United States

Department

of Agriculture

Forest Service

Rocky Mountain

Research Station

General Technical

Report RMRS-GTR-42-

volume 2

December 2000

Wildland Fire in

Ecosystems

Effects of Fire on Flora
Abstract _____________________________________

Brown, James K.; Smith, Jane Kapler, eds. 2000. Wildland fire in ecosystems: effects of fire on flora. Gen.

Tech. Rep. RMRS-GTR-42-vol. 2. Ogden, UT: U.S. Department of Agriculture, Forest Service, Rocky

Mountain Research Station. 257 p.

This state-of-knowledge review about the effects of fire on flora and fuels can assist land managers

with ecosystem and fire management planning and in their efforts to inform others about the ecological

role of fire. Chapter topics include fire regime classification, autecological effects of fire, fire regime

characteristics and postfire plant community developments in ecosystems throughout the United States

and Canada, global climate change, ecological principles of fire regimes, and practical considerations

for managing fire in an ecosytem context.

Keywords: ecosystem, fire effects, fire management, fire regime, fire severity, fuels, habitat, plant

response, plants, succession, vegetation

The volumes in “The Rainbow Series” will be published from 2000 through 2001. To order, check the box or boxes below, fill in

the address form, and send to the mailing address listed below. Or send your order and your address in mailing label form to one

of the other listed media. Your order(s) will be filled as the volumes are published.

RMRS-GTR-42-vol. 1. Wildland fire in ecosystems: effects of fire on fauna.

RMRS-GTR-42-vol. 2. Wildland fire in ecosystems: effects of fire on flora.

RMRS-GTR-42-vol. 3. Wildland fire in ecosystems: effects of fire on cultural resources and archeology.

RMRS-GTR-42-vol. 4. Wildland fire in ecosystems: effects of fire on soil and water.

RMRS-GTR-42-vol. 5. Wildland fire in ecosystems: effects of fire on air.

Send to: ________________________________________________________________________________

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Cover photo—Arnica and fireweed flowers, Bob Marshall Wilderness, MT, 2 years after crown fire.

Photo by Melanie Miller.

Wildland Fire in Ecosystems

Effects of Fire on Flora

Authors

R. James Ansley, Plant Physiologist, Texas A&M University

System, Texas Agricultural Experiment Station, Vernon, TX

76385

Stephen F. Arno, Research Forester (Emeritus), Fire Sciences Laboratory, Rocky Mountain Research Station, U.S.

Department of Agriculture, Forest Service, Missoula, MT 59807

Brent L. Brock, Research Associate, Division of Biology,

Kansas State University, Manhattan, KS 66506

Patrick H. Brose, Research Forester, Northeastern Research

Station, U.S. Department of Agriculture, Forest Service, Irvine,

PA 16329

James K. Brown, Research Forester, Systems for Environmental Management, Missoula, MT 59802 (formerly with Fire

Sciences Laboratory, Rocky Mountain Research Station, U.S.

Department of Agriculture, Forest Service)

Luc C. Duchesne, Research Scientist, Canadian Forestry

Service, Great Lakes Forestry Centre, Sault Ste Marie, ON P6A

5M7

James B. Grace, Research Ecologist, National Wetlands

Research Center, U.S. Geological Survey, Lafayette, LA 70506

Gerald J. Gottfried, Research Forester, Southwest Forest

Sciences Complex, Rocky Mountain Research Station, U.S.

Department of Agriculture, Forest Service, Flagstaff, AZ 86001

Sally M. Haase, Research Forester, Riverside Forest Fire

Laboratory, Pacific Southwest Research Station, U.S. Department of Agriculture, Forest Service, Riverside, CA 92507

Michael G. Harrington, Research Forester, Fire Sciences

Laboratory, Rocky Mountain Research Station, U.S. Department of Agriculture, Forest Service, Missoula, MT 59807

Editors

James K. Brown, Research Forester, Systems for Environmental Management, Missoula, MT 59802 (formerly with Fire Sciences Laboratory, Rocky Mountain Research Station, U.S. Department of Agriculture, Forest Service).

Jane Kapler Smith, Ecologist, Fire Sciences Laboratory, Rocky

Mountain Research Station, U.S. Department of Agriculture,

Forest Service, Missoula, MT 59807.

Brad C. Hawkes, Fire Research Officer, Canadian Forestry

Service, Pacific Forestry Centre, Victoria, BC V8Z 1M5

Greg A. Hoch, Graduate Research Assistant, Division of

Biology, Kansas State University, Manhattan, KS 66506

Melanie Miller, Fire Ecologist, Bureau of Land Management,

National Office of Fire and Aviation, Boise, ID 83705

Ronald L. Myers, Director of National Fire Management Program, The Nature Conservancy, Tall Timbers Research Station, Tallahassee, FL 32312

Marcia G. Narog, Ecologist, Riverside Forest Fire Laboratory,

Pacific Southwest Research Station, U.S. Department of Agriculture, Forest Service, Riverside, CA 92507

William A. Patterson III, Professor, Department of Forestry

and Wildlife Management, University of Massachusetts,

Amherst, MA 01003

Timothy E. Paysen, Research Forester, Riverside Forest Fire

Laboratory, Pacific Southwest Research Station, U.S. Department of Agriculture, Forest Service, Riverside, CA 92507

Kevin C. Ryan, Project Leader of Fire Effects Unit, Fire

Sciences Laboratory, Rocky Mountain Research Station, U.S.

Department of Agriculture, Forest Service, Missoula, MT 59807

Stephen S. Sackett, Research Forester (Emeritus), Riverside

Forest Fire Laboratory, Pacific Southwest Research Station,

U.S. Department of Agriculture, Forest Service, Riverside, CA

92507

Dale D. Wade, Research Forester, Forestry Sciences Laboratory, Southern Research Station, U.S. Department of Agriculture, Forest Service, Athens, GA 30602

Ruth C. Wilson, Professor of Biology, California State University, San Bernardino, CA 92407
Preface _____________________________________

In 1978, a national workshop on fire effects in Denver, Colorado, provided the impetus

for the “Effects of Wildland Fire on Ecosystems” series. Recognizing that knowledge of

fire was needed for land management planning, state-of-the-knowledge reviews were

produced that became known as the “Rainbow Series.” The series consisted of six

publications, each with a different colored cover, describing the effects of fire on soil,

water, air, flora, fauna, and fuels.

The Rainbow Series proved popular in providing fire effects information for professionals,

students, and others. Printed supplies eventually ran out, but knowledge of fire effects

continued to grow. To meet the continuing demand for summaries of fire effects knowledge,

the interagency National Wildfire Coordinating Group asked Forest Service research leaders

to update and revise the series. To fulfill this request, a meeting for organizing the revision was

held January 4-6, 1993, in Scottsdale, Arizona. The series name was then changed to “The

Rainbow Series.” The five-volume series covers air, soil and water, fauna, flora and fuels, and

cultural resources.

The Rainbow Series emphasizes principles and processes rather than serving as a

summary of all that is known. The five volumes, taken together, provide a wealth of information

and examples to advance understanding of basic concepts regarding fire effects in the United

States and Canada. As conceptual background, they provide technical support to fire and

resource managers for carrying out interdisciplinary planning, which is essential to managing

wildlands in an ecosystem context. Planners and managers will find the series helpful in many

aspects of ecosystem-based management, but they will also need to seek out and synthesize

more detailed information to resolve specific management questions.

–– The Authors

October 2000

Acknowledgments____________________________

The Rainbow Series was completed under the sponsorship of the Joint Fire Sciences

Program, a cooperative fire science effort of the U.S. Department of Agriculture, Forest Service

and the U.S. Department of the Interior, Bureau of Indian Affairs, Bureau of Land Management,

Fish and Wildlife Service, National Park Service, and U.S. Geological Survey. We thank Marcia

Patton-Mallory and Louise Kingsbury for persistence and support.

The authors wish to thank the following individuals for their suggestions, information, and

assistance that led to substantial technical and editorial improvements in the manuscript:

Stephen Arno, Andrew Applejohn, David Bunnell, Tammy Charron, Lisa Clark, Scott Collins,

Bonni Corcoran, Luc Duchesne, Colin Hardy, Mick Harrington, Janet Howard, Bill Leenhouts,

Jim Menakis, Melanie Miller, Penelope Morgan, Rob McAlpine, Carmen Mueller-Rowat, Ron

Myers, Phil Omi, Pat Outcalt, Tim Paysen, Kevin Ryan, Dennis Simmerman, Jim Snyder, Peter

Stickney, Ann Murray Strome, Fred Swanson, David VanLear, Dale Wade, Phil Weatherspoon,

Mike Weber, and John Zasada.

ii
Summary ........................................................................... vi

Chapter 1: Introduction and Fire Regimes ..................... 1

by James K. Brown

Flora and Fuel Volume ........................................................ 2

Fire Regimes ....................................................................... 3

Chapter 2: Fire Autecology .............................................. 9

by Melanie Miller

Plant Mortality ..................................................................... 9

Aerial Crown Mortality .................................................. 10

Stem Mortality .............................................................. 10

Root Mortality ............................................................... 16

Fire Resistance ............................................................ 16

Vegetative Regeneration .................................................. 16

Sprouting of Woody Plants ........................................... 17

Sprouting of Forbs ........................................................ 20

Sprouting of Grasses ................................................... 21

Seedling Establishment .................................................... 22

Seed Supply and Dispersal .......................................... 22

Seedbank ..................................................................... 24

Seed Environment ........................................................ 25

Fire Stimulated Germination ........................................ 26

Burn Severity and Seed Regeneration ......................... 27

Seasonal Influences .......................................................... 28

Carbohydrates .............................................................. 28

Flowering ...................................................................... 29

Phenology .................................................................... 30

Burning Conditions ....................................................... 31

Discussion ......................................................................... 32

How This Applies to Management ............................... 34

Chapter 3: Fire in Northern Ecosystems ...................... 35

by Luc C. Duchesne and Brad C. Hawkes

Mixed Fire Regime ............................................................ 35

Major Vegetation Types ............................................... 35

Fire Regime Characteristics .................................... 35

Fuels ........................................................................ 39

Postfire Plant Communities .......................................... 40

Aspen ...................................................................... 40

White and Red Pine ................................................ 40

Jack Pine ................................................................. 41

Stand-Replacement Fire Regimes .................................... 42

Major Vegetation Types ............................................... 42

Fire Regime Characteristics .................................... 42

Fuels ........................................................................ 46

Postfire Plant Communities .......................................... 47

Black Spruce ........................................................... 47

White Spruce ........................................................... 48

Balsam Fir ............................................................... 48

Red Spruce ............................................................. 49

Conifer Bogs ............................................................ 50

Tundra ..................................................................... 50

Chapter 4: Fire in Eastern Ecosystems ........................ 53

by Dale D. Wade, Brent L. Brock, Patrick H. Brose,

James B. Grace, Greg A. Hoch, and William A.

Patterson III

Understory Fire Regimes .................................................. 53

Major Vegetation Types ............................................... 53

Fire Regime Characteristics .................................... 56

Fuels and Fire Behavior .......................................... 62

Postfire Plant Communities .......................................... 65

Longleaf pine ........................................................... 65

Slash Pine ............................................................... 69

Loblolly Pine ............................................................ 71

Shortleaf Pine .......................................................... 73

Oak-Hickory Forests ................................................ 73

Mixed Fire Regimes .......................................................... 75

Major Vegetation Types ............................................... 75

Pines........................................................................ 75

Hardwoods .............................................................. 76

Fire Regime Characteristics .................................... 76

Fuels and Fire Behavior .......................................... 78

Postfire Plant Communities .......................................... 78

Pitch Pine ................................................................ 78

Virginia Pine ............................................................ 81

Pond Pine ................................................................ 81

Mixed Mesophytic Forest ........................................ 81

Northern Hardwoods ............................................... 82

Bottomland Hardwoods ........................................... 84

Stand-Replacement Fire Regimes .................................... 84

Major Vegetation Types ............................................... 84

Wet Grasslands ....................................................... 84

Prairie ...................................................................... 85

Bay Forests ............................................................. 85

Conifers ................................................................... 86

Fire Regime Characteristics .................................... 86

Fuels and Fire Behavior .......................................... 88

Postfire Plant Communities .......................................... 89

Salt and Brackish Marshes ...................................... 89

Fresh and Oligohaline Marshes .............................. 91

Prairie ...................................................................... 92

Bay Forests ............................................................. 94

Sand Pine ................................................................ 95

Table Mountain Pine ................................................ 95

Spruce-Fir................................................................ 96

Atlantic White Cedar ................................................ 96

Chapter 5: Fire in Western Forest Ecosystems ........... 97

by Stephen F. Arno

Understory Fire Regimes .................................................. 97

Major Vegetation Types ............................................... 97

Fire Regime Characteristics .................................... 97

Fuels ...................................................................... 100

Postfire Plant Communities ........................................ 100

Contents________________________________________________

Page Page

iii
Ponderosa Pine/Jeffrey Pine and

Ponderosa Pine-Mixed Conifer .......................... 100

Redwood ............................................................... 105

Oregon Oak Woodlands ........................................ 105

Mixed Fire Regimes ........................................................ 106

Major Vegetation Types ............................................. 106

Fire Regime Characteristics .................................. 106

Fuels ...................................................................... 107

Postfire Plant Communities ........................................ 108

Coast Douglas-fir and Douglas-fir/Hardwoods ...... 108

Red Fir and Sierra/Cascade Lodgepole Pine ........ 109

Interior Douglas-fir, Larch, Rocky Mountain

Lodgepole Pine .................................................. 110

Whitebark Pine .......................................................111

Ponderosa Pine ..................................................... 112

Stand-Replacement Fire Regimes .................................. 113

Major Vegetation Types ............................................. 113

Fire Regime Characteristics .................................. 113

Fuels ...................................................................... 114

Postfire Plant Communities ........................................ 115

Coast Douglas-fir ................................................... 115

Coastal True Fir-Mountain Hemlock ...................... 115

Interior True Fir–Douglas-fir–Western Larch ......... 116

Rocky Mountain Lodgepole Pine ........................... 117

Western White Pine-Cedar-Hemlock ..................... 118

Spruce-Fir-Whitebark Pine .................................... 119

Aspen .................................................................... 120

Regimes Where Fire is Rare ........................................... 120

Chapter 6: Fire in Western Shrubland,

Woodland, and Grassland Ecosystems ............. 121

by Timothy E. Paysen, R. James Ansley, James K. Brown,

Gerald J. Gottfried, Sally M. Haase, Michael G.

Harrington, Marcia G. Narog, Stephen S. Sackett,

and Ruth C. Wilson

Understory Fire Regimes ................................................ 121

Major Vegetation Types ............................................. 121

Fire Regime Characteristics .................................. 121

Fuels ...................................................................... 124

Post Fire Plant Communities ...................................... 126

Southwestern Ponderosa Pine .............................. 126

Mixed Fire Regimes ........................................................ 128

Major Vegetation Types ............................................. 128

Fire Regime Characteristics .................................. 130

Fuels ...................................................................... 131

Postfire Plant Communities ........................................ 132

Pinyon-Juniper ...................................................... 132

Western Oaks ........................................................ 134

Texas Savanna ...................................................... 135

Stand-Replacement Fire Regimes .................................. 136

Major Vegetation Types ............................................. 136

Grasslands ............................................................ 136

Shrublands ............................................................ 137

Fire Regime Characteristics .................................. 142

Fuels ...................................................................... 146

Postfire Plant Communities ........................................ 150

Plains Grasslands ................................................. 150

Mountain Grasslands ............................................ 151

Cheatgrass ............................................................ 152

Annual Grasslands ................................................ 152

Desert Shrublands ................................................. 153

Sagebrush ............................................................. 153

Blackbrush ............................................................. 154

Saltbush-Greasewood ........................................... 155

Creosotebush ........................................................ 155

Creosotebush-Bursage ......................................... 156

Mesquite ................................................................ 156

Paloverde-Cactus Shrub ....................................... 156

Southwestern Shrubsteppe ................................... 157

Trans-Pecos shrubsteppe...................................... 157

Chaparral—Mountain Shrub .................................. 158

Chapter 7: Fire in Tropical and Subtropical

Ecosystems ........................................................... 161

by Ronald L. Myers

Understory Fire Regimes ................................................ 161

Major Vegetation Types ............................................. 161

Fire Regime Characteristics .................................. 163

Fuels ...................................................................... 164

Postfire Plant Communities ........................................ 164

Pine Flatwoods and Pine Rocklands ..................... 164

Pondcypress Wetlands .......................................... 167

Cabbage Palm Savannas and Forests .................. 169

Live Oak Forests ................................................... 169

Mixed Fire Regimes ........................................................ 169

Major Vegetation Types ............................................. 169

Fire Regime Characteristics and Vegetation

Response ........................................................... 169

Stand-Replacement Fire Regimes .................................. 170

Major Vegetation Types ............................................. 170

Fire Regime Characteristics and Vegetation

Response ........................................................... 171

Chapter 8: Global Change and Wildland Fire ............. 175

by Kevin C. Ryan

Changes Over Time ........................................................ 175

Climate, Weather, and Fire Interactions ......................... 177

Climate and Vegetation Interactions ............................... 178

Fire and Vegetation Interactions ..................................... 180

Uncertainty of Interactions: Can We Predict

the Future? .................................................................. 182

Chapter 9: Ecological Principles, Shifting Fire

Regimes and Management Considerations ....... 185

by James K. Brown

Ecological Principles ....................................................... 185

Fire Recurrence.......................................................... 185

Biodiversity ................................................................. 186

Plant Response to Fire .......................................... 186

Community and Landscape Responses to Fire .... 187

Ecological Processes ................................................. 189

Successional Pathways ......................................... 189

Decomposition ....................................................... 189

Fuel Accumulation ................................................. 190

Human Influences ...................................................... 192

Shifting Fire Regimes ...................................................... 193

iv
v

Forests and Woodlands ............................................. 193

Grasslands and Shrublands ....................................... 194

Managing Fire ................................................................. 195

Strategies and Approaches ........................................ 195

Historical Range of Variability ................................ 196

Process Versus Structure Goals ........................... 197

Landscape Assessment ........................................ 198

Restoration of Fire ...................................................... 198

Prescribed Fire and Silviculture ............................. 199

Understory Fire Regime Type .................................... 200

Mixed and Stand-Replacement Regimes ................... 200

Grazing and Exotic Plants .......................................... 200

Fire Prescriptions ....................................................... 202

Research Needs ............................................................. 203

Characteristics of Fire Regimes ................................. 203

Effects of Fire on Ecosystem Processes

and Biodiversity ...................................................... 203

Restoration of Ecosystems ........................................ 203

Development of Ecosystem Evaluation

Methodologies ........................................................ 203

References ..................................................................... 205

Appendices .................................................................... 239

Appendix A: Common and Scientific Names of

Plant Species .............................................................. 239

Appendix B: Succession Simulation Models ................... 247

Appendix C: Glossary ..................................................... 248

Index ............................................................................... 250
Summary

This state-of-knowledge review about the effects of fire

on flora and fuels can assist land managers in planning

for ecosystem management and fire management, and

in their efforts to inform others about the ecological role

of fire. Chapter 1 presents an overview and a classification of fire regimes that is used throughout the report.

Chapter 2 summarizes knowledge of fire effects on

individual plants, including susceptibility to mortality of

aerial crowns, stems, and roots; vegetative regeneration;

seedling establishment from on-site and off-site seed

sources; seasonal influences such as carbohydrates and

phenological stage; and factors affecting burn severity.

Five chapters describe fire regime characteristics such

as fire severity, fire frequency, and fire intensity, and

postfire plant community responses for ecosystems

throughout the United States and Canada. Typical fuel

compositions, fuel loadings, and fire behavior are described for many vegetation types. Vegetation types

including Forest-Range Environmental Study (FRES),

Kuchler, and Society of American Foresters (SAF) types

are classified as belonging to understory, mixed, or stand

replacement fire severity regime types. The severity and

frequency of fire are described for the pre-Euro-American settlement period and contrasted with current fire

regimes. Historic fire frequencies ranged from a fire

every 1 to 3 years in some grassland and pine types to a

fire every 500 to 1,000 years in some coastal forest and

northern hardwood types. In many vegetation types characterized by understory fire regimes, a considerable shift

in fire frequency and fire severity occurred during the past

century. Successional patterns and vegetation dynamics

following disturbance by fire, and in some cases related

grazing and silvicultural treatments, are described for

major vegetation types. Management considerations are

discussed, especially for the application of prescribed fire.

A chapter on global climate change describes the

complexity of a changing climate and possible influences

on vegetation, fuels, and fire. The uncertainty of global

climate change and its interactions with vegetation means

expectations for fire management are general and tentative. Nonetheless, manipulation of wildlands and disturbance regimes may be necessary to ensure continual

presence of some species.

The last chapter takes a broader, more fundamental

view of the ecological principles and shifting fire regimes

described in the other chapters. The influences of fire

regimes on biodiversity and fuel accumulation are discussed. Strategies and approaches for managing fire in

an ecosystem context and sources of technical knowledge that can assist in the process are described. Research needs are broadly summarized.

vi
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 1

James K. Brown

Chapter 1:

Introduction and Fire

Regimes

At the request of public and private wildland fire

managers who recognized a need to assimilate current

fire effects knowledge, we produced this state-of-theart integrated series of documents relevant to management of ecosystems. The series covers our technical understanding of fire effects, an understanding

that has expanded considerably since about 1980,

along with an awareness that fire is a fundamental

process of ecosystems that must be understood and

managed to meet resource and ecosystem management goals. The “Rainbow Series” of documents stresses

concepts and principles, provides an entry into the

relevant literature, and discusses management implications. The volumes in the series are intended to be

useful for land management planning, development of

environmental assessments and environmental impact statements, training and education, informing

others such as conservation groups and regulatory

agencies, and accessing the technical literature.

Knowledge of fire effects has attained increased

importance to land managers because fire as a disturbance process is an integral part of the concept of

ecosystem management. Fire—a disturbance that initiates change—affects the composition, structure, and

pattern of vegetation on the landscape. Disturbance is

necessary to maintain a diversity of living things and

processes. The old idea of plant communities and their

broader ecological systems reaching equilibrium is

being rejected by modern ecologists and resource managers (Botkin 1990; Morgan and others 1994).

Aldo Leopold (1949) recognized the principle of ecosystems decades ago, but it has only recently received

widespread recognition as an important process to

guide management of wildlands. Although the variety

of governmental and private organizations responsible for management of natural resources disagree on

an exact definition of ecosystem management, the goal

of sustainability is central to most approaches

(Christensen and others 1996). This goal focuses on

delivery of goods and services. Ecosystem management defined by Christensen and others (1996) is

management driven by explicit goals, executed by policies, protocols, and practices, and made adaptable by

monitoring and research based on our best understanding of the ecological interactions and processes

necessary to sustain ecosystem structure and function.

This definition puts the primary focus on sustainability of ecosystem structures and processes necessary to

deliver goods and services.
2 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Brown Chapter 1: Introduction and Fire Regimes

Ecosystem management broadens the focus of management from patches or stands to landscapes of

variable scale. It moves the focus from individual

ecosystem parts such as timber, water, range, fish,

and wilderness, to how the parts fit together and

function as a whole (Bormann and others 1994). It

embodies other concepts such as conservation of biodiversity, sustained yield of multiple resources, and

ecosystem health (Salwasser 1994). A guiding premise

for sustaining ecosystems and protecting biodiversity

put forth by Kaufmann and others (1994) is to manage

ecosystems to conserve the structure, composition,

and function of all elements, including their frequency,

distribution, and natural extinction. Fire effects are

woven through all aspects of this premise. An ecosystem can be defined as simply a place where things live

(Salwasser 1994) or in more detailed terms that relate

the interaction of organisms and physical environment through a flow of energy (Bormann and others

1994). Ecosystems contain components such as plants,

vegetative communities, and landforms, and processes

such as nutrient cycling. The dynamic nature of ecosystems and the scale of landscape patterns and processes are fundamental ecosystem characteristics that

managers must consider in integrating knowledge of

fire into the management of ecosystems.

Fire is a dynamic process, predictable but uncertain,

that varies over time and the landscape. Fire has

shaped vegetative communities for as long as vegetation and lightning have existed on earth (Pyne 1982).

Recycling of carbon and nutrients depends on biological decomposition and fire. In regions where decay is

constrained by dry and cold climates, fire plays a

dominant role in recycling plant debris. In warmer,

moist climates, decay plays the dominant role (Harvey

1994).

Lightning as a cause of fire over geologic time is

widely appreciated. But humans also have been a

major source of ignition, having used fire for various

purposes during the past 20,000 years (Wright and

Bailey 1982) and beyond. The pervasive influence of

intentional burning by Native Americans during the

past several centuries is probably not fully appreciated (Denevan 1992; Gruell 1985a). Human influence

was particularly significant in grasslands and those

communities bordering grasslands (Wright and Bailey

1982). Historically, fire caused by all ignition sources

occurred over large areas covering more than half of

the United States at intervals of 1 to 12 years; and fire

occurred at longer intervals over most of the rest of the

country (Frost 1998).

The land manager faces a complex challenge in

managing fire to achieve beneficial effects and avoid

unwanted results. Even attempts to eliminate harmful fire can over the long term cause undesirable

consequences, such as increased risk of damaging fire

and declining ecosystem health (Covington and others

1994; Mutch and others 1993). Thus, it is imperative

that the immediate and long-term effects of fire be

understood and integrated into land management

planning.

Flora and Fuel Volume ___________

The purpose of the Flora and Fuel volume is to assist

land managers with ecosystem and fire management

planning and in their efforts to inform others about the

role of fire and the basis for including fire as an

ecosystem management practice. The geographic area

covered in this series volume includes Canada and all

of the United States and adjoining Caribbean areas.

The contents focus on principles, generalities, and

broad scale fire effects on flora rather than on detailed

site specific responses. Vegetative response to individual fires can vary substantially depending on a host

of factors involving characteristics of the fire, existing

vegetation, site conditions, and postfire weather. The

value and indeed challenge in preparing this volume

was in providing a summary of fire effects that was

meaningful over broad areas even in view of highly

variable responses. Those wishing a more detailed

explanation of fire effects on flora are referred to

several textbooks on the subject (Agee 1993; Johnson

1992; Wein and MacLean 1983; Wright and Bailey

1982).

 Chapter 2 covers autecological effects of fire. Chapters 3 through 7 are about regional fire regime characteristics and postfire development of plant communities. Chapter 8 reviews the potential for climate change

and implications for fire management. Chapter 9

provides an overview of the ecological principles underlying fire regimes, shifts in fire regimes, and related management considerations.

The regional fire regime chapters are organized by

the following biogeographic regions:

• Northern ecosystems

• Eastern United States forests and grasslands

• Western forests

• Western shrublands, woodlands, grasslands

• Subtropical ecosystems

Each plant community chapter is organized by fire

regime type (understory, mixed, stand-replacement)

and similar subheadings. First, the fire regime characteristics are described, including fire severity, fire

frequency, fire size and pattern, and fuels and fire

behavior. This emphasizes the commonality of vegetation types that have similar fire regime characteristics

based on the dominant vegetation undergoing similar structural changes. Next, postfire plant communities are discussed with emphasis on temporal

changes in vegetation and fuels. Pre-1900 and post-1900
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 3

Chapter 1: Introduction and Fire Regimes Brown

subheadings are often used to help distinguish between succession occurring before and after organized

fire suppression. The 1900 date is an approximation of

when fire suppression effectively reduced the extent of

wildfire. In some vegetation types knowledge is insufficient to determine whether successional patterns

differ between the two periods. Last, management

considerations are described involving silvicultural

practices, prescribed fire, ecosystem restoration, and

other aspects of planned disturbance for maintaining

healthy ecosystems.

The word “fuels” refers to live and dead vegetation

that can potentially contribute to combustion. Fuel

quantities can vary from a small portion to all of the

aboveground biomass depending on a number of fuel

properties especially particle size, moisture content,

and arrangement. Although vegetation biomass increases predictably with time because of perpetual

photosynthesis, changes in fuel biomass over time can

be highly irregular due to the tradeoff between annual

increment and decay and properties affecting fuel

availability. In this volume fuels are described generally in terms of accumulation and flammability. Some

information on fuel loadings is presented primarily to

show typical values or a range in values that characterize various vegetation types. Fuel loading models

for major vegetation types can be found in the First

Order Fire Effects Model (FOFEM), which provides

quantitative predictions of tree mortality, fuel consumption, smoke emissions, and soil heating

(Reinhardt and others 1997). More detailed knowledge can be found in the literature referenced throughout this volume.

Plant community fire effects are discussed for broad

vegetation types. Forest and Range Environmental

Study (FRES) ecosystem types (Garrison and others

1977), which cover the 48 contiguous States, are used

at the broadest scale. Society of American Forester

cover types (Eyre 1980) are also used especially for

Canada and Alaska, and where more resolution of fire

effects knowledge is needed. FRES ecosystem types

are based on an aggregation of Kuchler’s (1964) Potential Natural Vegetation classes (American Geographical Society 1975). FRES, SAF, and Kuchler types that

have similar fire regime characteristics are grouped to

show synonymy (tables 3-1, 4-1, 5-1, 6-1, 7-1). Due to

the subjective nature of defining vegetation types,

some overlap occurs among types, particularly Kuchler

and SAF types. For example, some SAF types may be

reported in more than one Kuchler type. The correspondence of Kuchler and SAF types with FRES types

is also described in the Fire Effects Information System User’s Guide (Fischer and others 1996).

Scientific names of plant species referred to by

common names throughout the volume are listed

in appendix A. Appendix B describes the succession

simulation models. And the glossary of fuel, fire, and

plant reproduction terms is in appendix C.

Fire Regimes ___________________

“Fire regime” refers to the nature of fire occurring

over long periods and the prominent immediate effects

of fire that generally characterize an ecosystem. Descriptions of fire regimes are general and broad because of the enormous variability of fire over time and

space (Whelan 1995). Classification of fire regimes

into distinct categories faces the same difficulties and

a dilemma that underlie any ecological classification.

One difficulty is that putting boundaries around segments of biological processes that vary continuously

involves some degree of arbitrariness. The dilemma is

that for a classification to be useful to managers it

must be practical and easily communicated, thus free

of complexity. Yet to accurately reflect the nature of a

biological process, such as response to fire, it must

account for a complexity of interacting variables. A

tradeoff between practicality and accuracy or between

simplicity and complexity is required. The fire regime

concept brings a degree of order to a complicated body

of fire behavior and fire ecology knowledge. It provides

a simplifying means of communicating about the role

of fire among technical as well as nontechnical

audiences.

 Classifications of fire regimes can be based on the

characteristics of the fire itself or on the effects produced by the fire (Agee 1993). Fire regimes have been

described by factors such as fire frequency, fire periodicity, fire intensity, size of fire, pattern on the landscape, season of burn, and depth of burn (Kilgore

1987). The detail of a classification determines its best

use. The more detailed classifications are primarily

useful to ecologists and fire specialists attempting to

describe and understand the more intricate aspects of

fire. The simpler classifications are more useful for

broadscale assessments and for explaining the role of

fire to nontechnical audiences.

 Heinselman (1978) and Kilgore (1981) produced the

first classifications of fire regimes directed at forests.

Two factors, fire frequency and intensity, formed the

basis for their commonly referenced fire regime classifications (fig.1-1). A difficulty with fire intensity is

that a wide range of intensities, including crown fire

and surface fire, can cause stand-replacement because

mortality to aboveground vegetation is complete or

nearly complete. Fire intensity relates only generally

to fire severity. Severity of fire reflects (1) the immediate or primary effects of fire that result from intensity

of the propagating fire front and (2) heat released

during total fuel consumption. Plant mortality and

removal of organic matter are the primary fire effects.

Kilgore emphasized fire severity in his modification of
4 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Brown Chapter 1: Introduction and Fire Regimes

is very sensitive to fire frequency and its effects on

understory herbaceous species. Sensitivity to frequency

is provided by recognizing four frequency classes between 0 and 25 years. This separates some eastern and

western vegetation into different fire regime types.

The fire regime classification employed in this volume is based on fire severity. Characteristic fire frequencies are reported but not combined with fire

severity into classes. Use of fire severity as the key

component for describing fire regimes is appealing

because it relates directly to the effects of disturbance,

especially on survival and structure of the dominant

vegetation. It is intended for broadscale applications

and for communication about fire’s role among resource managers and others interested in natural

resources.

Detailed information available about past fire regimes is based mostly on biophysical evidence, written

records, and oral reports that encompass the period

from about 1500 to late 1800, a time before extensive

settlement by European-Americans in many parts of

North America, before intense conversion of wildlands

for agricultural and other purposes, and before fire

suppression effectively reduced fire frequency in many

areas. In this volume, we refer to the fire regimes of the

Heinselman’s fire regimes by referring to mortality of

the primary tree cover as stand-replacement.

