyte cover on canopy 

branches was slightly lower at hard edges (possibly because 

of the microclimate effects discussed above), about the same 

at soft edges, and slightly higher at natural edges. There was 

a large difference in the density of trees with potential nest 

platforms between hard edges and forest interiors (1.5 versus 

6.4 platform trees per acre [0.6 versus 2.6 per hectare]); the 

difference was less marked at soft edges (6.5 versus 10.8 platform trees per acre [2.6 versus 4.4 per hectare]) and negligible 

at natural edges. The authors concluded that the creation of 

artificial edges by forest fragmentation would have negative 

consequences for epiphytic development for 20 to 30 years, 

and this might reduce nesting habitat for murrelets. 

Natural forest edges bordering openings produced by 

streams, avalanche chutes, and wetlands generally do not provide adverse conditions for nesting murrelets, and if temperature and moisture regimes are favorable, such edges might 

be more suitable for murrelets than interior forests (Harper 

et al. 2005, Van Rooyen et al. 2011). Despite the evidence of 

negative microclimates and bryophyte development near hard 

edges, murrelet nests have been observed within 150 ft (46 m) 

of such edges, suggesting that conditions there are not always 

an absolute deterrent to the birds. We do not know if they 

avoid hard edges, i.e., whether nest densities at hard edges are 

lower than those elsewhere in old-growth forests. On balance, 

however, the evidence suggests that the creation of small 

patches and hard edges can be detrimental in areas where 

maintenance of nesting murrelets is a priority. Occurrence 

of nests along edges may, as noted above, be a result of site 

fidelity and a tendency to nest at previously used locations 

even when disturbances have created edges near those sites.

Microclimates within old-growth forests differ from 

those in clearcuts or young regenerating forests. In general, 

extremes of temperature and solar radiation are minimized, 

and humidity in summer is higher and more stable in oldgrowth forests than in recent clearcuts (Chen et al. 1999, Frey 

et al. 2016). Changes in microclimates can have both direct 

and indirect effects on nesting murrelets. Direct effects 

include thermal stress (both hot and cold) and dehydration if 

adults or chicks are exposed to direct sunlight or increased 

winds. Indirect effects are most likely to occur through 

changes to the availability of moss pads and other epiphyte 

growth on which most murrelet nests have been found. 

Analysis across the Plan area indicates that the prevalence of fog is a strong contributor to predictive models of 

suitable nesting habitat for murrelets (Raphael et al. 2016a). 

In areas where fog is frequent, it might mitigate some edge 

effects, by promoting epiphyte growth and ameliorating 

stressful solar radiation. However, there is some evidence of 

reduced fog frequency, at least in California, over the past 

century (Johnstone and Dawson 2010).

Landscape-level relationships between nesting habitat 

and populations—

Data from radar surveys—In this section, a landscapelevel spatial scale considers entire watersheds and similar 

large areas in contrast to smaller stand- and patch-level 

analyses. Counts of murrelets entering watersheds obtained 

by detections from radar equipment have been instrumental 

in showing that murrelet numbers are strongly correlated 

with available areas of suitable old-growth nesting habitat 

(Burger 2001, Burger et al. 2004, Raphael et al. 2002a). In 

addition, Raphael et al. (2002a) also tested for the effects of 

habitat fragmentation in watersheds sampled with radar on 

the Olympic Peninsula, Washington. In their 3-year study, 

numbers of murrelets detected increased as the amount of 

core-area old-growth (defined as interior forest more than 

300 ft [92 m] from an edge) increased (r2

 = 0.69, 0.82, and 

0.76 in 1998, 1999, and 2000, respectively, p < 0.01), but decreased with increasing amounts of edge in late-seral patches. Numbers of murrelets were not correlated with patch 

density (number of patches per hectare), mean patch size, or 

spacing (proximity) of late-seral patches, nor with the overall diversity of all forest cover types within the landscape. 
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Cortese (2011) compared radar counts of murrelets 

entering watersheds with forest cover parameters within 

watersheds in three regions of British Columbia: southwest 

Vancouver Island, and the central and southern mainland 

coasts. One goal of the study was to investigate the effects 

of forest fragmentation within the watersheds. As expected 

from previous radar studies (Burger 2001, Burger et al. 

2004, Raphael et al. 2002a), total area of old-growth forest 

was included in the top predictive models for all three 

regions, which explained 11 to 35 percent of the variability 

in radar counts. Measures of mature forest edge density 

(including “hard” edges with clearcuts 0 to 20 years old, 

and “soft” edges with regenerating forest 21 to 140 years 

old) also were included in most predictive models, but 

there were marked regional differences in whether these 

were positive or negative associations. In the central and 

southern mainland coast regions, hard edges had a positive 

association with murrelet numbers, although there was 

high uncertainty in the model selection for the latter region. 

Cortese (2011), following Zharikov et al. (2006, 2007a), 

attributed this result to the preference by murrelets and the 

logging companies for the same patches of old-growth forest. Much of the old-growth forest in the watersheds studied 

in these regions has already been removed (Zharikov et al. 

2006), and therefore murrelets tend to nest in the remaining 

forests where there is active logging and hence fragmentation. By contrast, murrelets in southwestern Vancouver 

Island, where a greater proportion of murrelet nesting 

habitat remains, showed a negative association with the 

density of hard edges and a strong negative association with 

the density of soft edges, and these edge factors were more 

important predictors in this region than in the other two 

regions (Cortese 2011). 

Data from at-sea surveys—Comparison of murrelet counts 

at sea with forest nesting habitat parameters emphasizes the 

value of tracts of suitable old-growth forest close to marine 

foraging areas (e.g., Falxa and Raphael 2016, Miller et al. 

2002, Ronconi 2008, Raphael et al. 2015). In addition to the 

total area of accessible nesting habitat, Miller et al. (2002) 

found that nesting habitat patch size (r = 0.91) and contiguity of old-growth forest (r = 0.95) were the strongest predictors of murrelet densities at sea in northern California and 

southern Oregon. Raphael et al. (2016b) analyzed 13 years 

of data (2000–2012) from marine surveys in nine geographic strata across three states (Washington, Oregon, and 

California). Murrelet abundance at sea was most strongly 

correlated with the amount of higher suitability nesting 

habitat in the adjacent terrestrial environment (r2 = 0.324), 

but there was considerable variance that was not explained 

by the factors included in the analysis. In addition, cohesion 

(an index of nesting habitat pattern in which higher values indicate more contiguous and less fragmented nesting 

habitat) was strongly and positively correlated (r2 = 0.76) 

with murrelet abundance within the survey strata. We note, 

however, that amount of nesting habitat and cohesion of that 

habitat cannot be considered independent; cohesion tends to 

increase as amount of nesting habitat increases. Although 

the unexplained variance indicates that other factors also 

influence murrelet distribution and abundance, the results of 

Miller et al. (2002) and Raphael et al. (2015, 2016b) indicate 

that fragmentation of nesting habitat has negatively affected 

murrelet populations across the large, diverse, and highly 

modified NWFP area.

Nesting habitat configuration and risk of nest predation—Breeding success in murrelets tends to be low (typically less than 35 percent of nests fledge chicks). A study 

using museum specimens indicated that historical breeding 

success about a century ago was sufficient to maintain stable murrelet populations, but that contemporary reproductive success is not (Beissinger and Peery 2007). Predation 

is the highest known cause of nest failure in recent decades 

and is likely to limit murrelet populations in many areas. 

Corvids (crows, ravens, and jays) are the nest predators 

most commonly documented, but owls, diurnal raptors, and 

arboreal mammals (squirrels and mice) (Bradley et al. 2003; 

Malt and Lank 2007, 2009) are also likely to be important 

predators. Although definitive demographic studies testing 

the effects of predation are limited to the edge of the species 

range in central California (Peery and Henry 2010; Peery 

et al. 2004, 2006a), those studies and cumulative evidence 

from across the species range indicate that nest predation 

is a limiting factor on murrelet populations (McShane et al. 

2004, Nelson and Hamer 1995, Piatt et al. 2007). Studies 

in several parts of the species range show that only about a 
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GENERAL TECHNICAL REPORT PNW-GTR-966

third of murrelet nests result in fledging, e.g., 0.33 fledglings 

per nesting attempt rangewide, n = 124 nests (McShane 

et al. 2004), and 0.23 to 0.46 in British Columbia (Burger 

2002). Research using radiotelemetry found failure rates 

of 54 percent in British Columbia (Bradley 2002), 68 to 86 

percent in northern California (Hébert and Golightly 2006), 

84 to 100 percent in central California (Peery et al. 2004), 

80 percent in southeast Alaska (Barbaree et al. 2014), and 

31 percent in south-central Alaska (Kissling et al. 2015). It 

is possible that nesting success results from radiotelemetry studies are affected by the method: Peery et al. (2006b) 

found that radio-tagged murrelets had a lower survival 

rate, and Ackerman et al. (2004) found that radio-tagging 

reduced reproductive success in another small alcid, the 

Cassin’s auklet (Ptychoramphus aleuticus). 

Predation is the greatest known cause of failure at 

78 percent, or 29 of 37 nests with known outcomes in a 

rangewide analysis (McShane et al. 2004). In southern 

British Columbia, Malt and Lank (2007) found no difference between the survival of 57 actual versus 40 artificial 

murrelet nests and were able to document predator 

discovery at 40 percent of 136 artificial nests. In northern 

California, Hébert and Golightly (2006, 2007) attributed 

a minimum of 51 percent of nest failures across 3 years 

to predation, and documented egg predation by ravens 

(Corvus corax) and Stellar’s jays (Cyanocitta stelleri). In 

central California, rates of nest predation were consistently high (67 to 81 percent)(Peery et al. 2004). 

Several studies across the southern part of the murrelet’s range have investigated nest success relative to forest 

edges and habitat fragmentation (table 5-6). As in many 

studies of habitat fragmentation, separating the effects of 

proximity to edge to the related effects of patch size and 

habitat configuration is often difficult (Harper et al. 2005, 

Lindenmayer and Fischer 2007). Nelson and Hamer (1995) 

found that successful nests were significantly further from 

forest edges (mean 510 ± SE 241 ft [155 ± 73 m], n = 9) than 

nests that failed (mean 90 ± SE 20 ft [27 ± 6 m], n = 8), and 

all successful nests, except one, were more than 180 ft (55 

m) from the forest edge. For 58 nests with known locations 

from Oregon and British Columbia, Manley and Nelson 

(1999) (see also Burger 2002) reported that the success of 

nests within 150 ft (46 m) of a forest edge was 38 percent (n 

= 29) and for those more than 150 ft from an edge, success 

was 55 percent (n = 29), but this difference is not statistically 

significant. Successful nests were significantly further from 

edges (mean 462 ft [141]) than failed nests (mean 184 ft [56 

m]). Predation was responsible for the failure of 60 percent 

of all active nests in these samples, and predation rates were 

higher within 150 ft of edges than farther into the forest 

interior. All 13 nests that were more than 450 ft (137 m)

from an edge were successful or failed from reasons other 

than predation. There was a trend for successful nests from 

Oregon and British Columbia to occur in larger stands (mean 

1,212 ac [491 ha]) than unsuccessful nests (mean 694 ac [281 

ha]), although this was not statistically significant.

Bradley (2002) analyzed the success of nests found by 

telemetry in Desolation Sound, British Columbia, relative 

to their proximity to forest edges. Successful nesting was 

assumed if the radio-tagged adult visited the nest up to the 

midpoint in the chick-rearing period and was confirmed 

at some nests by tree climbing after the chick had fledged. 

Bradley (2002) conducted two analyses. One was from 

ground-based measures of distance from edge and nest 

success from 37 accessible nest sites, analyzed at 150 and 

300 ft (46 and 91 m) distances from edge. At both distances, 

there were no significant differences in nest success at sites 

adjacent to or far from forest edges. Most nests were located 

adjacent to natural edges rather than artificial ones. Comparing nest success at natural and artificial edges was difficult, 

because only two nests were located directly adjacent to artificial edges (both were successful). Bradley’s (2002) second 

analysis was a coarse-scale geographic information system 

(GIS) analysis using 98 nest sites, looking at edge type 

within 600 ft (182 m) of sites based on 1:250,000 landscape 

classification maps. In this analysis, the proportions of sites 

adjacent to edges versus interior were similar to those in 

the first ground-based sample. As in the first analysis, many 

nest sites were adjacent to natural edges, predominantly 

avalanche chutes, and most of these nesting attempts were 

successful (79 percent, n = 42). Nest success near artificial 

edges (61 percent, n = 23) and in forest interiors (48 percent, 

n = 33) was lower. Nests adjacent to natural edges had significantly higher success than those in the forest interior, but 
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there were no significant differences between nests adjacent 

to artificial versus natural edges and artificial edges versus 

interior forest. In summary, Bradley’s (2002) analysis did 

not support the hypothesis that nesting near forest edges was 

harmful to murrelets, but could not resolve whether natural 

or artificial edges produced differences in nest success. 

Bradley’s (2002) study was limited because only 38 percent 

of the nests were accessible for ground-based measures and 

tree climbing, and proximity to edges for most nests was 

inferred from coarse-scale global positioning system (GPS) 

locations with ±100 m (328 ft) accuracy. The more detailed 

study by Malt and Lank (2007, 2009) in the same area and 

using some of the same nest data did find significant negative 

edge effects and differences between edge types (see below).

Table 5-6—Summary of studies investigating the effects of habitat fragmentation, small patches, and forest 

edges on the success of marbled murrelet nesting

Study Location Type of study Conclusions

Nelson and Hamer 1995 Rangewide Review of early studies Successful nests significantly farther from 

forest edges than failed nests. Corvid 

predation important.

Manley and Nelson 1999 Oregon and British 

Columbia—using some 

of same data as above

Review of early studies 38-percent success in nests <150 ft; 55 

percent success in nests >150 ft. Predation 

responsible for at least 60 percent of 

failures. 

Bradley 2002; see also 

Burger 2002

Desolation Sound, 

British Columbia

Nest success based on 

telemetry and postfledging evidence

No negative effect of natural edges (e.g., 

avalanche chutes); insufficient data to test 

effects of clearcut edges.

Luginbuhl et al. 2001, 

Marzluff et al. 2000, 

Raphael et al. 2002b

Olympic Peninsula, 

Washington, and 

Oregon

Artificial nests with 

mimic eggs and 

chicks in natural nest 

locations

No consistent effects of forest fragmentation 

on nest survival. Proximity to human 

activity increased predation rates. Corvid 

predation important. Maturing forest 

bordering old-growth nesting habitat 

reduced predation risk.

Malt and Lank 2007 Southwestern British 

Columbia

Artificial nests with 

mimic eggs and 

chicks in natural nest 

locations

Predator visits significantly higher at 

edges (<150 ft) than in forest interior 

(>450 ft from edges), but no difference 

between “hard,” “soft,” and natural edges. 

Predatory corvids more likely at “hard” 

edges. 

Malt and Lank 2009 Southwestern British 

Columbia

Artificial nests with 

mimic eggs and 

chicks in natural nest 

locations

Predator disturbance 2.5 times more likely 

at hard edges than in forest interior. Soft 

and natural edges not so. Corvid predation 

important. Maturing forest (20 to 40 years 

old) bordering old-growth nesting habitat 

reduced avian predation risk. 

Hébert and Golightly 

2006, 2007; Peery et al. 

2004, 2006 

Central and northern 

California

Telemetry and nest 

observations showing 

nest success in highly 

fragmented forests

84-percent nest failure; 67 to 81 percent 

of nests predated. Corvid predation 

important. Repeated use of same nest site 

associated with high predation. 

Zharikov et al. 2006 Desolation sound, 

British Columbia

Nest success based on 

telemetry evidence only 

(new analysis using 

Bradley 2000 data)

Breeding success was greater in areas with 

recent clearcuts and lower in areas with 

much regrowth.
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A later analysis by Zharikov et al. (2006) studied 

habitat selection and breeding success at nest sites located 

with telemetry in Desolation Sound (heavily logged; 121 

nests) and Clayoquot Sound on the west coast of Vancouver 

Island (relatively intact; 36 nests). Comparing nest sites with 

randomly located points in these same areas, they found that 

murrelets used either old-growth fragments proportionately 

to their size frequency distribution (more intact landscape) 

or tended to nest in disproportionately smaller fragments 

(heavily logged landscape). Nests were closer to clearcut 

edges than expected, with mean distances to forest edges of 

1.2 and 1.5 mi (1.9 and 2.4 km) at nest sites and randomly 

chosen points, respectively). Breeding success, as inferred 

from nest attendance patterns by radio-tagged parents, was 

modelled in Desolation Sound, where sample sizes were 

sufficient (Zharikov et al. 2006). They found that breeding 

success was greater in areas with recent clearcuts and lower 

in areas with much regrowth, implying that marbled murrelets can continue nesting in highly fragmented old-growth 

forests, successfully using patches of about 25 ac (10 ha) or 

greater. However, they cautioned that breeding success in 

fragmented areas may decrease as adjacent clearcuts overgrow, and that their findings imply that the same stands of 

old-growth forest may be equally attractive to marbled murrelets and logging companies, versus a murrelet preference 

for forest fragmented by logging (Zharikov et al. 2006). 