Two recent fire regime classifications have proven

useful for mapping extensive areas of forest, shrubland, and grassland vegetation at 1 km resolution.

Morgan and others (1998) mapped historical and current fire regimes in the Interior Columbia River Basin

based on four fire severity and five fire frequency

classes (fig. 1-1). Hardy and others (1998) mapped fire

regimes of the Western United States using fire severity and fire frequency combined into five classes. They

keyed the fire regime classes to spectral images and

biophysical data including elevation, hydrologic units,

Kuchler’s vegetation types, and Bailey’s (1995) sections. Results were used to prioritize allocation of

funds and resources as part of a national strategy for

prescribed fire. For example, high priority for restoration using prescribed fire was assigned to areas where

current and historical fire regimes have departed

significantly such as in the ponderosa pine type.

Readers who wish to view a more complex ecological

classification are referred to the detailed classification

of fire regimes developed by Frost (1998). It incorporates periodicity of the fire cycle, primary season of

burn, fire frequency, and fire effects on vegetation. It

Heinselman

Frequent,

light surface fires (2)

Infrequent, 

light surface fires (1)

Infrequent,

severe surface fires (3)

 

Short-return interval,

Crown fires (4)

Very long-return interval,

Crown fires (6)

Long-return interval,

Crown fires (5)

No natural fires (0)

Morgan & Others

Nonlethal fires

(forest)

Nonlethal fires

(grassland)

Standreplacement

fires

(forest & shrublands)

Mixed-severity fires

(forest)

Rarely burns

Kilgore

Frequent,

low-intensity surface fires (1)

Infrequent, 

low-intensity surface fires (2)

Infrequent,

high-intensity surface fires (3)

Short-return interval,

stand-replacement fires (4)

Very long-return interval,

stand-replacement fires (6)

Variable: Frequent,

low-intensity surface &

long return-interval

stand-replacement fires (5) 

Flora Volume

Understory fires

(forest) 

Standreplacement

fires

(any vegetation type)

Mixed-severity

fires (forest)

Nonfire regimes

Hardy & Others

<35 yr.

Low-severity fires

(forest)

35-100+ 

Mixed-severity fires

(forest)

<35 yr. Standreplacement fires

(any vegetation type)

35-100+ yr. Standreplacement fires

(any vegetation type)

200+ yr. Standreplacement fires

(forest)

No burn

Figure 1-1—Comparison of fire regime classifications by Heinselman (1978), Kilgore (1981), Hardy and others (1998), Morgan and

others (1998), and the Flora and Fuel Volume. Lines connect similar fire regime types. In parentheses, forest includes woodlands

and grassland includes shrublands.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 5

Chapter 1: Introduction and Fire Regimes Brown

past several centuries as “presettlement” fire regimes.

The following describes the fire regime types used in

the Flora and Fuel Volume:

1. Understory fire regime (applies to forests and

woodlands)—Fires are generally nonlethal to the

dominant vegetation and do not substantially

change the structure of the dominant vegetation.

Approximately 80 percent or more of the

aboveground dominant vegetation survives fires.

2. Stand-replacement fire regime (applies to forests, woodlands, shrublands, and grasslands)—

Fires kill aboveground parts of the dominant

vegetation, changing the aboveground structure

substantially. Approximately 80 percent or more

of the aboveground dominant vegetation is either consumed or dies as a result of fires.

3. Mixed severity fire regime (applies to forests and

woodlands)—Severity of fire either causes selective mortality in dominant vegetation, depending on different tree species’ susceptibility

to fire, or varies between understory and standreplacement.

4. Nonfire regime—Little or no occurrence of natural fire.

In this volume, we consider all ecosystem types

other than forest and woodland to have stand-replacement fire regimes because most fires in those ecosystem types either kill or remove most of the aboveground

dominant vegetation, altering the aboveground structure substantially. Most belowground plant parts survive, allowing species that sprout to recover rapidly.

This is true of tundra, grasslands, and many shrubland ecosystems. Morgan and others (1998) consider

grasslands to have “nonlethal” fire regimes based on

the criterion that structure and composition of vegetation is similar to the preburn condition within 3 years

after a burn (fig. 1-1). Because fire radically alters the

structure of the dominant vegetation for at least a short

time, however, we consider grassland ecosystems to

have stand-replacement fire regimes. Because grassland, tundra, and many shrublands are stand-replacement fire regime types, a more interesting aspect of fire

regimes in these ecosystems is fire frequency, which can

vary substantially and have a major influence on vegetation composition and structure.

The understory and mixed severity fire regimes

apply only to forest and woodland vegetation types.

The mixed severity fire regime can arise in three ways:

• Many trees are killed by mostly surface fire but

many survive, usually of fire resistant species and

relatively large size. This type of fire regime was

described as the “moderate severity” regime by

Agee (1993) and Heyerdal (1997).

• Severity within individual fires varies between

understory burning and stand-replacement, which

creates a fine-grained pattern of young and older

trees. This kind of fire regime has not been recognized in previous classifications. It probably occurs because of fluctuations in weather during

fires, diurnal changes in burning conditions, and

variation in topography, fuels, and stand structure within burns (see chapters 5 and 6). Highly

dissected terrain is conducive to this fire regime.

In actuality, a blend of these two mixed severity

types probably occurs.

• Fire severity varies over time with individual fires

alternating between understory burns and standreplacement. Kilgore (1987) described this as the

“variable” regime and applied it to redwood forests. It also fits red pine forests (chapter 3).

The fire regime types were simplified from the classifications reported by Heinselman (1978) and Kilgore

(1981). They are identical to the fire severity component utilized by Hardy and others (1998) except we use

“understory” instead of “nonlethal” to depict that fire

regime. We chose the term understory as a fire regime

name because the term nonlethal is more easily misinterpreted when considering forest and nonforest

ecosystems. Our fire regime classification is similar to

that reported by Morgan and others (1998). To show

how all of these classifications are related, equivalent

or similar fire regime types are connected by lines in

figure 1-1. The primary ecological knowledge imparted

by fire regime types is whether fires leave the dominant aboveground vegetation standing and alive or

result in stand-replacement. To reflect this, the fire

regime types used in this volume, are characterized as

nonlethal understory fire, stand-replacement fire, and

mixed severity fire.

Fire severity is defined by what happens on areas

that actually burned. In reality, unburned islands and

patches of variable size and shape occur within the

perimeter of fires. In studies of historical fire, it is

difficult to separate burned from unburned patches.

Thus, in applying the classifications, some nonlethal

effects of fire can be attributed to unburned patches.

Forests of all types can be grouped into the understory, mixed, or stand-replacement fire regimes, which

correspond to low, moderate, and high fire severity

types described by Agee (1993). Some forest types

occurring over a wide range of environmental conditions can fall into two fire regime classes. For example,

most lodgepole pine and jack pine forests were characterized by stand-replacement fire. But some of the

forests, typically on drier sites, reflect a mixed fire

regime history. Evidence (Arno and others [in press];

Frost 1998) indicates that the mixed fire regime type

was more prevalent than previously thought especially in coniferous forests. As fire moves across the

landscape its behavior and effects can change dramatically due to variability in stand structure, fuels,
6 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Brown Chapter 1: Introduction and Fire Regimes

topography, and changing weather elements. This can

result in highly variable tree mortality and survival

patterns within a fire’s boundary. Generally, the severity and intensity of fire are inversely related to fire

frequency (Swetnam 1993). For example, stand-replacement fires tend to occur in forests with low

frequency, and understory to mixed severity fires tend

to occur in forests with high fire frequency. Considerable variability exists within this generalization.

In this volume we consider grasslands and tundra

fire regimes to be essentially all stand-replacement

because the aboveground dominant vegetation is either killed or removed by fire. Also, many shrubland

ecosystems are stand-replacement fire regime types

because the dominant shrub layer is usually killed

back to growing points in or near the ground. Standreplacement fire in grass and sedge dominated ecosystems may be either lethal or nonlethal to aboveground

vegetation. It is nonlethal if vegetative parts have

already cured and exist as dead fuel, which is often the

case in Western United States. But it is lethal if some

of the aboveground grasses and sedges are living and

are killed by fire as is commonly the case in marshes

of eastern North America and in tundra. Fire is usually nonlethal to belowground plant parts allowing

species that sprout to recover rapidly.

The natural role of fire can be understood and

communicated through the concept of fire regimes.

Significant changes in the role of fire due to management actions or possible shifts in climate can be

readily described by shifts in fire regimes. It is increasingly recognized that knowledge of fire regimes is

critical to understanding and managing ecosystems.

To assist in this, fire regime types are identified for the

major vegetation types in the United States and Canada

(tables 3-1, 4-1, 5-1, 6-1, 7-1). The prevalence of each

fire regime type within an ecosystem is characterized

as being of major or minor importance. Fire frequency

classes defined by Hardy and others (1998) are also

tabulated along with a range in fire frequencies where

there was sufficient knowledge.

To illustrate the extent and juxtaposition of various

fire regimes across the landscape, presettlement fire

regime types showing fire severity and fire return

intervals were mapped for the lower United States

(fig. 1-2). The mapping was based on a digitized atlas

of Kuchler’s Potential Natural Vegetation Types (Hardy

and others 1998) and the fire regime types ascribed to

the Kuchler types in chapters 3 to 7. In interpreting

the map, keep in mind that Kuchler types represent

broad classes; vegetative cover types and fire regime

types can vary within the Kuchler types. In the figure

legend, the overlapping of fire frequency classes such

as 0 to 10 and 0 to 35 years means that the broader

class encompasses more variability in fire return intervals and uncertainty of estimation.

The map illustrates the great expanse occupied by the

short return-interval, understory fire regime type in

the Eastern United States. It is important to note that

much of the presettlement oak-hickory type was a

savanna classified as forest having an understory fire

regime, but it reasonably could have been classified as

prairie having a short return-interval, stand-replacement fire regime. The pattern and frequency of the

mixed fire regime type varies substantially between

western conifers and eastern hardwoods. Although the

mixed regime mortality is similar, the fire behavior and

species fire resistance differ. Fires in conifers typically

are more intense than in hardwoods, but conifers have

a higher resistance to fire injury. Mapping of fire regime

types and changes between current and historical periods can be useful for broad-scale fire management

planning and for communicating with non fire managers about landscape fire ecology.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 7

Chapter 1: Introduction and Fire Regimes BrownFigure 1-2—Fire regime types based on Kuchler’s Potential Natural Vegetation types (prepared by Jim Menakis). Understory fires 0 to 10 years Understory fires 0 to 34 years Mixed severity fires 35 to 200 years Mixed severity fires 201 to 500 years Mixed severity fires 500+ years Stand replacement fires 0 to 34 years Stand replacement fires 500+ years Forest and Woodland Types Stand replacement fires 0 to 10 years Stand replacement fires 0 to 34 years Stand replacement fires 101 to 500 years Grass and Shrub Types Water Mixed severity fires 0 to 34 years Stand replacement fires 35 to 200 years Stand replacement fires 201 to 500 years Stand replacement fires 35 to 100 years Other Mixed severity fires 0 to 34 years
8 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Notes

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USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 9

Melanie Miller

Chapter 2:

Fire Autecology

F ̇ire is a key ecological process within most ecosystems in the United States and Canada. An understanding of factors controlling the initial response of

vegetation to fire is essential to its management. Fire

effects on plants can vary significantly among fires

and on different areas of the same fire. Fire behavior,

fire duration, the pattern of fuel consumption, and the

amount of subsurface heating all influence injury and

mortality of plants, and their subsequent recovery.

Postfire responses also depend upon the characteristics of the plant species on the site, their susceptibility

to fire, and the means by which they recover after fire.

This chapter describes the key elements that explain fire effects on vascular plants, those plants with

specific structures for gathering and transporting water

and nutrients. Effects on mosses, lichens, liverworts,

algae, and fungi are not discussed. The chapter addresses plant survival, resprouting, and seedling establishment in the initial stages of postfire recovery.

Factors that affect a species presence or absence in the

immediate postfire community will be described, but

not those that affect productivity, such as changes in

soil nutrient availability. The adaptations that allow

survival, and the methods by which plants recover, are

common to species found in almost all of the ecosystems discussed in this volume. The chapter describes

principles in a general way, and provides specific

examples from different ecosystems, although no

attempt has been made to present examples from

every system. Mechanisms operate in the same way no

matter where they occur.

Plant Mortality __________________

The likelihood of plant tissue being killed by fire

depends upon the amount of heat it receives. The heat

received by a plant is determined by the temperature

reached and the duration of exposure. Most plant cells

die if heated to temperatures between about 122 to

131 °F (50 to 55 °C) (Wright and Bailey 1982). Plant

tissue withstands heat in a time-temperature dependent manner. Mortality can occur at high temperatures after a short period (Martin 1963), while death at

lower temperatures requires a longer exposure (Ursic

1961). Additionally, some plant tissues, particularly

growing points (meristems or buds) tend to be much

more sensitive to heat when they are actively growing

and their tissue moisture is high, than when their

moisture content is low (Wright and Bailey 1982). The

concentration of other compounds that vary seasonally such as salts, sugars, and lignins may also be

related to heat tolerance of plants. Plant mortality

depends on the amount of meristematic tissues killed.

Susceptible tissue may not be exposed to heating by

fire because it is protected by structures such as bark

or bud scales, or is buried in duff or soil.
10 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Miller Chapter 2: Fire Autecology

Plant mortality is often the result of injury to several

different parts of the plant, such as crown damage

coupled with high cambial mortality. Death may not

occur for several years and is often associated with the

secondary agents of disease, fungus, or insects. The

resistance of plants to these agents is often lowered by

injury, and wound sites provide an entry point for

pathogens in conifers (Littke and Gara 1986) and

hardwoods (Loomis 1973). A plant weakened by

drought, either before a fire or after wounding, is also

more likely to die.

Aerial Crown Mortality

A woody plant’s structure affects the probability

that the aboveground portion will be killed by fire.

Important aerial crown characteristics include branch

density, ratio of live to dead crown material, location

of the base of the crown with respect to surface fuels,

and total crown size (Brown and Davis 1973). Height

enhances survival, as the aerial portions of small

stature plants are almost always killed. Species of

trees that self-prune their dead lower branches, such

as red pine, are less likely to have a fire carry into their

crowns (Keeley and Zedler 1998). Small buds are more

susceptible to lethal heating than large buds because

of their small mass (Byram 1948; Wagener 1961).

Large buds, such as on some of the pines, are more heat

resistant. The small diameter twigs and small buds of

most shrub species make them fairly susceptible to

fire. For conifers, long needles provide more initial

protection to buds than short needles that leave the

bud directly exposed to heat from the fire (Wagener

1961). Whether leaves are deciduous or evergreen

affects crown survival in that deciduous trees are

much less susceptible during the dormant than growing season.

In order for the aerial crown to survive fire, some

buds and branch cambium must survive. For conifers

with short needles and trees and shrubs with small

buds, crown scorch is often equivalent to crown death

because small buds and twigs do not survive (Wade

1986). The upper portions of the crown may survive on

taller trees. Large buds shielded by long needles can

survive fires that scorch adjacent foliage (Ryan 1990;

Wade 1986). The large shielded buds of ponderosa

pine, lodgepole pine, western white pine, and western

larch can survive at a 20 percent lower height than

that where foliage is killed (Ryan 1990). Crown consumption is a better indicator of crown mortality than

scorch for fire-resistant conifers such as longleaf pine,

which has long needles, large well protected buds, and

thick twigs (Wade 1986). Crown characteristics that

affect survival of trees after fire are listed in table 2-1.

The scorching of a tree crown is primarily caused by

peak temperature heat fluxes associated with the

passage of the flaming fire front (Van Wagner 1973).

Long-term heating caused by burnout of fuel concentrations after the flaming front has passed can also

scorch crowns. Whether the heat generated by fire is

lethal to foliage also depends on the ambient air

temperature (Byram 1958). For example, at a 90 °F air

temperature without wind, the height of foliage scorch

can be approximately 25 percent higher than it would

be at 77 °F, because at higher air temperatures less

additional heat is required to raise the foliage temperature to a lethal level (Albini 1976). Scorch is also

affected by the degree to which heat is dissipated by

wind (Van Wagner 1973). In western conifers, the

percent of crown volume with scorched foliage is a

better predictor of crown mortality than scorch height

because it is a better measure of the amount of remaining live foliage (Peterson 1985). In southern pine

species, nearly all trees can survive 100 percent crown

scorch except during the fall when survival is about 95

percent (Wade 1985; Weise and others 1990). Heatcaused needle damage is detectable within a few days,

sometimes within hours, and becomes more obvious

over the next several weeks (Ryan and Wade 2000).

Stem Mortality

In fires where aerial crowns are not burned, trees

and shrubs can be killed by girdling, caused by lethal

heating of the cambial layer, the active growth layer

just beneath the bark. Fire resistance of tree stems is

most closely related to bark thickness, which varies

with species, tree diameter and age, distance above

the ground, site characteristics, and health and vigor

of the tree (Gill 1995). Some species with thin bark

have a fairly thick collar of bark at the base of the bole

(Harmon 1984). The insulating quality of bark is also

affected by its structure, composition, density, and

moisture content (Hare 1965; Reifsnyder and others

1967), factors that vary among species. For example,

among central hardwoods, bark of silver maple has a

high specific gravity and thermal conductivity, and

can transmit heat to cambial layers in less time than

bark with a low specific gravity and conductivity, such

as bur oak and eastern cottonwood (Hengst and Dawson

1994). Flame length (Brown and DeByle 1987), flaming residence time (Wade 1986), and stem char height

(Regelbrugge and Conard 1993; Regelbrugge and Smith

1994) can be related to the amount of mortality of thinbarked trees. The cambium layer of thin-barked trees

such as lodgepole pine and subalpine fir is usually

dead beneath any charred bark (Ryan 1982). For

Northwestern conifers in natural fuel situations, minimum bark thickness associated with consistent tree

survival is about 0.39 inches (1 cm) (Ryan 1990). Wade

and Johansen (1986) noted that bark as thin as 0.5

inch (1.25 cm) could protect young loblolly and slash

pines during dormant season fires with low fireline
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 11

Chapter 2: Fire Autecology Miller Table 2-1

—Tree characteristics important to surviving fire and an overall species resistance to fire rating.a

Basal

bark Ability to Sizeb

thickness, regenerate when fire Fire

mature Branch Size Length vegetatively resistance resistance

Species trees density of buds of needles after fire is gainedc at maturity

Conifers

Pines

Digger pine Medium Low Medium Long None None Medium

E.white pine Thick Low Medium Medium None Mature Medium

Jack pine Thin Low Medium Short None None Low

Jeffrey pine Thick Medium Medium Long None Pole High

Loblolly pine Thick Medium Medium Long Root crownd Sapling High

Longleaf pine Thick Medium Large Long Root crownd Seedling High

Pinyon pine Thin Low Large Short None None Low

Pitch pine Thick Medium Medium Medium Root crown, Mature Medium

Stump sprouts

Pond pine Thick Medium Medium Long Root crown, Pole High

Stump sprouts

Ponderosa pine Thick Medium Large Long None Sapling/Pole High

Red pine Thick Low Large Medium None Pole Medium

Rocky Mt. Very Thin Low Medium Short None Mature Medium

Lodgepole pine

Sand pine Thin Medium Medium Medium None Mature Low

Shore pine Thin Medium Medium Medium None None Low

Shortleaf pine Thick Medium Medium Medium Root crown d Sapling High

Slash pine Thick Medium Large Long None Sapling High

Sugar pine Thick Medium Medium Medium None Mature Medium

Virginia pine Thin Medium Medium Short None None Low

W. white pine Medium Medium Medium Medium None Mature Medium

Whitebark pine Very Thin Medium Medium Medium None Mature Medium

Firs

Balsam fir Thin High Small Short None None Low

Douglas-fir, coast Very Thick High Medium Medium None Pole/Mature High

Douglas-fir, Thick High Medium Medium None Pole High

Rocky Mountain

Grand fir Medium High Medium Medium None Mature Medium

Noble fir Medium Medium Medium Medium None Mature Medium

Pacific silver fir Medium Medium Medium Medium None None Low

Subalpine fir Very Thin High Medium Medium None None Very Low

White fir Medium High Medium Medium None Mature Medium

(con.)
12 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Miller Chapter 2: Fire Autecology Junipers Alligator juniper Thin/Med Low Small Short Root crown, Mature Low/Med Stump sprouts, Roots E. redcedar Thin High Small Short None None Low Oneseed juniper Thin/Med Low Small Short None Mature Low/Med Utah juniper Thin/Med Low Small Short None Mature Low/Med W. juniper Thin Low Small Short None Mature Low/Med Other conifers Alaska-cedar Very Thin Medium Small Short None None Low Black spruce Medium High Small Short None Mature Low/Med Blue spruce Thin High Medium Medium None None Low Engelmann spruce Thin High Medium Medium None None Low Giant sequoia Very Thick Medium Small Short None Pole Very High Incense-cedar Thick High Small Short None Mature Medium Mt. Hemlock Medium High Small Short None None Low Redwood Very Thick Medium Small Short Root Crown, Sapling Very High Stump Sprouts Sitka spruce Thin Medium Medium Medium None None Low Tamarack Medium Medium Small Short None Mature Medium W. hemlock Medium High Small Short None None Low W. larch Very Thick Low Small Medium None Pole High W. redcedar Thin High Small Short None Mature Medium White spruce Medium High Small Short None Mature Medium Hardwoods Oaks Black oak Thin/Med ——— Root crown, Mature Low/Med Stump Sprouts Blackjack oak Thin/Med ——— Root crown, Mature Low/Med Stump Sprouts Blue oak Thin ——— Root crown, Mature Low/Med Stump Sprouts Bur oak Medium ——— Root crown, Mature Medium Stump Sprouts (con.) Table 2-1

—Con.

Basal

bark Ability to Sizeb

thickness, regenerate when fire Fire

mature Branch Size Length vegetatively resistance resistance

Species trees density of buds of needles after fire is gainedc at maturity
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 13

Chapter 2: Fire Autecology Miller (con.) California black oak Thin/Med ——— Root crown, Mature Low/Med Stump sprouts Canyon live oak Medium ——— Root crown, Mature Medium Stump sprouts Gambel oak Thin ——— Root crown, rhizomes None Low Northern red oak Medium ——— Root crown, Mature Medium Stump sprouts Oregon white oak Thin/Med ——— Root crown, Pole Medium Stump sprouts Post oak Thin/Med ——— Root crown, Mature Low/Med Stump sprouts Southern red oak Thin ——— Root crown, Mature Low/Med Stump sprouts Turkey oak Medium ——— Root crown, Mature Medium Stump sprouts White oak Thin/Med ——— Root crown, Mature Low/Med Stump sprouts Other hardwoods American beech Thin ——— Root crown, None Low Stump sprouts Roots American elm Thin/Med ——— Root crown, Mature Low/Med Stump sprouts Aspen Medium ——— Roots, Root collar Mature Low/Med Basswood Thin ——— Root crown, None Low Stump sprouts Bigleaf maple Thin ——— Root crown, None Low Stump sprouts Bitternut hickory Thin ——— Root crown, None Low Stump sprouts Black cottonwood Medium ——— Stump sprouts, Mature Low/Med Root crown Blackgum Thin ——— Root crown, Mature Low/Med Stump sprouts Eastern cottonwood Medium ——— Root crown, Mature Low/Med Stump sprouts Table 2-1

—Con.

Basal

bark Ability to Sizeb

thickness, regenerate when fire Fire

mature Branch Size Length vegetatively resistance resistance

Species trees density of buds of needles after fire is gainedc at maturity
14 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Miller Chapter 2: Fire Autecology Honey mesquite Thin ——— Root crown, Roots None Very low Mockernut hickory Thin ——— Root crown, None Low Stump sprouts Pacific madrone Thin ——— Root crown None Low Paper birch Medium ——— Root Collar None Low Persimmon Thin/Med ——— Root crown, Mature Low/Med Stump sprouts Pignut hickory Thin/Med ——— Root crown, Mature Low/Med Stump sprouts Red alder Thin ——— Stump sprouts Mature Low/Med Red maple Thin/Med ——— Root crown, Mature Low/Med Stump sprouts Southern magnolia Thin/Med ——— Root crown, Mature Med/High Stump sprouts Sugar maple Thin ——— Root crown (rarely) None Low Sweetgum Thin ——— Root crown, None Low Stump sprouts Tanoak Medium ——— Root crown, Pole Medium Stump sprouts White ash Thin ——— Root crown, None Low Stump sprouts Yellow-poplar Thin/Med ——— Root crown, Pole Med/High Stump sprouts

a The ratings of physical attributes are relative among the range of conditions observed for all tree species based on reviews of literature. b Sizes are defined as follows: seedlings, <1 inch dbh; saplings, 1 to 4 inch dbh; poles, 5 to 10 inch dbh; mature, >11 inch dbh. c Size when medium or high fire resistance is gained.

d For seedlings (loblolly, longleaf, and shortleaf pines) and saplings (loblolly and shortleaf pines). Shortleaf pine is a fairly strong sprouter; loblolly is weaker, and longleaf is the

Table 2-1

weakest of the three species (Wade 2000).

—Con.

Basal

bark Ability to Sizeb

thickness, regenerate when fire Fire

mature Branch Size Length vegetatively resistance resistance

Species trees density of buds of needles after fire is gainedc at maturity
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 15

Chapter 2: Fire Autecology Miller

intensity. A summary of tree bark characteristics

related to fire survival is in table 2-1.

Cambium that grows beneath thick bark layers

typically found on mature Douglas-fir, western larch,

and ponderosa (fig. 2-1), Jeffrey, longleaf, slash, and

loblolly pines is insulated from heat released by the

flaming front. However, the cambium can be killed by

long-duration heating, such as from burnout of logs

and smoldering combustion in deep litter and duff

layers (fig. 2-2). Complete basal girdling is generally

only caused by smoldering ground fires because the

amount and distribution of dead woody fuels is rarely

adequate to lethally heat the entire circumference of a

thick-barked tree (Ryan and Reinhardt 1988). The

deeper the basal mound of dry duff that is consumed,

the more likely that tree cambium is killed (Harrington

and Sackett 1992; Ryan and Frandsen 1991). In thickbarked trees, crown injury is more often the cause of

mortality than bole damage (Ryan and Reinhardt

1988).

Fire scars occur where the cambium is killed and

often are not evident until the dead bark sloughs from

the tree (Smith and Sutherland 1999). Because charring doesn’t happen unless the bark actually burns,

charring often doesn’t occur until a subsequent fire

burns the exposed surface. Once tree cambium is

injured by fire or mechanical damage, it is often more

susceptible to additional fire scarring, both because

the bark is thinner near the scar, and because of pitch

that is often associated with wounds. Fire scars can

become infected by wood-inhabiting microorganisms

Figure 2-1—Scorched boles on surviving ponderosa pine, including decay fungi. The survival of chestnut and

Selway-Bitterroot Wilderness, Idaho.

Figure 2-2—Smoldering and

glowing combustion in duff

can lethally heat tree boles

and roots such as in this Douglas-fir/western larch stand,

Lubrecht Experimental Forest, Montana.
16 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Miller Chapter 2: Fire Autecology

black oaks after surface fires in Eastern hardwood

forests has been attributed to their ability to rapidly

and effectively compartmentalize the wound, forming

a boundary around the injured and decayed tissue

that reduces the spread of infection (Smith and

Sutherland 1999).

Many large hardwoods survive fire but have charred

bark on the lee side, which in thin-barked species is a

telltale sign that the underlying cambium has been

killed. Even though the bark often remains intact for

1 or 2 years, the damaged sapwood begins to decay,

reaching the heartwood in several years and then

progressing upward at a more rapid rate. Height of

decay is directly correlated to age of wound (Kaufert

1933). On fast-growing bottomland hardwoods, wounds

less than 2 inches (5 cm) wide usually heal over before

rot enters, but larger wounds are nearly always infected, ruining the butt log (Toole 1959). Decayed

sapwood disintegrates rather quickly, creating the

hollow found on many old growth hardwoods in the

South. Most hollow trees also develop an enlarged

buttress. Toole (1959) found that bottomland hardwoods that initially survive fire suffer considerable

mortality over the next several years from breakage of

decay weakened stems. Loomis (1973) presented methodology for predicting basal wound size and mortality

to surviving trees in oak-hickory stands.

Root Mortality

Structural support roots growing laterally near the

surface are more susceptible to fire damage than those

growing farther beneath the surface. Roots found in

organic layers are more likely to be consumed or

lethally heated than those located in mineral soil

layers. The locations of structural roots are summarized for important tree species in table 2-1.

Feeder roots collect most of a tree’s water and nutrients, are small in diameter, and are usually distributed near the surface. Feeder roots located in organic

soil layers are more subject to lethal heating and

consumption than those located in mineral soil. Loss of

feeder roots may be a more significant cause of tree

mortality than structural root damage (Wade 1993).

Feeder root death may not always kill the tree, but it

can place the tree under significant stress. Increased

amounts of root damage can result from fires that

smolder in accumulations of litter beneath trees

(Herman 1954; Sweezy and Agee 1991; Wade and

Johansen 1986). This can be a critically important

factor if most of the feeder roots are located in thick

duff layers, caused by the exclusion of fire or a regime

of dormant season prescribed burning that consumed

hardly any surface organic matter. There may be

enough root injury or death to kill trees and shrubs,

even though little or no damage is apparent to their

aerial crowns (Geiszler and others 1984). While tree

crown mortality can be related to fireline intensity,

mortality of buried plant parts depends much more on

the duration of all phases of combustion that regulates

the downward heat pulse, than on the duration of the

flaming front (Wade 1986).

Fire Resistance

Tree resistance to fire generally increases with age.

Crowns become larger and for some species, the height

to the base of the live crown increases, either from self

pruning or removal of basal branches by surface fires.

Bark thickness and stem diameter increase. A suppressed tree may develop fire resistance characteristics at a much slower rate than a vigorous tree of the

same age and species resulting, for example, in a much

thinner bark in suppressed loblolly pine (Wade 1993).

The growth stage at which important species of trees

become fire resistant and the degree of resistance of

mature trees are summarized in table 2-1.

Vegetative Regeneration _________

Sprouting is a means by which many plants recover

after fire. Shoots can originate from dormant buds

located on plant parts above the ground surface or

from various levels within the litter, duff, and mineral

soil layers (fig. 2-3). The type of plant parts that

support dormant buds and where they are located in or

above the soil are species-specific characteristics (Flinn

and Wein 1977). The plant structures that give rise to

regenerating shoots are summarized for different life

forms of North American native plants in table 2-2.

Figure 2-3—Various plant parts that regenerate new shoots

and their location in and above the soil.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 17

Chapter 2: Fire Autecology Miller

Table 2-2—Type and location of buds that regenerate new shoots after fire by broad plant groups (X = common, u = uncommon).

Trees Herbs

Broad-leaf Perennial

Bud type Location Conifer evergreen Deciduous Shrubs forbs Grasses

Epicormic Aerial u u

Stolon Above soil & duff X X

Root collar In & above soil u X X X

Root crown In & above soil u X X X X

Caudex In soil or duffa X

Root In soil or duffa XXX X

Rhizome In soil or duffa XXX

Bulb, corm In soil X

a Budding organs may grow into duff after it accumulates to suitable depth.

Figure 2-4—Sprouts of paper birch that developed from root

collar, Frenchman Lake, Alberta.

Sprouting of Woody Plants

Dormant Bud Locations—Many woody plant species have dormant buds located in the tissue of stems,

above or below the surface of the ground. These plants

sometimes sprout from the root collar, the point where

roots spread out from the base of the stem. Such

species include antelope bitterbrush, bigleaf maple,

rabbitbrush, mountain mahogany, turkey oak, northern red oak, and paper birch (table 2-1, fig. 2-4).

Epicormic sprouts develop in species such as eucalyptus, pond pine, and pitch pine from buds buried in

woody tissue of tree stems, or from bud masses present

in branch axils. Lignotubers, burls, and root crowns

are names for masses of woody tissue from which roots

and stems originate, and that are often covered with

dormant buds (James 1984). These buds may be deeply

buried in wood, and may be located far below the

surface if the tissue mass is large. Plants with this

commonly occurring structure include white sage

(Keeley 1998), chamise, willow, serviceberry, alder,

and tanoak.