The finding by Zharikov et al. (2006) that murrelets can 

nest successfully in highly fragmented old-growth forests 

differs somewhat from results of other studies from British 

Columbia (Burger 2002); Burger and Page (2007) suggested 

that the spatial resolution and scale of the Zharikov et al. 

(2006) analyses were not sufficient to test edge effects (see 

Zharikov et al. 2007b for their response).

Because of the difficulties in locating and monitoring 

murrelet nests, several studies have resorted to using artificial nests with eggs or chicks mimicking those of the murrelet. Justification for this approach for studying murrelets 

is provided by Raphael et al. (2002b) and Malt and Lank 

(2007, 2009). “Predation” and disturbance by predators 

at artificial nests was based on removal, photographic or 

video evidence, movements detected by implanted motion 

sensors, or bite and peck marks made on wax coatings 

of eggs or chicks (Luginbuhl et al. 2001, Malt and Lank 

2007, 2009). Artificial murrelet nests do not, of course, 

have an attendant parent, which might affect the rates of 

predation, although incubating adults have been attacked 

by ravens, and adults do leave eggs unattended for periods 

of several hours (Nelson and Hamer 1995). Murrelet chicks 

are brooded by adults for only a few days after hatching. 

The use of artificial nests to test predation effects has been 

criticized (e.g., Faaborg 2004), but their use has also been 

supported as allowing more rigorous and controlled quantitative experiments (Batáry and Báldi 2004). In the only 

study to compare the success of real and artificial marbled 

murrelet nests at various edge types, Malt and Lank (2007, 

2009) found that artificial nests had significantly lower 

probabilities of disturbance (0.18 ± 0.05) than the probabilities of failure at real nests (0.35 ± 0.07), but the patterns of 

disturbance/failure were similar across edge types for real 

and artificial nests (reviewed below). If these results apply 

generally, then artificial nests seem unlikely to overestimate 

predation rates, and there is support for their application for 

studying edge effects in murrelets.

Intensive research on the likely impacts of forest 

structure and landscape contiguity on murrelet nest predation was undertaken by Marzluff and his team in the 

Olympic Peninsula, Washington, and in Oregon (Luginbuhl 

et al. 2001, Marzluff et al. 2000, Raphael et al. 2002b). Their 

experiments used painted plastic eggs and dark chicken 

chicks placed high in forest canopies to mimic those of 

murrelets. Video monitoring and marks on wax coatings 

identified predators, and field studies were supplemented 

with laboratory studies to test whether potential predators 

would attack eggs or chicks. Their field trials were focused 

on determining the effects of forest structure (simple or 

complex and of different ages), landscape contiguity (classified as fragmented when plots were more than 75 percent 

surrounded by clearcuts or contiguous when plots were more 

than 75 percent surrounded by mature forest), and proximity 

to human activities (near, less than 0.6 mi [1 km]), or far, 

more than 3.1 mi (5 km), from towns, farms, campgrounds, 

dumps, highways, etc.). Survival of simulated nests differed 

relatively little among the various forest cover types, and 

there were no consistent effects of forest fragmentation on 

nest survival but proximity to human activity increased predation rates. At locations far from human activity, predation 
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rates were greater in continuous stands than in fragmented 

stands, but close to human activity, predation rates were 

similar in continuous and fragmented stands. The highest 

nest survival occurred in mature forest with simple structure, which were either contiguous and near human activity 

or fragmented and far from humans. Densities of corvids 

were lowest in contiguous, simple-structured maturing forests, regardless of proximity to humans, and corvid numbers 

differed little among the other forest cover categories. It is 

difficult to infer generalizations from these results, apart 

from negative effects of proximity to human activities, but 

Marzluff et al. (2000) suggested that old-growth stands used 

by murrelets for nesting might be best buffered by surrounding the stands with maturing, simple-structured forests in 

which there were relatively few predators.

In the same study, Luginbuhl et al. (2001) reported a 

strong negative correlation between survival of simulated 

murrelet eggs and corvid abundance at the landscape level (2 

to 20 mi2 [5 to 52 km2] scale). Corvid abundance explained 

69 percent of the variance in predation of simulated murrelet 

eggs. This trend was not evident at smaller plot-level scales 

(60 to 120 ac [24 to 49 ha]). The cause of this scale-sensitive 

relationship was likely due to the large home range of some 

of the corvid species (ravens and crows). For monitoring and 

management purposes, this result implies that such negative 

correlations might not be evident unless large spatial scales 

are considered. Artificial nests in areas with high use by 

Steller’s jays lasted only half as long as those in low-use 

areas (Vigallon and Marzluff 2005).

Malt and Lank (2007, 2009) used artificial nests with 

painted eggs and stuffed quail chicks to study predation 

rates likely to apply to murrelets relative to edges in four 

sites in British Columbia. Avian predators caused more 

than half of the disturbances, with squirrels and mice also 

frequent. Artificial eggs were disturbed more frequently 

than nestling mimics, and birds and squirrels disturbed eggs 

more than nestlings, but the reverse was true for mice. In 

their first study (Malt and Lank 2007), disturbances of nest 

contents by all predators was significantly higher at edges 

(less than 150 ft [46 m]) than in the forest interior (more 

than 450 ft [137 m] from edges), but there was no difference 

between “hard,” “soft,” and natural edges. In both studies, 

predation of eggs by birds (mainly corvids) was always 

higher at hard edges than in interior forest, but soft or 

natural edges did not show this effect. Predation on nest 

contents by squirrels and mice was more variable regionally 

and with forest type, but generally predation by mice was 

not strongly affected by edges (although higher at natural 

edges than in adjacent interiors). They found no edge effects 

from squirrel predation in their first study (Malt and Lank 

2007), but in their second study squirrel predation was 

higher at all edge types than in adjacent interior forest (Malt 

and Lank 2009). 

At the landscape scale, Malt and Lank (2009) found 

that avian predation risk was negatively affected by the 

percentage of regenerating forest 20 to 40 years old; i.e., 

the risk of egg predation decreased by more than half if 

the bordering regenerating forest increased from 1 to 40 

percent. This matches the conclusions by Marzluff et al. 

(2000) on the buffering effects of regenerating younger forest. Malt and Lank (2007) also reported higher predation in 

landscapes with a higher proportion of old-growth forests, 

which might indicate that recent clearcuts and regenerating 

forests supported fewer predators overall.

Some important trends emerge from the work of Malt 

and Lank (2007, 2009). Predation risk from avian predators 

was considerably higher than from mammals, and the 

birds were more likely to target eggs than nestlings. This 

risk from avian predators was particularly high at hard 

edges, but much less likely at soft or natural edges, and the 

landscape-level analysis indicates that this is likely due to 

reduced predation risk as the regenerating matrix changes 

from clearcut to young (20- to 40-year-old) forest. They 

also found strong edge effects among squirrels, which 

is contrary to the general belief that squirrels are less 

attracted to edges than birds, such as corvids (Marzluff and 

Restani 1999). 

The reduction and fragmentation of old-growth forests 

can also lead to the undesirable situation in which murrelets 

and some of their predators (especially old-growth-dependent species such as goshawks) are restricted to using 

the same small patches. This could lead to greater risk of 

predation. If adult murrelets are put at risk in this way, it 

would have serious consequences for populations. 

Nesting murrelets and their eggs and chicks are at 

risk to a formidable array of potential predators, and the 
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murrelet’s cryptic and widely spaced nest sites, secretive and 

crepuscular visits to nests, and camouflaged breeding plumage are all obvious adaptations to reducing predation risk. 

Although it is difficult to estimate the predation impacts 

of the complete suite of predators (birds and mammals) in 

any area, it is clear that corvids, especially Steller’s jays 

and common ravens, are the most common nest predators 

across the murrelet’s range (McShane et al. 2004, Nelson 

and Hamer 1995, Piatt et al. 2007). Both of these species 

and, in some situations, other predators like squirrels (Malt 

and Lank 2007, 2009), exhibit strong affinities with forest 

edges (Marzluff et al. 2000). Murrelets nesting at edges, and 

especially hard edges bordering open areas like clearcuts, 

appear to be at greater risk of predation than in the forest 

interior. Given that nest predation appears to be a dominant 

demographic driver for the murrelet (McShane et al. 2004; 

Nelson 1997; Peery et al. 2004, 2006a; Piatt et al. 2007), 

any forest alteration that increases predation risk is likely to 

have a negative and perhaps serious impact on local murrelet 

populations. Reducing predator risks by minimizing edge 

habitats and controlling corvid access to garbage and human 

food (e.g., at campsites) is also likely to benefit murrelets in 

modified landscapes.

The situations in northern California, documented by 

Hébert and Golightly (2006, 2007) and in central California 

by Peery et al. (2004, 2006b), illustrate how massive nesting 

habitat loss and limited nesting options for murrelets lead to 

a classic habitat trap situation (Battin 2004). Murrelets nesting in those regions are concentrated in the relatively small 

patches of suitable redwood forests remaining, and reuse of 

the same trees and nest sites is higher than what is recorded 

elsewhere (Burger et al. 2009; Hébert and Golightly 2006, 

2007). These trees and nest sites are repeatedly visited by 

corvids (Steller’s jays and common ravens), and consequently nesting success is extremely low in conservation 

zone 6 at the southern end of the murrelet’s breeding range, 

where 84 percent of nests fail and predation rates at nests 

are 67 to 81 percent (Peery et al. 2004). Along with periodic 

food shortages linked to oceanic variability, nest predation 

is considered to limit this population, which appears to be 

sustained by immigration (Peery et al. 2004, 2006a, 2007). 

Reducing corvid populations (Peery and Henry 2010) and 

aversion conditioning to reduce nest predation by Steller’s 

jays (Gabriel and Golightly 2014) are potential management 

strategies to help maintain this marginal population of murrelets. This extreme situation might not be typical of other 

less-modified parts of the murrelet’s range, but is likely 

similar in northwest Oregon, southwest Washington, and 

northern California, and on Bureau of Land Management 

lands in Oregon where the landscape is highly fragmented. 

These situations indicate the risks of excessive habitat 

reduction and fragmentation.

In summary, this review shows that many factors affect 

the risks to murrelets when they nest near forest edges or in 

small forest patches, including the type of edge, the type of 

habitat bordering the edge, the suite of predators likely, and 

proximity to human activity (table 5-6). In most situations, 

particularly where ravens and jays are common, nesting near 

(<150 ft [46 m]) “hard” edges (i.e., the bordering regenerating forest is less than 20 to 40 years old) will increase 

predation risk. 

Marine habitat—

The NWFP is tightly linked to the status and trends of 

murrelets because its lands provide the majority of suitable 

nesting habitat within the species’ listed range in Washington, 

Oregon, and California. Recent analyses indicate that nesting 

habitat conditions best explain the abundance and trends of 

murrelets at sea off the NWFP area during the breeding season (Raphael et al. 2015, 2016b). A breeding-season pattern 

of murrelets tending to occur offshore of nesting habitat is 

consistent with nesting murrelets behaving as central-place 

foragers, subject to energetic constraints that limit them 

to foraging within some radius of their nest location—the 

“central place” (Raphael et al. 2015). Murrelets depend 

entirely on marine prey, and because of this, prey conditions 

such as abundance and quality, and the underlying factors 

affecting prey conditions, are important to the future of the 

murrelet in the Plan area and elsewhere. Thus, the juxtaposition of productive foraging habitat offshore of nesting habitat 

may be important to murrelet conservation. Notably, reviews 

of murrelet biology (McShane et al. 2004, Piatt et al. 2007) 

indicate that the distribution of foraging murrelets is strongly 

influenced by patterns of prey availability (and perhaps 

juxtaposition to nesting habitat), while other studies found 
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that prey quality or availability influence breeding success 

(Becker et al. 2007, Gutowsky et al. 2009, Norris et al. 2007).

Below, we summarize those recent analyses that evaluated the relative contributions of marine conditions and 

nesting habitat conditions to murrelet status and trend in the 

Plan area, and review the larger body of scientific information on the relationships between marine habitat conditions 

and murrelet biology throughout the species’ entire range.

To understand the murrelet’s marine habitat, it is 

helpful to introduce some key features of that habitat. First, 

most of the marine waters off the NWFP area are within 

the California Current system, the southward-moving 

surface current of colder water from the north Pacific. A 

key characteristic of the system is wind-driven upwelling 

of cooler and typically nutrient-rich waters to the surface 

in nearshore areas, particularly in spring and summer. This 

upwelling of nutrients results in increased productivity 

(Batchelder et al. 2002), and may be key to maintaining 

cold, productive marine conditions favorable to murrelets 

south of Washington state, in areas that would have warmer 

sea temperatures in the absence of the current system and 

upwelling (McShane et al. 2004).

The Puget Sound/Salish Sea region differs from elsewhere in the Plan area; it is not dominated by the California 

Current, and it has a more complex nearshore geography 

shaped by glaciation and with many islands, like many areas 

to the north, which creates local currents and tidal patterns 

that concentrate prey.

Marbled murrelet prey—Marbled murrelets prey on a 

wide range of marine fish and invertebrates (Burkett 1995, 

Nelson 1997). Murrelets appear to have a flexible foraging 

strategy, exploiting the prey species and foraging locations 

that maximize energy gain (Piatt et al. 2007). For example, murrelets selected less abundant, higher value Pacific 

herring (Clupea pallasii) at times over other, more abundant 

species (Ostrand et al. 2004), and sometimes foraged in 

deeper waters than normal, where local conditions created 

prey concentrations near breeding areas (Kuletz 2005).

Species composition of available prey changes across 

the murrelet’s range, perhaps most notably between the 

California Current system and the Alaska Current system, 

which dominates the species’ range north of the NWFP 

area. Common murrelet prey species include sand lance 

(Ammodytes hexapterus) and smelt (family Osmeridae), 

which are taken by murrelets in many areas, as are small 

herring and krill (Thysanoessa spp. and Euphausia spp.), 

where available. As one moves north, and particularly north 

of the California Current area, sand lance, capelin (Mallotus 

villosus), and small Pacific herring are frequent murrelet 

prey (Bishop et al. 2014, McShane et al. 2004, Piatt et al. 

2007); all three of these species are of moderate to high 

quality in terms of energy content (Anthony et al. 2000). 

Of these, capelin do not occur from the Olympic Peninsula southward, and sand lances become scarce in some 

areas to the south of the peninsula. Within the California 

Current, northern anchovy (Engraulis mordax) and, in 

spring, juvenile rockfish (Sebastes spp.) are dominant small 

schooling fish in nearshore waters (McShane et al. 2004), 

and are taken by murrelets (Burkett 1995). Although fish 

tend to dominate the murrelet diet and exclusively comprise 

prey brought to the nest, invertebrates, particularly krill, are 

taken at times by adults throughout the murrelet’s range.

Marine proxies for prey abundance in the NWFP area—

As part of the 20-year monitoring report and related work, 

the NWFP effectiveness monitoring program analyzed the 

relative influences of marine and terrestrial factors on murrelet distribution and population trends during the first two 

decades of the NWFP (Raphael et al. 2015, 2016b). Although 

the murrelet diet has been studied to the north of the Plan 

area, particularly in Alaska (summarized in McShane et al. 

2004 and Piatt et al. 2007), few studies have been conducted on the murrelet diet south of Canada, and monitoring 

data for murrelet prey species from waters off NWFP lands 

are equally sparse. For these reasons, Raphael et al. (2015, 

2016b) used physical and biological attributes of marine habitat as proxies for local prey abundance in their analyses. The 

attributes that the authors measured included chlorophyll 

“a” concentration in ocean surface waters and sea surface 

temperature, which have been used in comparable analyses 

by others (e.g., Ainley and Hyrenbach 2010, Hazen et al. 

2012), and are available at relatively fine temporal and spatial 

scales. The idea is that cooler water is rich in nutrients. This 

in turn leads to a more robust food chain, ultimately leading to a more robust supply of small fishes and invertebrates 
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that murrelet prey upon. Cooler waters are enriched with 

nutrients compared with warmer waters, and are frequently 

associated with upwelling. Chlorophyll “a” concentration has 

for decades been a proxy for phytoplankton abundance and 

primary productivity, and performs well in this role (Huot et 

al. 2007). In the northeast Pacific (Ware and Thomson 2005) 

and California Current (Reese and Brodeur 2006), chlorophyll “a” concentration was positively associated with fish 

abundance, as was phytoplankton abundance in the North 

Sea (Frederiksen et al. 2006). In the California Current, 

chlorophyll “a” peak abundance was a strong predictor of 

seabird abundance and hotspots of seabird density (Suryan et 

al. 2012). For these reasons, Raphael et al. (2015, 2016b) hypothesized that murrelet prey abundance would be positively 

associated with primary productivity.

Marbled murrelet prey availability is likely to be 

affected by broader Pacific Ocean conditions, including the 

Pacific Decadal Oscillation (PDO) (Mantua et al. 1997) and 

the El Niño Southern Oscillation (ENSO) (Trenberth 1997), 

which have widespread effects on marine productivity and 

food webs, as well as on seabird populations, including 

other diving seabirds in the California Current system (Ainley and Hyrenbach 2010). Therefore, the 20-year NWFP 

analyses also included factors to account for variability 

in PDO and ENSO conditions (Raphael et al. 2016b). The 

ENSO is a pattern of periodic changes (events), typically 

lasting about 9 to 18 months, that produce (1) El Niño events 

with increased sea surface temperatures and reduced coastal 

upwelling, and (2) La Niña events that result in colder, more 

nutrient-rich waters than usual (Mestas-Nunez and Miller 

2006, Schwing et al. 2002). The PDO represents long-term 

(20 to 30 years) climate variability in the north Pacific 

Ocean, in which there are observed warm and cool phases, 

or “regime shifts” with corresponding patterns of weaker or 

strong upwelling (Mantua et al. 1997). Later (see “Climate 

Change Considerations” below), we discuss potential effects 

of climate change on murrelet prey and these proxies. 

Associations with marine habitat and prey—Although 

prey and foraging habitat conditions differ across the 

murrelet’s wide range, murrelets forage and rest mostly in 

shallow nearshore waters associated with the continental 

shelf (Nelson 1997). Murrelets often use sheltered waters 

when available (Nelson 1997), but most of the coast in the 

Plan area (except for the Puget Sound area) lacks the complex structure and sheltered areas found farther north in the 

glaciated fjords and abundant islands of Alaska and British 

Columbia. In the Plan area, data from the at-sea work of the 

NWFP effectiveness monitoring program shows that most 

murrelet foraging during the breeding season occurs in 

water depths of 80 ft (24 m) or less, except for the San Juan 

Islands and northern Puget Sound, where murrelets used 

waters up to 130 ft (40 m) depth (Raphael et al. 2016b).

Analyses for the 20-year NWFP murrelet report examined variation in murrelet abundance in relation to dominant 

shoreline substrate within the Plan area, and found that 

murrelet abundance was greater offshore of fine- to medium-grained sand beaches and was also greater offshore 

of estuaries and marshes, compared to other substrates 

(Raphael et al. 2016b). In an earlier study of murrelet habitat 

use off southern Oregon, murrelets were most abundant 

near ocean bays, river mouths, sandy shores, and submarine 

canyons (Meyer and Miller 2002). Similarly, murrelet densities off British Columbia were highest over sandy substrate, 

near estuaries, and where waters are coolest (Burger 2002,, 

Piatt et al. 2007, Ronconi 2008, Yen et al. 2004). In a study 

at the southern end of the murrelet’s range near Santa Cruz, 

California, Becker and Beissinger (2006) found that foraging murrelets appeared to prefer cooler waters associated 

with areas of recent upwelling.

In their review of murrelet ecology, Piatt et al. (2007) 

concluded that physical and biological oceanographic 

processes that concentrate prey (such as upwellings and rip 

currents) have an important influence on where murrelets 

forage. Although that conclusion is largely based on work 

in Alaska and British Columbia (e.g., Burger 2002, Day 

and Nigro 2000, Kuletz 2005), it is supported by work in 

Washington, Oregon, and California (Ainley et al. 1995, 

Nelson 1997, Strong et al. 1995). This suggests that, at the 

finer scale, across their range, murrelets select foraging areas 

based on similar topographic and oceanic factors associated 

with higher prey densities in shallower waters. This pattern is 

consistent with the often strong positive relationship between 

forage fish abundance and the abundance of fish-eating birds 

(e.g., Durant et al. 2009, Furness and Tasker 2000).
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Changes in foraging habitat conditions—There is some 

information from analyses of stable isotopes in murrelet tissues indicating long-term declines in murrelet 

diet quality in portions of its range in central California 

(Becker and Beissinger 2006), the Salish Sea, including 

northern Washington (Gutowsky et al. 2009), and British 

Columbia (Norris et al. 2007). At least one of these studies suggested that murrelet foraging success along the 

Pacific Coast is sensitive to climate variability, and that 

cooler ocean waters and resulting prey conditions are 

associated with greater reproductive success (Becker et 

al. 2007). Further, though murrelets have flexible foraging and life history strategies that presumably evolved 

in an environment of varying prey conditions, there is 

evidence that declines in murrelet diet quality may have 

contributed to reduced murrelet reproductive success in 

the Salish Sea (Gutowsky et al. 2009), and that foraging 

flexibility in murrelets (Ronconi and Burger 2009) and 

other alcids (Schrimpf et al. 2012) may not be sufficient 

to avoid low reproductive success when environmental 

conditions are extremely poor. Adult survival in murrelets appears less vulnerable to poor forage conditions 

than does reproductive success (Beissinger and Peery 

2007, Peery et al. 2006a, Ronconi and Burger 2008), and 

Ronconi and Burger (2008) proposed that murrelets likely 

have a life history strategy in which adults do not initiate 

nesting, or abandon nesting attempts, to maximize their 

own survival when available forage is inadequate. Piatt et 

al. (2007) concluded that climate-related changes in marine ecosystems, in addition to human activities (logging, 

gill net bycatch, oil pollution), were the likely reasons for 

the wide-scale declines in murrelet populations in British 

Columbia and Alaska.

Environmental conditions, particularly El Niño events, 

have been shown to markedly reduce prey availability for 

some seabirds in California, leading to poor reproductive 

success (Ainley et al. 1995). Although El Niño events 

appear to reduce overall seabird prey availability, their 

effect on murrelets are not well known. Inner coastal 

waters in the Puget Sound and the Strait of Juan de Fuca, 

as well as estuarine areas along the outer coast, appear 

less influenced by El Niño conditions because of mixing 

and nutrients from sources other than outer coastal waters 

(USFWS 1997). In addition to ENSO variation, it is known 

that fish populations and zooplankton in the California 

Current generally do better during “cold” than in “warm” 

phases of the PDO, while in the more northerly Alaska 

Current, some fish populations such as salmon behave 

oppositely (Hallowed et al. 2001).

The 20-year NWFP analysis found only one forage-fish dataset from the Plan area of interest, and which 

spanned the period of that analysis (Raphael et al. 2016b). 

Those data provided abundance of forage fish from two 

transects located just north and south of the Columbia 

River. For this limited area, the authors found some 

evidence in the year-to-year variation of a positive relationship between forage fish and murrelet abundance, and 

concluded that direct measures of forage-fish abundance as 

predictors of murrelet abundance need additional investigation (Raphael et al. 2016b). 

Climate Change Considerations

The Intergovernmental Panel on Climate Change (IPCC) 

is a scientific body that was set up in 1988 by the World 

Meteorological Organization and United Nations Environment Program to inform policymakers about the causes of 

climate change, its potential environmental and socioeconomic consequences, and the adaptation and mitigation 

options to respond to it. In 2014, the IPCC published its 

Fifth Assessment Report, which is widely considered the 

most comprehensive compendium of information on actual 

and projected global climate change currently available. 

Although the extent of warming likely to occur is not 

known with certainty at this time, the IPCC has concluded 

that warming of the climate system is unequivocal: that 

the atmosphere and ocean have warmed, sea level has 

risen, and continued greenhouse gas emissions will cause 

further warming (IPCC 2014). Ocean warming accounts 

for more than 90 percent of the energy accumulated and 

stored in the climate system between 1971 and 2010 (IPCC 

2014). Although the report does not focus on changes at 

the scale of the NWFP, it did find with high confidence 

new evidence for decreasing spring snowpack in western 

North America.
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Climate change and terrestrial nesting habitat—

Although murrelets spend most of their time in the marine 

environment, murrelets require suitable forest cover for 

nesting. The U.S. Fish and Wildlife Service reviewed 

potential threats to murrelet nesting habitat in its last status 

review (USFWS 2009). The agency concluded that, based 

on climate model projections, the future conditions of 

forests where murrelets nest will be largely unfavorable 

for maintaining current forest structure and composition. 

Projections suggest that increases in annual temperature 

changes within the range of the murrelet will be greatest 

in the summer and lowest in the spring, but predicted that 

temperature increases near the coast will be generally lower 

than in the rest of the Plan area (Dalton et al 2013). Already 

in the Pacific Northwest, tree mortality rates in unmanaged 

old forests have increased in recent decades at a rate equivalent to doubling over 17 years (van Mantgem et al. 2009), 

a change the authors suggested was likely due, at least in 

part, to documented regional warming and drought stress 

associated with climate change. With respect to drought 

stress, Johnstone and Dawson (2010) found evidence of a 33 

percent reduction in fog frequency over the past century in 

the coast redwood forest zone of northern California, which 

includes most of the nesting habitat in conservation zone 5 

and in the California portion of conservation zone 4. Based 

on tree physiological data, they suggested that redwood and 

other western coastal forest ecosystems may experience 

increasing drought stress as a result of reduced summer fog 

and greater evaporative demand.

During the next 20 to 40 years, climate projections 

for the Pacific Northwest indicate likely decreases in 

Douglas-fir growth from drier summers (Littell et al 2010). 

Heat extremes and heavy precipitation events are likely to 

become more frequent (Loehman and Anderson 2009). With 

these changes, the potential exists for increased fire frequency and severity, even in the coastal forests where murrelets nest (Millar et al. 2006). In North America broadly 

(Dale et al. 2001) and the Pacific Northwest specifically 

(Kliejunas et al. 2009; Littell et al. 2009; Mote et al. 2003, 

2010), climate changes may also alter forest ecosystems via 

the frequency, intensity, duration, and timing of other disturbance factors such as drought, introduced species, insect 

and pathogen outbreaks, windstorms, ice storms, landslides, 

and flooding. Evidence for an increased role of fire within 

the range of the murrelet is mixed, with some models 

projecting increases and others projecting decreases (see 

chapter 2, “Climate,” and chapter 3, “Vegetation Change”), 

but the historical occurrence of large, high-severity fires 

suggests the potential for losses in nesting habitat if fires 

do occur (Agee 1993). Overall, the evidence is substantial 

that climate change will result in changes to forest habitats 

where murrelets nest. The magnitude of those changes 

is less known, as is how nesting murrelets and murrelet 

populations will respond to forest changes. However, to 

the extent that changes such as increased tree mortality, 

decrease in canopy epiphytes, and increased severity and 

frequency of fires reduce the number of potential nest trees, 

impacts to murrelets appear likely to be negative.

Climate change and marine habitat—

In addition to influencing the quality and abundance of 

nesting habitat, as discussed above climate change is likely 

to result in changes to the murrelet’s marine environment, 

with effects on murrelet food resources the most likely 

mechanism. Given the large body of climate change literature, we focus our review here on such potential effects on 

prey resources. 

The U.S. Fish and Wildlife Service reviewed the potential effects of climate change on murrelets south of Canada, 

and concluded that—although predicting climate change 

effects on marine resources is complex and has many uncertainties—taken as a whole, the evidence from models and 

other sources suggested that few changes are likely to benefit 

murrelets, with many more having the potential for neutral 

or adverse effects (USFWS 2009). The same review found 

it most likely that the murrelet prey base will be adversely 

affected to some extent by climate change, and noted that 

although seabirds generally have life-history strategies 

adapted to variable marine environments, ongoing and future 

climate change could present changes of a rapidity and scope 

outside the adaptive range of murrelets (USFWS 2009). 

Marine changes already observed may be attributable to 

climate change. El Niño events have become more frequent, 

persistent, and intense during the last decades of the 20th

century (Snyder et al. 2003). There is general agreement that 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

sea surface temperatures will increase as a result of climate 

change, with evidence that they have already increased in 

murrelet marine habitat off the NWFP area by 0.5 to 1.0 

°C (about 1 to 2 °F) over the last half century, both in the 

California Current system (Di Lorenzo et al. 2005) and in 

the Strait of Juan de Fuca (Rucklehaus and McClure 2007). 

In the murrelet’s nearshore environment, upwelling of cold 

waters may moderate some level of sea-surface-temperature changes, but differences in the timing, intensity, and 

duration of upwelling can affect productivity, resulting in 

considerable uncertainty regarding the ultimate effects of 

marine changes on murrelet foraging conditions. Climate 

models show inconsistent projections for the future of 

coastal upwelling in the Pacific Northwest (Melillo et al. 

2014). Illustrating the complexities of making such projections, Sydeman and others (2014) conducted a meta-analysis 

of the literature on wind intensification in coastal upwelling 

marine systems over the prior six decades. They found 

support for wind intensification in the California Current 

system and noted that this could increase nutrient input 

and benefit marine populations if primary production is 

nutrient limited. However, they emphasized the complexity 

of forecasting the consequences of wind intensification in 

coastal ecosystems because the ecological effects are likely 

sensitive to diverse factors including phenology of upwelling-favorable winds, patterns of nutrient transport offshore, 

differing responses of food web species, and potential for 

increased stratification resulting from increased water 

temperatures (Sydeman et al. 2014).

Climate change is anticipated to result in sea-level rise 

and a decrease in the pH of marine waters, with unknown 

effects in both the California Current system and Puget 

Sound. Increasing acidification of marine waters caused by 

increased absorption of carbon dioxide from the atmosphere 

may have significant impacts on marine food webs. This 

is because acidification reduces the availability of calcium 

ions for the formation of calcium carbonate, an essential 

component of the skeletons of marine plankton, shellfish, 

and other organisms (Doney et al. 2012, Feely et al. 2008). 

In the Pacific Northwest, which includes Oregon and 

Washington, projected marine changes include increasing 

but variable acidity, more increases in surface water 

temperature, and possible changes in storminess (Melillo 

et al. 2014). In Puget Sound, changes in the timing and 

amount of freshwater inflow may produce fresher waters 

during winter and saltier waters during summer, resulting 

in stronger stratification in winter and weaker stratification 

in the summer (Rucklehaus and McClure 2007). 

Although physical changes to the marine environment 

appear likely, much remains to be learned about the magnitude, geographic extent, and temporal and spatial patterns 

of change, and their effects on murrelets (USFWS 2009). 

However, we do know that climate variability can strongly 

influence the foraging and reproductive success of seabirds, 

including the murrelet (Becker et al. 2007, Grémillet and 

Boulinier 2009, Norris et al. 2007). Shifts in the intensity 

of upwelling influence nutrient availability and primary productivity in coastal waters, with cascading effects at higher 

trophic levels (Thayer and Sydeman 2007). For example, 

El Niño events have been associated with poor seabird 

survival and recruitment in the eastern Pacific (Bertram et 

al. 2005, Hodder and Graybill 1985). Some species respond 

more strongly to either the ENSO or PDO phases, but not 

both (Black et al. 2011, Sydeman et al. 2009), and the local 

effect of regional patterns such as the ENSO and PDO is 

modified by undersea topography, trophic interactions, 

bird movements to track prey, and food web impacts from 

commercial fisheries harvest (Doney et al. 2012). Although 

many seabirds have flexible foraging strategies, chronic 

food scarcity can compromise long-term breeding success 

(Cury et al. 2011) and reduce adult survival and fecundity 

(Kitaysky et al. 2010). 

With respect to foraging strategies, Lorenz et al. 

(2017) reported on marbled murrelet movements during the 

breeding season, based on the radio-tracking of 157 birds 

between 2004 and 2008 in northwestern Washington. The 

authors did not find oceanographic conditions to substantively explain variation in movements of foraging murrelets. 