An unusual trait shared by pitch, pond, and shortleaf pines is the formation of a basal crook that enhances the ability of these species to produce basal

sprouts when the stems are topkilled by fire. When

seedlings are small, they fall over (presumably from

their own weight), grow prostrate, and then resume

vertical growth, which results in a basal crook at the

soil surface (Little and Somes 1956). Primary needles

with their axillary buds form just above the hypocotyl

and just below the second bend of the crook (Stone and

Stone 1954; Walker and Wiant 1966). Rootlets also

form from the uppermost root tissue close to the bud

cluster anchoring the stem in place. Buds on the lower

side of the crook are thus well protected from fire. If

fire topkills a seedling or sapling, these dormant buds

sprout and the same growing process is repeated.

Because sprouts originating after the second or third
18 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Miller Chapter 2: Fire Autecology

fire have a well-developed root system, their height

growth is more rapid than that of the original seedling.

Dormant buds are often located on laterally growing

stems or roots of woody plants. Some woody species,

such as aspen and horsebrush, have dormant buds or

bud primordia located along roots from which new

shoots can originate. Rhizomes are the horizontal

underground stems that have a regular network of

dormant buds that can produce new shoots and adventitious roots (Welsh and others 1987; Zasada and

others 1994). Woody rhizomatous species include blue

huckleberry, bog blueberry, thimbleberry, white spirea,

Gambel oak, creeping barberry, chokecherry, and Labrador tea (fig. 2-5).

Sprouting Process—Postfire sprouting in woody

plants is a process that is regulated by the same factors

that control vegetative regeneration after other types

of disturbances. Consider the physiological interactions that produce new aspen shoots, a model summarized by Schier and others (1985) that likely applies to

other plants with buried regenerating structures. The

growth of most dormant buds or bud primordia is

controlled by a phenomenon called apical dominance.

Growth hormones, particularly auxin, a plant hormone manufactured in actively growing stem tips and

adjacent young leaves, are translocated to dormant

buds, which prevent them from developing into new

shoots. If stem tips and leaves are removed, the source

of growth hormones is eliminated. The balance of

plant hormones within the buds changes. Growth

substances in roots, particularly cytokinins, are translocated upward to the buds and can cause the dormant

buds to sprout, or stimulate bud primordia to

differentiate into shoots. Cytokinins may already be

present in buds, and a decrease in the ratio of auxins

to cytokinins provides the stimulus for bud outgrowth.

Fire initiates regeneration from buds by killing

surface plant parts that inhibited their growth. The

buds that become shoots are usually those nearest to

the part of the plant killed by the fire. If dormant buds

are destroyed, new buds may differentiate from wound

tissue, called callus, and subsequently produce shoots

(Blaisdell and Mueggler 1956). Once new shoots are

actively growing, they produce growth hormones that

are translocated to other dormant buds that are farther away from the point of damage, suppressing their

growth (Schier 1972) (fig. 2-6).

The reduced understory cover and thickness of organic layers following fire can increase light near the

surface, and in turn promote an increase in sprouting

because light can cause rhizome tips to turn upward

and develop leafy shoots once they reach the surface

(Barker and Collins 1963; Trevett 1956). This possibility suggests that some postfire shoots may develop

from rhizome tips, not dormant buds. Schier (1983)

found that decapitating a rhizomatous plant caused

laterally growing rhizomes to turn upward and become shoots. Additional rhizomes often form in response to vigorous aerial plant growth (Kender 1967),

and may subsequently produce aboveground shoots

(fig. 2-7). Sprouts from new rhizomes may recolonize

areas where old rhizomes and other reproductive plant

parts were killed by a fire. Plants may sprout soon

after a fire, or not until the following spring if the fire

occurs after the plants have become dormant (Miller

1978; Trevett 1962). Warmer soil temperatures following fire may enhance the amount of sprouting that

Figure 2-5—New shoot growth

from a rhizome of Labrador tea,

Seward Peninsula, Alaska.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 19

Chapter 2: Fire Autecology Miller

Figure 2-6—Suppression of bud outgrowth farther down the

stem by actively growing new shoot of blue huckleberry, Lubrecht

Experimental Forest, Montana.

Figure 2-7—New rhizomes formed on postfire aster sprout that

may colonize adjacent areas, East Kootenay Mountains, British

Columbia.

occurs (Zasada and Schier 1973). The initial energy

required to support growth until the sprout is photosynthetically self-sufficient comes from carbohydrates

and nutrients stored in the regenerating structures or

in adjacent roots (James 1984).

Postfire sprouting ability can vary with plant age.

Young plants that have developed from seed may not

be able to sprout until they reach a certain age, which

varies by species (Smith and others 1975; Tappeiner

and others 1984). Older plants of some species may be

able to produce few, if any, sprouts that survive. Older

plants (80+ years) of other species such as pitch pine

(Little and Somes 1956) can produce stump sprouts

prolifically. Minimal postfire root sprouting such as

documented in deteriorating aspen stands may be

caused by a combination of root system dieback and

continued inhibition of sprouting by residual stems

(Schier 1975). The ability of selected tree species to

vegetatively regenerate after a fire, and the structure

from which the sprouts develop are summarized in

table 2-1.

Sprouting and Burn Severity—Burn severity

(also called depth of burn and ground char, see glossary) is a measure of the amount of fuel consumption

and associated heating at and below the ground surface (also see fire severity in glossary). It is a function

of the duration of the fire, and relates closely to the

amount of surface fuel, litter and duff consumption,

and their moisture content. Severity classes have been

defined by Viereck and others (1979) and Ryan and

Noste (1985). A strong relationship exists between

subsurface heating and postfire sprouting in forested

areas (Dyrness and Norum 1983; Miller 1977; Morgan

and Neuenschwander 1988; Ryan and Noste 1985),

and in rangeland shrubs (Zschaechner 1985), which

can be related to the distribution of buried buds (Gill

1995). Figure 2-8 depicts the relationship between the
20 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Miller Chapter 2: Fire Autecology

amount of postfire sprouting of a rhizomatous shrub

and the depth of lethal heat penetration.

A low severity fire (lightly burned, short duration,

low ground char) that only consumes some of the

surface fuels may kill laterally growing rhizomes or

roots near the surface, or stem buds that are not well

protected. It has little effect on most buried plant parts

and can stimulate significant amounts of postfire

sprouting.

A moderate severity fire (moderately burned, moderate duration, moderate ground char) consumes the

litter layer, and partially consumes both large woody

debris and the duff layer. It incinerates plant structures in litter and the upper duff layer, such as shallow

rhizomes, and may kill buds on portions of upright

stems that are beneath the surface, and buds on the

upper part of root crowns (fig. 2-9). Sprouting occurs

from buds in deeper duff or soil layers. Johnston and

Woodard (1985) found that mortality of rhizomes of

beaked hazel and red raspberry only occurred in areas

of relatively high surface fuel loading, but sprouting

still occurred from more deeply buried rhizomes that

survived the fire. Moderate severity fires frequently

cause the greatest increase in stem numbers (fig. 2-8)

of root sprouters such as aspen (Brown and Simmerman

1986) and of rhizomatous shrubs (Miller 1976). When

heat prunes rhizomes below the surface where rhizome density is high (fig. 2-8b), shoots develop from

buds along the rhizome, and appear as separate plants

above the surface. Several shoots can replace what

was previously one shoot. Other new shoots may

develop from rhizome tips that are stimulated to turn

to the surface and become shoots.

A high severity fire (heavily burned, long duration,

deep ground char) removes the duff layer and most of

the large woody debris, particularly rotten material. It

can eliminate species with regenerative structures in

the duff layer, or at the duff-mineral soil interface, and

may lethally heat some plant parts in upper soil

layers, particularly where concentrations of heavy

fuels or thick duff layers are consumed. Any resprouting

that does occur on heavily burned microsites can only

occur from stolons and rhizomes that recolonize from

adjacent areas or from deeply buried plant parts (fig.

2-8c). Miller (1977) observed that sprouting was delayed on a severely burned microsite, with a single

huckleberry sprout not emerging until the third growing season from a rhizome about 9 inches below the

surface. Abundant vegetative regeneration can still

develop from species with deep roots such as aspen, or

deep rhizomes such as Gambel oak.

Sprouting of Forbs

In North America north of the tropics, perennial

forbs are broad leaved species that completely regrow

their leaves and stems each year, after dying back

during winter cold or summer drought. Some of these

plants can persist in closed canopy environments that

develop in the years after fire, producing a few leaves

at the beginning of each growing season, which are

often eaten by animals, or die back as the soils dry out

(Christensen and Muller 1975).

Forbs have regenerative structures that are similar

to those in woody plants, but also some that are

unique. Stolons are stems of herbaceous species that

grow on or near the surface of the ground, producing

plants with roots at the node apex such as a series of

strawberry plants (Benson 1957; Welsh and others

1987), and twinflower (McLean 1969). Dormant buds

of fireweed and bracken fern are located on roots.

Western yarrow, heartleaf arnica, showy aster, wild

sasparilla, and star-flowered Solomon’s seal are rhizomatous forb species.

A caudex is a largely underground, often woody stem

base that persists from year to year and produces

leaves and flowering stems (Benson 1957; Welsh and

others 1987). This structure is found in species such as

Indian paintbrush, lupine, wild columbine, and

arrowleaf balsamroot. Other buried reproductive structures include bulbs (buried buds covered with thick

fleshy leaves) found in common camas and deathcamas

and corms (bulb-like, short thickened stems) of glacier

lily and gayfeather species.

Figure 2-8—Effects of subsurface heating on postfire sprouting of a rhizomatous shrub: (a) Network of unburned rhizomes.

(b) Moderate severity fire lethally heated the rhizomes near the

surface, but many postfire sprouts grew from more deeply

buried rhizomes. (c) High severity fire killed most of the rhizomes, but a few sprouts grew from a deep, surviving rhizome.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 21

Chapter 2: Fire Autecology Miller

Figure 2-9—Sprouts originating from root crown of serviceberry, after a moderate severity fire killed buds on uppermost exposed

surfaces, Piceance Basin, northwestern Colorado.

Leaves and stems suppress outgrowth of subsurface

dormant buds. If fire occurs during the growing season, death of the apical meristems removes inhibition

of subsurface buds and new shoots form. If a fire occurs

when herbaceous plants are seasonally dormant, fire

does not remove the source of inhibition because

aboveground leaves and stems are cured and sometimes already decomposed. Dormant structures will

grow new leaves after the occurrence of appropriate

seasonal cues of temperature and moisture. Whether

herbaceous plants recover after fire depends largely

on whether their regenerative structures are exposed

to lethal temperatures. Similar to woody plants, their

survival depends on depth below the surface, whether

they are located in combustible material, and the

subsurface moisture regime at the time of the fire.

While the stolons of strawberries make them susceptible to even low severity fire, the deep rhizomes of

showy aster allow survival of some plants after fairly

severe fires. Lupine and timber milkvetch can regenerate even when the entire plant crown is consumed

(McLean 1969). Fireweed and bracken fern can produce

significant numbers of sprouts after high severity fires

because many buds far below the surface can survive

even severe fire treatments (Frye 1934; Moss 1936). In

California, 57 of 58 herbaceous perennial species

resprouted after wildfire in chaparral (Keeley 1998).

Sprouting of Grasses

New leaf tissue of grasses forms at the meristems

during the active growing period, and resumes after

summer quiescence or winter dormancy. New growth

also may occur by “tillering,” branching from dormant

axillary buds in the plant crown or on rhizomes. Grass

plants are killed when all meristems and buds are

lethally heated. Cool-season grasses that green up

early in the growing season can be killed by the

burning litter of associated warm-season grasses that

are still dormant and more heat resistant (Anderson

1973). Perennial grasses also may be killed if fire

burns in the cured litter of annual grasses while

perennials are still actively growing (Wright and

Klemmedson 1965).
22 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Miller Chapter 2: Fire Autecology

Stoloniferous and Rhizomatous Grasses—

Stoloniferous grasses are frequently killed by fire

because most stolons are at or near the surface. Whether

rhizomatous grasses are stimulated or killed by fire

depends on rhizome depth below the surface, whether

rhizomes are located in mineral or organic soil layers,

the moisture content of these layers, and the amount

and duration of heat generated by the surface fire.

Rhizomatous grasses such as western wheatgrass

often respond positively to rangeland fires because

meristems and buds are usually protected by soil, and

a long duration source of surface heat over large,

contiguous areas is rarely present (Wright and Bailey

1982). In forested areas, grass rhizomes are more

likely to be located in litter or duff layers or in association with dead woody fuels. On sites where fire consumes the duff layer, grass rhizomes located in the

duff, such as pinegrass may be killed (S. A. Snyder

1991). However, some rhizomes often survive in deep

mineral soil layers and can rapidly re-colonize severely burned areas.

Bunchgrasses—Meristems and dormant buds of

different bunchgrass species can be located within the

bunch above the level of the soil, or at various depths

below the soil surface. Buds and meristems can be

readily exposed to lethal temperatures, or be fairly

well protected if deeply buried in unburned organic

materials or in soil. For example, buds in the fairly

compact root crown of Idaho fescue lie at or above the

surface of the ground and are easily killed (Conrad and

Poulton 1966), while basal meristems of bottlebrush

squirreltail, Thurber’s needlegrass (Wright and

Klemmedson 1965), and wiregrass lie about 1.5 inches

(4 cm) below the mineral soil surface and are more fire

resistant (Uchytil 1992) (figs. 2-10, 2-11).

The moisture content of bunchgrass plants and

adjacent fuels affect the amount of heat that the

meristems receive. If plants are actively growing,

their foliar moisture content can be too high to allow

fire to enter the stand of plants. If bunchgrass crowns

are moist, it is unlikely that they will ignite and burn.

If the dead center of a bunchgrass plant is dry, it can

ignite, smolder, and burn, killing most or all growing

points. Heat from burning shrubs can dry and preheat

adjacent bunchgrass clumps to ignition temperature,

causing higher bunchgrass mortality than on a similar

site with few shrubs that burned under the same

conditions (Zschaechner 1985).

Dry bunchgrass crowns that are ignited are not

always consumed. Midsummer fires in northwest Colorado burning under high windspeeds charred only the

tops of the crowns of bluebunch wheatgrass and Indian

ricegrass plants that were 4 to 5 inches (10 to 13 cm) in

diameter. In this case, despite dry conditions, fire may

have moved through the grass litter too quickly to

ignite the root crowns, resulting in little plant mortality (Petersburg 1989).

Wright (1971) discussed the relationship between

stem coarseness and the rate at which a bunchgrass

clump burns. Fire tends to pass fairly quickly through

coarse-stemmed bunchgrasses, which do not have much

fuel concentrated at their base near reproductive structures. Fine-stemmed grasses with a dense clumping of

basal stems can burn slowly and generate a fair

amount of heat that can be transferred to meristems

and buds. Fires tend to burn more rapidly through

small-diameter bunches compared to large-diameter

bunches. Larger bunches usually have more dead fuel

and are thus more likely to produce enough heat to kill

growing points (Wright and Klemmedson 1965). These

relationships support Petersburg’s (1989) observation

that high rate of fire spread may have contributed to

the survival of moderate diameter bunches of fine

textured grass in northwestern Colorado that otherwise may have suffered fairly high mortality.

A rangeland bunchgrass grows where little surface

fine fuel surrounds the plants, other than the dead

grass blades immediately associated with the grass

crown. The amount of postfire sprouting in this environment can be related to the amount of growing point

mortality (Conrad and Poulton 1966). For those species having meristems above the mineral soil surface,

the highest postfire sprouting potential usually is

found in those plants with only some surface litter

removed. Sprouting decreases as the amount of basal

litter consumption increases, with new shoots tending

to appear only from the outside edge of the bunch when

little unburned stubble remains. Mortality is most

likely to occur if all plant material above the root

crowns is consumed.

Seedling Establishment __________

Seedling establishment is affected by the amount of

seed present and conditions required to induce germination and provide a favorable environment for initial

seedling growth. The interaction between the seed and

its environment determines whether it successfully

germinates and establishes. Requirements for successful germination and establishment can differ significantly among species.

Seed Supply and Dispersal

The supply of seeds of a given species is greatly

influenced by annual seed production, which can vary

significantly (Zasada and others 1978). Conifer regeneration may be limited if cone crops are poor during the

time when exposed mineral soil seedbed is present.

Surviving plants on or near the burned area may be too

young (Barney and Frischknecht 1974; Zasada 1971)
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 23

Chapter 2: Fire Autecology Miller

Figure 2-10—Sand dropseed producing postfire growth from root crown meristems at the soil surface,

Guadalupe Mountains National Park, western Texas

Figure 2-11—Basin wildrye growing after a fire from meristems below the soil surface, Bighorn Mountains,

Wyoming.
24 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Miller Chapter 2: Fire Autecology

or too old to produce much viable seed. Species of pines

occurring in high fire frequency habitats generally

begin producing cones earlier than other pine species

(Keeley and Zedler 1998).

The timing of seed dispersal is a species-specific

characteristic that varies with elevation and latitude

(Zasada 1986). The occurrence of fire with respect to

the dispersal of seeds can determine the rapidity of

regeneration. Heat from fire may kill seeds that have

recently fallen to the ground, preventing establishment of that species until after the next year’s seedfall.

Seeds that are available to recolonize a burned site

may have originated on-site or been dispersed from

off-site after the fire. On-site seeds may come from

surviving trees, from plants that grow after the fire, or

from seed stored in the soil before the fire. The amount

of off-site seed dispersal from unburned areas depends

on the amount of available seed, the distance of the

seed source from the burned area, the prevailing wind

direction, and the type of seed. Seed dispersal mechanisms vary. Light seeds may be carried aloft while

heavier seeds may skid across the surface of the snow.

Some seeds have wing-like structures that enhance

their movement through the air. Seeds with barbs or

hooks may be carried in fur or feathers. Hard-coated

seeds ingested along with their fruit pass through the

bird or animal, sometimes with an enhanced likelihood of germination. Mature capsules of some species

explosively release their seed (Parker and Kelly 1989).

Animals and birds can disperse seeds at great distances from the parent plant, with Clark’s nutcrackers

(Nucifraga columbiana) observed to carry pinyon pine

seed up to about 9 miles (Chambers and others 1999).

After dispersal, many seeds remain on or near the

surface, although gravity, freezing and thawing,

litterfall, and foraging activities of mammals, birds,

and insects can deeply bury seeds (West 1968; Tomback

1986). Clark’s nutcrackers, pinyon jays (Gymnorhinus

cyanocephalus), squirrels, and mice cache a significant proportion of seed of certain species below the

surface. There, seeds may exist within a matrix of soil,

organic material, or a mixture of both.

Seedbank

The seedbank, the supply of seeds present on a site,

is composed of transient and persistent seeds, which

may be in litter and soil layers and in the tree canopy

(table 2-3). There may be an enormous reserve of seed

in the seedbank. Seed supply of various species and

inherent seed longevity both affect the numbers of

viable seeds. Some plants produce seeds that are a

transient part of the seedbank, such as willow, which

may remain viable for only a few weeks. There is little

or no annual carryover of pine seed in soil seedbanks

(Pratt and others 1984), and few conifer seeds are

present in the forest floor of a mature forest (Archibold

1989; Ingersoll and Wilson 1990; Kramer and Johnson

1987). Many seeds, particularly large ones, are lost

from the seedbank by predation.

Soil-Stored Seed—In a ponderosa pine community, viable seeds of most grass and annual forb species

were found mostly in the litter layer, indicating short

longevity or recent dispersal, while seeds of perennial

forb species were found mostly in mineral soil, and

probably were fairly long-lived (Pratt and others 1984).

Seeds of some species persist in the soil for years after

dispersal. Seeds of pincherry can survive in the seedbank for up to 100 years after the parent trees have

died out of the overstory (Whittle and others 1997),

while snowbrush ceanothus seeds remain viable for

200 to 300 years or more (Noste and Bushey 1987).

Species present in the seedbank of mature deciduous (Pickett and McDonnell 1989) and coniferous

forests (Archibold 1989) are often shade-intolerant,

early seral species, which may not be present in the

overstory or understory. Few large seeded or shade

tolerant species reside for long in the deciduous forest

seedbank (Pickett and McDonnell 1989). In many

grassland communities, there is a distinct difference

between the species growing on the site and those

present in the seedbank (Rice 1989). Seedbanks tend

to contain more annual than perennial species, more

forbs than grasses, many leguminous species, and

more weedy species that colonize disturbed sites. In

contrast, the chaparral seedbank generally reflects

the composition of the standing vegetation with large,

persistent seed banks of many species of dominant

shrubs, although significant numbers of “fire-following” annuals may also be present (Parker and Kelly

1989). The life-forms of plants likely to have soil and

canopy stored seed are shown in table 2-3.

Canopy-Stored Seed—Serotinous cones of species

such as lodgepole, jack, pitch, Table Mountain, and

pond pines retain some of their seeds because of the

presence of a resin bond between scales on some of

their cones. Serotinous cones slowly open and release

their seeds after they are heated to at least 113 to

122 °F (45 to 50 °C), a temperature that melts the resin

bond (Lotan 1976). Cones protect a significant portion

of pitch pine seeds from the high temperatures reached

during fire (Fraver 1992). Lodgepole pine seeds survived in cones heated in flames for a length of time

typical of crown fires (Despain and others 1996). Numerous viable lodgepole pine seeds are dispersed even

after a long duration crown fire. There is considerable

variation in the amount of lodgepole cone serotiny,

both on individual trees (fig. 2-12), and geographically. A high degree of cone serotiny is likely to occur

where there are large, stand-replacement fires; relatively short, fire-free intervals; and fire sizes large
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 25

Chapter 2: Fire Autecology Miller

enough to limit seed dispersal from unburned areas

(Muir and Lotan 1985; Parker and Kelly 1989).

The semiserotinous cones of black spruce open and

release their seeds over a period of years (Zasada

1986). Cones are usually bunched near the top of the

tree, which shields some cones from heating and provides a postfire seed source. An additional on-site seed

source from canopies may be immature cones that

survive a fire and continue to ripen. This has been

observed in ponderosa pines with scorched foliage

(Rietveld 1976), in white spruce with boles completely

girdled by fire (Zasada 1985), and in scorched cones of

western larch and Douglas-fir where the tips of the

seed wing had been singed (Stickney 1999).

Seed Environment

The seed environment describes the microsite in

which a seed rests after it has been dispersed from the

parent plant, including seedbed, temperature, humidity, shade, and potential competition from other plants.

Moss, litter, and duff are poor seedbeds in many

Table 2-3—Occurrence of transient seeds and persistent seeds consisting of soil and canopy stored and

fire stimulated seed germination for broad plant species groups in the United States and

Canada.

Persistent seed

Soil Canopy Fire

Species group Transient seed stored stored stimulated

Conifer trees X X

Broad-leaf evergreen trees X

Deciduous trees X

Shrubs X X X

Annual forbs X X X

Perennial forbs X X X

Grasses X

Figure 2-12—Jack pine branch with serotinous and nonserotinous cones. Jack pine is ecologically similar

to and hybridizes with lodgepole pine, Acadia National Park, Maine.
26 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Miller Chapter 2: Fire Autecology

Table 2-4—Seedling establishment on moderate and heavily burned seedbeds in an

Interior Alaska black spruce forest.

Number of germinants Number of year 3 survivors

Moderately Heavily Moderately Heavily

Species burned burned burned burned

Alder 33 160 0 65

Paper birch 72 875 6 527

Balsam poplar 17 71 0 39

Bebb willow 105 144 0 46

climates because they frequently dry out in the summer, resulting in seedling death if roots have not yet

reached mineral soil. Organic seedbeds, even rotting

logs, may be able to successfully support seedling

establishment and survival if water is not limiting

during the growing season (Zasada 1971). Other attributes of organic seedbeds such as the presence of

allelopathic chemicals may inhibit seedling establishment. Oak litter has been observed to be a mechanical

barrier to Table Mountain pine regeneration, although

it enhances germination and survival of oak seedlings

(Williams and Johnson 1992).

Fire creates significant changes in site conditions,

which can vary substantially within the burned area

depending on the severity and pattern of the fire.

Consumption of fuel, especially the forest floor, is an

important determinant of postfire conditions, because

it controls the amount and distribution of good seedbed conditions. Where bare mineral soil seedbeds are

created, any allelopathic chemicals are volatilized

(McPherson and Muller 1969; Everett 1987b). Nutrients may be more readily available in ash, and the

mineral soil does not dry out as readily as organic

material. Moisture was more readily available at 12

inch depths beneath exposed mineral soil than below

organic layers in late summer, allowing better growth

of seedlings that can develop taproots such as ponderosa pine (Harrington 1992). The blackened surface

causes warmer soil temperatures that enhance nutrient cycling and can favor growth, particularly in cold

limited environments such as the boreal forest (Viereck

and Schandelmeier 1980). After a severe fire, there is

less competition from sprouting plants, seedlings, and

trees if feeder roots and seeds stored in the duff and

soil were killed. There may be little shade in the first

few postfire years because of plant mortality. The

length of time that a seed environment retains these

characteristics after fire determines the number of

postfire years that establishment of certain species

from seed can take place (Shearer and Stickney 1991).

The physiological requirements of individual species determine whether postfire conditions are favorable for seedling establishment. For most species that

develop from seeds dispersed after fire, the best

seedbeds are microsites where most or all of the

organic layer has been removed by fire because they

provide the greatest chance for seedling survival. For

some shade intolerant species, this is the only time

that seedlings can establish, but these conditions can

result in abundant regeneration, notably of western

larch and many species of pine. Some perennial forbs

resprout after fire, flower, and produce abundant

seeds that establish in the second and subsequent

postfire years (Keeley 1998). Wiregrass produces copious amounts of viable seed only after late spring and

early summer burns.

If some residual organic matter remains, species

with rapidly elongating roots may be favored over

species that grow more slowly. Small seeded species

are more likely to establish where little organic matter

remains. Because seedlings originating from small

seeds may be quite limited in their ability to grow

through organic layers to mineral soil (Grime 1979),

species with large seeds may be favored over smallseeded species where duff layers still exist.

In an Alaskan black spruce/feathermoss (Schreber’s

and mountain fern mosses) stand, germination and

first year survival of black spruce and seven species of

deciduous trees and shrubs occurred on both moderately and severely burned seedbeds. However, by the

third year, seedlings survived almost exclusively on

severely burned surfaces with no residual organic

matter (Zasada and others 1983) (table 2-4).

Some species that establish from seed may be temporarily eliminated from a burned area because the

postfire environment does not favor their establishment. In chaparral communities, species such as

Nuttall’s scrub oak, hollyleaf cherry, and toyon recover by sprouting after fire and thus remain on the

site. However, seedlings of these species do not establish until the canopy closes and a deep litter layer

forms (Keeley 1992).

Fire Stimulated Germination

Dormant seeds will not germinate when exposed to

appropriate temperature and moisture conditions (Keeley

1995). Dormancy is maintained by environmental
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 27

Chapter 2: Fire Autecology Miller

conditions such as high and low temperature, low

moisture, and inadequate amounts or quality of light

(Baskin and Baskin 1989); or it can be imposed by an

impermeable seed coat (Stone and Juhren 1953). Some

species, such as chamise and hoaryleaf ceanothus,

produce a proportion of seeds that remain dormant,

while other seeds from the same plant will germinate

under any suitable moisture and temperature conditions (Christensen and Muller 1975).

Fire can induce germination of dormant seeds of

some species, resulting in an abundance of seedlings of

these species in the first postfire year. Because essentially no seed germination occurs in subsequent years,

annual plants flower, set seed, and are gone after the

first year. Perennial seedlings that mature will flourish, depending on their inherent longevity, for as long

as the site meets their specific environmental requirements. Eventually, they may persist only as seeds.

Germination of hard seeds can occur only after fire

ruptures seed coat fissures or causes cracks to form in

the seed coat, allowing water to enter (Keeley 1987;

Rasmussen and Wright 1988). Requirements for optimum germination may be specific. Redstem ceanothus

seed has the highest percentage of germination after

exposure to moist heat at 176 °F (80 °C) (Gratkowski

1973), followed by stratification through a period of

exposure to cold, wet conditions (Quick 1959). Fire

stimulated germination has been documented for other

hard-seeded genera including Cassia, showy partridgepea (Martin and others 1975; Tesky 1992);

Iliamna, particularly wild hollyhock (Brown and

DeByle 1989); Lotus or trefoil species (Keeley 1991);

Rubus including blackberries and raspberries (Morgan and Neuenschwander 1988; Rowe 1983; Stickney

1986); Ribes such as gooseberry and currant (Lyon

and Stickney 1976); and Prunus (Morgan and

Neuenschwander 1988). Plant life-forms with firestimulated seed germination are shown in table 2-3.

Dormancy of species without hard seed coats can be

broken by exposure to smoke and to chemicals leached

from charred materials, although some species will

germinate only in association with additional stimuli,

such as cold stratification or burial (Keeley and

Fotheringham 1998). This phenomenon has been studied most intensely in Mediterranean ecosystems, such

as chaparral, and in Australia. Germination of chamise

and many herbaceous species of the California chaparral can be induced by these treatments (Keeley 1991).

Smoke exposure requirements varied significantly

among species, with the duration of exposure that is

optimum for germination for some species being lethal

to others (Keeley and Fotheringham 1998). This suggests that seed germination in chaparral, both pattern

and species, may be relative to different types of fire

behavior and levels of fuel consumption, because these

can result in significant variation in the amount and

duration of smoke. Fire behavior may also relate to

establishment of lodgepole pine. The greatest proportion of germination of lodgepole pine seeds from serotinous cones was enhanced by exposure of cones to a

duration of flaming that most commonly occurs in

crown fires (Despain and others 1996).

Seed germination for some chaparral species is

adapted to wildfires that normally occur during fairly

hot, dry late summer or fall conditions. Some seeds

require dry heat to induce germination, but are killed

by lower temperatures if they have imbibed moisture.

Other seeds require higher temperatures for a longer

duration to induce germination than generally occur

under spring burning conditions (Parker 1989). If

chaparral sites are burned under moist spring conditions, germination of both of these types of seeds is

often much reduced. This is a particular concern for

maintaining seedbanks of fire-following annuals,

shrubs, and perennial forbs that can only reproduce

from stored seed (Parker 1987a).

Burn Severity and Seed Regeneration

Variation in burn severity including its pattern and

associated effects on seed mortality, seed stimulation,

and seedbed quality can cause considerable variation

in seedling numbers and species after a fire. The

influence of burn severity on regeneration depends

partly on where seeds are located. Viable seeds are

characteristically present at different depths within

the duff and soil profile. Seed produced by short-lived

species that are stimulated to germinate by the heat of

the fire, or that establish only on bare mineral soil, are

usually found at the base of the forest floor layer, on

top or near the surface of the mineral soil (Stickney

1991). Seeds of longer lived early seral species will be

present near the duff mineral soil interface but will

also have some vertical distribution within the duff

layer. On sites without disturbance, these plants may

have died out, and their seeds may not be present in

the uppermost layers. Transient seeds are only present

on and near the surface of the litter layer.

While fire kills most seeds within the surface litter

layer, the temperature and duration of subsurface

heating controls the amount of mortality and heatstimulation of buried seed (Morgan and Neuenschwander

1988; Weatherspoon 1988). The pattern of severity

relates to the pattern of fuel consumption and can

cause variable mortality or stimulation of seeds around

a burned site. Where little soil heating occurs, few

heat-requiring seeds may germinate. Redstem

ceanothus seedlings tended to occur on severely burned

microsites within a matrix of less severely burned

sites (Morgan and Neuenschwander 1988). If the lethal temperature isotherm penetrates below the level

at which most duff and soil stored seeds occur, much
28 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Miller Chapter 2: Fire Autecology

less postfire plant establishment from the soil

seedbank will occur than after a fire of more moderate

severity (Weatherspoon 1988). However, there can

still be a significant amount of regeneration from

seedbank species where seeds in lower layers were

heat stimulated but not lethally heated. Generally,

the dryer the fuels, the more severe the fire, and the

more seedbank mortality will occur. Fires of high

burn severity also create more bare mineral soil

seedbed, opening the site to colonization by seed

dispersed from on-site or off-site after the fire. For

those obligate, early seral species that only establish

on bare mineral soil seedbeds, fires of high burn

severity favor their regeneration.

Seasonal Influences _____________

Carbohydrates

Carbohydrates, primarily starches and sugars, are

manufactured by plants and provide energy for metabolism, and structural compounds for growth (Trlica

1977). Energy and material needed for initial plant

growth following fire are provided by carbohydrates

stored in undamaged plant parts, usually belowground

structures. The timing of a fire, and its relationship to

a plant’s carbohydrate balance, can be a factor in

postfire recovery because the rate and amount of

regrowth is related to carbohydrate reserves (Trlica

1977).

The importance of carbohydrate reserves to plant

regrowth after fire depends on survival of photosynthetically capable material, such as leaf blades and

sheath leaves on grass stubble (Richards and Caldwell

1985). If some photosynthetic tissue remains or new

tillers rapidly regenerate, newly grown leaf material

soon manufactures all the carbohydrates that the

plant needs for growth and respiration (Caldwell and

others 1981). Evidence from clipping and grazing

studies has shown that the recovery of grass plants is

more related to the removal of growing points than to

the carbohydrate level at the time of defoliation

(Caldwell and others 1981; Richards and Caldwell

1985). However, fire may have a greater impact on

grass plants than severe defoliation because it kills all

photosynthetic material and elevated meristems. New

growth must be supported by stored reserves.

There is a seasonal cycle of depletion and restoration

of total nonstructural carbohydrates related to the

growth cycle of the plant. The most rapid depletion

usually occurs during periods of rapid growth, but

carbohydrates may also be used for flower and fruit

development, cold-acclimation (“hardening off” for

winter), respiration and cellular maintenance during

winter dormancy, and warm weather quiescence (Trlica

1977). Restoration of carbohydrates occurs when production by photosynthesis exceeds demands for growth

and respiration. The timing of fluctuations in the

annual cycle of total nonstructural carbohydrates

(TNC) differs among species because of variability in

plant growth cycles and growing season weather

(Zasada and others 1994).

The limited survival of chamise sprouts after spring

prescribed fires has been attributed to low winter and

spring carbohydrate reserves because of high spring

demand for growth, flowering, and fruiting (Parker

1987b). Number and dry weight of shoots of salmonberry were lowest on rhizome segments collected from

May through July, which was also the seasonally

lowest level of stored TNC (Zasada and others 1994).

Salmonberry is most susceptible to physical disturbance during this time (Zasada and others 1994). For

other species, the effects are most negative if the plant

is burned late in the growing season because the plant

uses a considerable amount of stored carbohydrates to

sprout, but does not have enough time to restore

reserves before winter dormancy (Mueggler 1983;

Trlica 1977).

Severely burned chamise root crowns produced fewer

sprouts than plants that experienced less heating,

probably because more dormant buds were killed.

Subsequent death of plants and limited sprouting may

occur because insufficient carbohydrates are produced

to sustain the root mass (Moreno and Oechel 1991).

Root system dieback after excessive defoliation (Moser

1977) is considered to be a significant cause of plant

mortality in grasses.

Repeated burning during the low point of a plant’s

carbohydrate cycle can increase any negative effects of

treatment. Reduced density, canopy cover, and frequency of Gambel oak in southwestern Colorado, after

two summer burns 2 years apart, were attributed to an

inability to restore spent carbohydrate reserves for the

9 months after top-killing and resprouting (Harrington

1989). In the Southeast, annual summer burning

nearly eliminated understory hardwood vegetation in

a loblolly pine stand (Waldrop and Lloyd 1991). Burning when carbohydrates were low eventually killed or

weakened root systems. Annual winter burning resulted in significant increases in numbers of small

diameter sprouts on these same plots, because burning occurred when reserves were fairly high and sprouts

had a full growing season to restore reserves before the

next treatment.

If burning occurs in close association with heavy use

of the plant community by livestock or wildlife, either

before or after the burn, plant recovery may be delayed

or prevented because of the excessive demand on

stored reserves. Heavy postfire grazing or browsing of

perennial plants in the first growing season after a fire

is likely to cause the most harm, particularly in arid

and semiarid range communities (Trlica 1977).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 29

Chapter 2: Fire Autecology Miller

Flowering

Burning has long been used as a tool to enhance

flower and fruit production of blueberry. Flowering of

grasses such as pinegrass and wiregrass has been

noted to increase significantly after burning (Brown

and DeByle 1989; Uchytil 1992) (fig. 2-13). Burning

during the growing season of April to mid-August

causes profuse flowering of wiregrass in Florida, a

marked contrast to a paucity of flowering that follows

dormant season burning (Myers 1990b). Warmer soil

temperature resulting from litter removal in these

months may be the flowering stimulus (Robbins and

Myers 1992). Increased light resulting from removal of

the chaparral canopy stimulates flowering in golden

brodiaea, a perennial forb that produces only vegetative growth in the shade (Stone 1951). This has also

been observed in the Northern Rocky Mountains, in

heartleaf and broadleaf arnicas, showy aster, and

pinegrass. Increased availability of soil moisture and

soluble nutrients also stimulates increased flowering.

In response to late spring fires, Henderson and

others (1983) observed significantly greater flowering

in big bluestem, little bluestem, sideoats grama, and

Indian grass, all Wisconsin warm-season grasses. Increased flowering was attributed to higher levels of

carbohydrate production caused by improved growing

conditions, such as mulch removal. Grass flowering

and seed production draw heavily upon carbohydrate

reserves. Higher net photosynthate production was

observed in big bluestem after spring burning. In

contrast, cool-season grasses that were actively growing during late spring experimental fires showed a

marked reduction in flowering, possibly because growth

initiated after the fires further depleted carbohydrate

reserves already drawn down by early growth. There

also may have been more damage to meristematic

tissue because plants were actively growing at the

time of burning.

Fires enhanced flowering of dominant forb and shrub

species in longleaf pine forests on the Florida panhandle (Platt and others 1988), with the most significant effects resulting from growing season fires. These

fires increased the number of flowering stems, decreased the average flowering duration per species,

and synchronized the period of peak flowering of

Figure 2-13—Abundant flowering of Thurber’s needlegrass the first growing season after an October prescribed fire, near Carey, Idaho.
30 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Miller Chapter 2: Fire Autecology

Figure 2-14—Postfire seed production by arnica, Bob Marshall Wilderness, Montana.

herbaceous plants, particularly fall flowering composites with a clonal growth form. Fire killed the elevated

apical meristems, which no longer suppressed dormant buds on rhizomes, roots, and stolons. Multiple

stems were initiated from these buds at times of the

year when photoperiod strongly induced flowering.

Dormant season fires had little effect on flowering

periods because apical meristems were located at or

below the ground surface, were little affected by fire,

and continued to suppress secondary meristems the

next growing season.

These mass flowering events are a means by which

plants that regenerate after the fire redistribute themselves within the stand (fig. 2-14). For plants that

germinate from soil stored seed, their profuse flowering in the first few years after fire resupplies the

seedbank and ensures their presence after the next

fire (Stickney 1990) (fig. 2-15).

Phenology

Plant growth stage at the time of a fire can result in

different plant responses. Fire effects can vary substantially during a specific season, such as spring, because

several phenological stages can occur in that 3 months.

Phenology and the accompanying variation in plant

condition, not season, leads to observed differences in

plant response to fire. Phenological differences that

affect plant responses to fire include varying levels of

stored plant carbohydrate, presence of elevated herbaceous meristems that are more susceptible to fire

because of their location, and presence of actively

growing tissues that are more sensitive to high temperatures than when they are dormant or quiescent.

The seasonal growth pattern that is characteristic of

each species can be significantly modified by temperature and moisture in a specific year.

As an example of seasonal influences, ponderosa

pine trees scorched in late October survived higher

percentages of crown damage than trees scorched in

early June and mid-August. The increased survival of

fall burned trees was attributed to reduced physiological activity, lower bud tissue moisture contents, bud

protection by fully developed bud scales and needles,

and replenished carbohydrate stores, allowing adequate reserves to support spring shoot and root growth

(Harrington 1987a, 1993).

Phenology also affects flammability. Moisture content of 1 year and older foliage of Western conifers is

lower in the early part of the growing season than later

in the summer (Chrosciewicz 1986; Jameson 1966;

Philpot and Mutch 1971), and may contribute to higher

spring crowning potential (Norum 1975). Seasonal

differences in the moisture content of surface vegetation can determine whether the vegetation is a heat

sink or is dry enough to be a heat source and thus
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 31

Chapter 2: Fire Autecology Miller

Figure 2-15—Flowering of wild hollyhock, developed from seeds stimulated to germinate by fire, Caribou

National Forest, Idaho.

contribute to fire spread. Seasonal curing of herbaceous vegetation changes it from live to dead fuel.

Burning Conditions

Fuel and soil moisture conditions have a major

influence on upward and downward heat flows that

affect plant responses. Seasonal fluctuations in temperature and precipitation cause a progression of

moisture content in dead woody fuels, litter, duff, soil

organic layers, and soil. For a given vegetation and

fuel type, burning conditions vary seasonally according to a general pattern; and the response of individual

plant species to fires occurring under typical seasonal

fuel and soil moisture conditions are fairly predictable

based on their life-form.

Yearly variations in weather and associated departures from average moisture conditions can cause

substantial variation in fire behavior and fire effects.

For example, a winter and spring of above average

precipitation results in wet woody fuels and duff in

higher elevation forests that limit fire spread and fuel

consumption. Below average precipitation in winter

and spring can create dry enough conditions in forests

to create potential for fires with high fuel consumption, and significant amounts of heat release both

above and below the surface. Dry large fuels, duff, and

mineral soil increase the potential for significant

amounts of surface and subsurface heating, with concomitant mortality of roots, buried regenerative structures and seeds, and tree cambium.

Differences in seasonal weather in shrub/grass types

can result in a large range in grass production, particularly annual species, which creates different fire

behavior potentials. A wet year in the Great Basin

leads to much more herbaceous biomass in sagebrush/

grass communities, a greater likelihood of ignition,

and larger sizes of fires that do occur. However, whether

these higher fuel loadings relate to greater consumption of basal fuels and higher mortality of bunchgrasses and sprouting shrubs has not been documented.

The pattern of fire effects across the landscape

varies with burning conditions. Areas of tree crown

consumption, crown scorch, and little crown damage

can be intermixed (fig. 2-16). Heavily burned areas of

the forest floor where significant amounts of fuel were

consumed and most buried plant parts were killed can

be adjacent to areas where prefire fuel loading was

low, and little subsurface heating occurred (fig. 2-17).

On rangelands, the pattern can vary between areas of

significant heat release associated with consumption

of shrubs and accumulated litter and other areas

where little heat was generated due to sparse fine

fuels (fig. 2-18).

During a dry season, especially in a drought, a much

higher percentage of forest canopy is apt to be scorched
32 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Miller Chapter 2: Fire Autecology

or consumed. Lethal temperatures may be driven to

greater depths because fuel and duff consumption is

fairly complete. During a wet year or early in the year

before significant drying has occurred, less canopy will

be killed and consumed and few buried plant parts will

be killed.

Discussion _____________________

Plant response to fire is a result of the interaction

between severity of the fire and characteristics of the

plants in the fire, both their inherent resistance to

injury and ability to recover. Fuel quantity and arrangement, fuel moisture content, topography,

windspeed, and structure of the plant community

itself cause the lethal heat zone created by fire to vary

significantly in time and space. Fire can cause dramatic and immediate changes in vegetation, eliminating some species or causing others to appear where

they were not present before the fire. However, in

burned areas with a high component of surviving trees

and resprouting understory vegetation, within a few

Figure 2-17—Burn pattern in a forest floor, with different plant species present on moderate and severely

burned areas, Washington Creek, interior Alaska.

Figure 2-16—Fire mosaic in a forest canopy,

Selway-Bitterroot Wilderness, Idaho.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 33

Chapter 2: Fire Autecology Miller

Figure 2-18—Burn pattern on rangeland, north of

Boise, Idaho.

years it can be difficult to determine that a fire recently occurred.

For the vascular plant groups discussed in this

chapter, the recovering plant community, in the first

few years after a fire, comprises individuals from the

following categories:

• Plants that survived the fire with their form intact

• Sprouts or suckers that grew from the base or

buried parts of top-killed plants

• Plants that established from seed

Seedlings can be further described as:

• Plants that re-established from seed dispersed

from surviving plants, usually trees

• Plants that re-established from seed dispersed

from off of the burned site

• Plants that re-established from fire stimulated

seed within the seedbank

• Plants that re-established from seed that developed on plants that resprouted after the fire

Certain species can only recover after fire by a single

means. Some will only be present after fire if regenerative structures survive and produce sprouts, because

their seedlings are unlikely to survive in postfire

environments. Species of plants that cannot resprout

after top-killing must establish from seed. However,

some species can successfully recover from fire both by

resprouting and by seedling establishment. Severity

of the fire largely determines whether new plants are

sprouts or seedlings.

Where fire top-kills plants that can vegetatively

regenerate, sprouting will be a significant source of

postfire vegetation. Where lethal temperature penetrates deeply enough to kill many regenerative structures, sprouting may be limited, but some buried seed

may receive the proper stimulus to germinate and

produce significant numbers of seedlings. A microsite

that sustained lethal heat deeply enough to kill all

stored seed probably had enough fuel and duff consumed to prepare areas of bare mineral soil seedbed.

Reproduction on these sites occurs from seeds dispersed onto these burned surfaces. The availability of

canopy-stored seeds depends on the height to which

lethal temperatures reached into the tree crowns,

while dispersal from wind-carried offsite seeds depends on distance and direction of prevailing winds.

Surviving trees and resprouting plants may produce

seeds that can establish within the next few growing

seasons while suitable seedbed exists.

The immediate response of plants can differ within

the same fire because of variations in the pattern of

burn severity. For example, chamise and redstem

ceanothus can sprout after a low to moderate severity

fire treatment. High severity fires kill existing plants

but they are replaced by new plants that develop from

fire-stimulated seeds. The postfire community may

contain both sprouts and seedlings of these species

with the proportion related to the severity of the fire.

Postfire species composition is usually an assemblage of many of the species that were growing on the

site and represented in the seedbank at the time of the

fire. Vegetative regeneration is common to many species and can make a major contribution to the

postdisturbance community (Ingersoll and Wilson

1990), contrary to the commonly held notion that seed

reproduction is dominant. Resprouts from rhizomes,

root crowns, or protected meristems can account for a

substantial proportion of postfire recruitment (Lyon

and Stickney 1976). Many of the seedlings present in

the first few postfire years may have grown from seeds

formed on resprouting species such as fireweed and

heartleaf arnica. The only locations in which new

species are likely to be added to the plant community

are microsites that are severely burned and receptive

to germination and establishment of seeds from species dispersed from off of the site (Stickney 1999).
34 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Miller Chapter 2: Fire Autecology

How This Applies to Management

Knowledge of plant response to fire can be critical to

successful application of prescribed fire. Designing fire

prescriptions requires knowledge of fire behavior, fire

severity, species survival mechanisms, and methods of

postfire vegetation recovery. Fire prescriptions should

describe a set of weather and fuel moisture conditions

that will control the rate and amount of fuel consumption by different size classes. Properly conducted, the

prescribed fire will result in the desired amount of

mortality, injury, resprouting, and seedling establishment from target species. To ensure that a desirable

range of plant species establishes after prescribed fire,

it is helpful to acquire predictive means of assessing

how different prescriptions can produce different fire

and plant responses (Whittle and others 1997). A fire

prescription that takes both the surface and subsurface

heat regime into account, thereby regulating its severity, is most likely to achieve desired fire effects.

How can you learn about the effects of fire and

postfire response for individual plant species? The

Fire Effects Information System is an Internet database that contains specific information about 900 plants

from the United States and Canada (see chapter 9).

The best way to learn about fire effects in a specific

location is to visit areas recently burned under different fire prescriptions and fuel and soil moisture conditions. Answers to the following check-list questions

 Manager’s Checklist ✎

• Which species are present?

• How much injury did surviving plants sustain?

• Are there new plants of different species sprouts or

seedlings?

• From what depths and structures did sprouting plants

originate?

• Are seedlings rooted in organic or mineral

substrate?

• Is there a pattern of sprouts and seedlings that is

related to fuel and duff consumption?

• Is the species composition of seedlings related to

canopy mortality or proximity to unburned areas?

can help you understand the type of fire treatment

that will create the effects you desire:
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 35

Luc C. Duchesne

Brad C. Hawkes

Chapter 3:

Fire in Northern Ecosystems

Mixed Fire Regime ______________

Major Vegetation Types

Major forest types include those where aspen, eastern white and red pine stands, and jack pine stands

are found either as fire-maintained seral types or

exceptionally as climax stands (see table 3-1 for FRES,

Kuchler, and SAF cover type designations). This includes extensive areas from Newfoundland across to

Alaska and the Great Lakes region. Fire regime characteristics varied extensively for these major vegetation types, reflecting local and regional topography,

fuel type and climate. Fire regime characteristics

(summarized in table 3-1) are discussed by first describing distribution and site features of the major

vegetation types followed by discussion of the nature

of fire.

Fire Regime Characteristics

Aspen—With its continental distribution, aspen is

the most widely distributed forest type in North

America extending from Newfoundland to Alaska,

then southward through the western mountains of

Canada and United States to Mexico (Eyre 1980;

Farrar 1995). Aspen-dominated ecosystems are

generally found as seral and more rarely as climax

ecosystems.

Aspen occurs mostly as a pioneer type (fig. 3-1, 3-2)

on burns or clear-cut areas on a wide variety of soil

types excluding only the driest sands and the wettest

swamps (Eyre 1980). In eastern Canada, aspen is

found in association with sugar maple, balsam fir,

speckled alder, eastern white pine, and paper birch

(Eyre 1980). The transition area between the northern

boreal forest and the central North American grasslands is dominated by aspen in central Canada where

most even-aged stands have a fire origin (Jones and

DeByle 1985). This forest type, frequent in Manitoba

south of Lake Winnipeg and in North Dakota and

northwestern Minnesota, occurs as well stocked stands

in the northern portion of its range, and dwindles into

patches surrounding moist depressions and streams

farther south. Aspen is found in association with bur

oak and in wetter locations with balsam poplar. On

alluvial soils of eastern Saskatchewan and Manitoba,

aspen is associated with white elm, green ash, Manitoba

maple, and eastern cottonwood. Farther west, pure

stands of aspen are found in association with the

chernozen soil zones of Saskatchewan and Alberta

(Corns and Anna 1986). Within the southern part of the

grasslands, patches of aspen parkland are found in

moist depressions and on bluffs and hills. In its western

range, aspen is found most frequently as pure stands or

in association with various conifers such as Engelmann

spruce, lodgepole pine, ponderosa pine, and Douglas-fir.
36 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Duchesne Chapter 3: Fire in Northern Ecosystems Table 3-1

—Occurrence and frequency of presettlement fire regime types by Forest and Range Environmental Study (FRES) ecosystems, Kuchler potential natural vegetation

classes (1975 map codes), and Society of American Foresters (SAF) cover types. Occurrence is an approximation of the proportion of a vegetation class represented

by a fire regime type. Frequency is shown as fire interval classes defined by Hardy and others (1998) followed by a range in fire intervals where data are sufficient.

The range is based on study data with extreme values disregarded. The vegetation classifications are aligned to show equivalents; however, some corresponding

Kuchler and SAF types may not be shown.

Fire regime types

Understory Mixed Stand-replacement

FRES Kuchler SAF Occura Freqb Occur Freq Occur Freq Nonfire

Aspen-birch 19 Aspen Parklands c Aspen 16 M 2 M 2

Paper birch 252, 18 M 2

W. aspen d W. spruce-fir K015 Aspen 217 m 1,2 M 1,2

White-red-jack pine 10 Great Lakes pine K095 Red pine 15 M 2

White pine 21 M 2

White pine-hemlock M 2

White pine-red oak-red maple 20 M 1,2

Jack pine 1 m 1 M 2

Spruce-fir 11 Great Lakes Spruce-fir K093 Balsam fir 5 M 2,3

Northeastern spruce-fir K096 White spruce 107 M 2

Red spruce 32 M 3

Red spruce-balsam fir 33 M 2

Paper birch-red spruce-balsam fir 35 M 2

— Black spruce c Black spruce 12 M 2

— Conifer bog K094 Black spruce-tamarack 13 M 2

Tamarack 38 M 2

— Tundra c M 2

aM: major, occupies >25% of vegetation class; m: minor, occupies <25% of vegetation class.

bClasses are 1: <35 year, 2: 35 to 200 years, 3: >200 years.

cThis type occurs primarily in Canada and was not defined by Kuchler.

dAdded subdivision of FRES.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 37

Chapter 3: Fire in Northern Ecosystems Duchesne

Paradoxically, aspen stands do not ignite easily and

specific site and climatic conditions are necessary

before fire can ignite and spread (Jones and DeByle

1985; Peterson and Peterson 1992; Wright and Bailey

1982). Generally, fires in young aspen stands are low

intensity surface fires unless there is a great deal of

fuel on the forest floor. In older stands, particularly

those that are breaking up, abundant fuel can lead to

higher intensity fires (Peterson and Peterson 1992).

White and Red Pine—Eastern white and red pine

associations were generally fire-maintained seral types

and existed occasionally as self-perpetuating climax

under mixed fire regimes (fig. 3-3). The former include

an extensive area associated with the Great-Lakes

and St-Lawrence River ranging from Newfoundland

to eastern Manitoba in Canada and New England,

New York, Minnesota, Wisconsin, and Michigan in the

United States (Eyre 1980; Farrar 1995; Roberts and

Mallik 1994). Red and white pine were found in pure

stands, or either species comprised the majority of

stocking with jack pine, red oak, red maple, aspen, pin

cherry, white spruce, and balsam fir. Pure red pine

stands are mostly associated with dry, coarse soils

whereas white pine tends to dominate in lighter structured soils (Eyre 1980).

Before fire protection, white and red pine stands

were subjected to a mixture of lethal and nonlethal

fires (Burgess and Methven 1977; Cwynar 1977, 1978;

Figure 3-1—Successional pathways involving seral aspen,

fire, and conifers (Brown 1985).

Pure aspen stands are particularly susceptible to

mortality of aboveground stems from fire of low intensities, an unusual characteristic for a forest type so

well adapted to regeneration after fire (Jones and

DeByle 1985; Mutch 1970). Occasionally, fires of intensities greater than 30 Btu/ft/s (100 kW/m) have

been shown not to damage mature trees (Quintillio

and others 1989).

Figure 3-2—Seral aspen stand in Ontario (Canadian Fire Danger Rating Group photo point).
38 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Duchesne Chapter 3: Fire in Northern Ecosystems

Frissell 1973; Heinselman 1981; Maissurow 1935,

1941; Methven 1973; Methven and Murray 1974; Van

Wagner 1970; Wendel and Smith 1990). Pine stands

on dry sites underwent a cycle of light to moderate fires

every 20 to 40 years, then high intensity fires every

100 to 200 years. On mesic sites, white pine experienced mostly high intensity fires. Postfire survival

depended on fire behavior and tree age. This burning

pattern for white and red pine stands led to a negative

exponential age-class distribution with approximately

55 percent of forest stands older than 125 years, and

the pattern applies to landscapes at least 960,000

acres (400,000 ha) (Baker 1989), suggesting a large

proportion of so-called old growth forests.

On dry sites, fire frequency varied from 20 to 300

years with an average return interval of approximately 100 years for the entire Great Lakes and St.

Lawrence forest region (Bergeron and Brisson 1990;

Burgess and Methven 1977; Cwynar 1977, 1978;

Duchesne and Gauthier; Frissell 1973; Heinselman

1981; Maissurow 1935, 1941; Methven and Murray

1974; Rowe 1972; Van Wagner 1970) and from 15 to 30

years in Newfoundland (Roberts and Mallik 1994).

Mesic sites colonized mainly by white pine were characterized by longer fire intervals, probably 200 to 300

years or more (Heinselman 1983).

Jack Pine—Jack pine forests (fig. 3-4) are a widespread ecosystem type extending across Canada from

Newfoundland to the Mackenzie river valley in the

Northwest Territories (Eyre 1980; Farrar 1995), where

they are found as fire-dominated pioneer stands. In

the United States, jack pine stands are mainly found

around the Great Lakes States, and less often in

northern New England and New York. Typically, jack

pine constitutes the majority of the stocking with

several possible associate species. Associate tree species in the boreal forest are aspen, paper birch, balsam

fir, and black spruce whereas, in the Great Lakes, they

are northern pin oak, red pine, aspen, paper birch,

balsam fir, and white pine (Eyre 1980). In general,

jack pine stands are found on eskers, sand dunes, rock

outcrops, clays, and organic soils. On the better sites,

they exist primarily as pioneer stands but they tend to

persist on xeric sites that are periodically visited by fire.

Eyre (1980) recognized six variant types of jack pine:

1. Jack pine-balsam fir-black spruce in the mixed

wood and hardwood forest zones of the boreal forest.

2. Jack pine-feather moss on clay and sands of the

boreal forest.

3. Jack pine-sheep laurel on coarse sand and rocky

outcrops of the mixed wood and coniferous forest

zones.

4. Jack pine-Sphagnum on poorly drained organic

deposits of the mixed wood and coniferous forest zones.

5. Jack pine-bog labrador tea on well-drained deposits of a different nature in the coniferous forest zone.

Figure 3-3—White and red pine stand in Ontario (Canadian Fire Danger Rating Group photo point).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 39

Chapter 3: Fire in Northern Ecosystems Duchesne

6. Jack pine-lichen in the taiga on various kinds of

deposits.

Fires in jack pine forests, as on most of the boreal

forest, were dominated by crown fires or high intensity

surface fires that cover broad areas (Van Wagner

1983) because of the climate, topography, and continuity of fuel. Fires were frequent on dry sites with

average return intervals as low as 15 to 35 years in

eastern Canada (Bergeron 1991; Heinselman 1983)

and the Great Lakes. Fire intervals were greater in the

West (Heinselman 1983). Exceptionally, jack pine is

submitted to nonlethal understory fire regimes in

insular situations (Bergeron 1991).

Fuels

Aspen—Aspen stands are only flammable in the

spring, late summer, and fall when they are leafless

due to the drying effect of sun and wind on the leaf

litter. Furthermore, in the fall the herbaceous plant

and shrub component of the understory is dead and

dried out, forming a continuous layer of loosely organized fine fuel. In general flammability depends largely

on the amount of herbaceous and shrub fuels present

in the stand. Herbaceous fuels can vary from as little

as 0.04 tons/acre to 0.75 tons/acre (0.08 t/ha to 1.60 t/

ha). Shrub fuels range from 0 to over 2.25 tons/acre (0

to 4.80 t/ha) (Brown and Simmerman 1986). Forest

floor litter layer typically ranges from 0.38 to 2.9 tons/

acre (0.80 to 6.60 t/ha) (Brown and Simmerman 1986;

Quintilio and others. 1989).

White and Red Pine—White and red pine stands

are fire-prone. Van Wagner (1970) reports that high

density red pine stands are the most flammable fuel

type of all northern cover types. The organic layer of

white and red pine stands is generally 4 to 6 inches (10

to 15 cm) deep (Forestry Canada Fire Danger Group

1992). Light to moderate fires consumed the litter and

shrubs along with invading shade tolerant conifers,

but only scarred trees older than 30 to 75 years

(Heinselman 1983), whereas high intensity fires killed

most trees. In addition to age, white and red pine

survival to fires of all intensity varied according to

rooting, fuel loading, density, and fire behavior. Trees

with more than 75 percent crown damage were more

likely to die within the first postfire year (Van Wagner

1970). Fuel loadings in white pine stands include

approximately 7.1 and 11.1 tons/acre (15.9 and 24.8 t/

ha) of woody surface fuel and forest floor material,

respectively, and 5.3 and 10.5 tons/acre (11.8 and 23.6

t/ha) of woody surface fuel and forest floor material,

respectively, for red pine (Stocks and others 1990).

Jack Pine—Immature jack pine stands are stocked

with 4,000 to 12,000 stems/acre (10,000 to 30,000

stems/ha) and heavy thinning mortality results in a

Figure 3-4—Mature jack pine stand (Canadian Fire Danger Rating Group photo point).
40 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Duchesne Chapter 3: Fire in Northern Ecosystems

large quantity of standing dead stems and down woody

fuel, creating both horizontal and vertical continuity

in the fuel. Fully stocked mature stands 400 to 800

stems/acre (1,000 to 2,000 stems/ha) have reached

canopy closure and the base of the crown is separated

from the ground. Surface woody fuel loadings vary

from 2.9 to 17.2 tons/acre (6.5 to 38.6 t/ha) and forest

floor loadings vary from 7.0 to 51.4 tons/acre (15.7 to

115 t/ha) (Quintilio and others 1977; Stocks 1989;

Stocks and others 1990).

Postfire Plant Communities

Aspen

Pre-1900 Succession—Before settlement by EuroAmericans, large expanses of western aspen and aspen parkland existed in both the Canadian and American West. These stands were often perpetuated as a

shrublike cover by light to moderate intensity fires

that swept across the prairie grasslands and ignited

the aspen stands on a regular 3 to 15 year basis. Aspen

regenerated well after fire. Settlement of the West in

the late 1800s and early 1900s increased fire frequency because of land clearing fires, slash burning,

and railway traffic (Murphy 1985). In more recent

times, following the implementation of rigorous fire

protection programs, lack of fire has threatened the

continued existence of aspen in the West (Brown and

DeByle 1987, 1989; Peterson and Peterson 1992).

In the Rocky Mountains, low intensity fires caused

thinning and encouraged all-aged stands whereas

high intensity fires resulted in new even-aged stands.

In early postfire communities aspen may be dominant

but replacement of seral aspen by conifers is gradual

and may take 200 to 400 years or more (Bartos and

others 1983), depending on the potential for establishment and growth of conifers. Forage production declines two- to four-fold during succession.

Post 1900 Succession—Fire suppression since the

19th century has altered dynamics in aspen stands in

the mountainous Western United States, changing

fire frequencies from as little as 10 years (Meinecke

1929 in DeByle and others 1987) to approximately

12,000 years (DeByle and others 1987). Without the

occurrence of disturbance, aspen clones mature in

about 80 to 100 years (Schier 1975) and regeneration

for this species is threatened. The dying back of those

stands is now favoring shade-tolerant conifers or in

some case grasses, forbs, or shrubs, depending on the

availability of seed sources (Beetle 1974; Bergeron and

Danserau 1993; DeByle 1976; DeByle and others 1987;

Krebill 1972; Schier 1975).

Aspen stands may be replaced by conifers if subjected only to low intensity fires that do not kill conifer

regeneration. Such conversion may require the absence of moderate to high intensity fire for as little as

one aspen generation to as long as 200 to 1,000 years

in the southern boreal forest or in the Rocky Mountains (Bergeron and Dubuc 1989; J. K. Brown 1985;

Brown and Simmerman 1986; Perala 1990).

Management Considerations—In those areas

where the perpetuation of pure aspen stands is desired, prescribed burning can be an economical and

ecologically sensitive silvicultural tool since it closely

mimics the natural disturbance and regeneration patterns (Brown and DeByle 1987, 1989). Wildland fire

use in wilderness areas and prescribed fire are the

most acceptable management tools for maintaining

aspen stands in the American West (DeByle and

others 1987) and can be useful in aspen stands in the

East as well (Perala 1974; Weber 1990). The timing

and resulting intensity of the fires is critical because

high intensity burns can reduce site productivity

(Weber 1990). But also, low intensity fires may not

spread adequately and with enough heat to kill most

aspen resulting in a poor response of aspen suckers

(Brown and DeByle 1987). The prescription window

for successfully burning aspen tends to be narrow due

to inherently low flammability. However, flammability varies substantially among aspen stands. Appraising flammability to recognize good burning opportunities can increase chances of success (Brown and

Simmerman 1986). Thus, prescribed burners must be

ready to act quickly when opportunities arise. Fuel

enhancement by cutting some trees especially conifers

to increase surface fuel loading and continuity can be

helpful in some situations. In the Western United

States, grazing by domestic stock and sometimes wild

ungulates must be controlled to assure successful

regeneration of aspen (Bartos and others 1994).

White and Red Pine

Pre-1900 Succession—Full succession of plant

communities was not possible because of the fire

regime associated with white and red pine stands.

Recurrent nonlethal fires eliminated shade-tolerant

competing vegetation and prepared seedbeds for trees.

Red pine regeneration was probably more closely associated with high intensity fires than white pine regeneration because of red pine’s greater intolerance to

shade and humus. However, shading is required for

optimal regeneration of both white and red pine. The

resulting forest is an even-aged stand for both white

and red pine. The within stand age-class distribution

is greater for white pine than red pine based on

repeated observations that white pine can regenerate

well under a nurse crop if the seedbed is suitable (Eyre

1980).

Post-1900 Succession—As white and red pine

stands mature, their regeneration fails to establish under cover, and tolerant species slowly invade. Without
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 41

Chapter 3: Fire in Northern Ecosystems Duchesne

fire, stands on dry sites evolve into mixed wood forests

dominated by balsam fir and black spruce, or on mesic

sites, white birch and aspen. White pine frequently

pioneers on abandoned agricultural land, particularly

in New England. However, in stands where fire is

delayed for excessive periods, invasion by shade-tolerant species is a widespread phenomenon (Day and

Carter 1990) leading to other forest types dominated

by shade-tolerant species. After a high intensity 1995

wildfire in mature white and red pine stands in Quetico

Provincial Park, red pine seedlings were present in

low numbers but white pine seedlings were rare

(Lynham and Curran 1998).