They did find low breeding propensity, large marine ranges, 

and long nest-sea commutes compared to studies elsewhere 

in the murrelet’s range, and hypothesized that this may 

indicate that marine habitat in their study area was lower 

quality compared to elsewhere in the species’ range. They 

also found, unexpectedly, that a recent widespread and 
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strong delay of the onset of spring upwelling in the California Current in 2005 did not appear to substantially affect 

murrelet movements or breeding propensity. This finding 

differs from that of Ronconi and Burger (2008), who linked 

reduced murrelet breeding productivity in southwestern 

British Columbia to the 2005 upwelling delay.

If recent warm-water events are an indicator of future 

effects of increased sea-surface temperatures, the murrelet 

prey base could be negatively affected. Studies of other 

diving seabirds such as Cassin’s auklets (Sydeman et al. 

2006), historical versus recent murrelet diet (Becker and 

Beissinger 2006), and recent annual variations in murrelet 

reproductive success (Becker et al. 2007) suggest that 

warmer coastal waters tend to adversely affect prey quality 

and result in lowered reproduction.

Research Needs, Uncertainties, 

Information Gaps, and Limitations

The challenges of accurately sampling such a mobile and 

patchily distributed species result in fairly large uncertainty 

around each year’s density and population estimates, as 

seen in the confidence intervals. The NWFP population 

monitoring data provide 15 years from which to assess population trends and, based on the observed sampling error, 

power analysis indicates that 15 or more years of population 

estimates are required to detect an annual rate of decline of 

2 percent (Falxa et al. 2016). Even with these constraints, 

the population monitoring data for 2000 through 2015 

indicate a marked decline in Washington, no evidence of 

a trend in Oregon, and an increasing trend in California. 

Additional years of population monitoring will increase the 

power to detect ongoing trends, such as those of 2 percent or 

less per year. Conversely, population trajectories can change 

over longer monitoring periods, resulting in nonlinear 

trends, which adds temporal variability and complexity in 

describing trends. 

A major source of uncertainty is whether the murrelet 

population is closed or open. That is, existing population 

models (such as McShane et al. 2004) assume there is little 

or no recruitment of adults or juveniles from outside the 

study population, and little or no emigration out of the 

study population. For example, the local population may be 

declining but is being supplemented by immigrants, perhaps 

from Alaska or British Columbia, where murrelets are more 

numerous. Recruitment of birds from outside the local range 

has been proposed as the most likely explanation for the 

seemingly stable population estimates in central California 

(Peery et al. 2006a), despite demographic models that 

predict a decline (Peery 2004). The open population hypothesis, at least for their range from southern Alaska through 

northern California, is supported by genetic analyses (Piatt 

et al. 2007), and recent studies showing long-distance 

movements of murrelets tracked by satellite (e.g., Bertram 

et al. 2015b). However, it is not known if movements of 

murrelets are sufficient to affect population estimates and 

trends within the NWFP area.

Future population trends are difficult to predict because 

of uncertainties in the timing and extent of risk factors. 

Catastrophic loss of nesting habitat from uncharacteristically 

severe wildfire is an ever-present risk. Among factors other 

than habitat loss, murrelets at sea are subject to risk from 

large oil spills at sea (USFWS 1997); oil spills killed an 

estimated 872 to 2,024 murrelets between 1977 and 2008 

in California, Oregon, and Washington (USFWS 2009). A 

recent review concluded that spills continue to be a threat and 

can cause severe localized impacts owing to direct mortality 

from oiling, as well as reductions in reproductive success 

through changes in prey base, marine habitat, and disturbance (USFWS 2009). Gill net mortality was cited as a factor 

for listing the murrelet in 1992. Since then, this risk has been 

substantially reduced in the NWFP area, with no mortality in 

California and Oregon because of gill net bans, and reduced 

mortality in Washington as a result of measures implemented 

to reduce seabird mortality (McShane et al. 2004, USFWS 

2009). Gill net mortality remains a threat to the north in British Columbia and Alaska, however, and could be a risk to the 

NWFP murrelet population to the extent that murrelets move 

between waters off the Plan area and marine waters to the 

north. Future energy development, both at sea and on land, 

could also pose a local threat to murrelets, such as potential 

collisions with wind turbines (USFWS 2009).

Changes in prey base present risks as well. As discussed earlier, studies have found evidence that murrelet 

reproductive success is influenced by prey availability, and 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

future prey resources can be affected by fishing as well 

as changes in ocean conditions, including those linked to 

climate change. In some other seabird species (e.g., Montevecchi and Myers 1997), changes in ocean currents can 

have profound effects on forage fish, leading to starvation 

in addition to breeding inhibition. For murrelets, one study 

found that adult survival appears less vulnerable to prey 

shortages (Ronconi and Burger 2008). To date, disease has 

not been found to be a significant threat to murrelets (Piatt 

et al. 2007, USFWS 2009), but pathogens new to the region 

could cause direct mortality to nesting birds, and could also 

have indirect effects (USFWS 2009). For example, the West 

Nile virus is documented to kill jays, crows, and ravens, and 

if mortality of these species resulted in appreciable reductions in their densities, this might increase nest success of 

murrelets by reducing nest depredation.

Raphael et al. (2016a) describe sources of uncertainty 

in estimating the amount and distribution of nesting habitat 

of the murrelet. But one additional source warrants further 

mention. Because murrelet nesting behavior is so cryptic, 

biologists have found very few actual nests of the species. 

To supplement actual nesting observations, biologists 

rely on locations of “occupied behaviors” to infer nesting 

activity. Occupied behaviors are observations of murrelets 

flying into the canopy, circling very close above the 

canopy, or landing in trees. These behaviors are typically 

associated with nesting, but some sites where occupied 

behaviors are observed may not be true nest sites. To the 

extent that false positives may be included in the murrelet 

database used to build models, these models may be less 

accurate than if all locations were based on verified nests 

(Plissner et al. 2015). A more reliable modeling solution 

would be to conduct intensive research to identify more 

known nest sites across a broad sampling of regions within 

the NWFP area, then build models exclusively from 

training sites that represent actual murrelet nests. Such 

intensive surveys would also help our understanding of 

spacing and density of nesting activity in relation to forest 

stand characteristics. 

Some uncertainty also exists in the distance that 

murrelets fly inland to nest and how that varies within the 

Plan area. The Forest Ecosystem Management Assessment 

Team designated two inland zones within the area in which 

murrelets nest: Inland zone 1 formed the area closer to 

the marine environment, and inland zone 2 was further 

inland, extending to the eastern boundary of the species’ 

nesting range (see fig. 5-4). Nesting was assumed to occur 

mostly in zone 1. Recent survey-based studies in some 

areas have led to local contractions of zone 2, especially 

in northern California and southern Oregon (Alegria et al. 

2002, Hunter et al. 1998, Schmidt et al. 2000). Agencies 

in those areas have redefined the eastern boundary of the 

area in which surveys for murrelets are required prior to 

timber harvest, bringing it farther to the west to match 

study results. This revised boundary has not been formally 

implemented in the NWFP agency maps; to date this 

revision applies only to survey requirements for management units where the studies were conducted. This strategy 

adds uncertainty in the calculation of amounts of nesting 

habitat to the extent that acres classified as nesting habitat 

may actually fall outside the species’ true breeding range. 

This uncertainty is reduced in the most recent analysis by 

the NWFP monitoring program, which did not model or 

estimate suitable murrelet nesting habitat in inland zone 

2 in California or Oregon; this is because of the lack of 

inland zone nest sites in those states with which to train 

the nest habitat models (Raphael et al. 2016a).

We found no studies documenting the response of 

murrelets to silvicultural activities designed to accelerate 

expression of mature forest conditions, and this remains 

an area in which much further research is needed. Foresters have conducted studies using experimental thinning 

prescriptions, but none of these has incorporated responses 

of murrelets to these treatments. 

Perhaps the most important area of uncertainty is the 

relationship between murrelet population size and trend 

and the influences of either amount and trend of nesting 

habitat versus variation and trends in ocean conditions that 

affect foraging habitat. The studies that we summarize 

point toward nesting habitat as the primary driver, but all 

these studies concede that relationships are correlational. 

Cause-effect relationships have not been established, so 

further work will be needed to confirm whether these 

correlations reflect true underlying causes.
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Conclusions and Management 

Considerations

Are NWFP Assumptions Still Valid?

Nesting habitat status and trend—

The NWFP has played a pivotal role in the fate of murrelet 

nesting habitat on federal lands. The Plan has been highly 

successful in conserving existing murrelet nesting habitat, 

and little nesting habitat has been lost to timber harvest 

on federal lands. Some loss of nesting habitat, especially 

in federal reserves, was caused by fire. Loss of murrelet 

nesting habitat to catastrophic events will always be a risk, 

and such losses were expected. The NWFP has less control 

over the risk of such losses, except to the extent that active 

management in fire-prone areas might reduce the risk of 

fire in younger forests in proximity to murrelet nesting 

habitat, and by reducing vegetation that could transmit fire 

to the canopy of murrelet nesting trees, such as in forests 

with scattered nest trees within younger forest. One caution 

should be recognized: managing forest cover to reduce 

fire risk could also lead to better habitat for corvids (nest 

predators); silvicultural practices near suitable murrelet 

nesting habitat may need to be fine-tuned to ensure they 

do not inadvertently impair nesting success of murrelets 

by increasing the rate of nest depredation. In addition to 

active fire management, another area for potential reduction 

of nesting habitat loss on federal lands is management to 

reduce the risk of windthrow associated with the creation 

of hard edges. In this case, the greatest potential benefit to 

murrelets would be in (1) creating and maintaining forested 

buffers adjacent to existing known and suitable murrelet 

nesting habitat, and (2) developing nesting habitat within 

reserves plus in adjacent buffers.

The fate of nesting habitat on nonfederal lands is 

beyond the scope of the NWFP; 67 percent of habitatcapable forest is in nonfederal ownership, as is 41 percent of 

suitable murrelet nesting habitat. The rate of loss of suitable 

nesting habitat on nonfederal lands (1.5 percent per year) 

has been far more rapid than on federal lands (0.1 percent 

per year).

The requirement for preproject surveys on federal land 

was assumed to prevent the loss of any occupied sites from 

timber harvest. We are not able to test this assumption 

because we have no way to assess whether sites on federal 

land were classified as unoccupied when they might actually 

have been occupied. Occupied behaviors are not observed 

at every visit to a site; a finite likelihood exists of failing 

to detect occupied behaviors even if the site is occupied. 

The protocol used to determine site occupancy (Evans 

Mack et al. 2003) sets the numbers of visits required to 

have a high likelihood (set at 0.95) of observing occupied 

behavior at an occupied site. Under this protocol, a 5-percent chance of failing to detect occupied behavior exists, 

so a small number of sites might be mistakenly classified 

as unoccupied and released for timber harvest. The Pacific 

Seabird Group (a society of professional seabird researchers 

and managers dedicated to the study and conservation of 

seabirds) is considering a revision of the current survey 

protocol (Evans Mack et al. 2003), which would use the best 

available science to ensure that the 5 percent criterion is met 

by the protocol. We can say that sites classified as occupied 

were, in fact, set aside and managed as LSR3 reserves. 

There apparently have been some differences among NWFP 

management units in applying the NWFP standards and 

guidelines to occupied sites, with some reserves including 

all forest within a 0.5-mi (0.8 km) radius (which provides a 

larger block and more protection), and others including only 

contiguous forest within the radius that is existing suitable 

or recruitment murrelet nesting habitat (USDI BLM 2016).

Population status and trends—

Murrelet populations are affected by a variety of factors, 

only some of which are under the NWFP’s direct influence. The Plan most directly affects populations through 

its provisions for conservation and restoration of nesting 

habitat, but even then its influence extends only to federal 

lands. Although NWFP forest management may have 

minor or local effects on marine habitats, such as through 

altered input of sediment and coarse wood, overall the Plan 

has little to no influence on marine conditions affecting 

murrelet populations (including marine food sources) or on 

sources of mortality at sea, such as oil spills and gillnetting. 

This makes it more difficult to relate changes in murrelet 

populations to land management under the NWFP. With 

the NWFP conserving nesting habitat as expected, murrelet 

populations could still fall because of adverse marine 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

conditions or because of nesting habitat loss on nonfederal 

lands. Despite this uncertainty, circumstantial evidence 

suggests that inland nesting habitat conditions are the major 

driver setting murrelet population size at this time. This 

point is illustrated in Raphael et al. (2016b), in which the 

authors found a positive correlation with the total amount of 

nesting habitat and size of adjacent murrelet population for 

segments of the murrelet range. In addition, Raphael et al. 

(2015, 2016b) constructed a model to assess relative contribution of marine and terrestrial habitat attributes toward 

abundance and trend of murrelets throughout their range in 

Washington, Oregon, and California south to San Francisco 

Bay. In that model, amount and pattern of nesting habitat 

made the strongest contribution to predictions of spatial 

distribution and temporal trends of murrelet populations 

at sea; marine factors such as sea surface temperature and 

chlorophyll, as well as ENSO and PDO indices, had little 

effect. Murrelet nesting habitat seems to be the primary 

driver of murrelet population status and trend, at least in 

recent decades, but that relationship has not been tested 

empirically and a cause-effect relationship has not been 

established. Raphael et al (2016b) suggest that one test 

of this relationship will be whether murrelet populations 

are observed to increase when the net amount of suitable 

nesting habitat increases at some point in the future.

The fundamental assumptions of the NWFP were 

that the rate of loss of murrelet nesting habitat in reserves 

would slow or stop, and that unsuitable forest cover types 

would recover. Available data support this assumption and 

show that rates of loss on NWFP lands are low, and that 

forest stands in reserves are on a trajectory toward higher 

nesting habitat suitability. Conservation and restoration of 

murrelet nesting habitat are essential to population viability of the species.

Although federal protection of nesting habitat is essential to murrelet viability, it may not be sufficient given 

the cumulative effects of other influences on population 

viability. Research has documented that murrelet viability 

depends on a variety of factors, many of which (e.g., 

supply of ocean prey) are not under the control or influence 

of the NWFP. Nesting habitat loss on nonfederal lands, 

marine conditions, and threats from disease, oil spills, and 

gillnetting could reduce the likelihood of population viability despite the habitat protections built into the NWFP. 

Past timber harvest was hypothesized to have lingering 

effects on murrelet carrying capacity and nesting success. 

We are aware of no new data to challenge this hypothesis. 

Recent research shows that murrelet population size is 

reduced as nesting habitat is lost, and that birds do not 

pack into remaining suitable nesting habitat (Burger 2001, 

Raphael et al. 2002a). 

A major premise of the NWFP is that large reserves 

will support more murrelets, eventually leading to stationary or increasing populations. Because of the long period 

of time required to recruit new nesting habitat in reserves, 

thus forming larger blocks of nesting habitat, it is too soon 

to fully evaluate this premise, but trends on Forest Service 

lands in the Oregon Coast Range suggest that this may be 

starting to occur there. 

Fahrig (1997) suggested that habitat loss tends to 

far outweigh the spatial configuration of habitat (fragmentation) as a risk to species. Although habitat loss and 

limitation appear to best explain the observed patterns 

of murrelet distribution and population trends in the Plan 

area, spatial configuration of nesting habitat is also a factor. 

As discussed above (see “Landscape-level relationships 

between nesting habitat and populations”), fragmentation 

of nesting habitat and the associated greater amounts of 

habitat edge may increase the risk of breeding failure due 

to nest predation. 

Also, as summarized above, nest depredation seems 

to be a major limiting factor on murrelet populations, and 

nesting habitat configuration may affect predation risk. 

More than half of known murrelet nests whose fate has 

been determined failed because eggs or chicks were lost to 

predators, primarily jays, crows, and ravens (Manley and 

Nelson 1999, and other papers cited above). The relationship 

of predation risk and forest configuration appears to be 

complex. Increased edge resulting from forest fragmentation 

appears to have negative effects on murrelets. For example, 

some research has found higher densities of nest predators 

near edges (primarily jays), particularly where edges are near 

human development such as campgrounds (Goldenberg 2013, 

Marzluff and Neatherlin 2006) or include berry-producing 
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GENERAL TECHNICAL REPORT PNW-GTR-966

plants (Masselink 2001). Other research suggests that predator numbers are high in old-growth forests with complex 

forest structure, such as those expected to develop in NWFP 

reserves, but lower in mature forests with simpler structure 

(Marzluff et al. 2000, Raphael et al. 2002b). At the plot scale 

(90 to 260 ac), one study found predator densities higher and 

nest success lower in plots with a variety of tree ages intermixed with young tree/brush habitats (Luginbuhl et al. 2001). 