Management Considerations—Traditionally

white and red pine stands were clear-cut (Mayall

1941) or clear-cut with residual seed trees at one to

eight seed trees/acre (one to 20 seed trees/ha). However, competing vegetation, particularly beaked hazel

and balsam fir on mesic sites, invade easily and prevent white and red pine from regenerating properly.

This has led to a normal-shaped age-class distribution, contrasting strongly with a natural negative

exponential distribution. Moreover, average age of

white and red pine forests is now 72 years whereas

pre-settlement average age was approximately 230

years. Shelterwood cutting is now the preferred silvicultural method for white pine management because

it favors regeneration of white pine and to a lesser

extent red pine, provided that mineral soil is bared.

Contrary to southern pine associations, fuel in white

and red pine associations stabilizes after several years

and there is no need for fuel control to prevent catastrophic fires. Furthermore, understory invasion by

shade tolerant species has a negative impact on fire

behavior by creating a layer of less flammable material (Van Wagner 1970).

Because white and red pine ecosystems are fireprone and so dependent on fire for their regeneration,

prescribed fire is an ideal silvicultural tool for regeneration and removal of competing vegetation (McRae

and others 1998). Parameters of the Canadian Forest

Fire Weather Index System (FWI) (Van Wagner 1987)

that provide optimal results from prescribed burning

in white and red pine stands have been determined to

be (1) fine fuel moisture code 90 to 95, (2) initial spread

index 8 to 16, (3) build up index up to 52, and (4) fire

weather index 12 to 24 (Van Wagner and Methven

1977). Restoration forestry on rich sites may require

two successive burns to eradicate resilient understory

competition, particularly beaked hazel, and to prepare

seed beds.

In the United States, low-intensity prescribed fire is

the only practical treatment currently available to

control white pine cone beetle that can ruin close to

100 percent of the cone crop in white pine seed orchards

if not controlled. Wade and others (1990) developed

guidelines for using prescribed fire in seed orchards

throughout the Eastern United States.

Jack Pine

Pre-1990 Succession—Owing to its serotinous

cones, jack pine regenerated well following fire and its

shade-tolerant understory associates were eliminated.

Jack pine is a shade-intolerant short-lived, pioneer

species that almost always grew in even-aged stands

(Eyre 1980; Farrar and others 1954; Heinselman 1981).

Seeds germinate best on mineral soil or highly reduced

organic soil. Postfire growth was found to be favored

on thin organic soils in Ontario (Weber and others

1987). However, in northern Quebec, postfire organic

horizon thickness did not affect seedling growth

(St-Pierre and others 1991). Without fire, jack pine

forests are succeeded by balsam fir stands on xeric to

mesic morainic surface deposits in the southern part of

the Quebec boreal forest (Bergeron and Dubuc 1989).

Alternatively, jack pine forests may be replaced by

white spruce or black spruce in western Canada (Rowe

and Scotter 1973; Wein 1983) and by balsam fir, aspen,

paper birch, and eastern white cedar in eastern Canada

(Bergeron and Dubuc 1989).

Post-1900 Succession—Fire suppression has a

dramatic effect on the long-term productivity of jack

pine associations. In the Northwest Territories, Wein

(1983) described the possible impact of the absence of

fire in boreal jack pine stands as follows. As soil

organic matter builds up, jack pine becomes less

vigourous and black spruce tends to move into the jack

pine habitat until finally the organic matter becomes

too deep or nutrients become too limiting for black

spruce regeneration and black spruce is replaced by

willow. In eastern Canada, old-aged stands of the

boreal forest suffer from high levels of dieback and are

transitory to associations dominated by black spruce

or balsam fir (Cogbill 1985). In Michigan, the loss of

extensive areas of dense, fire regenerated jack pine

stands that existed prior to 1900 is the primary reason

for the decline of the Kirtland’s warbler, a species whose

breeding habitat requires large tracts of 8 to 20 year-old

jack pine associations (Radke and others 1989).

Management Considerations—Clear-cutting is

the preferred harvesting method for jack pine. However, competing vegetation (particularly on the richer

sites), insufficient seed rain, and thick humus can

prevent development of proper stocking (Burns and

Honkala 1990; Chrosciewicz 1990). In Quebec, and

probably throughout the entire jack pine range, only 4

percent of jack pine stands have sufficient pre-established growth before clear-cutting (Doucet 1988). On

the other hand, jack pine can regenerate well in

clearcuts if 40 to 50 percent of treated areas display

bare mineral soils and if branches are scattered evenly
42 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Duchesne Chapter 3: Fire in Northern Ecosystems

over the terrain (Ball 1975; Chrosciewicz 1990). If

planting is necessary, jack pine must be planted on

site-prepared terrain (Beland and Bergeron 1993;

Sutton and Weldon 1993).

Slash burning, together with seed tree systems,

featuring approximately eight seed trees/acre (20 trees/

ha) is the best method of regenerating jack pine after

cutting (Chrosciewicz 1988, 1990). In addition, direct

seeding has been shown to be successful on sites that

have received prescribed burning (Cayford and McRae

1983). In practice, it was found that maximum regeneration can be achieved when burning occurs under

the following conditions of the Canadian Forest Fire

Danger Rating System: fire weather index of 4, duff

moisture code of 37, fine fuel moisture code of 80, build

up index of 58, and initial spread index of 1.2

(Chrosciewicz 1988). Regeneration of jack pine is best

achieved by high intensity fires (>434 Btu/ft/s or

>1500 kW/m) that consume most of the humus layer

(Weber and others 1987).

Stand-Replacement Fire

Regimes _______________________

Major Vegetation Types

Major vegetation types experiencing stand-replacing fire regimes include stands dominated by black

spruce, white spruce, red spruce, and balsam fir along

with conifer bogs and tundra (table 3-1). All these

types are characterized by stand-replacement fire

regimes having different fire cycles that vary according to climate and topography. Stand-replacement

fires are the most common type of fires in northern

forests, sometimes becoming as large as 2.8 million

acres (1.4 million ha) (Murphy and Tymstra 1986).

Fire Regime Characteristics

Black Spruce—Black spruce associations are widespread throughout the northern biome from Alaska to

Newfoundland and from the northern limits of the tree

line south into the northeastern and North-Central

United States. This species is found on a wide variety

of sites ranging from moderately dry (fig. 3-5) to very

wet. At the southern portion of its range, black spruce

is found primarily on wet organic soils, but farther

north its abundance on uplands increases. In the

Lakes States and in New England, black spruce is

mostly found in peat bog and swamps.

In southern areas of the black spruce range, it is

commonly associated with a variety of other tree

species including balsam fir, paper birch, tamarack,

and aspen (Hare 1954; Rowe 1972). The stocking and

associated vegetation on organic peatland soils depends largely on water sources and movement (Eyre

Figure 3-5—Black spruce stand (Canadian Fire Danger Rating Group photo point).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 43

Chapter 3: Fire in Northern Ecosystems Duchesne

1980). On moist sites, well stocked to dense stands of

black spruce grow above a well-developed mat of

feather mosses such as Schreber’s moss, mountain

fern moss, and knight’s plume moss) (Eyre 1980;

Meades and Moores 1989). On very wet mineral

or organic soils, the moss carpet is dominated by

Sphagnum mosses. Typically, moist to very dry nutrient-poor sites are characterized by open stands of

black spruce with the ground covered by species of

Cladina lichens. The shrub layer in these forests

consists mainly of ericaceous plants such as sheep

laurel and bog Labrador tea. The organic layer may

exceed a depth of 8 to 12 inches (20 to 30 cm) and

comprises branches, other woody debris, and a variety

of slowly decomposing plant material (less than 2

percent of organic matter decays per year) (Van Cleve

and others 1979). In interior and south central Alaska,

black spruce is extremely common on cold, poorly

drained sites (Viereck and Dyrness 1980).

Fires in boreal black spruce ecosystems are large

and frequent owing to the flammable nature of the

forest, continuity of fuel sources and continental climate conditions (Heinselman 1981; Viereck 1983).

Most fires in black spruce associations are either

severe ground fires or crown fires of sufficient intensity to damage aboveground vegetation including the

black spruce overstory and all of the associated understory shrubs. Some of the organic soil component

usually remains, but high intensity summer fires,

during dry, windy climatic conditions may achieve

sufficient intensity to burn off this entire layer and

completely expose the mineral soil. When drought

conditions are extreme the entire forest structure

becomes highly flammable and burns readily. In less

extreme conditions, the lichen-dominated black spruce

forest burns while the moister and older feather mossdominated stands or deciduous mixed wood areas

remain unburned (Foster 1983).

In general, the fire rotation of black spruce forests

tended to be relatively short in the West while it was

longer eastward, in response to a moister climate.

Estimates of fire rotations range from 50 to 100 years

in northwestern Canada and Alaska (Heinselman

1983; Rowe and others 1974; Viereck 1983) and for

eastern Canada, 500 years in southeastern Labrador

(Foster 1983; Viereck 1983), and 480 years in western

Newfoundland (Wilton and Evans 1974). The longer

rotation in eastern areas of the Canadian boreal forest

are probably the result of higher levels of precipitation, and the occurrence of natural firebreaks such as

lakes and extensive areas of moist peatlands.

White Spruce—White spruce associations have a

transcontinental range from Newfoundland and Labrador, west across Canada along the northern tree

line to Hudson Bay, Northwest Territories, Yukon,

and Alaska. From British Columbia the range extends

east through Alberta and Manitoba and south and

east through northern Minnesota and Wisconsin, central Michigan, northeastern New York, and Maine. In

the eastern forest, white spruce grows from sea level to

elevations of about 5,000 feet (1,520 m) in pure stands

or in mixed stands where it is the major component.

Associated species include black spruce, paper birch,

aspen, red spruce, and balsam fir. In eastern Canada

and northern New England white spruce is a climax

species. Closed pure white spruce stands and sprucebalsam fir occur sporadically throughout the boreal

forest on moist and fresh soils, particularly on intermediate and low slopes (Burns and Honkala 1990;

Eyre 1980).

Because mature white spruce forests accumulate

large amounts of organic matter consisting of feather

mosses, woody fuels, flaky barks, and shrubs, they are

highly susceptible to fire. Also the crown and canopy

structure with tree crowns extending nearly to the

ground is ideal for ignition and propagation of crown

fires (Rowe and Scotter 1973; Van Wagner 1983). In

the boreal forest, the fire regime was characterized by

crown fires or severe surface fires with a return interval averaging 50 to 150 years. Lightning caused most

fires, especially in the months of July and August

(Heinselman 1981). Fires could cover areas of 3 to 250

acres (1 to 100 ha), but the most ecologically significant fires were large, often thousands of acres. One of

the longest fire cycles, 300 years, characterized the

flood plain white spruce forests. In the East cycles may

average 150 to 300 years. In British Columbia, Hawkes

and others (1997a) reported fire cycles ranging from

794 to 2,083 years for the wet, cool sites of white/

englemann spruce–subalpine fir forests in the northern Rocky Mountains. The dry, cool sites of white

spruce/aspen forests (Kluane Section, B.26d, of the

boreal region; Rowe 1972) in Kluane National Park,

Yukon, had mean fire return intervals ranging from

113 to 238 years while individual stand intervals

ranged from 9 to 403 years. Francis (1996) reported

that white spruce dominated forests of the Shakwak

Trench, Yukon, were characterized by large-scale,

infrequent fire disturbances on north slopes while the

south-slopes had small to medium-scale, frequent

disturbances.

Balsam Fir—Balsam fir associations are found

throughout eastern and central Canada as a band

stretching from Newfoundland and central Labrador

south to New York in the east and to central and

northern Alberta in the west (Hosie 1973). Balsam fir

is found in pure stands, and also in association with

many species common to moist and wet sites where the

ground-covering feather mosses build layers from less

than an inch (a few centimeters) to more than 1 foot

(30 cm) in thickness depending on stand maturity.
44 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Duchesne Chapter 3: Fire in Northern Ecosystems

Balsam fir is a shade tolerant and late successional

species. In the Eastern forests important associates

were red spruce (mainly in New Brunswick and Maine)

and paper birch. In the boreal forest region, it was

more commonly found with black spruce, white spruce,

jack pine, paper birch and trembling aspen. Balsam fir

occurred on a wide range of soils including heavy

clays, loams, and sandy glacial till. It was a climax

association on upper slopes and tops of mountains

(Burns and Honkala 1990; Eyre 1980).

In the high precipitation areas of eastern Canada

and the Northeastern States, fire cycles were between

150 to 300 years (Heinselman 1981; Wein and Moore

1977, 1979). Stand-killing crown fires of high intensity

or severe surface fires occur often after long droughts

(Heinselman 1981; Rowe 1983; Van Wagner 1983).

Unique spatial arrangements and dry conditions are

responsible for several smaller, nonstand replacing

fires of lower intensities, every 20 to 30 years in

northwestern Quebec (Dansereau and Bergeron 1993).

The periodic outbreak of the spruce budworm

(Choristoneura fumiferana) causes heavy tree mortality (fig. 3-6), leading to a large fire potential peaking 5

to 8 years after tree death. However, the flammability

of such forests decreases gradually after that period as

balsam fir fuel starts to decompose and understory

vegetation proliferates (Stocks 1987).

Red Spruce—Red spruce was one of the more

important conifers in the Northeastern United States

and adjacent Canadian provinces. Red spruce was

most abundant in the Maritime Provinces, neighboring portions of Quebec, south-central Ontario and in

parts of Eastern North America but especially in

Maine. It grows best in a cool, moist climate, in either

pure stands or as a major component of the growing

stock. In the northern part of its range, red spruce

grows at elevations from near sea level to about

4,600 feet (1,400 m). In mixed stands other common

associates were red maple, eastern hemlock, eastern

white pine, white spruce, eastern white cedar, and

black spruce (on wet sites). Red spruce stands are

found over a range of sites including moderately well

drained to poorly drained flats and the thin-soiled

upper slopes. On acidic tills it is considered climax. It

is present on fresh and moist acidic outwash and on

well drained slopes and varying acidic soils in abandoned fields and pastures where it is usually subclimax (Eyre 1980) being replaced by shade-tolerant

hardwoods such as sugar maple and beech.

The ground cover in dense red spruce stands consisted mostly of bryophytes, lichens, tree litter, and

patches of young conifer germinants that rarely survive over 2 or 3 years. As stands opened up and light

conditions improved additional arboreal species,

Figure 3-6—Aerial photograph of a Balsam fir stand after infestation by spruce budworm (Canadian Fire

Danger Rating Group photo point).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 45

Chapter 3: Fire in Northern Ecosystems Duchesne

shrubs, and herbs developed. The three main associations—Hylocomium/Oxalis, Oxalis/Cornus, and Viburnum/Oxalis—indicate increasing site productivity and

increasing hardwood competition. The Oxalis/Cornus

association is considered the best for growing conditions (Eyre 1980). Presettlement fire rotations in red

spruce associations averaged 100 to 150 years and

higher (Heinselman 1981; Wein and Moore 1977,

1979).

Conifer Bogs—Conifer bogs are found over a wide

range of soil and climatic conditions in Canada, and

the North-Central United States and Alaska. Tree

height, stand density, and especially floristic composition can differ considerably. Stand density varies from

open to well-stocked stands. Composition can range

from pure black spruce, tamarack, or eastern white

cedar stands to black spruce in association with tamarack, balsam fir, eastern white cedar, paper birch,

speckled alder, black ash, poplar, and willow (Armson

1982; Bergeron and Dubuc 1989; Eyre 1980; Johnston

1975, 1976).

Black spruce-dominated bogs, the most abundant

bog type, are widely distributed throughout the boreal

forest from Newfoundland to Alaska and south into

the Great Lakes-St. Lawrence Region. Tamarack bogs,

or wetlands stocked mainly with tamarack, are found

from Quebec across the boreal forest to northwestern

Alberta, south into the Lake States, and large portions

of Minnesota, New York and New England. Pure

eastern white cedar or eastern white cedar dominated

bogs are limited to southern Ontario, Quebec and New

Brunswick, the Lake States, and Northeastern States

(Eyre 1980).

In general, conifer bogs are not as fire-prone as other

forest stand types because of their wetter nature

(Heinselman 1973, 1981; Payette and others 1989;

Rowe and Scotter 1973). The high water table in the

spring, a green dense understory, and a more humid

environment render conifer bogs unsusceptible to forest fire except in severe drought years. These conditions, favoring low decomposition rates, enable the

formation of an organic layer that can become greater

than 3 feet (1 m) thick (Heinselman 1981). In turn, the

thick organic layer of conifer bogs maintains a high

moisture content that further contributes to the longer

fire cycle of bog sites (Heinselman 1981; Rowe and

Scotter 1973).

A second characteristic of bogs that tends to exclude

fire from them is their tendency to occupy depressions

and lowland areas. Because conifer bogs are mainly

located in convex depressions, they are avoided by fire,

which seeks convex landforms such as hilltops and

slopes (Payette and others 1989; Rowe and Scotter

1973). Consequently, bog sites are spared from even

large, high intensity forest fires, leaving unburned

pockets of forests that become important seed sources

to accompanying uplands (Bergeron and Dubuc 1989;

Payette and others 1989).

Fire in conifer bogs occurs mostly in July, August, or

September during severe drought years when a low

water table allows the forest floor to become thoroughly desiccated. Under these circumstances with

sufficient wind, spruce, tamarack, and eastern white

cedar trees of conifer bogs can sustain major crown

fires. Heinselman (1981) estimated the fire rotation to

be 100 to 150 years for large forested spruce bogs in

Minnesota. He found that the average fire rotation of

spruce bogs is longer than those of nearby upland

forests.

Tundra—Tundra ecosystems can be separated into

three main types: (1) arctic tundra occurring at high

altitudes and low elevations, (2) alpine tundra occurring at higher elevations farther south, and (3) the

sedge tussock-mixed shrub in Alaska, throughout the

Seward Peninsula and interior. Whereas vegetation is

similar in the first two types of tundra, they differ

mostly in climatological factors such as maximum

summer temperatures, extremes of day length, and

intensity of solar radiation (Barbour and others 1980).

In the three tundra types, the upper to 6 to 24 inches

(15 to 60 cm) of soil is subjected to a seasonal freeze and

thaw cycle. This layer, called the active layer, lies

above the permafrost layer (Barbour and others 1980).

Vegetation is short, often only 6 inches (15 cm) tall,

and is dominated by perennial forbs, grasses, sedges,

dwarf shrubs, mosses, and lichens (Barbour and others 1980; Brown 1983). Herbaceous perennials make

up 95 to 99 percent of the tundra flora and 60 percent

of these are hemicryptophytes (perennials with perennating buds just at the soil surface).

In general, distribution of vegetation over the tundra biome is characterized by a patchy occurrence of

dense vegetation, sparse vegetation, and bare ground

offering an interrupted fuel bed (Payette and others

1989). However, in the sedge tussock-mixed shrub

tundra, the fuel layer, made up mostly of sheathed

cottonsedge, is dense and continuous leading to large,

fast spreading conflagrations (Racine and others 1987).

The transition zone between the upper limit of the

boreal forest and the tundra has been distinguished as

forest tundra and is characterized by a fragmented

cover of scattered forest stands of lichen-spruce and

lichen-heath communities on well-drained sites. The

forest tundra is further delineated into the forest

subzone and shrub subzone (Payette and others 1989;

Sirois and Payette 1991). In both the forest tundra and

tundra, the ground vegetation is the primary carrier of

fire. Lichens resemble dead tissue more than live

tissue in their susceptibility to fire and often serve as

the initial point of ignition. The low shrub component

can provide a high percentage of fine, dry fuel with a

low temperature of ignition (Auclair 1983).
46 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Duchesne Chapter 3: Fire in Northern Ecosystems

Burning patterns of tundra ecosystems generally

are characterized by moderate intensity surface fires

that may kill all aboveground plant parts but seldom

destroy underground parts (Bliss and Wein 1972;

Van Wagner 1983; Viereck and Schandelmeier 1980).

The fire cycle may be as short as 100 years, but it is

usually much longer (Viereck and Schandelmeier 1980).

According to Sirois and Payette (1991) and Payette

and others (1989), the modern fire rotation period

increases from 100 years in the upper boreal forest to

180 to 1,460 years in the forest subzone and to 9,320

years in the shrub subzone of the forest tundra of

northern Quebec. Tundra west of Hudson Bay is more

influenced by a continental climate and is more likely

to burn and therefore have a shorter fire rotation

(Payette and others 1989). In addition, the tussocksedge tundra of Alaska may be a fire-dominated ecosystem, although the fire interval has yet to be determined (Racine and others 1987).

The long fire rotations in the tundra are probably

related to the prevalence of cold, humid summers,

saturated peat profiles, and the absence of continuous

vegetation cover. These features serve to restrict fire

spread over a large area (Payette and others 1989). In

northern Quebec, small sized fires (<120 acres or <50 ha)

occurred more frequently in the tundra than in the

forest tundra; and forest tundra fires were smaller, on

average, than boreal forest fires. Some large fires

overlapped the limit between the biomes of the northern boreal forest and the forest tundra, but there was

no overlapping between the forest tundra and the

shrub tundra (Payette and others 1989).

Fuels

Black Spruce—The Canadian Forest Service’s Fire

Behaviour Prediction system recognizes two fuel types

associated with spruce forests (Forestry Canada Fire

Danger Group 1992). In spruce-lichen woodlands (fuel

type C1), the lichens that dominate the ground cover

are exposed to sunlight and, with no vascular tissue,

dry out easily (Auclair 1983). These dry lichens form a

mat 0.75 to 1.5 inches (2 to 4 cm) thick and combine

with a variety of woody debris and shrubs to form a

highly flammable ground fuel (Johnson 1992). Furthermore, spruce-lichen woodlands without a closed

canopy permit the passage of wind, thus enhancing

the rate of fuel drying; and once a burn is initiated,

wind passage carries fire and augments radiant heat

transfer (Foster 1983). Alexander and others (1991)

reported that wind tilts the flame easily in the lichen

layer but a high-intensity flame radiation surface fire

does not develop. The combination of the lichen layer

and scattered black spruce tree clumps produce highintensity fires with spread rates from 2.0 to 168 ft/min

(0.6 to 51 m/min) and fire intensities of nearly 9,540

btu/ft/s (33,000 kW/m).

The boreal spruce fuel type (C2), the second black

spruce fuel type, consists of a continuous mat of feather

moss, lichens, moderate amounts of woody debris and

a well developed ericaceous shrub layer. This fuel type

can become highly flammable when it is dry. However,

when forming a wet peatland, it is not readily

flammable and can act as a firebreak (Foster 1983).

This fuel type typically occurs in old, closed canopy

black spruce stands where the mineral soil may remain frozen for most of the year (Foster 1983) because

it is covered by a thick (4 to 8 inches; 10 to 20 cm) moss

layer lying over a substantial accumulation of humus

(6 to 12 inches; 15 to 30 cm). These sites are less prone

to drying than the spruce-lichen woodlands because

the ground fuel receives less radiant energy and less

wind (Foster 1983; Viereck 1983). The tendency of

black spruce to hold dead branches along the entire

length of their stems provides a fuel ladder for fire to

spread from the flammable shrub layer up into the

crown in both fuel types (Viereck 1983).

In Alaska, open canopy black spruce with feather

moss, lichens, low shrubs, and little woody debris is a

common forest type (Viereck and Dyrness 1980). In

such associations, the shrub layer is not an important

part of the fuel ladder since fire bridges from the mosslichen layer into the spruce crowns. In fuel studies

surface woody fuels varied from 0.67 to 20.0 tons/acre

(1.50 to 44.9 t/ha) and forest floor varies from 5.4 to

23.8 tons/acre (12.2 to 53.4 t/ha) (Dyrness and Norum

1983; Kiil 1975; Stocks and others 1990).

White Spruce—Mature northern white spruce

stands have well developed moss layers, mostly mountain fern moss, Schreber’s moss, knights plume moss,

dicranum moss, and sphagnum. In northern Ontario,

surface woody fuels varied from 10.9 to 13.1 tons/acre

(25.3 to 32.4 t/ha) and forest floor varied from 19.6 to

111.1 tons/acre (45.6 to 53.8 t/ha) (Stocks and others

1990). In the British Columbia Rocky Mountains,

surface woody fuels varied from 17.1 to 69.1 tons/acre

(38.3 to 155.5 t/ha) and forest floor varied from 23.3 to

61.5 tons/acre (52.3 to 137.7 t/ha) (Hawkes 1979;

Hawkes and others 1997b).

Balsam Fir—Fire behavior in budworm-killed

stands differs greatly between spring and summer. To

illustrate, spring fires resulted in sustained crowning

and spotting with extremely fast spread rates whereas

the summer fires were unable to sustain themselves

and did not spread. The major reason for this difference seems to be a proliferation of lush green understory vegetation in the summer due to crown openings

(Stocks 1987; Stocks and Bradshaw 1981).

Spring fires in balsam fir stands within the boreal

and Great Lakes-St. Lawrence forest regions fires

were common. They behaved explosively with continuous crowning, high spread rates, and severe downwind spot fires that killed and regenerated entire
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 47

Chapter 3: Fire in Northern Ecosystems Duchesne

stands. But they seldom burned much of the organic

layer because it is still cold and wet (Heinselman 1981;

Rowe and Scotter 1973). Summer fires were much more

likely to consume the organic layer. Surface woody fuels

varied from 6.53 to 17.4 tons/acre (14.5 to 38.7 t/ha) and

forest floor loading varies from 6.53 to 32.7 tons/acre

(18.7 to 52.8 t/ha) (Stocks and others 1990).

Red Spruce—Under drought and extreme fire

weather conditions, fires of high intensity covering

large areas or severe surface fires were possible. In old

stands where red spruce was associated with balsam

fir the periodic outbreak of the spruce budworm caused

heavy tree mortality. This made these stands more

susceptible to wildfires due to crown breakage and the

proliferation of highly flammable fine fuels such as

needles, dry twigs, and bark (insect-wildfire hypothesis) (Furyaev and others 1983). Fire potential is

greatest 5 to 8 years after tree mortality. During this

period, fires of great intensities tend to spread quickly

due to evenly distributed fuel. To our knowledge, fuel

loadings have not been reported in the literature

for red spruce ecosystems. Presumably they are comparable to those of black spruce and white spruce

ecosystems.

Conifer Bogs and Tundra—Fuel loadings in conifer bogs are highly variable because of the multiple

species combinations found in this forest type. Fuel

loadings for tundra are extremely variable, ranging

from nearly nothing in tundra barrens to a maximum

of 375 lb/acre (400 kg/ha) in Alaskan sheathed cottonsedge communities (Ranice and others 1987).

Postfire Plant Communities

Black Spruce

Pre-1900 Succession—The semiserotinous cones

of black spruce, located near the main stem, act as a

potential seed bank from which seeds are dispersed

after fire (Johnson 1992). Once open the cones release

seed for 2 to 3 years following fire. Black spruce is

relatively slow-growing and is best suited to mineral

soils. However, the seedling and sapling stages are

relatively shade tolerant, allowing black spruce to

establish well even on organic soil.

Revegetation following fire usually follows one of

two basic paths. On moderate to moist relatively

fertile soils, black spruce feather moss succession

dominates (Black and Bliss 1978; Viereck 1983; Wein

1975). After high severity fires that expose mineral

soil, bryophytes and herbs dominate while black spruce

seedlings establish. On the less severely burned sites,

herbs and shrubs develop from rhizomes that were not

killed by fire. This initial re-establishment lasts for up

to 4 years, and although black spruce becomes established at this time, the shrub layer may dominate the

site for the first 25 postfire years. Eventually, the

shade-tolerant black spruce, often growing at densities up to 12,000 stems/acre (28,800 stems/ha) (Viereck

1983) outcompete the shrub layer and forms the canopy.

The second regeneration path begins with the establishment of black spruce with mosses, lichens, and in

some cases grasses. After 10 to 60 years, the ground

cover on such sites is dominated by fruticose lichens

along with low ericaceous shrubs (Black and Bliss

1978; Viereck 1983). During this phase, the black

spruce component eventually forms an open canopy

until approximately 100 years after fire, a typical

black spruce lichen association is developed (Viereck

1983).

Post-1900 Succession—In much of the Canadian

and Alaskan boreal forest, fire regimes and postfire

communities remain similar to the presettlement period due to remoteness and lack of access. Most fire

disturbed stands eventually regenerate back into black

spruce ecosystems just as they would have before

effective fire suppression. In areas undergoing commercial harvest, slash material and improved access

(and resulting human activities) may combine to cause

more frequent and more severe fires than in the period

preceding fire suppression (Heinselman 1981). Fire

protection has lengthened fire rotation from 120 to 200

years in spruce lichen forest and from 100 to 150 years

in closed black spruce stands of Northwestern Canada

and Alaska (Viereck 1973). In Ontario and Quebec,

fire protection has increased fire rotations from 50 to

100 years to approximately 90 to 150 years (Heinselman

1981).

Management Considerations—Large-scale commercial harvest of black spruce and its associated

species in Canada is one of the single most important

industrial activities in the boreal forest. The resultant

emphasis on protection of black spruce stands from

insect and fire damage has generally increased the fire

rotation in black spruce, particularly in protected

forest reserves (Heinselman 1981; Viereck 1983).

Harvest scheduling of high risk stands (high fir component; older or senescent stands) is a strategy that

has also been used to reduce fire risks. In Alaska, black

spruce stands occur mostly on permafrost with low

stand productivity and are a low priority for fire

protection.

Where fire exclusion is practiced, the flammable

nature of black spruce ecosystems combined with

their remoteness present major problems to fire managers (Viereck 1983). In spruce-lichen woodlands, drying occurs rapidly (Auclair 1983; Johnson 1992) causing high fire hazard throughout the spring, summer,

and fall. In spruce-feather moss ecosystems, particularly in the East, higher moisture levels reduce fire

hazards. However, periodic droughts allow large tracts

of this forest type to burn (Heinselman 1981).
48 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Duchesne Chapter 3: Fire in Northern Ecosystems

Prescribed fire in black spruce cannot be used in

areas with shallow soil profiles as soil conservation is

a major concern in such sites. In low-lying Sphagnumdominated sites, fire is usually not required for seedbed

preparation as black spruce seeds easily germinate on

the Sphagnum bed itself (Archibald and Baker 1989).

In lowland spruce-feather moss stands, moderate to

high severity fires are required to remove the feather

moss layer, slash, and litter to prepare the seedbed.

When heavy competition from alder or other boreal

hardwoods is present, high severity burns are required to kill off the subterranean rhizomes of the

unwanted species, although if herbicides are applied

before the burn, low to moderate severity fires may

suffice (Archibald and Baker 1989).

The weather indices used for prescribed burns vary

with silvicultural goals. In northwestern Ontario, low

severity fires designed to burn fine fuels without

removing large amounts of duff require a duff moisture code of 4 to 10 and a buildup index of 15 to 19.

Moderate severity burns for the purpose of removing

fine fuels, some heavy fuels and patches of duff require

a duff moisture code of 8 to 16 and build up index of 18

to 28. Severe fires removing both fine and heavy fuels

as well as 0.5 to 1.5 inch (1 to 4 cm) of duff require a duff

moisture code of 13 to 25 and a build up index of 24 to

45 (Archibald and Baker 1989). Quantitative slash

and forest floor consumption can be predicted using

tables developed for lowland and upland black spruce

forest by McRae (1980).

White Spruce

Pre-1900 Succession—Depending on seed rain,

white spruce stocking may either decrease or increase

in postfire communities. Seed production varies greatly

from year to year with good seed crops every third year

or so. Cones mature in late August or September and

the majority of the seeds fall soon thereafter. So postfire

regeneration of white spruce is increased when fire

occurs in late summer of a good seed year. Otherwise,

the only available seeds come from unburned patches or

the edge of the burn (Dix and Swan 1971).

Feather mosses and lichens are eliminated wherever the moss layer has burned but reappear 10 to 15

years after fire and often reach high ground coverage

within 50 years. If only white spruce is present in the

prefire community, and seed trees are eliminated over

large areas, an initial establishment of aspen seedlings may occur, thus slowing the reestablishment of

white spruce (Van Cleve and Viereck 1981).

Post-1900 Succession—Past fires, logging, and

land-clearing in the Northeastern forests caused white

spruce associations to be replaced with aspen. The

successional trend for such stands is generally toward

spruce-fir and, where soil and climatic conditions

permit, toward northern hardwoods. Fire suppression

and control in Alberta allow stands to mature and

succession favors shade-tolerant species such as balsam fir and tolerant hardwoods (Weetman 1983).