The relationship between nest predator density and predation 

risk may also depend on the scale of observation. Luginbuhl 

and others (2001) found that nest predation risk was much 

better predicted by corvid abundance at the landscape level (2 

to 20 mi2 [5 to 52 km2] scale) than at a finer scale (60 to 120 

ac [24 to 49 km2]), likely because of the large home range of 

some corvids (ravens, crows).

Forest fragmentation will decline as young patches 

within reserves mature, creating more contiguous canopy 

cover, and where rates of nest predation would decrease 

as forests became less fragmented. Murrelet populations 

may not grow at the rate predicted from recovery of nesting 

habitat in reserves because nest depredation could suppress 

successful reproduction. We lack understanding of the full 

suite of factors that affect nest success, which increases 

uncertainty about the relations between amounts of nesting 

habitat and murrelet populations. 

Research indicates that maintaining older, maturing 

forest adjacent to nesting habitat also reduces predation 

risk (table 5-6). Taken as a whole, research to date suggests 

that, apart from increasing the amount of nesting habitat 

and reducing its fragmentation, managing forest structure 

to reduce nest predation risk should be approached with 

consideration of local factors that might affect predator 

densities (e.g., overstory thinning that might result in 

increased abundance of berry-producing early-seral shrubs 

that attract corvids).

Although habitat loss and fragmentation lead as factors 

influencing murrelet numbers and trends, birds in the NWFP 

area are also affected by marine factors. Murrelets are subject 

to risk from large oil spills at sea, which killed an estimated 

872 to 2,024 murrelets between 1977 and 2008 in California, 

Oregon, and Washington, and continue to be a threat, as they 

can cause severe localized impacts such as direct mortality 

through oiling, as well as other less direct effects (USFWS 

2009). Gill net mortality in the Plan area has been reduced 

substantially since 1994, with California and Oregon banning 

gill net use near shore, and measures taken in Washington 

to reduce seabird mortality (McShane et al. 2004, USFWS 

2009). As discussed above (“Marine Habitat,” “Changes in 

foraging habitat conditions,” and “Climate Change Considerations”), murrelet reproductive success is influenced by prey 

quality and availability, which can be affected by fishing as 

well as changes in ocean conditions, including those linked to 

climate change. Future energy development, both at sea and 

on land, could also pose a local threat to murrelets, such as 

potential collisions with wind turbines (USFWS 2009).

Cumulative effects—

Wildlife population trends reflect the cumulative effects of 

multiple interacting factors. Nesting habitat conditions on 

federal lands are but one of those factors, albeit the one over 

which the NWFP has the most direct influence. Monitoring 

both nesting habitat trends and population trends is of 

value: monitoring nesting habitat trends tells managers how 

well the Plan is meeting its primary objectives; monitoring 

population trends tells managers if the NWFP is having 

the desired effects. Ideally, population trends will track 

nesting habitat trends, but we may observe diverging trends. 

In such cases, we can dig deeper to discover whether our 

understanding of nesting habitat relationships is mistaken 

or whether other, perhaps unmeasured, factors are driving 

population trends. Research to date, as noted above, does 

support the idea that population trends track nesting habitat 

trends, but the evidence is still based on correlations and 

has not established cause-effect relationships.

Carrying capacity is a measure of the potential population size that can be supported by a given amount and 

distribution of suitable nesting habitat. The actual population may be lower than the carrying capacity owing to a 

variety of other factors such as hostile weather, interactions 

with other species, nesting habitat conditions outside of the 

planning area, disease, or other factors that might depress 

a population. Observing a declining population in the face 

of habitat conservation does not mean that habitat is not 

important or that habitat conservation is not important. It 

means we have to look at options to manage some of the 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

other factors that might be driving the population trend. 

Until we have more robust models of wildlife habitat 

relationships, which include these other factors, continued 

monitoring of both population and habitat trends will be 

important to evaluate how well the NWFP is meeting its 

intended objectives.

Efficacy of large reserves for murrelet conservation—

A central tenet of the NWFP was that the system of large, 

late-successional reserves would largely suffice to provide 

for species and biodiversity components associated with 

late-successional and old-growth forest ecosystems. We 

have found that, to an extent, this is true with respect to 

murrelets. However, the degree to which late-successional 

reserves—along with the set of other NWFP land allocations (e.g., riparian reserves in matrix lands)—suffice 

differs considerably by species. Our review has highlighted 

the importance of large contiguous blocks of nesting habitat 

in meeting the nesting needs of the murrelet, and reserves 

seem an essential way to create such landscapes.

One of the management dilemmas is that optimal 

habitat conditions differ among species. Creating shrubby 

foraging habitat will be good for the northern spotted owl 

in the southern parts of its range, but such habitat will also 

be good for jays and crows, which depredate nests of the 

murrelet. In this case, what is good for the owl may be bad 

for the murrelet (see chapter 12 for further discussion of 

interactions among NWFP goals and objectives). 

Management Considerations

Some key points emerge from this synthesis:

• Maintaining and increasing the area and cohesion 

(creating larger blocks) of suitable nesting-habitat 

area on federal lands will likely contribute to stabilizing and eventually recovering murrelet populations. Within NWFP lands, the current NWFP 

reserve system (including riparian buffers and other 

set-asides) appears well designed to accomplish 

this. Because it can take many decades for murrelet 

nesting habitat to develop, protection of existing 

habitat for the next several decades will continue to 

be key to minimizing habitat losses, both within and 

outside of reserves.

• Defining the inland limit of the murrelet nesting 

range will require additional survey work and a synthesis of existing observations. A refined range will 

better meet management objectives and avoid problems with managing for murrelets in areas where 

none are really expected to exist.

• Conservation of existing nesting habitat on federal 

lands may not be sufficient to conserve murrelet 

populations in the short term. Contributions from 

nonfederal lands may help the NWFP or its successor to achieve objectives for the murrelet, and the 

larger goal of murrelet conservation and recovery. 

This might be approached by collaborative programs 

to increase murrelet conservation on nonfederal 

lands, particularly those adjacent to NWFP lands, 

and in key areas (such as southwest Washington and 

northwest Oregon) where few federal reserves exist.

• Restoration of old-forest/murrelet nesting habitat 

in reserves may be accelerated by active management toward that end. Active management actions 

could include thinning in plantations to accelerate 

growth of potential nest trees and development 

of nesting platforms, but care will be needed to 

prevent simultaneously increasing numbers of nest 

predators attracted to more diverse understory 

conditions. Moreover, such management should 

also be careful to not increase the suitability of 

older forests to harbor barred owls (Strix varia), 

which may prey on murrelets and also reduce forest 

suitability for northern spotted owls (see chapter 4). Development and implementation of forest 

management practices that protect (short term) 

and develop (long term, e.g., over many decades) 

suitable murrelet nesting habitat on NWFP lands 

within the murrelet range would be beneficial in 

recovering murrelet populations (see chapter 3 for 

examples of restoration treatments).

• To guide management and increase its effectiveness 

in achieving nesting habitat expansion, modeling 

tools are needed to help forecast site-specific future 

nesting habitat development and structural characteristics of potential murrelet nesting habitat.
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GENERAL TECHNICAL REPORT PNW-GTR-966

• Restoration in plantations and younger natural forests can benefit murrelets by incorporating an understanding of relations among stand shape, extent of 

higher-contrast edges, and populations of potential nest predators, including corvids. Proximity 

of nest and occupied habitat should be considered. 

Treatments that consider risk to existing suitable 

nesting habitat along exposed edges from windthrow 

would also contribute to conservation of existing 

nesting habitat.

• Forest planning and management can positively 

affect murrelet status by managing human recreation 

activities that might promote murrelet nest predator 

populations (e.g., ravens, crows, and jays in campgrounds). The greatest benefit would be expected 

in areas within and near existing and developing 

murrelet nesting habitat. Implementing education 

programs, limiting garbage, and controlling predators could have positive effects.

• Future management and design of reserves will 

benefit from accounting for climate change, including increased risks to murrelet nesting habitat from 

fire and other natural disturbances. Boundaries of 

reserves (including making them larger) may be 

reconsidered if revised boundaries might better 

conserve nesting habitat in the face of anticipated 

effects of climate change. 

• Maintaining a broad distribution of large nesting 

habitat blocks over the NWFP landscape will likely 

help to minimizing the risk to the population from 

nesting habitat loss to fire, wind or other disturbance agents.

The NWFP remains the boldest effort ever undertaken 

by federal agencies to meet large-scale biodiversity objectives. The Plan had a short-term objective for murrelets: 

conserve much of the best remaining nesting habitat. The 

NWFP has been very successful in meeting this objective. 

The NWFP also has a long-term objective: create a system 

of reserves containing desired sizes and distributions of 

large blocks for suitable nesting habitat. Evidence suggests 

that nesting habitat trends on federal lands are on course 

toward this objective, but many more decades will be 

needed to observe whether the Plan is successful in achieving its goal to stabilize and increase murrelet populations 

by maintaining and increasing nesting habitat. We have 

shown that the NWFP has been remarkably successful 

in conserving nesting habitat over its first 20 years of 

implementation, but much work remains. Murrelet numbers 

continue to decline in the northern portion of the Plan area. 

Assuming no large fires, we believe that the current decline 

in amount of murrelet nesting habitat will reverse on federal 

lands, leading to a net increase in the amount of nesting 

habitat, and that murrelet populations should also increase 

in response. How many decades before this reversal in trend 

occurs is unknown, but at-sea monitoring suggests that the 

first step of possible population stabilization may be occurring in the southern Plan area. Lastly, climate change has 

emerged as an external force that may affect future murrelet 

populations, their nesting habitat, and, in particular, food 

resources for murrelets.

Acknowledgments

This work draws heavily on results of the 20-year effectiveness monitoring program, and we are indebted to the many 

team members and field assistants who contributed to that 

program for results presented herein. Portions of this document were derived from a report submitted to and funded 

by Environment Canada as part of Canada’s Marbled Murrelet Recovery Strategy, and that material is included with 

permission of that agency. We thank Kim Mellen-McLean 

and Ann Poopatanapong for comments on an earlier draft. 

We also thank the five anonymous peer reviewers whose 

comments helped improve the earlier draft. This work was 

funded by the U.S. Fish and Wildlife Service and the U.S. 

Forest Service Pacific Northwest Research Station.

Metric Equivalents

When you know: Multiply by: To find:

Feet (ft) 0.3049 Meters

Miles (mi) 1.61 Kilometers 

Acres (ac) 0.4049 Hectares 

Square miles (mi2

) 2.59 Square kilometers
339

Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

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Scientific and common names of plant species identified in this report

Scientific name Common name

Abies amabilis (Douglas ex Loudon) Douglas ex Forbes Pacific silver fir

Abies concolor (Gord. & Glend.) Lindl. ex Hildebr. White fir

Abies grandis (Douglas ex D. Don) Lindl. Grand fir

Abies lasiocarpa (Hook.) Nutt. Subalpine pine

Abies magnifica A. Murray bis California red fir

Abies procera Rehder Noble fir

Acer circinatum Pursh Vine maple

Acer macrophyllum Pursh Bigleaf maple

Achlys triphylla (Sm.) DC. Sweet after death

Adenocaulon bicolor Hook. American trailplant

Alliaria petiolata (M. Bieb.) Cavara & Grande Garlic mustard

Alnus rubra Bong. Red alder

Amelanchier alnifolia (Nutt.) Nutt. ex M. Roem. Saskatoon serviceberry

Anemone oregana A. Gray Blue windflower

Apocynum cannabinum L. Dogbane

Arbutus menziesii Pursh) Madrone

Arceuthobium M. Bieb. Dwarf mistletoe

Arceuthobium occidentale Engelm. Gray pine dwarf mistletoe

Arceuthobium tsugense Rosendahl Hemlock dwarf mistletoe

Arctostaphylos nevadensis A. Gray Pinemat manzanita

Brachypodium sylvaticum (Huds.) P. Beauv. False brome

Brodiaea coronaria (Salisb.) Engl. Cluster-lilies

Callitropsis nootkatensis (D. Don) Oerst. ex D.P. Little Alaska yellow-cedar

Calocedrus decurrens (Torr.) Florin Incense cedar

Cannabis L. Marijuana 

Carex barbarae Dewey and C. obnupta L.H. Bailey Sedges

Centaurea solstitialis L. Yellow starthistle

Chamaecyparis lawsoniana (A. Murray bis) Parl. Port Orford cedar

Chimaphila menziesii (R. Br. ex D. Don) Spreng. Little prince’s pine

Chimaphila umbellata (L.) W.P.C. Barton Pipsissewa

Clematis vitalba L. Old man’s beard

Clintonia uniflora Menzies ex Schult. & Schult. f.) Kunth Bride’s bonnet

Coptis laciniata A. Gray Oregon goldthread

Corylus cornuta Marshall var. californica (A. DC.) Sharp California hazel

Cornus canadensis L. Bunchberry dogwood

Cytisus scoparius (L.) Link Scotch broom

Disporum hookeri (Torr.) G. Nicholson var. hookeri Drops-of-gold

Fallopia japonica (Houtt.) Ronse Decr. var. japonica Japanese knotweed

Gaultheria ovatifolia A. Gray Western teaberry

Gaultheria shallon Pursh Salal
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Scientific name Common name

Gentiana douglasiana Bong. Swamp gentian

Geranium lucidum L. Shining geranium

Geranium robertianum L. Robert geranium

Goodyera oblongifolia Raf. Western rattlesnake plantain

Hedera helix L. English ivy

Heracleum mantegazzianum Sommier & Levier Giant hogweed

Hesperocyparis sargentii (Jeps.) Bartel Sargent’s cypress

Hieracium aurantiacum L. Orange hawkweed

Ilex aquifolium L. English holly

Iris pseudacorus L. Paleyellow iris

Juniperus occidentalis Hook. Western juniper

Lamiastrum galeobdolon (L.) Ehrend. & Polatschek Yellow archangel

Lilium occidentale Purdy Western lily

Linnaea borealis L. Twinflower

Lithocarpus densiflorus (Hook. & Arn.) Rehder Tanoak

Lonicera hispidula Pursh Honeysuckle

Lupinus albicaulis Douglas Sickle-keeled lupine

Lycopodium clavatum L. Running clubmoss

Lythrum salicaria L. Purple loosestrife

Mahonia nervosa (Pursh) Nutt. Cascade barberry

Malus fusca (Raf.) C.K. Schneid. Pacific crabapple

Notholithocarpus densiflorus (Hook. & Arn.) P.S. Manos, C.H. Cannon, & S.H. Oh Tanoak

Notholithocarpus densiflorus (Hook. & Arn.) P.S. Manos, C.H. Cannon, & S.H. Oh 

var. echinoides (R.Br. ter) P.S. Manos, C.H. Cannon & S.H. Oh

Shrub form of tanoak

Nuphar polysepala (Engelm.) Yellow pond lily

Nymphoides peltata (S.G. Gmel.) Kuntze Yellow floating heart

Osmorhiza chilensis Hook. & Arn. Sweetcicely

Phalaris arundinacea L. Reed canarygrass

Picea engelmannii Parry ex Engelm. Engelmann spruce

Picea sitchensis (Bong.) Carrière Sitka spruce

Pinus albicaulis Engelm. Whitebark pine

Pinus attenuata Lemmon Knobcone pine

Pinus contorta Douglas ex Loudon Lodgepole pine

Pinus contorta Douglas ex Loudon var. contorta Beach pine, shore pine

Pinus jeffreyi Balf. Jeffrey pine

Pinus lambertiana Douglas Sugar pine

Pinus monticola Douglas ex D. Don) Western white pine

Pinus ponderosa Lawson & C. Lawson Ponderosa pine

Populus trichocarpa L. ssp. trichocarpa (Torr. & A. Gray ex Hook) Brayshaw Black cottonwood

Potamogeton crispus L. Curly pondweed

Potentilla recta L. Sulphur cinquefoil
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Scientific name Common name

Prunus emarginata (Douglas ex Hook. D. Dietr.) Bitter cherry

Pseudotsuga menziesii (Mirb.) Franco Douglas-fir

Pteridium aquilinum (L. Kuhn) Brackenfern

Pueraria montana (Lour.) Merr. var. lobata (Willd.) Maesen & S.M. Almeida ex 

Sanjappa & Predeep

Kudzu

Pyrola asarifolia Sweet American wintergreen

Quercus agrifolia Née var. oxyadenia (Torr.) J.T. Howell Coastal live oak

Quercus berberidifolia Liebm. Scrub oak

Quercus chrysolepis Liebm. Canyon live oak

Quercus douglasii Hook. & Arn. Blue oak

Quercus garryana Douglas ex hook. Oregon white oak

Quercus kelloggi Newberry California black oak

Quercus lobata Née Valley oak

Rhamnus purshiana (DC.) A. Gray Cascara

Rhododendron groenlandicum Oeder Bog Labrador tea

Rhododendron macrophyllum D. Don ex G. Don Pacific rhododendron

Ribes lacustre (Pers.) Poir. Prickly currant

Rubus armeniacus Focke Himalayan blackberry

Salix exigua Nutt. Sandbar willow

Senecio bolanderi A. Gray Bolander’s ragwort

Sequoia sempervirens (Lamb. ex D. Don) Endl. Redwood

Smilacina stellata (L.) Desf. Starry false Solomon’s seal

Synthyris reniformis (Douglas ex Benth.) Benth. Snowqueen

Taxus brevifolia Nutt. Pacific yew

Thuja plicata Donn ex D. Don Western redcedar

Tiarella trifoliate L. Threeleaf foamflower

Trapa natans L. Water chestnut

Trillium ovatum Pursh Pacific trillium

Tsuga heterophylla (Raf.) Sarg. Western hemlock

Tsuga mertensiana (Bong.) Carrière Mountain hemlock

Typha latifolia L. Cattails

Umbellularia californica (Hook. & Arn.) Nutt. California bay laurel

Vaccinium alaskaense Howell Alaska blueberry

Vaccinium membranaceum Douglas ex Torr. Thinleaf huckleberry, big huckleberry

Vaccinium ovatum Pursh Evergreen huckleberry

Vaccinium oxycoccos L. Small cranberry

Vaccinium parvifolium Sm. Red huckleberry

Vancouveria hexandra (Hook.) C. Morren & Decne. White insideout flower

Xerophyllum tenax (Pursh) Nutt. Beargrass
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Glossary

This glossary is provided to help readers understand 

various terms used in the Northwest Forest Plan (NWFP) 

science synthesis. Sources include the Forest Service 

Handbook (FSH), the Code of Federal Regulations (CFR), 

executive orders, the Federal Register (FR), and various 

scientific publications (see “Glossary Literature Cited”). 

The authors have added working definitions of terms used 

in the synthesis and its source materials, especially when 

formal definitions may be lacking or when they differ 

across sources.

active management—Direct interventions to achieve desired outcomes, which may include harvesting and planting 

of vegetation and the intentional use of fire, among other 

activities (Carey 2003).

adaptive capacity—The ability of ecosystems and social 

systems to respond to, cope with, or adapt to disturbances 

and stressors, including environmental change, to maintain 

options for future generations (FSH 1909.12.5).

adaptive management—A structured, cyclical process for 

planning and decisionmaking in the face of uncertainty and 

changing conditions with feedback from monitoring, which 

includes using the planning process to actively test assumptions, track relevant conditions over time, and measure 

management effectiveness (FSH 1909.12.5). Additionally, 

adaptive management includes iterative decisionmaking, 

through which results are evaluated and actions are adjusted 

based on what has been learned.

adaptive management area (AMA)—A portion of the federal land area within the NWFP area that was specifically 

allocated for scientific monitoring and research to explore 

new forestry methods and other activities related to meeting the goals and objectives of the Plan. Ten AMAs were 

established in the NWFP area, covering about 1.5 million 

ac (600 000 ha), or 6 percent of the planning area (Stankey 

et al. 2003).

alien species—Any species, including its seeds, eggs, 

spores, or other biological material capable of propagating 

that species, that is not native to a particular ecosystem 

(Executive Order 13112). The term is synonymous with exotic species, nonindigenous, and nonnative species (see also 

“invasive species”).

allochthonous inputs—Material, specifically food resources, that originates from outside a stream, typically in the 

form of leaf litter. 

amenity communities—Communities located near lands 

with high amenity values.

amenity migration—Movement of people based on 

the draw of natural or cultural amenities (Gosnell and 

Abrams 2011).

amenity value—A noncommodity or “unpriced” value of 

a place or environment, typically encompassing aesthetic, 

social, cultural, and recreational values.

ancestral lands (of American Indian tribes)—Lands that 

historically were inhabited by the ancestors of American 

Indian tribes.

annual species review—A procedure established under the 

NWFP in which panels of managers and biologists evaluate new scientific and monitoring information on species to 

potentially support the recommendation of changes in their 

conservation status.

Anthropocene—The current period (or geological epoch) 

in which humans have become a dominant influence on the 

Earth’s climate and environment, generally dating from the 

period of rapid growth in industrialization, population, and 

global trade and transportation in the early 1800s (Steffen et 

al. 2007).

Aquatic Conservation Strategy (ACS) —A regional strategy applied to aquatic and riparian ecosystems across the 

area covered by the NWFP) (Espy and Babbit 1994) (see 

chapter 7 for more details).

at-risk species—Federally recognized threatened, endangered, proposed, and candidate species and species of conservation concern. These species are considered at risk of 

low viability as a result of changing environmental conditions or human-caused stressors.
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best management practices (BMPs) (for water quality)—Methods, measures, or practices used to reduce or 

eliminate the introduction of pollutants and other detrimental impacts to water quality, including but not limited to 

structural and nonstructural controls and to operation and 

maintenance procedures.

biodiversity—In general, the variety of life forms and their 

processes and ecological functions, at all levels of biological 

organization from genes to populations, species, assemblages, communities, and ecosystems. 

breeding inhibition—Prevention of reproduction in 

healthy adult individuals.

bryophytes—Mosses and liverworts.

canopy cover—The downward vertical projection from the 

outside profile of the canopy (crown) of a plant measured in 

percentage of land area covered.

carrying capacity—The maximum population size a specific environment can sustain.

ceded areas—Lands that particular tribes ceded to the 

United States government by treaties, which have been catalogued in the Library of Congress.

climate adaptation—Management actions to reduce vulnerabilities to climate change and related disturbances.

climate change—Changes in average weather conditions 

(including temperature, precipitation, and risk of certain 

types of severe weather events) that persist over multiple 

decades or longer, and that result from both natural factors 

and human activities such as increased emissions of greenhouse gases (U.S. Global Change Research Program 2017).

coarse filter—A conservation approach that focuses on 

conserving ecosystems, in contrast to a “fine filter” approach that focuses on conserving specific species. These 

two approaches are generally viewed as complementary, 

with fine-filtered strategies tailored to fit particular species 

that “fall through the pores” of the coarse filter (Hunter 

2005). See also “mesofilter.”

co-management—Two or more entities, each having legally 

established management responsibilities, working collaboratively to achieve mutually agreed upon, compatible objectives to protect, conserve, use, enhance, or restore natural 

and cultural resources (81 FR 4638).

collaborative management—Two or more entities working together to actively protect, conserve, use, enhance, or 

restore natural and cultural resources (81 FR 4638).

collaboration or collaborative process—A structured 

manner in which a collection of people with diverse interests share knowledge, ideas, and resources, while working 

together in an inclusive and cooperative manner toward a 

common purpose (FSH 1909.12.05).

community (plant and animal)—A naturally occurring 

assemblage of plant and animal species living within a defined area or habitat (36 CFR 219.19).

community forest—A general definition is forest land that 

is managed by local communities to provide local benefits 

(Teitelbaum et al. 2006). The federal government has specifically defined community forest as “forest land owned in 

fee simple by an eligible entity [local government, nonprofit 

organization, or federally recognized tribe] that provides 

public access and is managed to provide community benefits pursuant to a community forest plan” (36 CFR 230.2).

community of place or place-based community—A group 

of people who are bound together because of where they 

reside, work, visit, or otherwise spend a continuous portion 

of their time.

community resilience—The capacity of a community to 

return to its initial function and structure when initially 

altered under disturbance.

community resistance—The capacity of a community to 

withstand a disturbance without changing its function and 

structure. 

composition—The biological elements within the various 

levels of biological organization, from genes and species to 

communities and ecosystems (FSM 2020).
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congeneric—Organisms that belong to the same taxonomic 

genus, usually belonging to different species.

connectivity (of habitats)—Environmental conditions 

that exist at several spatial and temporal scales that provide landscape linkages that permit (a) the exchange of 

flow, sediments, and nutrients; (b) genetic interchange of 

genes among individuals between populations; and (c) the 

long-distance range shifts of species, such as in response to 

climate change (36 CFR 219.19).

consultation (tribal)—A formal government-to-government process that enables American Indian tribes and 

Alaska Native Corporations to provide meaningful, timely 

input, and, as appropriate, exchange views, information, 

and recommendations on proposed policies or actions 

that may affect their rights or interests prior to a decision. 

Consultation is a unique form of communication characterized by trust and respect (FSM 1509.05).

corticosterone—A steroid hormone produced by many species of animals, often as the result of stress.

cryptogam—An organism that reproduces by spores and 

that does not produce true flowers and seeds; includes fungi, 

algae, lichens, mosses, liverworts, and ferns. 

cultural keystone species—A species that significantly 

shapes the cultural identity of a people, as reflected in diet, 

materials, medicine, or spiritual practice (Garibaldi and 

Turner 2004).

cultural services—A type of ecosystem service that includes the nonmaterial benefits that people obtain from 

ecosystems through spiritual enrichment, cognitive development, reflection, recreation, and aesthetic experiences 

(Sarukhán and Whyte 2005).

desired conditions—A description of specific social, economic, or ecological characteristics toward which management of the land and resources should be directed.

disturbance regime—A description of the characteristic 

types of disturbance on a given landscape; the frequency, 

severity, and size distribution of these characteristic disturbance types and their interactions (36 CFR 219.19).

disturbance—Any relatively discrete event in time that 

disrupts ecosystem, watershed, community, or species 

population structure or function, and that changes resources, substrate availability, or the physical environment (36 

CFR 219.19).

dynamic reserves—A conservation approach in which protected areas are relocated following changes in environmental conditions, especially owing to disturbance.

early-seral vegetation—Vegetation conditions in the early 

stages of succession following an event that removes the 

forest canopy (e.g., timber harvest, wildfire, windstorm), 

on sites that are capable of developing a closed canopy 

(Swanson et al. 2014). A nonforest or “pre-forest” condition 

occurs first, followed by an “early-seral forest” as young 

shade-intolerant trees form a closed canopy.

ecocultural resources—Valued elements of the biophysical 

environment, including plants, fungi, wildlife, water, and 

places, and the social and cultural relationships of people 

with those elements.

ecological conditions—The biological and physical environment that can affect the diversity of plant and animal 

communities, the persistence of native species, invasibility, 

and productive capacity of ecological systems. Ecological 

conditions include habitat and other influences on species 

and the environment. Examples of ecological conditions 

include the abundance and distribution of aquatic and terrestrial habitats, connectivity, roads and other structural 

developments, human uses, and occurrence of other species 

(36 CFR 219.19).

ecological forestry—A ecosystem management approach 

designed to achieve multiple objectives that may include 

conservation goals and sustainable forest management and 

which emphasizes disturbance-based management and 

retention of “legacy” elements such as old trees and dead 

wood (Franklin et al. 2007).

ecological integrity—The quality or condition of an ecosystem when its dominant ecological characteristics (e.g., 

composition, structure, function, connectivity, and species 

composition and diversity) occur within the natural range of 
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variation and can withstand and recover from most perturbations imposed by natural environmental dynamics or 

human influence (36 CFR 219.19).

ecological keystone species—A species whose ecological 

functions have extensive and disproportionately large effects 

on ecosystems relative to its abundance (Power et al. 1996).

ecological sustainability—The capability of ecosystems to 

maintain ecological integrity (36 CFR 219.19).

economic sustainability—The capability of society to 

produce and consume or otherwise benefit from goods and 

services, including contributions to jobs and market and 

nonmarket benefits (36 CFR 219.19).

ecoregion—A geographic area containing distinctive ecological assemblages, topographic and climatic gradients, 

and historical land uses.

ecosystem—A spatially explicit, relatively homogeneous 

unit of the Earth that includes all interacting organisms and 

elements of the abiotic environment within its boundaries 

(36 CFR 219.19).

ecosystem diversity—The variety and relative extent of 

ecosystems (36 CFR 219.19).

ecosystem integrity—See “ecological integrity.” 

ecosystem management—Management across broad 

spatial and long temporal scales for a suite of goals, including maintaining populations of multiple species and 

ecosystem services.

ecosystem services—Benefits that people obtain from 

ecosystems (see also “provisioning services,” “regulating 

services,” “supporting services,” and “cultural services”).

ectomycorrhizal fungi—Fungal species that form symbiotic relationships with vascular plants through roots, typically 

aiding their uptake of nutrients. Although other mycorrhizal fungi penetrate their host’s cell walls, ectomycorrhizal 

fungi do not. 

endangered species—Any species or subspecies that the 

Secretary of the Interior or the Secretary of Commerce has 

deemed in danger of extinction throughout all or a significant portion of its range (16 U.S.C. Section 1532).

endemic—Native and restricted to a specific geographical 

area. 

El Niño Southern Oscillation (ENSO)—A band of anomalously warm ocean water temperatures that occasionally 

develops off the western coast of South America and can 

cause climatic changes across the Pacific Ocean. The extremes of this climate pattern’s oscillations cause extreme 

weather (such as floods and droughts) in many regions of 

the world.

environmental DNA (eDNA)—Genetic material (DNA) 

contained within small biological and tissue fragments that 

can be collected from aquatic, terrestrial, and even atmospheric environments, linked to an individual species, and 

used to indicate the presence of that species.

environmental justice populations—Groups of people who have low incomes or who identify themselves as 

African American, Asian or Pacific Islander, American 

Indian or Alaskan Native, or of Hispanic origin.

ephemeral stream—A stream that flows only in direct response to precipitation in the immediate locality (watershed 

or catchment basin), and whose channel is at all other times 

above the zone of saturation. 

epicormic—Literally, “of a shoot or branch,” this term implies growth from a previously dormant bud on the trunk or 

a limb of a tree. 

epiphyte—A plant or plant ally (including mosses and 

lichens) that grows on the surface of another plant such as a 

tree, but is not a parasite. 

even-aged stand—A stand of trees composed of a single 

age class (36 CFR 219.19).

fecundity—The reproductive rate of an organism or 

population.

federally recognized Indian tribe—An Indian tribe or 

Alaska Native Corporation, band, nation, pueblo, village, or 

community that the Secretary of the Interior acknowledges 
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to exist as an Indian tribe under the Federally Recognized 

Indian Tribe List Act of 1994, 25 U.S.C. 479a (36 CFR 

219.19).

fine filter—A conservation approach that focuses on conserving individual species in contrast to a “coarse filter” 

approach that focuses on conserving ecosystems; these 

approaches are generally viewed as complementary with 

fine-filtered strategies tailored to fit particular species that 

“fall through the pores” of the coarse filter (Hunter 2005). 