Management Considerations—Clearcutting is

the usual harvesting method for white spruce ecosystems. Clearcutting followed by broadcast slash burning results in the most marked changes in vegetation

and soil temperature. Experiments in Alaska showed

that many of the species present before disturbance

disappeared afterwards (Dyrness and others 1988).

Soil temperature increased on clearcut and shelterwood

sites, but not as much as on clearcut and burned areas.

Thinning had the least impact on vegetation and soil

temperature (Dyrness and others 1988). Prescribed

fire is effective in reducing heavy slash, seedbed preparation, controlling unwanted vegetation, and promoting vigor in desired species. Because white spruce is

particularly sensitive to environmental conditions

during germination and early growth, it is among the

most difficult conifer species to regenerate naturally.

However, summer logging or scarification that exposes mineral soil seems to improve regeneration

(Brand and Janas 1988; Eis 1965, 1967; Endean and

Johnstone 1974).

Natural white spruce stands can respond well to

cultural practices, especially releasing. White spruce

stands should be maintained at basal areas from 100

to 140 ft2

/acre (23.0 to 32.1 m2/ha) to provide maximum volume growth and good individual tree development. Below these levels, individual tree increment

and resistance to some pests are greatly increased,

whereas total volume production is reduced (Burns

and Honkala 1990).

Balsam Fir

Pre-1900 Succession—Owing to the high sensitivity of this species to fire, balsam fir survives only

extremely low intensity fires or in patches of unburned

forest. Even if trees are only damaged, fungal diseases

and insect attacks will quickly destroy the stand

structure. Following canopy opening by a first fire, a

second fire can cause even greater changes in species

composition (Little 1974).

Long fire cycles allow establishment of old-growth

stands dominated by shade-tolerant balsam fir regeneration, a condition that is most abundant about 100

years after fire (Heinselman 1973). Because balsam fir

has little fire tolerance, fire in balsam fir dominated

forests tends to eliminate most of the existing stems

and favors conversion to other tree species. Accordingly, fire in black spruce-balsam fir stands tends to

favor the black spruce component (Candy 1951; Damman

1964; MacLean 1960). In the absence of conifer seed

trees, postfire communities are dominated first by
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 49

Chapter 3: Fire in Northern Ecosystems Duchesne

aspen or paper birch seedlings that originate from

wind-borne seeds. Subsequently, white or red spruce

seedlings invade the burns, developing as an understory to the aspen-birch complex and eventually replace it. On some of the better sites, northern hardwoods such as sugar maple and beech eventually

replace white spruce. Alternatively, balsam fir dominates (Day 1972).

Post-1900 Succession—Fire protection, coupled

with clear-cut logging have favored the establishment

of pure balsam fir associations. The shade-tolerant

seedlings of balsam fir establish well in the shade of

balsam fir or spruce stands. Thus, following logging,

pre-established balsam fir generally dominates in the

regenerating stand. However, in some situations,

bracken fern, raspberry, and hardwood suckers may

represent intense competition to the balsam fir regeneration for the first 10 to 25 years after disturbance

(Burns and Honkala 1990). Repeated burns or cutting

and burning in Newfoundland can lead to the development of Kalmia barrens or heathland conditions

(Damman 1964), which once established are difficult

to reforest. Black spruce invades such sites due to poor

seed supplies of balsam fir and soil deterioration

(Damman 1964).

Management Considerations—Clearcut logging

has been the most used harvesting method for this

forest type. However, site characteristics, the time of

season, logging practices and size of the cut area are

important factors to prevent damage for the preestablished regeneration of most spruce and fir stands,

especially in the boreal region. A study in northwestern Quebec has shown that there is a deterioration in

stand quality and a severe softwood mortality (92

percent) after harvesting and a shift in species composition from softwood-dominated advance regeneration

to mixed or hardwood and shrub-dominated secondary forests. Often, it will be necessary to treat the area

if the softwood crop is desired. Silvicultural treatments include scarification, direct seeding, planting,

brush control, and subsequent thinning.

Overmature balsam fir forests (about 100 years of

age) are susceptible to spruce budworm infestations

(Heinselman 1973). Silvicultural options include harvest or protection against spruce budworm outbreaks,

which involves the use of chemical or biological insecticides (Dimond and others 1984). Budworm outbreaks

have serious economical and social impacts on forest

industry. Accordingly, there is interest in converting

fir forests in eastern Canada that are susceptible to

spruce budworm to less susceptible species, particularly black spruce. Species conversion requires intensive site preparation owing to the presence of understory fir regeneration that provides strong competition

to planted spruce seedlings on many sites. A costeffective method of site preparation is prescribed burning, which eliminates fir regeneration completely and

favorably prepares the site for spruce seedlings. Prescribed burning in balsam fir forests could also be used

for removal of slash prior to planting, or for reduction

of fire hazard (Furyaev and others 1983). Prescribed

burning in budworm-killed forests has been applied in

Ontario (Stocks 1987) and in Quebec (Robitaille 1994).

Red Spruce

Pre-1900 Succession—Red spruce was a late successional species forming old-growth stands in areas

with long fire cycles (Burns and Honkala 1990). Its

shallow root system, thin bark, and flammable needles

make trees of all ages susceptible to fire damage.

Natural reproduction of red spruce depends on seedling survival since seedlings have an exceptionally

slow-growing, fibrous, and shallow root system. A

critical factor for their survival and establishment is

the depth of the organic layer. Although a shadetolerant species, the relative tolerance of red spruce to

shade varies with soil fertility and climate. At first,

seed germination and initial seedling establishment

proceed best under cover. However, older seedlings

require 50 percent or more sunlight for optimum

growth (Burns and Honkala 1990).

Red spruce’s chief competition comes from balsam

fir and hardwoods such as beech and maple. Red

spruce seems to respond to release even after many

years of suppression (up to 100 years). Nevertheless,

many of its associated tree species such as balsam fir

and hemlock may outgrow red spruce after release

(Burns and Honkala 1990).

Post-1900 Succession—Clearcutting had a serious impact on red spruce ecosystems. Many former

spruce sites are now occupied by inferior tree species,

blackberries, and ferns. Dense growth of bracken,

raspberry, and hardwood sprouts are the chief competition for seedlings on cutover lands (Burns and

Honkala 1990). Logging and fire in the Maritimes has

left large areas of red spruce stands in poor conditions

in terms nutrient status and species composition,

leading to poor growth and conversion to heathlands.

In turn, this has caused shortages of saw timber and

pulpwood (Weetman 1983).

Management Considerations—Red spruce may

be grown successfully using even-aged silvicultural

prescriptions. Because red spruce is shallow-rooted, it

is highly sensitive to windthrow. To minimize the

danger of wind damage, it is recommended that no

more than one-fourth to one-half of the basal area,

depending on site, be removed in the partial harvest of

a spruce-fir stand (Frank and Blum 1978).
50 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Duchesne Chapter 3: Fire in Northern Ecosystems

Conifer Bogs

Pre-1900 Succession—There are two different

views on the effect of fire on cold northern bogs. Rowe

and Scotter (1973) describe how fire can contribute to

site paludification. Fire removes the transpiring

crowns, the water table rises, light levels increase on

the ground floor, which promotes the establishment of

Sphagnum mosses. Then moss layers build up, insulating the soil and lowering soil temperature. All of

these fire effects create poor seedbed and poor rooting

conditions for trees, thus excluding tree establishment and leading to a tundra type community. Conversely, Heinselman (1981) found that prolonged firefree intervals cause a rise in permafrost tables, impede

soil drainage and general site degradation. The result

is a depauperate black spruce muskeg (Heinselman

1981). Strang (1973) found the same successional

pattern in northern open black spruce-lichen

associations.

Spruce-Sphagnum bogs that have been destroyed by

fire are replaced by birch and other pioneer species,

then evolve to a stage where both birch and spruce are

present and, finally, to the spruce stage at which

paludification begins (Larsen 1980). However, recurrence of fire in conifer bogs prevents most stands from

reaching their climax stage (Bergeron and Dubuc

1989; Heinselman 1973, 1981).

Lightning was, and still is, a major cause of fire in

the boreal forest. Lightning strikes during rainless

storms during prolonged summer droughts ignite major

fires to which conifer bogs are most susceptible. Such

fires expose the mineral soil and create seedbed conditions conducive to the regeneration of conifer bog

species (Heinselman 1981).

Spring and fall fires become more common in the

boreal forest with increased human activity. Such

fires have relatively little effect on reducing the duff

layer and subsequently produce poor seedbed conditions for conifer bog species (Heinselman 1981; Viereck

and Johnston 1990). Lightning fires are probably still

responsible for most of the annual area burned.

Before 1900, fire prevented high concentrations of

dwarf-mistletoe on black spruce because this parasite

is temporarily eliminated when fire removes its host.

Therefore fire was able to eradicate centers of infestation, thus preventing this disease from spreading.

Current fire exclusion policies favor vast expansions of

mistletoe (Heinselman 1973).

Post-1900 Successions—Human influence on conifer bog sites has been somewhat limited because

most bogs have poor timber productivity. Most anthropogenic influence arises from the harvesting of surrounding stands and the side effects of fire protection

on those same higher value sites. The policy of fire

protection has lengthened the fire cycle for the boreal

forest and, therefore, the fire cycle of nearby bog sites

(Heinselman 1981).

Management Considerations—Management

practices in conifer bogs depend greatly on site productivity. Poorly drained, wet sites have such low wood

productivity that little silvicultural management is

done in these conifer bogs (Viereck and Johnston

1990). For the more productive peatland black spruce

sites, clear-cutting in strips or patches, combined with

broadcast burning of slash, followed by direct or natural seeding, is the best silvicultural treatment

(Johnston 1975; Viereck and Johnston 1990). This

same system works well for northern white cedar, but

shelterwood management is preferred as it supplies

partial shade for establishing seedlings (Johnston

1990b).

The shade-intolerance of tamarack dictates the use

of even-aged management, with some adaptation of

clearcutting or seed tree cutting generally considered

the best silvicultural system (Johnston 1990a). Uneven or all-aged management is best applied to poor

black spruce sites where stands are wind firm and

have abundant layering (Viereck and Johnston 1990).

Prescribed burning improves seedbed conditions and

kills back competing brush (Armson 1982; Johnston

1973, 1975, 1976). However, tamarack slash is difficult to burn (Johnston 1973, 1975). Seedling survival

is substantially greater on prescribed burn sites.

Tundra

Pre-1900 Succession—Fire effects in tundra ecosystems vary greatly according to the composition of

prefire plant communities (Bliss and Wein 1971).

Generally, fire favors rapidly growing species, particularly graminoids, and there is a decreased abundance of slow growing species such as evergreen shrubs

immediately following fire (Bliss and Wein 1971, 1972).

The recovery of mosses and lichens is slow as opposed

to that of sedges and grasses, and recovery of shrubs is

intermediate (Bliss and Wein 1972). Postfire regeneration is reliable because all shrubs and deciduous

tree species common to the tundra are capable of

reproducing vegetatively after fire (Auclair 1983;

Viereck and Schandelmeier 1980). Establishment of

pioneer species is mainly by wind-borne seeds from

adjacent plant communites (Auclair 1983). Most lichens establish by small thallus fragments within the

first several years following a burn, but their slow

growth limits their abundance for the first 25 to 30

years (Auclair 1983).

The forest tundra zone is transitional, depending on

fire cycle and fire severity (Viereck and Schandelmeier

1980). Sirois and Payette (1991) suggested that recurring fire progressively decreases the regenerative potential of trees and that sustained depletion of tree
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 51

Chapter 3: Fire in Northern Ecosystems Duchesne

populations following several fires is a key process in

the development of the forest tundra. In the shrub

subzone of the forest tundra where tree regeneration

is tenuous, a single fire event can eradicate the presence of trees and creates treeless tundra. The boreal

forest-forest tundra interface may be prone to sudden

fragmentation caused by fire and climate interactions

(Sirois and Payette 1991). Severe fires causing depletions of trees resulting in shifts from forest to foresttundra communities have been reported from subalpine sites in Western North America and subarctic

sites in northern Sweden and northern North America

(Sirois and Payette 1991).

Post-1900 Succession—There has been relatively

little human activity in the vast expanse of the arctic

tundra (Heinselman 1981). Most activity has been

since the late 1960s with oil and gas exploration and

the construction of pipelines. Disturbance to tundra

communities has been mostly in the form of cut-lines

(or seismic lines), winter roads, air strips, and fire

(Bliss and Wein 1971). Continuing investigation into

the effect of these anthropogenic disturbances has resulted in the use of technologies that have a less

deleterious impact on plant communities of the tundra.

Management Considerations—Efforts to contain

or stop the spread of fire in the tundra produce more

drastic long-term effects than the fire itself (Brown

1971; Viereck and Schandelmeier 1980). Construction

of firelines with bulldozers strips away all insulating

moss and peat layers and exposes bare mineral soil.

This allows the summer heat to penetrate directly into

the frozen ground, which in turn increases the depth of

the active layer under the fireguards compared to under

burned areas. This causes a more rapid and greater

degree of subsidence under the firelines than under the

burned areas due to the melting of ground ice (Brown

1971), erosion and gully formation (Brown 1983).
52 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Notes

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USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 53

Dale D. Wade

Brent L. Brock

Patrick H. Brose

James B. Grace

Greg A. Hoch

William A. Patterson III

Chapter 4:

Fire in Eastern Ecosystems

Prior to Euro-American settlement, fire was a ubiquitous force across most of the Eastern United States.

Fire regimes spanned a time-scale from chronic to

centuries. Fire severity varied from benign to extreme

(fig. 1-2). Today, fire is still a major force on the

landscape. In some ecosystems fire stabilizes succession at a particular sere, while in others, succession is

set back to pioneer species. The wide range in fire

regimes coupled with elevation and moisture gradients produce a myriad of plant communities that

continually change over time in both stature and

composition, although it is not uncommon for the

major species to remain dominant. Discussion is primarily about major vegetation types, for example,

oak-hickory. However, some minor types such as

spruce-fir and Table Mountain pine are also covered.

Vegetation types are discussed under the most representative fire regime type, recognizing that some vegetation types overlap two fire regime types (table 4-1).

Understory Fire Regimes _________

Major Vegetation Types

The presettlement understory fire regime (overstory survival typically exceeds 80 percent) primarily

applies to southern pine and oak-hickory forests

comprised of pine and pine-oak associations such as

Kuchler’s southern mixed forest, oak-hickory-pine,

Northeast oak-pine types, and oak-hickory associations. Kuchler potential natural vegetation classes are

listed by fire regime types in table 4-1 and crossreferenced to FRES ecosystems and Society of American Foresters cover types.

The extent of the various forest types at the time of

Euro-American colonization is difficult to reconstruct.

Considering old survey corner trees, Plummer (1975)

concluded that pine and post oak predominated in the

Georgia Piedmont. Based on soils, Nelson (1957) concluded that 40 percent of the Piedmont was in hardwoods, 45 percent was hardwood with varying degrees

of pine, and 15 percent was predominantly pine. Nomenclature used by early naturalists was often ambiguous including different names for the same species, or use of a species name to group several species

together such as the use of the term short-leaf pine to

differentiate all short-leafed pines from longleaf pine.

Furthermore, earlier works estimated that 80 percent

of this region had been cleared at some point in time

(Nelson 1957). Hamel and Buckner (1998) described

the “original southern forest” at three periods and

concluded that no specific period represents the “true”

original condition of the southern forest because it has

been shaped by humans for millennia.
54 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems(con.) Table 4-1

—Occurrence and frequency of presettlement fire regime types by Forest and Range Environmental Study (FRES) ecosystems, Kuchler potential natural vegetation

classes (1975 map codes), and Society of American Foresters (SAF) cover types. Occurrence is an approximation of the proportion of a vegetation class represented

by a fire regime type. Frequency is shown as fire interval classes defined by Hardy and others (1998) followed by a range in fire intervals where data are sufficient.

The range is based on study data with extreme values disregarded. The vegetation classifications are aligned to show equivalents; however, some corresponding

Kuchler and SAF types may not be shown.

Fire regime types

Understory Mixed Stand-replacement

FRES Kuchler SAF Occura Freqb Occur Freq Occur Freq Nonfire

Longleaf-slash pine 12 Southern mixed forest K112 Longleaf pine 70 M 1a:1-4

Longleaf-slash pine 83c M 1a:1-4

Slash pine 84 M 1a:3-8 M 1a:5-10

Loblolly-shortleaf pine 13 NE oak-pine K110 Virginia pine 79 M 1b

Pitch pine 45 M 1a M 1b

Oak-hickory-pine K111 Shortleaf pine 75 M 1a:2-15

Loblolly-shortleaf pine 80 M 1a:3-10

Loblolly pine 81 M 1a:3-8 m 1a:5-10

Pocosin K114 Pond pine 98 m 1a:3-8 M 1:6-25

Sand pine-scrub K115 Sand pine 69 M 1b:25-45

Oak-pine 14 Oak-hickory-pine K111 Shortleaf pine-oak 76 M 1a

Virginia pine-oak 78 M 1b

Table mountain pine M 1,2

Southern mixed forest K112 Loblolly pine-hardwood 82 c M 1

Slash pine-hardwood 85 m 1

Longleaf pine-scrub oak 71 M 1a:6-10

Spruce-fir 11 S.E. spruce-fir K097 Red spruce-Frazer fir 34 M 2,3 M

Oak-hickory 15 Oak-hickory K100 Northern pinoak 14 M 1

Mosaic of above and Bur oak 44 M 1a

bluestem prairie K082

Appalachian oak K104 Northern red oak 55 M 1b

Bear oak 43 M 1

Chestnut oak 44 M 1a:3-8

Black oak 110 M 1

White oak-black oak-n.red oak 52 M 1

Southern scrub oak 72 M 1 M1

Blackbelt K089 Yellow popular 57 M 1 M1

Cross timbers K084 Post oak-blackjack oak 40 M 1a

Oak-savanna K081 M 1a:2-14

Elm-ash-cottonwood 17 Northern floodplain forest K098 Black ash 39 M 1,2 M 1,2

Elm-ash forest K101 Silver maple-American elm 62 M 1,2 M 1,2

Cottonwood 63 M 1,2 M 1,2

Sugarberry-Am. elm-green ash 93 M 1,2 M 1,2

Sycamore-sweetgum-Am. elm 94 M 1,2 M 1,2
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 55

Chapter 4: Fire in Eastern Ecosystems Wade Table 4-1—Con. Fire regime types Understory Mixed Stand-replacement FRES Kuchler SAF Occura Freqb Occur Freq Occur Freq Nonfire Oak-Gum-Cypress 16 Southern floodplain forest K113 Alantic white-cedar 97 M 2,3 Maple-beech-birch 18 Maple-basswood K099 Sugar maple-basswood 26 M 3:>1000 m Beech-maple K102 Beech-sugar maple 60 M 3:>1000 m Black cherry-sugar maple 28 M 3:>1000 m N. hardwoods K106 Sugar maple-beech-yellow birch 25 M 3:>1000 m Sugar maple 27 M 3:>1000 m N. hardwoods-fir K107 Hemlock-yellow birch 24 M 3 N. hardwoods-spruce M 3:300- K108 500 Mixed mesophytic forestd Mixed mesophytic forest K103 M 2,3 Prairie 39 Bluestem prairie K074 M 1a Nebraska Sand Hills K075 M 1a Blackland prairie K076 M 1a Bluestem-sacahuista prairie M 1a K077 Cedar glades K083 M 1a:3-7 Fayette prairie K088 M 1a Wet grasslands 41 Tule marshes K049 M 1 N. cordgrass prairie K073 M 1a:1-3 S. cordgrass prairie K078 M 1a:1-3 aM: major, occupies >25% of vegetation class; m: minor, occupies <25% of vegetation class. bClasses in years are 1: <35, 1a: <10, 1b: 10 to <35, 2: 35 to 200, 2a: 35 to <100, 2b: 100 to 200, 3: >200. cThis type developed after European settlement. dThis type is a subdivision of FRES ecosystem types.
56 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

The Southern mixed hardwood forest occupies the

Atlantic and Gulf Coastal Plains and is often classified

as deciduous broadleaved forest even though it includes pine-hardwood and broadleaved evergreen

types. Southern pines are the predominant overstory

species. The potential hardwood climax, however, was

probably only reached on sites with a geomorphologic

position or hydrologic condition that protects such

sites from the chronic fire regime that characterizes

this region. Southern mixed hardwood forests were

confined to narrow strips between floodplain forests

and uphill longleaf pine forests before Euro-American

colonization (Batista and Platt 1997). Some authors

placed the former longleaf pine forests in this forest

type, but this seems illogical where time since disturbance, primarily by hurricanes and fire, and the elevation gradient (measured in inches) determine plant

community composition. Embedded within this overall forest type are numerous other ecosystems such as

nonalluvial depressional wetlands, including Carolina bays, limesinks, cypress ponds and savannas,

gum ponds, bay swamps, pitcherplant bogs, shrub

bogs, and spring seeps. Water stands in these depressions during at least part of the growing season, but

they also dry allowing fire to enter from adjacent

upland plant communities (Kirkman and others 1998;

Wharton 1977). These fires regulate numerous processes needed to maintain these communities

(Christensen 1977; Wade and others 1980; Wright and

Heinselman 1973). Periodic fire also knocks back the

wall of hardwoods that continually invades adjacent

upland sites (Barrett and Downs 1943; Chaiken 1949;

Chapman 1947; Harcombe and others 1993; Heyward

1939; Long 1888; Oosting 1942; Platt and Schwartz

1990; Wahlenberg 1949; Wells 1928). As the time span

between fires increases, the stature of the ever-present

hardwoods also increases, shading out the herbaceous

groundcover, thereby breaking up its continuity and

ability to support spread of fire; this allows the hardwoods to further expand until they eventually dominate the site (fig. 4-1). Tansley (1935) recognized this

situation and classified Southeastern pine forest as a

disturbance-dependent pyrogenic stable type. Wilhelm

(1991) extended this thinking to the Central Hardwoods Region and questions the whole concept of

succession and climax vegetation in the Midwest.

Over three dozen oaks and almost two dozen hickories form a myriad of potential overstory plant communities that comprise the oak-hickory forest (fig. 4-2).

We define it as all stands in which oaks and hickories,

singly or in combination, compose at least 50 percent

of the dominant trees (J. A. Barrett 1994; Braun 1950).

This forest type includes the oak-hickory-pine forest of

the mid-Atlantic States, the Appalachian oak forest,

the Northeast oak-hickory-pine forest, and stands

within the western mesophytic and mixed mesophytic

forest regions where oaks and hickories comprise a

majority of the dominant trees. Such stands occur

primarily on average to dry upland sites, although

they can be found on moist upland sites as well,

depending upon past disturbance history. Where other

hardwood species dominate within the western and

mixed mesophytic forest regions, they are discussed as

mixed fire regime types.

Fire Regime Characteristics

Southern Pine Forests—The regional climate is

characterized by long, hot growing seasons, abundant

rain punctuated by occasional multiyear droughts,

and the most frequent wind and lightning (Komarek

1964) storms in North America. Natural disturbance

events such as microbursts, tornadoes, and hurricanes can significantly impact forested lands at several scales and often set the stage for more intense

fires (Myers and Van Lear 1997). Lightning becomes

increasingly common as one moves from north to

south. Although the number of lightning fires peaks in

June and July, the vast majority of acreage now burns

in May and June in Florida (Robbins and Myers 1992)

and southern Georgia, before the summer thunderstorm pattern becomes fully entrenched; this probably

was the historic pattern as well.

The few remaining old-growth southern pine forest

relics are too small to experience natural fire regimes,

and neither dendrochronology nor palynology can be

used to determine fire history in this region (Landers

1991). Thus, historical southern pine fire regimes

must be inferred from interpretation of old records,

field observations, experimental studies, and species

traits. Frost’s (1998) generalized map of pre-EuroAmerican settlement fire frequency for the Southern

United States summarizes current thinking and shows

that all forest types mentioned above had a fire-return

interval of less than 13 years except pond pine pocosins in North Carolina.

Before Euro-American settlement, some fires were

undoubtedly far ranging because they were associated

with dry frontal passages. Growing season fires during severe drought years can burn for weeks or months,

particularly in organic soils that underlie the deeper

depressional wetlands, before being extinguished by

rain and a rising water table (Cypert 1961, 1973). Such

fires make frequent forays into adjacent upland communities. Most fires were, however, probably limited

in extent, at least once the summer convective weather

pattern set in. Nighttime humidities near 100 percent

are the norm during the summer; and such humid,

calm conditions tend to extinguish fires in light fuels.

The pattern of occasional high-intensity, wind-driven

fires and severe drought fires was superimposed on

the chronic lightning and Native American fire regime
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 57

Chapter 4: Fire in Eastern Ecosystems Wade

Figure 4-1—Fifty years of fire exclusion in an old-growth longleaf pine stand, Flomaton, Alabama. Photo

by John Kush, 1994.

Figure 4-2—The combined distribution of the Oak-Hickory

Forest Region, Mixed Mesophytic Forest Region, and Appalachian Oak Forest Region (Kuchler Types 100, 103, and 104) as

depicted by Martin and others (1993).

creating the open woodlands referred to by early EuroAmerican explorers (Barden 1997; Landers and others

1990; Olson 1996) (fig. 4-3). The journals of many of

these explorers also mention numerous smoke columns

and extensive smoke and haze often lasting for days.

Native Americans used fire extensively to shape the

vegetative mosaic for thousands of years, and for the

past 400 or more years Euro-Americans have subjected these same lands to varying degrees of fire use

and abuse, including the exclusion of fire. Whether

human ignitions substantially increased the total acres

treated by fire is uncertain, but based on the sheer

number of artifacts left behind, accounts by early

white explorers, and studies of depopulation ratios

following disease (Dobyns 1966; Jacobs 1974), Native

American populations must have been much larger

than is generally recognized. According to Kennedy

(1994), the population of Cahokia in central Illinois

was greater than that of London in the 13th century.

Native American-induced changes include the fact

that they ignited fires throughout the year, but before

their influence, many sites burned more frequently by

fires confined to the spring and summer thunderstorm

season (Martin and Sapsis 1992). A slight shift toward

reduced fire intensity and severity took place. Growing-season fires tend to be patchier than dormantseason fires because of different weather patterns and
58 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

greater variation in fine-fuel moisture once greenup

occurs.

Komarek (1982) pointed out that it is difficult to

understand the practice of prescribed fire in the South

without an appreciation of Southern history. It is also

a prerequisite to realizing the tremendous impact humans have had in shaping the present vegetative mosaic. For example, prior to the Civil War, over 75

percent of the white population were pastoral herdsmen (Owsley 1945) who came from the British Isles,

Spain, and France where fire was an integral part of

their livelihood. They brought this practice with them,

blended their knowledge with that of the Native Americans they displaced, and aggressively expanded the use

and frequency of fire throughout the South. A circa 1731

North Carolina law required the annual burning of all

pastures and rangelands every March (Carrier and

Hardison 1976). Eldredge (1911) described the turn-ofthe-century fire situation in north Florida as follows:

...the turpentine operator burns his woods and

all other neighboring woods during the winter

months, generally in December, January, or

February. The cattleman sets fire during March,

April, and May to such areas as the turpentine

operator has left unburned. During the summer

there are almost daily severe thunderstorms,

and many forest fires are started by lightning. In

the dry fall months hunters set fire to such

“rough” places as may harbor game. It is only by

chance that any area of unenclosed land escapes

burning at least once in two years.

There is little doubt that this pattern was typical

throughout the region, although not everyone was in

agreement with this ubiquitous use of fire. The debate

regarding the benefits of intentional burning versus

fire exclusion has been ongoing for more than 100

years as attested to by the following quote from the

1875 Atlanta Constitution (Blocker 1875).

I have seen an article in the Early County News

of June 12th, copied from the Columbus Sun

condemning the burning of woods, but the writer

is very badly mistaken in his views on the subject. I think there has been a great deal more

injury from not burning the woods at a proper

time, and when in proper condition, than has

ever been done by burning; for fire will accidentally get in some time, and often when it is very

dry, and do a great deal of injury.

Although some early Federal foresters thought fire

was universally bad (Schwarz 1907), others disagreed.

Chiefs of the United States Forest Service Gifford

Pinchot (1899) and Henry Graves (1910), Supervisor

of the Florida National Forests I.F. Eldredge (1911), as

well as university researchers Gifford (1908), Bryant

(1909), and Chapman (1912) recognized the invariable

result of attempted fire exclusion and advocated the

use of prescribed burning in Southern pines as a

hazard reduction measure. However, with passage of

the Clarke-McNary Act in 1924, which offered the

States Federal funding for fire suppression, thinking

within the Forest Service reversed itself, the use of fire

was condemned (Demmon 1929), and the Dixie Crusaders crisscrossed the Deep South preaching the

evils of woods burning. An excellent account of the

methods used to force compliance with this new fire

exclusion policy is given by Schiff (1962).

With attempted fire exclusion, dead fuels accumulated on the forest floor and a needle-draped understory of highly flammable shrubs such as saw palmetto, gallberry, and wax myrtle developed within 5

or 6 years on all but xeric sites. Given this receptive

fuel bed and human and natural ignition sources,

wildfire occurrence remained high. At first, these fires

were easy to extinguish, but as fuels accumulated on

unburned areas, fires became increasingly difficult

to suppress, and the probability of catastrophic,

Figure 4-3—Pine flatwoods near Tallahassee, Florida, around

1900. Note open understory. Florida Department of Environmental Protection archives.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 59

Chapter 4: Fire in Eastern Ecosystems Wade

high-intensity fire increased. Often, the postfire outcome was atypical successional pathways with concomitant declines in flora and fauna (fig. 4-4).

Many of these ecosystems can still be restored with

the judicious reintroduction of fire (fig. 4-5), sometimes in combination with other chemical and mechanical methods, because the long association between fire and Southern vegetation has evolved species

traits (table 2-1) that favor them in fire-prone ecosystems (Christensen 1977; Landers 1991). Providing a

certain threshold limit has not been reached, the

natural resiliency within these systems allows them to

recover (Vogl 1976). However, once this threshold

limit has been exceeded, nature can no longer rectify

the situation. Thus, many components of the original

ecosystem cannot survive long periods without fire

(Garren 1943).

Most students of fire history agree that typical

longleaf pine sites burned every 1 to 4 years prior to

the arrival of Europeans, and then every 1 to 3 years

until aggressive fire suppression activities began in

the 1920s and 1930s (Landers 1991; Landers and

others 1990). As typical upland sites grade toward

mesic (wet) or xeric (dry and thus low rates of fuel

accumulation), fire frequency decreases. Loblolly,

slash, and pond pines were historically confined to

wetter sites (Monk 1968) where a 3 to 4 year fire-free

interval allowed saplings to become large enough to

withstand low intensity fire. In developing a definition

for old-growth wet-pine forests, Harms (1996) emphasized the fact that fire is necessary to establish and

maintain these ecosystems. The historic shortleaf pine

fire return interval is thought to have ranged from

about 2 to 6 years on fertile, lower elevation sites to 6

to 15 years on drier, nutrient-poor sites where fuels

take longer to accumulate. Annual burning was extensively practiced throughout the shortleaf pine region

(Matoon 1915).

Fires in the depressional wetlands embedded in the

understory fire regime types are typically stand-replacement. Wharton (1977) gives fire cycles for these

types as follows: 3 to 9 years in herb bogs and highdiversity shrub bogs (including pocosins); 20 to 30

years in many cypress ponds, gum ponds and bog

swamp; 20 to 50 years in low diversity (titi) shrub

bogs; and 50 to 150 years in some cypress ponds and

bay swamps.

Perturbations to these chronic low-intensity Southern pine fire regimes occurred in the form of episodic

(often catastrophic) events such as blowdowns (fig. 4-6)

and drought, which were precursors to fires of much

higher intensity and severity (Myers and Van Lear

Figure 4-4—The forest floor accumulated from decades of fire exclusion completely consumed after an

extreme drought-year fire on Myakka River State Park, Florida. Fetterbush (center), which typically

resprouts prolifically after fire, was killed. Note weak resprouting of palmetto. Photo by Robert Dye, 1985.
60 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

Figure 4-5—Longleaf pine sandhill site following three prescribed fires during 6 years, Wekiwa Springs State

Park, Florida. Note wiregrass in flower 1-year post burn. Photo by Jim Stevenson, 1983.

Figure 4-6—Growing-season burn in a 3 year rough on Francis Marion National Forest, 8 months after

Hurricane Hugo. Photo by Ken Forbus, 1990.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 61

Chapter 4: Fire in Eastern Ecosystems Wade

1997). Explosive increases in southern pine beetle (Ips

spp. and Dendroctonus frontalis) populations and subsequent pine mortality often either preceded or followed these fires. In nature, such infrequent events

are often controlling factors that have profound effects

limiting vegetative development past a certain point

so that succession is a cyclic phenomenon (Vogl 1976).