See also “mesofilter.” 

fire-dependent vegetation types—A vegetative community that evolved with fire as a necessary contributor to its 

vitality and to the renewal of habitat for its member species. 

fire exclusion—Curtailment of wildland fire because of 

deliberate suppression of ignitions, as well as unintentional effects of human activities such as intensive grazing 

that removes grasses and other fuels that carry fire (Keane 

et al. 2002). 

fire intensity—The amount of energy or heat release 

during fire.

fire regime—A characterization of long-term patterns of 

fire in a given ecosystem over a specified and relatively long 

period of time, based on multiple attributes, including frequency, severity, extent, spatial complexity, and seasonality 

of fire occurrence.

fire regime, low frequency, high severity—A fire regime 

with long return intervals (>200 years) and high levels of 

vegetation mortality (e.g., ~70 percent basal area mortality 

in forested ecosystems), often occurring in large patches 

(>10,000 ac [4047 ha]) (see chapter 3 for more details).

fire regime, moderate frequency, mixed severity—A 

fire regime with moderate return intervals between 50 and 

200 years and mixtures of low, moderate, and high severity; high-severity patches would have been common and 

frequently large (>1,000 ac [>405 ha]) (see chapter 3 for 

more details).

fire regime, very frequent, low severity—A fire regime 

with short return intervals (5 to 25 years) dominated by 

surface fires that result in low levels of vegetation mortality 

(e.g., <20 percent basal area mortality in forested ecosystems), with high-severity fire generally limited to small 

patches (<2.5 ac [1 ha]) (see chapter 3 for more details). 

fire regime, frequent, mixed severity—A fire regime with 

return intervals between 15 and 50 years that burns with a 

mosaic of low-, moderate-, and high-severity patches (Perry 

et al. 2011) (see chapter 3 for more details).

fire rotation—Length of time expected for a specific 

amount of land to burn (some parts might burn more than 

once or some not at all) based upon the study of past fire 

records in a large landscape (Turner and Romme 1994).

fire severity—The magnitude of the effects of fire on ecosystem components, including vegetation or soils.

fire suppression—The human act of extinguishing wildfires (Keane et al. 2002). 

floodplain restoration—Ecological restoration of a stream 

or river’s floodplain, which may involve setback or removal 

of levees or other structural constraints.

focal species—A small set of species whose status is assumed to infer the integrity of the larger ecological system 

to which it belongs, and thus to provide meaningful information regarding the effectiveness of a resource management plan in maintaining or restoring the ecological conditions to maintain the broader diversity of plant and animal 

communities in the NWPF area. Focal species would be 

commonly selected on the basis of their functional role in 

ecosystems (36 CFR 219.19).

food web—Interconnecting chains between organisms in 

an ecological community based upon what they consume.

Forest Ecosystem Management Assessment Team 

(FEMAT)—An interdisciplinary team that included expert 

ecological and social scientists, analysts, and managers 

assembled in 1993 by President Bill Clinton to develop 

options for ecosystem management of federal forests within 

the range of the northern spotted owl (FEMAT 1993).
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forest fragmentation—The patterns of dispersion and 

connectivity of nonhomogeneous forest cover (Riitters et 

al. 2002). See also “landscape fragmentation” and “habitat 

fragmentation” for specific meanings related to habitat loss 

and isolation.

frequency distribution—A depiction, often appearing in 

the form of a curve or graph, of the abundance of possible 

values of a variable. In this synthesis report, we speak of the 

frequency of wildfire patches of various sizes.

fuels (wildland)—Combustible material in wildland areas, 

including live and dead plant biomass such as trees, shrub, 

grass, leaves, litter, snags, and logs. 

fuels management—Manipulation of wildland fuels 

through mechanical, chemical, biological, or manual means, 

or by fire, in support of land management objectives to control or mitigate the effects of future wildland fire.

function (ecological)—Ecological processes, such as energy flow; nutrient cycling and retention; soil development and 

retention; predation and herbivory; and natural disturbances 

such as wind, fire, and floods that sustain composition and 

structure (FSM 2020). See also “key ecological function.” 

future range of variation (FRV)—The natural fluctuation 

of pattern components of healthy ecosystems that might 

occur in the future, primarily affected by climate change, 

human infrastructure, invasive species, and other anticipated disturbances.

gaps (forest)—Small openings in a forest canopy that 

are naturally formed when one or a few canopy trees die 

(Yamamoto 2000).

genotype—The genetic makeup of an individual organism. 

glucocorticoid—A class of steroid hormones produced by 

many species of animals, often as the result of stress.

goals (in land management plans)—Broad statements of 

intent, other than desired conditions, that do not include expected completion dates (36 CFR part 219.7(e)(2)).

guideline—A constraint on project and activity decisionmaking that allows for departure from its terms, so long as 

the purpose of the guideline is met (36 CFR section 219.15(d)

(3)). Guidelines are established to help achieve or maintain a 

desired condition or conditions, to avoid or mitigate undesirable effects, or to meet applicable legal requirements.

habitat—An area with the environmental conditions and 

resources that are necessary for occupancy by a species and 

for individuals of that species to survive and reproduce.

habitat fragmentation—Discontinuity in the spatial distribution of resources and conditions present in an area at a 

given scale that affects occupancy, reproduction, and survival in a particular species (see “landscape fragmentation”).

heterogeneity (forest)—Diversity, often applied to variation in forest structure within stands in two dimensions: 

horizontal (e.g., single trees, clumps of trees, and gaps of no 

trees), and vertical (e.g., vegetation at different heights from 

the forest floor to the top of the forest canopy), or across 

large landscapes (North et al. 2009).

hierarchy theory—A theory that describes ecosystems at 

multiple levels of organization (e.g., organisms, populations, 

and communities) in a nested hierarchy.

high-severity burn patch—A contiguous area of highseverity or stand-replacing fire.

historical range of variation (HRV)—Past fluctuation or 

range of conditions in the pattern of components of ecosystems over a specified period of time.

hybrid ecosystem—An ecosystem that has been modified from a historical state such that it has novel attributes while retaining some original characteristics (see 

“novel ecosystem”).

hybrid—Offspring resulting from the breeding of two 

different species.

inbreeding depression—Reduced fitness in a population 

that occurs as the result of breeding between related individuals, leading to increased homogeneity and simplification of the gene pool.
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in-channel restoration—Ecological restoration of the 

channel of a stream or river, often through placement of materials (rocks and wood) or other structural modifications.

individuals, clumps, and openings (ICO) method—A 

method that incorporates reference spatial pattern targets 

based upon individual trees, clumps of trees, and canopy 

openings into silvicultural prescriptions and tree-marking 

guidelines (Churchill et al. 2013).

Interagency Special Status and Sensitive Species 

Program (ISSSSP)—A federal agency program, established under the U.S. Forest Service Pacific Northwest 

Region and Bureau of Land Management Oregon/

Washington state office. The ISSSSP superseded the Survey 

and Manage standards and guidelines under the NWFP and 

also addresses other species of conservation focus, coordinates development and revision of management recommendations and survey protocols, coordinates data management 

between the agencies, develops summaries of species biology, and conducts other tasks. 

intermittent stream—A stream or reach of stream channel 

that flows, in its natural condition, only during certain times 

of the year or in several years, and is characterized by interspersed, permanent surface water areas containing aquatic 

flora and fauna adapted to the relatively harsh environmental conditions found in these types of environments.

invasive species—An alien species (or subspecies) whose 

deliberate, accidental, or self-introduction is likely to cause 

economic or environmental harm or harm to human health 

(Executive Order 13112).

key ecological function—The main behaviors performed 

by an organism that can influence environmental conditions 

or habitats of other species.

key watersheds—Watersheds that are expected to serve as 

refugia for aquatic organisms, particularly in the short term, 

for at-risk fish populations that have the greatest potential 

for restoration, or to provide sources of high-quality water. 

land and resource management plan (Forest Service)—A 

document or set of documents that provides management 

direction for an administrative unit of the National Forest 

System (FSH 1909.12.5).

landform—A specific geomorphic feature on the surface of 

the Earth, such as a mountain, plateau, canyon, or valley.

landscape—A defined area irrespective of ownership 

or other artificial boundaries, such as a spatial mosaic of 

terrestrial and aquatic ecosystems, landforms, and plant 

communities, repeated in similar form throughout such a 

defined area (36 CFR 219.19).

landscape fragmentation—Breaking up of continuous 

habitats into patches as a result of human land use and 

thereby generating habitat loss, isolation, and edge effects 

(see “habitat fragmentation”).

landscape genetics—An interdisciplinary field of study 

that combines population genetics and landscape ecology to explore how genetic relatedness among individuals 

and subpopulations of a species is influenced by landscape-level conditions.

landscape hierarchy—Organization of land areas based 

upon a hierarchy of nested geographic (i.e., different-sized) 

units, which provides a guide for defining the functional 

components of a system and how components at different 

scales are related to one another.

late-successional forest—Forests that have developed after 

long periods of time (typically at least 100 to 200 years) following major disturbances, and that contain a major component of shade-tolerant tree species that can regenerate beneath a canopy and eventually grow into the canopy in which 

small canopy gaps occur (see chapter 3 for more details). 

Note that FEMAT (1993) and the NWFP also applied this 

term to older (at least 80 years) forest types, including both 

old-growth and mature forests, regardless of the shade tolerance of the dominant tree species (e.g., 90-year-old forests 

dominated by Douglas-fir were termed late successional).

leading edge—The boundary of a species’ range at which 

the population is geographically expanding through colonization of new sites.
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legacy trees—Individual trees that survive a major disturbance and persist as components of early-seral stands 

(Franklin 1990).

legacies (biological)—Live trees, seed and seedling banks, 

remnant populations and individuals, snags, large soil aggregates, hyphal mats, logs, uprooted trees, and other biotic 

features that survive a major disturbance and persist as 

components of early-seral stands (Franklin 1990, Franklin 

et al. 2002).

lentic—Still-water environments, including lakes, ponds, 

and wet meadows.

longitudinal studies—Studies that include repeated observations on the same response variable over time.

lotic—Freshwater environments with running water, including rivers, streams, and springs.

low-income population—A community or a group of individuals living in geographic proximity to one another, or 

a set of individuals, such as migrant workers or American 

Indians, who meet the standards for low income and experience common conditions of environmental exposure or 

effect (CEQ 1997).

managing wildfire for resource objectives—Managing 

wildfires to promote multiple objectives such as reducing 

fire danger or restoring forest health and ecological processes rather than attempting full suppression. The terms 

“managed wildfire” or “resource objective wildfire” have 

also been used to describe such events (Long et al. 2017). 

However, fire managers note that many unplanned ignitions are managed using a combination of tactics, including 

direct suppression, indirect containment, monitoring of fire 

spread, and even accelerating fire spread, across their perimeters and over their full duration. Therefore, terms that 

separate “managed” wildfires from fully “suppressed” wildfires do not convey that complexity. (See “Use of wildland 

fire,” which also includes prescribed burning).

matrix—Federal and other lands outside of specifically 

designated reserve areas, particularly the late-successional 

reserves under the NWFP, that are managed for timber production and other objectives.

mature forest—An older forest stage (>80 years) prior to 

old-growth in which trees begin attaining maximum heights 

and developing some characteristic, for example, 80 to 200 

years in the case of old-growth Douglas-fir/western hemlock forests, often (but not always) including big trees (>50 

cm diameter at breast height), establishment of late-seral 

species (i.e., shade-tolerant trees), and initiation of decadence in early species (i.e., shade-intolerant trees).

mesofilter—A conservation approach that “focuses on conserving critical elements of ecosystems that are important 

to many species, especially those likely to be overlooked 

by fine-filter approaches, such as invertebrates, fungi, and 

nonvascular plants” (Hunter 2005).

meta-analysis—A study that combines the results of multiple studies. 

minority population—A readily identifiable group of people living in geographic proximity with a population that is 

at least 50 percent minority; or, an identifiable group that 

has a meaningfully greater minority population than the 

adjacent geographic areas, or may also be a geographically 

dispersed/transient set of individuals such as migrant workers or Americans Indians (CEQ 1997).

mitigation (climate change)—Efforts to reduce anthropogenic alteration of climate, in particular by increasing 

carbon sequestration. 

monitoring—A systematic process of collecting information to track implementation (implementation monitoring), 

to evaluate effects of actions or changes in conditions or relationships (effectiveness monitoring), or to test underlying 

assumptions (validation monitoring) (see 36 CFR 219.19).

mosaic—The contiguous spatial arrangement of elements 

within an area. In regions, this is typically the upland vegetation patches, large urban areas, large bodies of water, and 

large areas of barren ground or rock. However, regional mosaics can also be described in terms of land ownership, habitat 
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patches, land use patches, or other elements. For landscapes, 

this is typically the spatial arrangement of landscape elements.

multiaged stands—Forest stands having two or more 

age classes of trees; this includes stands resulting from 

variable-retention silvicultural systems or other traditionally even-aged systems that leave residual or reserve 

(legacy) trees.

multiple use—The management of all the various renewable surface resources of the National Forest System so that 

they are used in the combination that will best meet the 

needs of the American people; making the most judicious 

use of the land for some or all of these resources or related 

services over areas large enough to provide sufficient latitude for periodic adjustments in use to conform to changing 

needs and conditions; that some land will be used for less 

than all of the resources; and harmonious and coordinated 

management of the various resources, each with the other, 

without impairment of the productivity of the land, with 

consideration being given to the relative values of the various resources, and not necessarily the combination of uses 

that will give the greatest dollar return or the greatest unit 

output, consistent with the Multiple-Use Sustained-Yield 

Act of 1960 (16 U.S.C. 528–531) (36 CFR 219.19).

natal site—Location of birth.

native knowledge—A way of knowing or understanding the 

world, including traditional ecological, and social knowledge 

of the environment derived from multiple generations of indigenous peoples’ interactions, observations, and experiences with their ecological systems. This knowledge is accumulated over successive generations and is expressed through 

oral traditions, ceremonies, stories, dances, songs, art, and 

other means within a cultural context (36 CFR 219.19).

native species—A species historically or currently present 

in a particular ecosystem as a result of natural migratory or 

evolutionary processes and not as a result of an accidental 

or deliberate introduction or invasion into that ecosystem 

(see 36 CFR 219.19).

natural range of variation (NRV)—The variation of ecological characteristics and processes over specified scales of 

time and space that are appropriate for a given management 

application (FSH 1909.12.5).

nested hierarchy—The name given to the hierarchical 

structure of groups within groups used to classify organisms.

nontimber forest products (also known as “special forest products”)—Various products from forests that do not 

include logs from trees but do include bark, berries, boughs, 

bryophytes, bulbs, burls, Christmas trees, cones, ferns, firewood, forbs, fungi (including mushrooms), grasses, mosses, 

nuts, pine straw, roots, sedges, seeds, transplants, tree sap, 

wildflowers, fence material, mine props, posts and poles, shingle and shake bolts, and rails (36 CFR part 223 Subpart G).

novel ecosystem—An ecosystem that has experienced large 

and potentially irreversibly modifications to abiotic conditions 

or biotic composition in ways that result in a composition 

of species, ecological communities, and functions that have 

never before existed, and that depart from historical analogs 

(Hobbs et al. 2009). See “hybrid ecosystem” for comparison.

old-growth forest—A forest distinguished by old trees 

(>200 years) and related structural attributes that often (but 

not always) include large trees, high biomass of dead wood 

(i.e., snags, down coarse wood), multiple canopy layers, 

distinctive species composition and functions, and vertical 

and horizontal diversity in the tree canopy (see chapter 3). 