Intentional use of fire to manage vegetation again

became commonplace after World War II. Besides

rough reduction, vast acreages were burned to: (1)

improve native range (Halls and others 1952,1964;

Hilmon and Hughes 1965a; Lewis 1964; Shepherd and

others 1951), (2) control the relentless hardwood invasion into pine stands (Brender and Cooper 1968;

Chaiken 1952; Ferguson 1957; Harrington and

Stephenson 1955; Lotti 1955), and (3) manipulate wildlife habitat to favor herbs and forbs and replace unpalatable out-of-reach woody understory crowns with

succulent, nutritious sprouts (Brennan and others

1998; Harlow and Van Lear 1981, 1987; Harris 1978;

Wood 1982).

During the 1980s, an estimated 4 million acres (10

million ha) of forest land and 4 million acres of range

and agricultural land were treated with prescribed

fire each year in the Southern United States (Wade

and Lunsford 1989). Prescribed fire is used by Southern resource managers to meet widely varying objectives: restoration and maintenance of ecosystems,

reduction of hazardous fuels, reduction of wildfire size

and suppression costs, reduction of firefighter risks,

preparation of seedbeds and planting sites, facilitation of harvesting (fiber, plants, game animals, insects, and earthworms), enhancement of wildlife habitat, range improvement, control of insects and diseases,

thinning overly dense stands, maintenance of scenic

vistas, promotion of showy herbaceous species, manipulation of understory structure and composition,

creating a safer environment for woods workers and

equipment operators, favoring plant and animal species of special concern, eradication of vermin, control

of exotic species, stimulation of fruit and fiber production, disposal of crop residues, and recycling of nutrients. The majority of treated acreage is for hazard

reduction, wildlife habitat improvement, and range

management. Nearly 2 million acres (0.8 million ha) of

rangeland are burned annually in Florida alone. An

increasing area is burned each year to maintain the

function of fire in ecosystems, particularly on State

and Federal lands. Today, prescribed fire is used to

treat over 6 million acres annually, which is only a

small fraction of the land that once supported the vast

Southern pine forests.

Oak-Hickory Forests—The fire regime of the oakhickory forest has varied spatially and temporally

because of changing cultural influences before, during, and after Euro-American settlement. Before such

influence, the fire regime for oak-hickory forests was

dictated by the activity of Native Americans because

they were the primary ignition source (Abrams 1992;

Buckner 1983; Denevan 1992; Pyne 1997). Native

Americans in the Central Hardwoods Region (Delcourt

and Delcourt 1997, 1998; Olson 1996), the Appalachians and Piedmont (Van Lear and Johnson 1983),

and the Northeast (Buell and others 1954; Day 1953)

commonly used fire for numerous reasons throughout

the year. Surface fires predominated and burned over

large areas; only natural barriers or unfavorable

weather stopped them. Hough (1877) thought the “oak

openings,” “barrens,” and prairies east of the Mississippi resulted from Native American use of fire to

promote grass growth and attract game. He stated,

“Scarcely a year passes without the occurrence of fires

of sufficient extent to attract public notice.” Numerous

authors (DeViro 1991; Patterson and Sassamen 1988;

Stewart 1951, 1963; Van Lear and Waldrop 1989)

have discussed the vast extent to which Native Americans used fire. One historian (Russell 1983) agreed

that fire frequency was greater near camps and villages than would be expected by lightning, but found

no strong evidence that Native Americans burned

large areas in the Northeast.

Lightning fires are uncommon in many of these

regions because thunderstorms occur primarily during the growing season, usually accompanied by rain

(Barden and Woods 1974; Ruffner and Abrams 1998).

However, lightning fires regularly occur in some areas

such as the Piedmont of North Carolina (Barden

2000).

Presettlement fire frequencies are not known.

Delcourt and Delcourt (1997, 1998) believe they varied

considerably depending on closeness to Native American habitation. Other references point out that Native

Americans maintained an extensive trail system

throughout the East that was kept open with fire.

Euro-American explorers reported many areas treated

with annual and biennial fires (Barden 1997; Buckner

1983; Day 1953). Dendrochronological studies, which

give conservative estimates, suggest fire return intervals of 7 to 14 years in the mid-Atlantic and Ozark

regions (Buell and others 1954; Guyette and Day

1997). Cutter and Guyette (1994) reported a firereturn-interval of 2.8 years during 1740 to 1850 on a

ridgetop in the Mark Twain National Forest. Presettlement fires in southern New England generally

occurred during the spring and summer (Bromley

1935; Christianson 1969). Brown (1960) believes the

prevalence of oak in Rhode Island is the result of a long

history of fire. The fire regime was probably more

pronounced in Southern areas than in Northern areas

due to more favorable climatic conditions for ignition

and spread, greater populations of Native Americans,

and vegetation more conducive to burning.
62 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

The frequency and extent of Indian burning decreased

substantially after white contact, which introduced new

diseases and decimated their population by 90 percent or

more over the next 100 to 150 years (Denevan 1992;

Dobyns 1983; MacCleery 1993). As a result grasslands,

savannas, and woodlands succeeded to closed forest

(Buckner 1983; Denevan 1992; Dobyns 1983; MacCleery

1993, 1995; Pyne 1997) (fig. 4-7). Subsequent settlement

of the oak-hickory forests by Euro-Americans, who used

fire for many of the same reasons as the Native Americans, increased the frequency and extent of burning

(Abrams 1992; Pyne 1997; Van Lear and Waldrop 1989).

Fire return intervals were shortened to 2 to 10 years with

many sites burning annually (Cutter and Guyette 1994;

Guyette and Day 1997; Holmes 1911; Sutherland and

others 1995; Sutherland 1997). For example, the barrens

of Pennsylvania and Maryland were burned annually at

least through 1731 (Tyndall 1992).

Presently, the fire regime of oak-hickory forests is

characterized by infrequent, low-intensity surface fires

that occur during the spring and fall. They are caused

almost exclusively by humans, and burn small areas

(Pyne and others 1996). Lightning is a minor ignition

source (Barden and Woods 1974; Ruffner and Abrams

1998). Fire return intervals have lengthened from a

few years to several millennia (Harmon 1982), the

longest fire-free intervals in the history of the Central

Hardwoods Region (Ladd 1991).

Fuels and Fire Behavior

Southern Pine Forests—Most studies show that

the amount of dead understory and forest floor material less than 3 inches in diameter varies widely by

site, overstory basal area, and time since last fire

(table 4-2). Live and dead accumulated fuel loadings in

slash pine and longleaf pine stands can increase five to

ten fold from 1 year after a fire to 20 years later. In

mature Southern pine forests on the Atlantic Coastal

Plain, stand average forest floor fuel loadings ranged

from about 1.5 tons/acre (3.4 t/ha) under an annual

dormant-season fire regime to 13 tons/acre (29.1 t/ha)

after 40 years (Boyer and Fahnestock 1966; Bruce

1951; Geiger 1967; McNab and others 1978; Southern

Forest Fire Laboratory Staff 1976; Williston 1965).

Live groundcover and understory fuels less than

1 inch in diameter ranged from about 0.75 tons/acre

(1.68 t/ha) with annual burns, to over 11 tons/acre (25

t/ha) after 25 years (table 4-3). Halls (1955) reported

that grass production varied from about 300 lb/acre

(336 kg/ha) under dense longleaf-slash pine stands to

1,000 lb/acre (1,120 kg/ha) on open forest rangelands.

Equations for predicting forest floor and understory

fuel weights from stand factors and age of rough were

developed by Williston (1965), Geiger (1967), and

McNab and others (1978) that explain 70, 80, and 86

percent of the variation in fuel weight, respectively.

Figure 4-7—Oak-dominated stand with no recent history of fire on the Cumberland Plateau, eastern

Tennessee. Photo by Tom Waldrop, 1982.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 63

Chapter 4: Fire in Eastern Ecosystems Wade

Fuel loading trends are similar on Piedmont sites for

shortleaf pine and mixed pine-hardwood (Metz 1954);

shortleaf pine (Crosby 1961; Johansen and others

1981); loblolly pine (Southern Forest Fire Laboratory

Staff 1976); loblolly, shortleaf, and Virginia pines

(Metz and others 1970); loblolly pine and groundcover

(Brender and Williams 1976); longleaf, loblolly, and

shortleaf pines, mixed pine-hardwood, and wiregrass

(Albrecht and Mattson 1977). Fuel loadings were

highest in loblolly and longleaf pine stands, and appreciably lighter in shortleaf and Virginia pine stands.

Fuel loadings in shortleaf and mixed shortleaf pinehardwood stands in the mountains of North Carolina

were substantially heavier than those on the Piedmont, partly because of a heavier understory component (Albrecht and Mattson 1977). Prediction equations that explain 80 percent of the variation in forest

floor loadings and 90 percent for the groundcover were

developed by Brender and Williams (1976).

Photo series publications are available showing fuels and estimated loadings for various Southern pine

types (Lynch and Horton 1983; Ottmar and Vihnanek,

in press; Sanders and Van Lear 1988; Scholl and

Waldrop 1999; Wade and others 1993).

Resource managers usually prescribe conditions that

limit fuel consumption to 1 to 3 tons/acre (2.2 to 6.7 t/ha)

Table 4-2—Total litter loading (tons/acre ovendry basis) under slash pine and loblolly pine stands by stand basal area and age of

rough (Southern Forest Fire Laboratory Staff 1976).

Basal area Age of rough (years)

(sq ft/acre) 1 2 3 4 5 7 10 15 20

Slash pinea

30 1.5 2.5 3.4 4.2 4.8 5.9 7.0 8.1 8.4

50 1.6 2.8 3.8 4.7 5.4 6.6 7.9 9.0 9.4

70 1.8 3.2 4.3 5.2 6.1 7.4 8.8 10.1 10.5

90 2.1 3.5 4.8 5.9 6.8 8.3 9.9 11.3 11.7

110 2.3 4.0 5.4 6.6 7.6 9.3 11.1 12.7 13.2

130 2.6 4.4 6.0 7.3 8.5 10.4 12.4 14.2 14.7

150 2.9 5.0 6.7 8.2 9.5 11.6 13.9 15.9 16.5

175 3.3 5.7 7.7 9.5 11.0 13.4 16.0 18.3 19.0

200 3.8 6.6 8.9 10.9 12.6 15.4 18.4 21.1 21.9

Loblolly pine

30 1.4 2.2 2.9 3.4 3.8 4.5 4.7 4.7 —

50 1.5 2.4 3.2 3.9 4.3 5.0 5.3 5.3 —

70 1.6 2.8 3.7 4.3 4.8 5.6 5.9 5.9 —

90 1.9 3.0 4.1 4.8 5.4 6.3 6.6 6.6 —

110 2.1 3.5 4.6 5.4 6.0 7.1 7.4 7.4 —

130 2.4 3.8 5.1 6.0 6.7 7.9 8.3 8.3 —

150 2.7 4.3 5.7 6.7 7.5 8.8 9.3 9.3 —

175 3.0 5.0 6.6 7.8 8.7 10.2 10.7 10.7 —

200 3.5 5.8 7.6 8.9 10.0 11.7 12.3 12.3 —

a

Applies to stands with and without understory vegetation ( McNabb and Edwards 1976).

Table 4-3—Understory vegetative loading (tons/acre dry weight) in the palmetto-gallberry type related to age of rough

and understory height (Southern Forest Fire Laboratory Staff 1976).

Understory Age of rough (years)

height (ft) 1 2 3 5 7 10 15 20

1 0.4 0.4 0.5 0.6 0.9 1.4 2.6a 4.2a

3 2.6 2.6 2.7 2.8 3.1 3.5 4.7 6.4

4 4.5a 4.5 4.6 4.7 5.0 5.5 6.6 8.3

5 7.0a 7.0a 7.0 7.2 7.4 7.9 9.1 10.8

6 10.0a 10.0a 10.0a 10.2 10.4 10.9 12.1 13.8

a

A situation not likely to be found in nature.
64 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

during passage of the flame front (fig. 4-8). Residual

combustion can more than double these values, especially under drought conditions, or 5 to 6 years after a

major disturbance when large down woody fuels have

become punky. Percent litter reduction and energy

release from backfires in palmetto-gallberry can be

predicted given the weight and moisture content of the

total litter layer (Hough 1968). Available fuel can be

estimated knowing total fuel loading and moisture

content (Hough 1978).

Annual and biennial prescribed headfires on the

Atlantic and Gulf Coastal Plains typically move through

the herbaceous groundcover with fireline intensities

less than 500 Btu/ft/s (144 kW/m). Dead woody stems

killed in the previous fire are often still standing and

can substantially contribute to fire behavior. Where

the canopy is open, midflame wind speeds often approach 4 to 5 mph. Short distance spotting up to 10 to

15 feet (3 to 4.6 m) is fairly common in herbaceous fuels

such as wiregrass, broomsedge, and switchcane, especially as relative humidity drops below 35 percent.

Rate of spread is always less than midflame windspeed

on level ground (ignoring spotting) and typically falls

between 200 and 1,500 feet/hour (60 to 460 m/hr) for

heading fires and less than 150 feet/hour (46 m/hr) in

backing fires. Short runs where headfire rate-of-spread

is more than doubled are not infrequent under gusty

winds. Headfire flame length can vary widely, from

less than 0.5 feet (15 cm) to over 10 feet (3 m) in 2 year

palmetto-gallberry roughs depending primarily upon

windspeed, fine-fuel moisture content, fuelbed depth

and porosity, and ignition pattern. Where the objective is to topkill hardwoods, backfires are often utilized because they concentrate released heat energy

near the ground (Lindenmuth and Byram 1948).

In uniform 4 year old palmetto-gallberry roughs,

headfire spread rates can exceed 0.5 mph (0.8 km/hr)

with flame lengths in excess of 20 feet (6 m) and

fireline intensities of 2,000 Btu/ft/s (578 kW/m) under

“good” burning conditions and Lavdas Dispersion Index (Lavdas 1986) values above 70. Firebrands consisting primarily of dead palmetto fronds are common

and often ignite spot fires 10 to 30 feet (3 to 9 m) ahead

of the fire front as relative humidities drop below 35

percent. More detailed descriptions of fuels and fire

behavior can be found elsewhere (Cheney and Gould

1997; Hough and Albini 1978; Johansen 1987; Wade

1995; Wade and Lunsford 1989; Wade and others

1993).

Oak-Hickory Forests—Leaf litter is the primary

fuel that sustains fire. Loading and thickness vary

depending on site, stand age, and time of year (Albrecht

and Mattson 1977; Blow 1955; Crosby and Loomis

1974; Kucera 1952; Loomis 1975; Metz 1954). Hardwood leaves tend to cup and hold water after a rain,

although the leaves of some species of oak tend to curl

Figure 4-8—Dormant-season prescribed fire backing through 12-year palmetto-gallberry rough in Collier

County, Florida. Photo by Dale Wade, 1977.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 65

Chapter 4: Fire in Eastern Ecosystems Wade

and dry quickly in comparison to other hardwoods,

allowing fire to run through oak litter when other

hardwood fuel types are too wet to burn. Fuel loadings

in like stands on comparable sites vary little longitudinally, but increase northward because decreasing

mean temperatures slow decomposition. Litter loading and depth are greatest immediately after leaf fall

in the autumn and decline until the following autumn.

Most stands have litter loadings ranging from 1 to 4

tons/acre (2.5 to 9.9 t/ha) and depths of 1 to 5 inches

(2.5 to 12.7 cm), depending on time of year. For example, the forest floor under a 150 year old hickory

stand (91 sq.feet basal area) on the Piedmont of South

Carolina averaged 3.1 tons/acre (7.0 t/ha) and 4.1 tons/

acre (9.2 t/ha) under a nearby 50 year old oak stand (62

sq.feet basal area) (Metz 1954). Annual litter production in fully stocked oak stands in Missouri averaged

2.1 tons/acre (4.7 t/ha) (Loomis 1975). The total forest

floor loading (exclusive of material larger than 0.5 inch

diameter) in the same stands averaged 8.3 tons/acre

(18.6 t/ha), but 6.4 tons/acre (14.3 t/ha) in a younger

stand (Crosby and Loomis 1974). In hardwood stands

of several Northeastern States, the L, F, and H layers

averaged 1.0, 3.3, and 5.9 tons/acre (2.2, 7.4, and 13.2

t/ha), respectively (Mader and others 1977). Generally

duff contributes little to fire spread but does influence

fire effects, especially during droughts.

Savannas typically contained grasses that were 3 to

6 feet (1 to 2 m) high according to early explorers

(Barden 1997; Buckner 1983; Denevan 1992). Based

on grass height, loadings may have ranged from 2 to 5

tons/acre (4 to 11 t/ha). After decades, fire exclusion

allows the mid-story forest canopy to close, and herbaceous fuels are shaded out and become relatively unimportant as a fire carrier.

Small woody fuels are often abundant in young

stands originating after a major disturbance. Woody

fuels are less abundant in mid-successional and mature stands and increase in old-growth stands due to

accumulation of large woody material. When present,

ericaceous shrubs such as mountain laurel and rhododendron can burn with extreme fire behavior resulting

in a mixed-severity or stand-replacement fire (Waldrop

and Brose 1999). Many of the firefighter fatalities in

hardwood forests have occurred because of the explosive nature of these fuels.

Postfire Plant Communities

Current and pre-Euro-American settlement vegetation types may have little in common because most of

the original vegetation was either logged or cleared for

agriculture prior to World War I. Many Eastern States

currently have more acres in forest than they did 150

years ago.

Longleaf pine

Vegetation Dynamics—The longleaf pine ecosystem is distinguished by open pine forests, woodlands,

and savannas. It is found on the Coastal Plain from

Virginia to Texas, and in the Piedmont and Appalachian Highland (both Blue Ridge and Valley) physiographic provinces of Alabama and Georgia (Boyer

1990a; Wahlenberg 1946) (fig. 4-9). Longleaf pine tolerates a wide range of sites from wet, boggy flatwoods

underlain with tight clays across xeric, deep-sand hills

to thin stony soils on southerly-facing mountain slopes.

Surface soils are typically acidic, tend to dry quickly

after precipitation (especially the Quartzipsamments),

and are characterized by a lack of organic matter and

low fertility (Landers and Wade 1994). Vegetation

that dominates these oligotrophic sites is resistant to

biological decomposition because of its high C to N

ratio but is amenable to thermal decomposition (Hon

Tak Mak 1989). Key plants exhibit pronounced fire

tolerance, are long lived, and are efficient at gathering

nutrients and water, which reinforces their dominance and restricts the rate and spatial scale of vegetation change including species turnover. A feedback

loop thus exists that includes climate, key plants, and

vast expanses that are quick drying and topographically susceptible to disturbance; these factors all interact to maintain a chronic fire regime. Recurrent fire is

Western limit of

saw palmetto

Western limit

of wiregrass

Northern limit of

saw palmetto

Northern limit

of wiregrass

Figure 4-9—The longleaf pine forest region (after Stout and

Marion 1993). The shaded area shows the distribution of

longleaf pine (Critchfield and Little 1966). A dotted line indicates

the landward extent of wiregrass and the western and northern

limits of its distribution. A dashed line indicates the northern limit

of saw palmetto.
66 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

crucial to perpetuation of longleaf pine ecosystems as

noted by Andrews (1917) and many others since.

These ecosystems persist and maintain their diversity

because of, rather than in spite of, constant disturbance and infertile soils (Christensen 1993a; Landers

and Wade 1994; Landers and others 1995; Wells and

Shunk 1931).

The flora and fauna that dominate this ecosystem

are those with well-developed adaptations to chronic

fire. This fire regime maintained a two-tiered structure comprising an open longleaf pine overstory and

an unusually diverse groundcover dominated by bunchgrasses (fig. 4-3). According to Peet and Allard (1993),

this is one of the most species-rich ecosystems found

outside the tropics. Overviews of the plant and animal

communities that form this ecosystem can be found in

Bridges and Orzell (1989), Harcombe and others (1993),

Myers and Ewel (1990), Platt and Rathbun (1993),

Skeen and others (1993), Stout and Marion (1993), and

Ware and others (1993).

Longleaf pine has numerous traits that make it

highly tolerant of fire (table 2-1). After an initial grass

stage it puts on a growth spurt (called bolting) that

quickly gets its terminal bud above potential flames.

Bark thickens rapidly during the first year of height

growth and protects stems from light surface fires.

The large buds have a high heat capacity and are

protected by a surrounding sheaf of long needles.

Fire results in small-scale temporary changes in

light, water, nutrients, and space along the elevationmoisture gradient. These sites are further modified by

frequent treefalls resulting from lightning, beetle activity, and faunal excavations primarily by the gopher

tortoise (Gopherus polyphemus) and pocket gopher

(Geomys spp.). Hermann (1993) described the importance of these microsites. Strong wind events and

hydrologic extremes introduce ephemeral resource

features over much broader areas. Thus a wide array

of habitats is continuously available to vegetative

change, which allows opportunistic plant species to

coexist with more permanent ones. Without frequent

fire, woody species overtop the herbs, strongly acidic

pine needles accumulate forming a duff layer, nutrient

dynamics change, and soil fertility increases, which all

favor extrinsic species at the expense of endemic

residents. About 191 taxa of vascular plants associated with the longleaf-bunchgrass system are threatened and endangered (Walker 1993), primarily because: (1) habitat has been paved over, (2) intensive

site preparation associated with row crops took place

prior to agricultural abandonment or when directly

converting to other pine species, (3) frequent fire has

been excluded, and (4) water tables are dropping. The

fauna also contains many endemics that are either

Federally or State listed as threatened or endangered.

Natural stands of longleaf pine may average from

just one or two canopy trees per acre on extremely dry,

deep sand ridges to over 150 per acre on good mesic

sites (Schwarz 1907). Glitzenstein and others (1995)

found that when burned on a 2 year cycle, longleaf pine

canopy tree damage is primarily related to variation in

fire intensity rather than month of application. Forest

openings resulting from death of overstory trees are

maintained by frequent fire until eventually colonized

by shade-intolerant longleaf pine whose seed require

a mineral soil seedbed. Germination takes place as

soon after seed fall as moisture becomes adequate. The

seedlings can withstand low-intensity fire the following growing season; fire is sometimes applied at this

time to destroy seedlings of competing species to increase the longleaf pine component. If the longleaf pines

are not yet fully established, they are often topkilled,

but sometimes resprout (Grace and Platt 1995). With

extended periods of fire exclusion, hardwoods dominate

forest openings (Gilliam and Platt 1999).

Longleaf pine seedlings undergo a grass stage lasting several years where no height growth occurs, but

a taproot develops and the rootcollar thickens. During

the grass stage, seedlings are often infected by a

fungal disease called brownspot needle blight (Scirrhia

acicola), which weakens the plants so they eventually

die. Fire controls this disease by consuming the infected needles and stimulating the seedlings to initiate height growth, perhaps due to the release of

nutrients (Christensen 1977).

The bunchgrass and longleaf pine fuel mixture can

burn within hours of a soaking rain so that an early

afternoon lightning strike in a dead snag often smolders in the accumulation of dead bark, litter, and

branches at its base resulting in a fire later that day or

the next day. Thus the longleaf pine type transforms

lightning strikes into fires, which then spread through

the highly flammable herbaceous groundcover and

needle litter throughout its ecosystem (Platt and others 1988b) and into neighboring ecosystems as well. It

is not surprising that lightning is the leading cause of

mortality of longleaf pine (Komarek 1968; Platt and

others 1988b). Kirkman and others (1998) characterized the ecotone between upland longleaf pine and

seasonally ponded wetlands. The hydric soil boundary

was consistently upslope from the vegetation boundary, which led them to conclude that the abrupt changes

in vegetation were likely related to fire periodicity.

Southern pines have been exploited since EuroAmerican colonization. During the first two decades of

the 20th century, the remainder of the “original”

Southern pine forest was heavily cutover, and then

indiscriminately burned at least every spring to promote forage for free-ranging cattle (Stoddard 1962).

These fires eliminated all pine regeneration except for

grass-stage longleaf pine. However, reforestation of

this species was prohibited by lack of an adequate

longleaf pine seed source (its seed is too heavy to be

effectively wind disseminated) and feral pigs that
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 67

Chapter 4: Fire in Eastern Ecosystems Wade

uprooted seedlings (Frost 1993). Fire exclusion was

seen as the only choice that would allow reforestation

of these lands. This policy allowed the aggressive, less

fire-tolerant loblolly and slash pines to move from the

more mesic sites where they had been confined under

the previous chronic fire regime (Landers and others

1995). The result was a dramatic shift in the makeup

of the emerging forest. Extensive type conversion by

the pulp and paper industry to genetically groomed

loblolly and slash pines with faster juvenile growth

than longleaf pine further exacerbated these species

composition shifts. The longleaf pine ecosystem, which

once dominated about 75 million acres (30 million ha)

(Chapman 1932), has been extirpated from over 95

percent of its historical area (Frost 1993; Outcalt and

Sheffield 1996) and is listed as critically endangered

by Noss and others (1995).

Associated Vegetation—Conifer associates of

longleaf pine include loblolly pine, slash pine, and

pond pine on wetter sites and shortleaf pine, Virginia

pine, and sand pine on drier sites depending upon the

fire return interval and fireline intensity, particularly

during the juvenile growth stage.The groundcover

comprises primarily bunchgrasses such as wiregrass

along the Atlantic seaboard (Lemon 1949; Lewis and

Hart 1972; Lewis and Harshbarger 1976) and little

bluestem and slender bluestem from central Alabama

westward (Grelen and Duvall 1966). Common woody

understory species include saw palmetto, gallberry,

and wax myrtle. The live foliage of all three shrubs is

unexpectedly flammable because of the release of

volatiles, which promote fire spread and higher fireline

intensities. Numerous studies describe the relationship between fire and various longleaf pine-dominated

plant assemblages (Harcombe and others 1993;

Huffman and Werner 2000; Landers 1991; Lewis and

Hart 1972; Peet and Allard 1993; Wells and Shunk

1931; Yahr and others 2000), between fire and indigenous fauna (Brennan and others 1998; Engstrom

1993; Folkerts and others 1993; Guyer and Bailey

1993; Jackson 1989), and between both flora and

fauna (Stout and Marion 1993; Ware and others 1993;

Weigl and others 1989).

Under natural fire regimes, dry-site oaks are generally the only hardwood associates to reach the midstory.

Common species include turkey oak, bluejack oak,

blackjack oak, and sand post oak. Based on current

information, Rebertus and others (1993) hypothesized

that both longleaf pine and these upland oaks are fire

dependent and that longleaf pine becomes more important relative to the oaks as fires become more

frequent and occur during the early growing season.

Greenberg and Simons (1999) presented evidence showing oaks have been an integral midstory component of

many high pine sites (Myers and Ewel 1990) in Florida

for at least several centuries.

Disrupted fire regimes result in an invasion of other

hardwoods such as sweetgum, oaks, hickories, common persimmon, and southern magnolia (Daubenmire

1990; Gilliam and Platt 1999) (fig. 4-10). These

Figure 4-10—Hardwoods are confined to the understory with annual dormant-season fire (on

left) and form a midstory with fire exclusion (on right), Okefenokee National Wildlife Refuge.

Photo by Ron Phernetton, 1992.
68 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

hardwoods form a midstory and prevent the shadeintolerant longleaf pine from reestablishing. Many of

these hardwoods are somewhat resistant to low-intensity fires when mature (Blaisdell and others 1974).

Understory trees are usually top-killed, but their

rootstocks are able to withstand all but annual growing-season fires (Waldrop and others 1987). Where

oaks form a midstory because of fire exclusion,

Glitzenstein and others (1995) found that late spring

and early summer burns increased oak topkill. Where

variations in microsite result in patchy growing-season burns, hardwoods occasionally reach the overstory, but not in sufficient numbers to assert dominance. For example, Plummer (1975) determined from

old records that the two most common hardwood

associates of the 18th century longleaf pine forest in

southwestern Georgia were oak and hickory, which

comprised 1.0 and 0.5 percent of the stems, respectively, and longleaf pine 91 percent. The influx of

exotics that are promoted by fire such as cogongrass

(Lippincott 1997), Japanese climbing fern, and

melaleuca (Wade 1981; Wade and others 1980) will be

an increasingly serious problem.

Management Considerations—Fragmentation of

the original landscape by conversion of sites to nonforest

uses and an extensive road network, coupled with

rapid, initial attack response times have dramatically

changed fire patterns in this ecosystem. Lightning

fires driven by shifting winds historically molded

habitats that were much more structurally diverse

than those managed today with rigidly scheduled linefires set under predictable dormant-season weather

conditions. The increasing use of variable growingseason fire regimes (currently about 15 to 25 percent

of the total acreage), point-source ignitions (both aerial

and ground), and “soft” firelines should help ameliorate this problem over time. Burn schedules that

incorporate variability into season, frequency, and

pattern of burn can be found in Robbins and Myers

(1992). Streng and others (1993) and Olson and Platt

(1995) suggested that many fires during the same

season over time are necessary to produce changes in

groundcover species composition.

Whenever the short fire-return interval is disrupted,

pines that are more prolific seeders or have faster

juvenile growth, outcompete the shade-intolerant

longleaf pine. Burning the site while these competitors are still seedlings or saplings will selectively favor

longleaf pine. However, once any of the multinodal

Southern pines attain a basal diameter of about 2

inches (in 3 to 4 years), they become fairly immune to

girdling by low-intensity fire (Wade 1993; Wade and

Johansen 1987). Assuming no bud damage, longleaf,

slash, and loblolly pines all usually survive complete

crown scorch except during late summer and early fall

when death often results, regardless of bud damage

(Storey and Merkel 1960; Weise and others 1990). Any

fire that kills buds during the growing season (crown

consumption is a good visual indicator) greatly diminishes a tree’s chances of survival (Wade 1985).

Desired fire-return intervals and timing of fire depend upon the objective. Early on, fire was advocated

to facilitate regeneration (Long 1889), reduce hazardous fuel accumulations (Mattoon 1915), eliminate invading hardwoods (Ashe 1910), control brownspot

(Siggers 1934), and manage rangelands (Wahlenberg

and others 1939). Quail plantations along the FloridaGeorgia border have burned annually to favor this

game bird for close to100 years (Stoddard 1931). Showy

plants such as orchids and pitcherplants (fig. 4-11)

that frequent acidic depressions can be maintained by

conducting annual late-spring burns (Komarek 1982).

Long-term repeated prescribed fire studies include

the 18 year effects of fire to control hardwoods in

southern Alabama (Boyer 1993), understory plant

community changes in a longleaf pine-wiregrass ecosystem in southern Georgia resulting from several

decades of dormant-season burns at 1, 2, and 3 year

intervals (Brockway and Lewis 1997), and understory

plant response to 30 years of prescribed burning at

various frequencies and seasons on the Santee Experimental Forest, South Carolina (Langdon 1981).

The effect of recurrent fire on Southern pine growth

is important when considering fiber production (Chambers and others 1986). Studies have shown reduced

growth after prescription fire (Boyer 1987; Boyer and

Miller 1994; Zahner 1989) as well as increased growth

after damaging fire (Johansen 1975). Grelen (1978)

reported no height growth effects after introducing

fire in 5 year old slash pine plantations and then

burning at various frequencies through 9 years of age.

Care must be exercised when interpreting the literature pertaining to the effects of fire on growth. Some

fires were conducted at inappropriate times, and the

methodology used to analyze results of some studies

was flawed (Streng and others 1993; Wade and

Johansen 1986). High fireline intensities that consume foliage (fig. 4-12) and high-severity fires that

cause root damage (by burning at low duff moistures)

most likely will cause mortality and reduce growth of

survivors (Wade 1985). Ill-timed growing season burns

and firing techniques will kill mature longleaf pine

(Boyer 1990b). A guide recommended for firing is to

consider a maximum air temperature of 99 °F, a

minimum relative humidity of 34 percent, and flank

firing technique. Reducing hazardous buildup of the

forest floor by alternative methods such as commercial

pine-straw raking (baled and sold on the retail landscape market as an ornamental mulch) will also reduce longleaf pine growth the following year (McLeod

and others 1979), because scarce nutrients are removed instead of recycled. The literature pertaining to
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 69

Chapter 4: Fire in Eastern Ecosystems Wade

Figure 4-11—Trumpet pitcherplant in a savanna burned the previous year on the Apalachicola National Forest,

Florida. Photo by Sharon Hermann.

Figure 4-12—Percent survival of 1 to 8 year old planted loblolly

pine after the first postfire year relates to percentage of live

foliage consumption (Wade 1985).

the effects of fire on Southern pine was summarized

and critiqued by Wade and Johansen (1986).