In dry, fire-frequent forests, old growth is characterized by 

large, old fire-resistant trees and relatively open stands without canopy layering. 

palustrine—Inland, nontidal wetlands that may be permanently or temporarily flooded and are characterized by the 

presence of emergent vegetation such as swamps, marshes, 

vernal pools, and lakeshores.

passive management—A management approach in which 

natural processes are allowed to occur without human intervention to reach desired outcomes.

patch—A relatively small area with similar environmental conditions, such as vegetative structure and composition. Sometimes used interchangeably with vegetation or 

forest stand.
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Pacific Decadal Oscillation (PDO)—A recurring (approximately decadal-scale) pattern of ocean-atmosphere 

—a stream or reach of a channel that flows continuously 

or nearly so throughout the year and whose upper surface 

is generally lower than the top of the zone of saturation in 

areas adjacent to the stream.

perennial stream—A stream or reach of a channel that 

flows continuously or nearly so throughout the year and 

whose upper surface is generally lower than the top of the 

zone of saturation in areas adjacent to the stream.

phenotype—Physical manifestation of the genetic makeup 

of an individual and its interaction with the environment.

place attachment—The “positive bond that develops 

between groups or individuals and their environment” 

(Jorgensen and Stedman 2001: 234).

place dependence— “The strength of an individual’s 

subjective attachment to specific places” (Stokols and 

Shumaker 1982: 157).

place identity—Dimensions of self that define an individual’s [or group’s] identity in relation to the physical 

environment through ideas, beliefs, preferences, feelings, values, goals, and behavioral tendencies and skills 

(Proshansky 1978).

place-based planning—“A process used to involve stakeholders by encouraging them to come together to collectively define place meanings and attachments” (Lowery and 

Morse 2013: 1423).

plant association—A fine level of classification in a hierarchy of potential vegetation that is defined in terms of a climax-dominant overstory tree species and typical understory 

herb or shrub species. 

population bottleneck—An abrupt decline in the size of 

a population from an event, which often results in deleterious effects such as reduced genetic diversity and increased 

probability of local or global extirpation.

potential vegetation type (PVT)—Native, late-successional (or “climax”) plant community that reflects the regional 

climate, and dominant plant species that would occur on a 

site in absence of disturbances (Pfister and Arno 1980).

poverty rate—A measure of financial income below a 

threshold that differs by family size and composition.

precautionary principle—A principle that if an action, 

policy, or decision has a suspected risk of causing harm 

to the public or to the environment, and there is no scientific consensus that it is not harmful, then the burden 

of proof that it is not harmful falls on those making that 

decision. Particular definitions of the principle differ, and 

some applications use the less formal term, “precautionary approach.” Important qualifications associated with 

many definitions include (1) the perceived harm is likely 

to be serious, (2) some scientific analysis suggests a significant but uncertain potential for harm, and (3) applications of the principle emphasize generally constraining 

an activity to mitigate it rather than “resisting” it entirely 

(Doremus 2007).

prescribed fire—A wildland fire originating from a 

planned ignition to meet specific objectives identified 

in a written and approved prescribed fire plan for which 

National Environmental Policy Act requirements (where applicable) have been met prior to ignition (synonymous with 

controlled burn).

primary recreation activity—A single activity that caused 

a recreation visit to a national forest.

probable sale quantity—An estimate of the average 

amount of timber likely to be awarded for sale for a given 

area (such as the NWFP area) during a specified period.

provisioning services—A type of ecosystem service that 

includes clean air and fresh water, energy, food, fuel, forage, wood products or fiber, and minerals.

public participation geographic information system 

(PPGIS)—Using spatial decisionmaking and mapping tools 

to produce local knowledge with the goal of including and empowering marginalized populations (Brown and Reed 2009).

public values—Amenity values (scenery, quality of life); 

environmental quality (clean air, soil, and water); ecological 
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values (biodiversity); public use values (outdoor recreation, 

education, subsistence use); and spiritual or religious values 

(cultural ties, tribal history).

record of decision (ROD)—The final decision document 

that amended the planning documents of 19 national forests 

and seven Bureau of Land Management districts within the 

range of the northern spotted owl (the NWFP area) in April 

1994 (Espy and Babbit 1994).

recreation opportunity—An opportunity to participate 

in a specific recreation activity in a particular recreation 

setting to enjoy desired recreation experiences and other 

benefits that accrue. Recreation opportunities include nonmotorized, motorized, developed, and dispersed recreation 

on land, water, and in the air (36 CFR 219.19).

redundancy—The presence of multiple occurrences of 

ecological conditions, including key ecological functions 

(functional redundancy), such that not all occurrences may 

be eliminated by a catastrophic event. 

refugia—An area that remains less altered by climatic and 

environmental change (including disturbances such as wind 

and fire) affecting surrounding regions and that therefore 

forms a haven for relict fauna and flora.

regalia—Dress and special elements made from a variety 

of items, including various plant and animal materials, and 

worn for tribal dances and ceremonies.

regulating services—A type of ecosystem service that 

includes long-term storage of carbon; climate regulation; 

water filtration, purification, and storage; soil stabilization; 

flood and drought control; and disease regulation.

representativeness—The presence of a full array of ecosystem types and successional states, based on the physical 

environment and characteristic disturbance processes.

reserve—An area of land designated and managed for a special purpose, often to conserve or protect ecosystems, species, 

or other natural and cultural resources from particular human 

activities that are detrimental to achieving the goals of the area.

resilience—The capacity of a system to absorb disturbance 

and reorganize (or return to its previous organization) so as 

to still retain essentially the same function, structure, identity, and feedbacks (see FSM Chapter 2020 and see also “socioecological resilience”). Definitions emphasize the capacity 

of a system or its constituent entities to respond or regrow after mortality induced by a disturbance event, although broad 

definitions of resilience may also encompass “resistance” 

(see below), under which such mortality may be averted.

resistance—The capacity of a system or an entity to withstand a disturbance event without much change.

restoration economy—Diverse economic activities associated with the restoration of structure or function to terrestrial 

and aquatic ecosystems (Nielsen-Pincus and Moseley 2013).

restoration, ecological—The process of assisting the recovery of an ecosystem that has been degraded, damaged, or 

destroyed. Ecological restoration focuses on reestablishing 

the composition, structure, pattern, and ecological processes necessary to facilitate terrestrial and aquatic ecosystems 

sustainability, resilience, and health under current and future conditions (36 CFR 219.19).

restoration, functional—Restoration of dynamic abiotic 

and biotic processes in degraded ecosystems, without necessarily a focus on structural condition and composition.

riparian areas—Three-dimensional ecotones (the transition zone between two adjoining communities) of interaction that include terrestrial and aquatic ecosystems that 

extend down into the groundwater, up above the canopy, 

outward across the floodplain, up the near slopes that drain 

to the water, laterally into the terrestrial ecosystem, and 

along the water course at variable widths (36 CFR 219.19).

riparian management zone—Portions of a watershed 

in which riparian-dependent resources receive primary 

emphasis, and for which plans include Plan components to 

maintain or restore riparian functions and ecological functions (36 CFR 219.19).

riparian reserves—Reserves established along streams and 

rivers to protect riparian ecological functions and processes 
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

necessary to create and maintain habitat for aquatic and riparian-dependent organisms over time and ensure connectivity 

within and between watersheds. The Aquatic Conservation 

Strategy in the NWFP record of decision included standards 

and guidelines that delineated riparian reserves.

risk—A combination of the probability that a negative outcome will occur and the severity of the subsequent negative 

consequences (36 CFR 219.19).

rural restructuring—Changes in demographic and economic conditions owing to declines in natural resource 

production and agriculture (Nelson 2001).

scale—In ecological terms, the extent and resolution in spatial 

and temporal terms of a phenomenon or analysis, which differs 

from the definition in cartography regarding the ratio of map 

distance to Earth surface distance (Jenerette and Wu 2000).

scenic character—A combination of the physical, biological, 

and cultural images that gives an area its scenic identity and 

contributes to its sense of place. Scenic character provides a 

frame of reference from which to determine scenic attractiveness and to measure scenic integrity (36 CFR 219.19).

science synthesis—A narrative review of scientific information from a defined pool of sources that compiles and 

integrates and interprets findings and describes uncertainty, including the boundaries of what is known and 

what is not known.

sense of place—The collection of meanings, beliefs, symbols, values, and feelings that individuals or groups associate with a particular locality (Williams and Stewart 1998).

sensitive species—Plant or animal species that receive 

special conservation attention because of threats to their 

populations or habitats, but which do not have special status 

as listed or candidates for listing under the Endangered 

Species Act.

sensitivity—In ecological contexts, the propensity of 

communities or populations to change when subject to 

disturbance, or the opposite of resistance (see “community resistance”).

sink population—A population in which reproductive rates 

are lower than mortality rates but that is maintained by immigration of individuals from outside of that population (see 

also “source population”). 

social sustainability—“The capability of society to support 

the network of relationships, traditions, culture, and activities that connect people to the land and to one another, and 

support vibrant communities” (36 CFR 219.19). The term is 

commonly invoked as one of the three parts of a “triple-bottom line” alongside environmental and economic considerations. The concept is an umbrella term for various topics 

such as quality of life, security, social capital, rights, sense 

of place, environmental justice, and community resilience, 

among others discussed in this synthesis.

socioecological resilience—The capacity of socioecological 

systems (see “socioecological system”) to cope with, adapt 

to, and influence change; to persist and develop in the face 

of change; and to innovate and transform into new, more 

desirable configurations in response to disturbance.

socioecological system (or social-ecological system)—A 

coherent system of biophysical and social factors defined 

at several spatial, temporal, and organizational scales that 

regularly interact, continuously adapt, and regulate critical 

natural, socioeconomic, and cultural resources (Redman et 

al. 2004); also described as a coupled-human and natural 

system (Liu et al. 2007).

source population—A population in which reproductive 

rates exceed those of mortality rates so that the population 

has the capacity to increase in size. The term is also often 

used to denote when such a population contributes emigrants (dispersing individuals) that move outside the population, particularly when feeding a sink population.

special forest products—See “nontimber forest products.”

special status species—Species that have been listed or 

proposed for listing as threatened or endangered under the 

Endangered Species Act.

species of conservation concern—A species, other than 

federally recognized as a threatened, endangered, proposed, 
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GENERAL TECHNICAL REPORT PNW-GTR-966

or candidate species, that is known to occur in the NWFP 

area and for which the regional forester has determined that 

the best available scientific information indicates substantial 

concern about the species’ capability to persist over the long 

term in the Plan area (36 CFR 219.9(c)).

stand—A descriptor of a land management unit consisting of 

a contiguous group of trees sufficiently uniform in age-class 

distribution, composition, and structure, and growing on a site 

of sufficiently uniform quality, to be a distinguishable unit.

standard—A mandatory constraint on project and activity 

decisionmaking, established to help achieve or maintain the 

desired condition or conditions, to avoid or mitigate undesirable effects, or to meet applicable legal requirements.

stationarity—In statistics, a process that, while randomly 

determined, is not experiencing a change in the probability 

of outcomes.

stewardship contract—A contract designed to achieve 

land management goals while meeting local and rural community needs, including contributing to the sustainability 

of rural communities and providing a continuing source of 

local income and employment.

strategic surveys—One type of field survey, specified 

under the NWFP, designed to fill key information gaps on 

species distributions and ecologies by which to determine 

if species should be included under the Plan’s Survey and 

Manage species list.

stressors—Factors that may directly or indirectly degrade 

or impair ecosystem composition, structure, or ecological 

process in a manner that may impair its ecological integrity, 

such as an invasive species, loss of connectivity, or the disruption of a natural disturbance regime (36 CFR 219.19).

structure (ecosystem)—The organization and physical 

arrangement of biological elements such as snags and down 

woody debris, vertical and horizontal distribution of vegetation, stream habitat complexity, landscape pattern, and 

connectivity (FSM 2020).

supporting services—A type of ecosystem service that 

includes pollination, seed dispersal, soil formation, and nutrient cycling.

Survey and Manage program—A formal part of the 

NWFP that established protocols for conducting various 

types of species surveys, identified old-forest-associated 

species warranting additional consideration for monitoring and protection (see “Survey and Manage species”), and 

instituted an annual species review procedure that evaluated 

new scientific and monitoring information on species for 

potentially recommending changes in their conservation 

status, including potential removal from the Survey and 

Manage species list. 

Survey and Manage species—A list of species, compiled 

under the Survey and Manage program of the NWFP, that 

were deemed to warrant particular attention for monitoring and protection beyond the guidelines for establishing 

late-successional forest reserves.

sustainability—The capability to meet the needs of the 

present generation without compromising the ability of future generations to meet their needs (36 CFR 219.19).

sustainable recreation—The set of recreation settings and 

opportunities in the National Forest System that is ecologically, economically, and socially sustainable for present and 

future generations (36 CFR 219.19).

sympatric—Two species or populations that share a common geographic range and coexist.

threatened species—Any species that the Secretary of the 

Interior or the Secretary of Commerce has determined is 

likely to become an endangered species within the foreseeable future throughout all or a significant portion of its 

range. Threatened species are listed at 50 CFR sections 

17.11, 17.12, and 223.102. 

timber harvest—The removal of trees for wood fiber use 

and other multiple-use purposes (36 CFR 219.19).

timber production—The purposeful growing, tending, 

harvesting, and regeneration of regulated crops of trees to 

be cut into logs, bolts, or other round sections for industrial 

or consumer use (36 CFR 219.19).

topo-edaphic—Related to or caused by particular soil 

conditions, as of texture or drainage, rather than by physiographic or climatic factors within a defined region or area.
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Synthesis of Science to Inform Land Management Within the Northwest Forest Plan Area

traditional ecological knowledge—“A cumulative body 

of knowledge, practice, and belief, evolving by adaptive 

processes and handed down through generations by cultural 

transmission, about the relationship of living beings (including humans) with one another and with their environment” 

(Berkes et al. 2000: 1252). See also “native knowledge.”

trailing edge—When describing the range of a species, the 

boundary at which the species’ population is geographically 

contracting through local extinction at occupied sites.

trophic cascade—Changes in the relative populations of 

producers, herbivores, and carnivores following the addition 

or removal of top predators and the resulting disruption of 

the food web.

uncertainty—Amount or degree of confidence as a result 

of imperfect or incomplete information.

understory—Vegetation growing below the tree canopy in a 

forest, including shrubs and herbs that grow on the forest floor.

use of wildland fire—Management of either wildfire or 

prescribed fire to meet resource objectives specified in land 

or resource management plans (see “Managing wildfire for 

resource objectives” and “Prescribed fire”).

variable-density thinning—The method of thinning some 

areas within a stand to a different density (including leaving 

dense, unthinned areas) than other parts of the stand, which 

is typically done to promote ecological diversity in a relatively uniform stand.

vegetation series (plant community)—The highest level 

of the fine-scale component (plant associations) of potential 

vegetation hierarchy based on the dominant plant species 

that would occur in late-successional conditions in the absence of disturbance.

vegetation type—A general term for a combination or 

community of plants (including grasses, forbs, shrubs, or 

trees), typically applied to existing vegetation rather than 

potential vegetation. 

viable population—A group of breeding individuals of a 

species capable of perpetuating itself over a given time scale. 

vital rates—Statistics describing population dynamics such 

as reproduction, mortality, survival, and recruitment.

watershed—A region or land area drained by a single 

stream, river, or drainage network; a drainage basin (36 

CFR 219.19).

watershed analysis—An analytical process that characterizes watersheds and identifies potential actions for addressing problems and concerns, along with possible management 

options. It assembles information necessary to determine the 

ecological characteristics and behavior of the watershed and 

to develop options to guide management in the watershed, 

including adjusting riparian reserve boundaries.

watershed condition assessment—A national approach 

used by the U.S. Forest Service to evaluate condition of 

hydrologic units based on 12 indicators, each composed of 

various attributes (USDA FS 2011).

watershed condition—The state of a watershed based on 

physical and biogeochemical characteristics and processes 

(36 CFR 219.19).

watershed restoration—Restoration activities that focus 

on restoring the key ecological processes required to create 

and maintain favorable environmental conditions for aquatic and riparian-dependent organisms.

well-being—The condition of an individual or group in social, economic, psychological, spiritual, or medical terms.

wilderness—Any area of land designated by Congress as 

part of the National Wilderness Preservation System that 

was established by the Wilderness Act of 1964 (16 U.S.C. 

1131–1136) (36 CFR 219.19).

wildlife—Undomesticated animal species, including amphibians, reptiles, birds, mammals, fish, and invertebrates 

or even all biota, that live wild in an area without being 

introduced by humans.

wildfire—Unplanned ignition of a wildland fire (such as a 

fire caused by lightning, volcanoes, unauthorized and accidental human-caused fires), and escaped prescribed fires.

wildland-urban interface (WUI)—The line, area, or zone 

where structures and other human development meet or intermingle with undeveloped wildland or vegetation fuels.
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Acknowledgments

We thank reviewers from U.S. Forest Service Regions 5 and 

6, who provided valuable input on earlier drafts of these 

chapters. We thank the many anonymous reviewers for 

their constructive comments. We also thank members of 

the public, Tribes, and other agencies who provided peer-reviewed literature for consideration, attended the public 

forums, or provided review comments for peer reviewers to 

consider. We thank Cliff Duke with the Ecological Society 

of America for organizing and coordinating the peer review 

process. We thank Lisa McKenzie and Ty Montgomery 

with McKenzie Marketing Group for coordinating and 

summarizing the extensive public input we received and for 

organizing and facilitating the public forums. We also thank 

Kathryn Ronnenberg for assisting with figures and editing 

for some chapters, as well as Keith Olsen for his work 

on figures. Sean Gordon is thanked for his creation and 

management of the NWFP literature reference database and 

for formatting all the citations in the chapters, which was a 

very large task. We really appreciate the efforts of Rhonda 

Mazza, who worked with the authors to write the executive 

summary. We appreciate the heroic efforts of the entire 

Pacific Northwest Research Station communications team 

to get the science synthesis edited and published to meet 

tight deadlines, especially editors Keith Routman, Carolyn 

Wilson, and Oscar Johnson and visual information specialist Jason Blake. Borys Tkacz and Jane Hayes are acknowledged for their diligent policy reviews of all chapters. We 

want to acknowledge the leadership of Paul Anderson and 

Cindy Miner and Yasmeen Sands for their communications 

coordination. Finally, the team wishes to thank Region 5, 

Region 6, and the Pacific Northwest and Pacific Southwest 

Research Stations for significant funding and other support 

for this effort. 
Pacific Northwest Research Station

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