Chronic burning regimes result in an open parklike

forest with a groundcover dominated by grasses and

forbs (fig. 4-13). Not only are such plant communities

aesthetically pleasing, but they also provide excellent

wildlife habitat and high quality forage for cattle and

deer. Many such landholdings are profitably managed

for the simultaneous production of cattle, wildlife, and

timber resources (Wade and Lewis 1987; Wade and

Moss 1999). Detailed restoration and management

options for this species across its range are presented

in Brockway and others (1998), Farrar (1990), Hermann

(1993), and Landers and others (1990, 1995); by accessing the Longleaf Alliance through the home page

for the Auburn University School of Forestry; or by

contacting the Tall Timbers Research Station in Tallahassee, FL, or the Joseph Jones Ecological Research

Center in Newton, GA.

Slash Pine

Vegetation Dynamics—Slash pine (var. elliottii)

is the chief conifer associate of longleaf pine on wet

Coastal Plain sites throughout its natural range, which

extends from coastal South Carolina through southern Georgia, the flatwoods of Florida, and across the

Gulf Coastal Plain into Mississippi and Louisiana
70 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

(Lohrey and Kossuth 1990). It is a prolific seeder, has

rapid juvenile growth, and becomes increasingly tolerant of fire with age. Because of these traits and the

ease of handling seedlings, slash pine has been extensively planted. A 1980 survey showed that almost half

of the 13 million acres (5.2 million ha) of slash pine

resource had been planted (Haeussler 1983). Slash

pine is perhaps the only native North American tree

whose commercial range has undergone a successful

large-scale expansion, through planting northward

into Tennessee and westward into Texas where it is

now naturalized (Stone 1983). Sheffield and others

(1983) estimated slash pine occupied over 40 percent

of the commercial forestland in Florida in 1980. A

variety of slash pine (densa), which goes through an

abbreviated grass stage somewhat like longleaf pine

and is almost as fire tolerant, is confined to southern

Florida and the Keys where it replaces longleaf pine

(see chapter 7). General descriptions of the slash pine

ecosystem can be found in Lohrey and Kossuth (1990),

Grelen (1980), Myers and Ewel (1990), Stone (1983),

and Stout and Marion (1993).

The most hydric slash pine sites are depressions

such as bays, bayheads, titi swamps, and cypress pond

margins embedded within the flatwoods matrix. Common associates are pond pine, loblolly pine, sweet bay,

loblolly-bay, swamp bay, pondcypress and swamp

tupelo, and bottomland hardwoods such as red maple,

Carolina ash, American elm, sweetgum, and water

oak. On such sites slash pine generally develops a

pronounced buttress (comprised mostly of bark) that

protects the tree from heat-girdling during drought

fires. Species diversity is low. Characteristic understory species include cyrillas, sweetpepperbush,

lyonias, and buckwheat tree. Little groundcover is

present except for sphagnum, ferns, and various

greenbriers.

Slash pine was historically confined to wet sites

because its susceptibility to fire when young kept it

from successfully competing with longleaf pine on

upland sites. With removal of the longleaf pine and

efforts to exclude fire, slash pine has successfully

invaded many of these drier sites. The suspected

historical successional scenario is that most fires favored slash pine. The occasional severe drought fire

topkilled the woody vegetation (although a few slash

pine survived), slash pine reestablished when a seed

source was nearby, hardwood rootstocks resprouted

(except in deep-burning peat fires), and the cycle

repeated over the next several decades (Clewell 1980;

Hodges 1980). Without fire the slash pine component

eventually disappeared because a good seedbed was

lacking and existing woody vegetation shaded out

established slash pine seedlings.

Boggy flatwoods border creek swamps and acidic

depressions. Fire enters these ecosystems during

Figure 4-13—Longleaf pine stand maintained by biennial growing-season fires near Thomasville, Georgia.

Lack of longleaf pine reproduction is an artifact of decades of annual dormant-season burns that ended in

1986. Photo by Ken Outcalt, 1993.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 71

Chapter 4: Fire in Eastern Ecosystems Wade

extended dry periods, which occur every couple of

decades. Longleaf pine is the primary overstory associate with an understory of fire-adapted shrubs such

as wax myrtle, gallberry, buckwheat tree, dahoon, and

yaupon, and a ground cover of pitcherplants. These

sites grade into slash pine flatwoods as the hydroperiod

(the length of time the water table is above the soil

surface) decreases with a several-inch rise in elevation

and concomitant increase in fire frequency; with another several-inch rise, sites grade into pure longleaf

pine stands where the fire regime is chronic. The more

mesic sites support a rank understory of flammable

vegetation, which produces some of the highest fuel

loadings encountered in the South (table 4-2). Fuel

consumption during extended dry-period wildfires can

exceed 15 tons/acre (37 t/ha) after a decade or so of fire

exclusion in fully stocked natural stands. Even highintensity prescription fires do not approach this figure

because burn plans almost invariably call for Keetch

Byram Drought Index (Keetch and Byram 1968) values below 500 to avoid killing the overstory pine feeder

roots that colonize the developing humus layer.

Extensive planting on old-field sites and decreased

fire frequency have led to development of many other

understory associations including saw palmetto, dwarf

huckleberry, ground blueberry, and grasses such as

Curtis’ dropseed, broomsedge, and chalky bluestem

and the creeping variety of little bluestem. Planting on

xeric, nutrient-poor deep sands (ancient sand dunes)

has resulted in associations with sand pine and numerous oaks such as post, blackjack, sand live, myrtle,

bluejack, and turkey (Lohrey and Kossuth 1990). Planting of slash pine outside its natural range has resulted

in many additional plant associations, virtually all of

which are also fire-adapted.

Management Considerations—Harvesting of

longleaf pine and fire exclusion policies in the first half

of the 20th century, followed by several decades of

extensive planting of slash pine coupled with dormant-season backing fires at wider than needed time

intervals, resulted in a significant invasion of slash

pine and its accompanying rank understory into

former mesic longleaf pine sites. Most landowners

were reluctant to use head fires, or in many cases

any fire, because of high fuel loadings. Low-intensity backing fires rarely penetrated into the more

mesic flatwoods because of high water conditions

during the winter from frequent rainfall and reduced

evapotranspiration.

During the 1980s, fire management of this pine

resource changed dramatically due to the advent of

aerial ignition, use of growing-season prescribed fires,

recognition that the aesthetically pleasing longleaf

pine ecosystems had virtually disappeared, and an

increased appreciation of ecosystem management. The

objectives included an increase in fire frequency to

reduce the potential for catastrophic fire and the

desire to directly or indirectly promote longleaf pine

and its associated biologically diverse groundcover.

On many lands, plowed firelines are no longer the

norm, the fire-free interval has been shortened and

high-intensity headfires are purposely run into the

encroaching walls of underbrush to push the ecotone

back toward historic boundaries (Ferguson 1998).

These ecotones are not static, and as water tables are

permanently lowered, they move further downslope.

The relationship between fire and major pests of

slash pine has received limited study. However, studies suggest that use of prescribed fire can reduce the

severity of annosus root rot (Froelich and others 1978)

and fusiform rust (Siggers 1949; Wade and Wilhite

1981), two serious diseases of slash pine. Southern

pine beetle infestations are not influenced by the

judicious use of fire, but fires that result in severe

crown scorch or root damage are often a precursor to

attack.

Fire management in slash pine stands is straightforward. Fire has to be withheld for 3 to 5 years until

sapling girth at ground level exceeds 2 inches (Johansen

and Wade 1987a). After 8 or 9 years, slash pine usually

recovers from complete crown scorch during all seasons except early fall (Weise and others 1990); however, its growth will be severely impacted (Johansen

and Wade 1987b). Slash pine will not recover if buds

are killed by complete consumption of foliage. Eyelevel winds within stands are necessary to give the

flame front direction and dissipate heat, but the controlling variable for effective use of fire is forest floor

moisture content; if it is too high, fires will not back,

whereas if too low, significant damage occurs to feeder

roots. The water table level as measured by the KeetchByram Drought Index (KBDI) should be considered in

planning prescribed fires because during severe

droughts (KBDI above about 600), fires will burn

through boggy flats and bayheads consuming underlying organic soils and destroying hardwood rootstocks

as well as those of overstory trees. General guides for

using fire in slash pine can be found in de Ronde and

others (1990) and Wade (1983).

Loblolly Pine

Vegetation Dynamics—Loblolly pine dominates

about 30 million acres (12 million ha) from New Jersey

south to Florida and then west to Texas with excursions into Oklahoma, Arkansas, and Tennessee interrupted only by the flood plain of the Mississippi River

(Baker and Langdon 1990). Loblolly pine historically

was confined to much of the same wet landscape as

slash pine for the same reason—its susceptibility to

fire when young. It is also common along stream

bottoms in the Piedmont where fire-free intervals

historically exceeded 5 to 6 years. With increased fire
72 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

protection, acreage dominated by this species dramatically increased as it seeded into former longleaf

pine sites and abandoned agricultural lands. Loblolly

pine was planted even more extensively than slash

pine. It is currently the leading commercial tree in the

Southern United States, comprising more than half of

the standing pine volume in the region (Baker and

Langdon 1990).

Loblolly pine is currently found on many of the same

sites as slash pine except deep sands, but because of its

greater range it occurs in association with many more

species, especially where fire frequency has been reduced. Wahlenberg (1960) lists over 60 principal associates. Descriptions of various ecosystems where

loblolly is a major component can also be found in

Baker and Langdon (1990), Crow (1980), and Skeen

and others (1993). On drier sites, longleaf, shortleaf,

and Virginia pines; southern red, white, post, and

blackjack oaks; mockernut hickory; and common persimmon are all common overstory associates. At the

other end of the elevation-moisture gradient, associates include slash pine and pond pine; water, willow,

southern red, swamp chestnut, and laurel oaks; southern magnolia; swamp tupelo; American elm; and red

maple.

Because loblolly pine favors sites where soil moisture is not limiting, plant community composition

usually includes a dense species-rich understory. Common understory hardwoods and vines usually resprout

after fire. Herbaceous groundcover is typically sparse

and includes bluestems, panicums, and longleaf uniola.

Oosting (1944) compared differences in vegetation

between an unburned area and two levels of fire

intensity 9 years after a wildfire in a 35 year old stand

of old-field loblolly pine. Species composition was about

the same on all three areas. Although the crown fire

killed virtually all overstory pines, many pine seedlings established and were successfully competing

with hardwood sprouts suggesting that the future

stand would again be mixed pine-hardwood. Twenty

years of prescribed burning on the Georgia Piedmont

at a 4 to 5 year cycle showed that the understory was

drastically reduced and wildlife habitat improved

(Wade and others 1989). Species composition in comparison to adjacent unburned stands was not altered.

The same trends were noted on the Coastal Plain of

South Carolina except where annual growing-season

fires extirpated some hardwood species (Langdon 1981;

Waldrop and others 1987).

Management Considerations—The key to managing loblolly pine is control of understory hardwoods,

which is usually accomplished with periodic fire

(fig. 4-14), mechanical methods, and chemicals

(Chaiken and LeGrande 1949; Chapman 1947; Grano

1970; Harshbarger and Lewis 1976; Trousdell 1970).

Where precipitation is a limiting factor, especially

toward the western edge of its range, hardwood control can increase the growth of overstory pine (Bower

and Ferguson 1968; Grano 1970). Loblolly pine is

easily killed by fire when young, although seedlings

will sometimes resprout at the root collar. On all but

the poorest sites, juvenile height growth is several feet

per year and its bark rapidly thickens, becoming

resistant to light surface fires by the time stem diameter at ground level reaches 2 inches (5 cm) (Wade

1993).

Operational underburning in loblolly pine is usually

not advocated until stands reach about 10 years of age,

but the sooner prescribed fire is used, the easier it is to

topkill competing hardwoods. If care is taken in selecting burning conditions and in executing the burn to

avoid crown scorch, plantations can be safely burned

at half this age. Fire has been advocated as an integral

part of the conversion from low-grade hardwood stands

that developed with fire exclusion, back to loblolly pine

or shortleaf pine. Planted loblolly exhibited increased

growth on sites where hardwoods were cleared and

burned (Applequist 1960). Numerous authors including Brender and Cooper (1968) and Cooper (1973b)

described the use of fire to prepare mineral soil seedbeds to facilitate natural regeneration of loblolly pine.

Figure 4-14—Growing-season prescribed fires topkilled more

hardwoods and larger ones than dormant-season fires. Data

collected on Brender Demonstration Forest, Georgia.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 73

Chapter 4: Fire in Eastern Ecosystems Wade

Low-intensity, low-severity fires can be used to prepare seedbeds on steep slopes without triggering erosion problems (Douglass and Van Lear 1982); Van

Lear and others 1985). The benefits to wildlife from

repeated prescribed fires in loblolly pine stands are

described well by Cushwa and Redd (1966) and Cushwa

and others (1966). Guides to the management of this

species include Brender (1973), Schultz (1997), and

Wahlenberg (1960).

Shortleaf Pine

Vegetation Dynamics—The range of this species

extends from southern New York southwest across

southern Pennsylvania, Ohio, Indiana, Illinois, and

Missouri, then south through eastern Oklahoma and

Texas, and all States to the east including northern

Florida (Mattoon 1915). It has the widest range of any

of the eastern pines, occurring in 22 States. It occupies

a wide variety of soils and environmental conditions

but does not tolerate poorly drained sites. Shortleaf

pine does well only on mineral soil seedbeds. It is a

prolific seeder, often forming dense sapling stands

that are favored over competing hardwoods by frequent fire.

Shortleaf pine can sprout repeatedly from its base if

the tree is topkilled (see chapter 2). Trees 30 years old

can produce basal sprouts (Little and Somes 1956);

however, Matoon (1915) and Wakeley (1954) reported

that this species looses its ability to sprout when less

than half that age. Many sprouts can arise (70 or

more), but virtually all die once a central leader

assumes dominance (Komarek 1982). In a 1912 survey

of shortleaf pine in Arkansas, few stands were found

that originated from seedlings because of pervasive

fire. There were more than three age classes of coppice

in many stands (Matoon 1915). Ability to resprout,

abundant seed crops, rapid juvenile growth (especially

of sprouts), and a low resin content of the wood make

this species markedly tolerant of fire (Mattoon 1915).

Trees larger than 0.5 inch (1.3 cm) dbh are somewhat

resistant to fire; mortality is negligible once trees

reach 4 inches (10 cm) dbh (Walker and Wiant 1966).

Like other Southern pines, trees over 5 feet (1.5 m) tall

rarely die when crown scorch is less than 70 percent

and buds are not killed by consumption of foliage.

Loblolly pine is the chief associate of shortleaf pine

at lower elevations throughout the Mid-South and

Southeast. Loblolly pine dominates the heavier, moist

soils while shortleaf pine dominates the lighter, drier

soils. Loblolly drops out at about 400 feet (122 m)

elevation in the Ozarks and Ouachitas, resulting in

pure stands of shortleaf pine up to about 2,000 feet

(610 m) on south-facing slopes, above which hardwoods begin to dominate with shortleaf pine disappearing at about 3,000 feet (914 m). In the Appalachians and Upper Piedmont, Virginia pine replaces

shortleaf pine on drier nutrient-poor sites east of the

Appalachian divide. On the New Jersey Coastal Plain,

pitch pine is the chief associate of shortleaf pine.

Common hardwood associates are oaks (red, black,

post, and chestnut), hickories (pignut and mockernut),

sweetgum, yellow poplar, red maple, common persimmon, flowering dogwood, and sassafras.

Management Considerations—Mineral soil seedbeds are preferred but not required for seed germination (Boggs and Wittwer 1993; Ferguson 1958; Loyd

and others 1978). Burns that expose mineral soil

seedbeds may compact the surface soil as observed in

the Arkansas mountains (Bower and Smith 1962).

Care should be taken when using fire in hilly terrain

because of the potential for erosion, especially on drier

sites in the western part of its range (Moehring and

others 1966). When low-severity fires are used to avoid

consuming the humus layer, soil movement is negligible, even on steep Appalachian sites (Van Lear and

Danielovich 1988). Shortleaf pine competes poorly

with other plants (Williston and Balmer 1980). Periodic low-intensity fires, herbicides, or both are necessary to control the relentless encroachment of hardwoods and improve growth of the pines (Bower and

Ferguson 1968; Crow and Shilling 1980; Ferguson

1957; Grano 1970; Hodgkins and Whipple 1963; Little

and Moore 1950; Rogers and Brinkman 1965; Yocom

1972). Understory hardwoods deplete soil water in

pine stands (Zahner 1958).

Somes and Moorhead (1950) showed that prescribed

fire does not reduce the yield from oak-pine stands in

New Jersey. Additional evidence showed that timber

harvest and prescribed fire are not nutrient depleting

practices and can enhance soil nutrient levels (Masters and others 1993). Phillips and Abercrombie (1987)

demonstrated that excellent stands of shortleaf pine

and mixed hardwoods could be produced by the fell

and burn technique. Guides for managing shortleaf

pine include Chen and others (1975, 1977), Nickels

and others (1981), and Walker and Wiant (1966).

Oak-Hickory Forests

Pre-1900 Succession—Frequent fires ignited by

Native Americans maintained open oak-hickory forests with a groundcover of grasses and forbs (fig. 4-15).

Oaks and hickories were favored because of their thick

bark. These species dominated the canopy as old,

large, fire-resistant trees. Densities of dominant trees

probably varied from 20 to 40 per acre. Shrubs, understory trees, and woody debris were rare (Barden 1997;

Buckner 1983; Denevan 1992; Pyne 1997). Hardwood

regeneration comprised seedling sprouts dominated

by oak and hickory because these species initially

emphasize root development over stem growth and

have the ability to sprout repeatedly (Barnes and Van
74 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

Lear 1998; Brown 1960; Van Lear 1991). With fire

excluded for a few years, the well-developed rootstocks

sent up vigorous stems that often developed sufficient

size and bark thickness to withstand future fires.

Where windstorms blew down trees over large areas,

the replacement stand was even aged. Consequently,

the forest was uneven aged, consisting of even aged

patches.

Post-1900 Succession—The exclusion of fire from

fire-dependent oak ecosystems should be considered a

catastrophic disturbance according to Packard (1993).

Reduction of fire has profoundly changed the oakhickory forest by allowing the forest to succeed to

mixed mesophytic and northern hardwood species

such as red maple, eastern white pine, sugar maple,

and beech (fig. 4-7). In the absence of fire, these species

become established in the understory, grow into the

midstory, and eventually change the composition of

the canopy. Stem densities are often hundreds per

acre. During the growing season, the dense shade from

these fire-sensitive species reduces the abundance

and richness of forbs and grasses and inhibits development of oak and hickory regeneration. Consequently,

when a dominant oak or hickory dies, its reproduction

is not capable of sufficient growth to capture the

canopy opening. Instead, the growing space is filled by

mesophytic and northern hardwood species (Abrams

and Downs 1990; Crow 1988; Lorimer 1985; McGee

1984). According to Olson (1996), the brushy character

of many sites is the result of an interruption in the

chronic fire regime that allows shrubs and hardwoods

to capture the site. When the area again burns several

years later, these stems are top killed producing a

dense growth of sprouts that can dominate the site for

decades, especially with occasional fire.

On drier mountainous sites, fire exclusion allows

ericaceous shrubs such as mountain laurel and rhododendron to move from riparian areas into upland

forests (Elliott and others 1999). These shrubs are

shade tolerant and evergreen, shading the forest floor

throughout the year. Hardwoods cannot regenerate

beneath them (Baker and Van Lear 1998), and without disturbance, these heath thickets are the climax

plant community on some sites. Although the forest

floor rarely dries enough to support surface fire, the

ericaceous shrub layer is flammable; and when it

burns, it typically supports intense, stand-replacement fires that alter successional pathways, reduce

site productivity, negatively impact involved streams,

and threaten human life and property (such slopes are

favored building sites). Altered fire cycles have also

impacted the “low elevation rocky summit” vegetation

type where fire historically maintained the hardwood

scrub savanna (Hallisey and Wood 1976). Fire exclusion over the past 50 years resulted in an increased

hardwood overstory and a dramatic decline in herbs

Figure 4-15—Oak-dominated stand after 10 annual spring burns on the Cumberland Plateau, eastern

Tennessee. Photo by Ivan Thor, 1976.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 75

Chapter 4: Fire in Eastern Ecosystems Wade

such as blazing star and some woody scrub species

such as bear oak (Barden 2000).

Management Considerations—Until recently,

foresters failed to appreciate the role of fire in maintaining open oak-hickory forests and in facilitating

regeneration of these species (Lorimer 1993). Regeneration of oak was attempted only with timber harvesting and herbicides, which generally hastened the

successional replacement of oaks by mixed mesophytic

species (Abrams and Scott 1989). Research indicates

that fire can be used in hardwood stands to establish

and release oak-hickory regeneration (Barnes and

Van Lear 1998; Brose and Van Lear 1998, 1999;

Christianson 1969).

Understory burning of mature, uncut hardwood

stands can help establish oak and hickory regeneration by preparing a seedbed (Barnes and Van Lear

1998). Acorns and hickory nuts are often buried by

wildlife, particularly squirrels and blue jays, which

prefer burned areas because of the thin root mat. The

fires also top-kill or eliminate many of the shrubs and

small trees that shade the forest floor. In a less shaded

environment, the acorns and hickory nuts germinate

and the new seedlings begin developing their root

systems. Eventually, the regeneration replaces canopy

trees. In this approach, fires are initially applied at a

frequent interval (annual or biennial) depending on

season of burn and severity of the shade. Once oak

seedlings are established, fire is withheld for a few

years (Cottam 1949), and then periodically reapplied

once or twice a decade. This minimizes mortality of the

oak regeneration by allowing time for root systems to

develop. This approach may take 15 to 20 years for

results to be apparent.

If oak and hickory regeneration is present in the

understory, a two-step shelterwood harvest combined

with a prescribed fire can be used for release (Brose

and others 1999a,b). The initial shelterwood cut reduces the basal area to about 50 sq. feet/acre (11.5 sq.

m/ha) removing low-value stems. The regeneration is

allowed to develop for 3 to 5 years. During this time,

oak and hickory regeneration develop large root systems but exhibit little height growth while their competitors do the opposite. When the root collar diameter

of the oak regeneration is about 0.75 inch (2 cm), a

growing-season prescribed fire with flame lengths of 3

to 4 feet (about 1 m) is used to kill the regeneration

layer. This treatment will completely kill the less firetolerant competing hardwoods (Christianson 1969),

invading eastern white pine (Blankenship and Arthur

1999), and rhododendron and mountain laurel. Few

oaks and hickories will be killed by the fire, and most

will sprout and grow vigorously. Regeneration should

be inventoried 2 to 3 years later to determine whether

additional fires are needed.

When using the shelterwood-burn technique, care

must be taken to protect dominant oaks from basal fire

damage. Directional felling during the logging operation is recommended so that the resultant slash is not

abutting the trees. Otherwise, slash must be removed

from the bases of dominant oaks to prevent fire damage. Generally, damage to dominant oaks is not a

problem when burning in uncut stands because fuel

loadings are considerably lighter. Graphs or equations

can be used to predict mortality of several oak species

after fires of varying intensity (Loomis 1973). The

shelterwood-burn technique appears to be a reasonable mimic to the disturbance regime of oak-hickory

forests before Euro-American influence. It has considerable value as a silvicultural method, a wildlife management tool, and a means for restoring habitats such

as oak savannas and open woodlands.

Mixed Fire Regimes _____________

Major Vegetation Types

The mixed fire regime best represents the presettlement fire history for several hardwood and conifer

dominated ecosystems. The conifers include pitch pine

and Virginia pine of Kuchler’s oak-pine association

and pond pine, a dominant tree of the pocosin association (table 4-1). The conifer types fit the mixed fire

regime because fire intensities are generally greater

than in the understory fire regime and cause mortality

ranging from 20 to 80 percent of the overstory. The

hardwood ecosystems comprise mesophytic hardwoods,

Northern hardwoods, and elm-ash-cottonwood ecosystems (table 4-1). Although the hardwoods are prone

to fire injury, many survive numerous fires before

eventually being girdled. These fires tend to be lowintensity due to less flammable fuels than found in

ecosystems having a substantial conifer component.

We believe that the low-intensity presettlement fires

that wounded and killed many trees did not cause

enough mortality (>80 percent according to our criteria) to be considered stand-replacement.

Pines

Pitch Pine—Pitch pine grows on poor, generally

sandy, gravelly, and shallow soils, primarily south of

the glaciated region in southern New England in a

fairly wide swath following the Appalachians and

Upper Piedmont into Georgia where it occurs below

3,000 feet (914 m) elevation (Little 1959; Little and

Garrett 1990). In New Jersey, pitch pine commonly

exists in two forms, as a tree interspersed with hardwood trees (Pine Barrens) or as a member of a scrub

oak community (Pine Plains).
76 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

Virginia Pine—The natural range of this species

stretches from New Jersey across southern Pennsylvania to Indiana, then southward into central Alabama, and then northeasterly up the eastern slope of

the Appalachians with more than half the standing

inventory in western Maryland, Virginia, and North

Carolina (Sternitzke and Nelson 1970). Virginia pine

characteristically occupies poor sites where it often

forms pure stands. Common associates include shortleaf, loblolly, and pitch pines; eastern redcedar; numerous oaks (SAF cover type Virginia pine-southern

red oak); and other hardwoods.

Pond Pine—This forest cover type (SAF 98)

stretches along the Coastal Plain from New Jersey to

Alabama dominating poorly drained sites characterized by organic soils such as pocosins, bays, and shrub

bogs where it often forms pure stands (Wenger 1958).

About 80 percent of the pond pine forest is in the

Carolinas (Sternitzke and Nelson 1970). Pond pine

communities are often referred to by the understory

vegetation such as shrub bogs or pocosins. The Native

American name pocosin means swamp on a hill and

they are just that. They occur on divides between

rivers and sounds but are not alluvial. They all have

long hydroperiods, burn periodically and are underlain by sandy humus or organic peat or muck soils

(Richardson and Gibbons 1993). Pocosins, shrub bogs,

and Carolina bays are often found within the loblolly

pine, slash pine, slash pine-hardwood, and sweetbayswamp tupelo-redbay cover types. On well-drained

sites, pond pine is usually a minor component.

Hardwoods

Mixed Mesophytic Hardwoods—These forests

occupy the transition zone from the oak-hickory forest

to the northern hardwood forest. They are among the

most diverse in the United States containing more

than 30 canopy tree species. This type lies west of the

Appalachians and transitions from the more northern

sugar maple-beech-birch forest in northern West Virginia, southwestern Pennsylvania, and southern Ohio

southward down the Allegheny Mountains, across the

Allegheny Plateau including all of the Cumberland

Plateau, and into northern Alabama where it transitions to the oak-hickory-pine type of the Southern

Mixed Hardwood Forest. Common overstory species

include sugar maple, red maple, basswood, northern

red oak, chestnut oak, white oak, yellow poplar, American ash, silverbell, yellow birch, southern magnolia,

Blackgum, black walnut, beech, yellow buckeye, and

butternut.

Northern Hardwoods—The maple-beech-birch

FRES ecosystem type, commonly known as Northern

hardwoods, occurs on mesic and fire protected sites in

the Lake States, Northeast, and Southeastern Canada

(fig. 4-16). The dominant hardwood species include

sugar maple, yellow birch, beech, and basswood in the

Midwest. Northern hardwoods mix with boreal spruce

and fir to the Northeast and with eastern hemlock,

eastern white pine and northern oaks to the south and

west. Component species, especially beech and sugar

maple, extend south at mid elevations in the Appalachian Mountains to western Virginia and North Carolina where they occur in Kuchler’s mixed mesophytic

forest type.

Bottomland Hardwoods—This is the FRES elmash-cottonwood ecosystem type that occurs in narrow

belts along major streams or scattered areas of dry

swamps. The major portion is on the lower terraces

and flood plains of the Mississippi, Missouri, Platte,

Kansas and Ohio Rivers (Garrison and others 1977).

This type comprises Kuchler’s Southern and Northern

flood-plain forest types and the elm-ash forest. Nineteen SAF forest cover types are included in these

bottomland hardwood forests (Shartz and Mitsch 1993).

Length of hydroperiod, which determines the anaerobic gradient (Wharton and others 1982), rather than

fire frequency, determines plant distribution. Common canopy species include numerous oaks, sugarberry, American elm, eastern cottonwood, green ash,

sweetgum, sycamore, and in deeper water, swamp and

water tupelos, and bald cypress.

Fire Regime Characteristics

Pitch and Virginia Pines—Mixed severity fires

were probably prevalent over much of the range of

pitch and Virginia pines. Where Native American

burning was common, pitch pine existed as an understory fire regime type. Understory fires were common

in pitch pine forests where burning by Native Americans resulted in a 2 to 10 year fire interval. This

Figure 4-16—The extent of northern hardwood forests in

northeastern North America. Redrafted from Bormann and

Likens (1979).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 77

Chapter 4: Fire in Eastern Ecosystems Wade

frequency maintained stands with relatively large

pines, scattered smaller pines and oaks, and little

understory besides low ericaceous shrubs and herbs

(Little 1946, 1973). Today, the mixed fire regime type

applies, at least in the New Jersey Pine Barrens,

because fire return intervals are longer and the majority of wildfires occur during the growing season when

damage is greater. The historical fire regime in Virginia pine is unknown but was probably less frequent

and resulted in higher mortality.

Pond Pine—Most pocosins burn on a 20 to 50 year

cycle (Christensen and others 1988). On better sites,

fire-return intervals range from about 3 to 10 years

and at the short end, can result in pine savanna with

a grass understory. On organic soil sites, such short

return intervals result in herb bogs. Historically, more

frequent fire in the adjacent longleaf-dominated uplands killed encroaching seedlings thereby confining

this species to wetter areas. These wet sites, however,

burned whenever they were dry enough. The rank

shrub layer characteristic of these mesic areas comprises many ericaceous evergreen shrubs that tend to

burn intensely, resulting in the topkill or death of all

vegetation except pond pine. Pond pine has the ability

to resprout from its base (fig. 4-17) and along its stem

and branches (Wenger 1958); thus, its aboveground

stem survives higher intensity fires than stems of

other pine species. This trait allows the species to

dominate wet areas such as pocosins, which support

intense fires. Summer fires during severe drought

usually eliminate the pond pine as well, because the

underlying organic soil burns, destroying the root

systems.

Mixed Mesophytic Hardwoods—Although little

is known about presettlement fire, it appears that fire

was much more common in the mesophytic forests

west of the Appalachian divide than in those to the

east. Harmon (1984) reported the fire return interval

on south-facing slopes in extreme western Great Smoky

Mountain National Park averaged 10 to 12 years

between 1850 and 1940 when fire exclusion was begun

in earnest. In a summary of fire in the Appalachians,

Van Lear and Waldrop (1989) stated, “Forests of the

Southern Appalachians probably did not burn as frequently as the pine-grasslands of the adjacent Piedmont. However, there can be no doubt that they did

burn periodically.” Buckner (1989) described fire’s

evolutionary importance in determining the vegetative mosaic of this region. Harmon and others (1983)

thought that fires in the mixed mesophytic region

were small and restricted to drier sites.

Northern Hardwoods—Although data are limited, evidence suggests that fires rarely occurred in

presettlement Northern hardwood forests (Foster and

Zebryk 1993; Patterson and Backman 1988). Fire

return intervals of many centuries are consistent with

Figure 4-17—Pond pine basal sprouts 1 year after wildfire topkilled the overstory. Photo by Dale Wade, 1972.
78 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

land survey records (Siccama 1971; Lorimer 1977) and

paleoecological data (Patterson and others 1983) in

the Northeast. Lorimer (1977) calculated a fire rotation of 806 years, but argued that this figure was too

low due to the effect of land clearing on his data. Based

on a study of maple-beech forests in Ohio, Runkle

(1990) concluded that the minimum fire-return interval was greater than a canopy generation. In the

Bigwoods maple-basswood forest of Minnesota and

Southeastern Ontario, fire probably occurred more

frequently due to Native American burning practices

that allowed prairie fires to spread eastward into the

Bigwoods (Grimm 1984; Vankat 1979).

Where conifers such as hemlock and spruce were

substantial components of the hardwood forests, standreplacement fires probably occurred more often

(Nichols 1913), in which case the stand-replacement

fire regime may be a better representation of presettlement fire. Portions of the Northern hardwood forest

were visited so infrequently by fire prior to settlement

that a strong case could also be made to place this

forest type in the nonfire regime (table 4-1).

Bottomland Hardwoods—The historical role of

fire in the bottomland hardwood ecosystem is unclear.

In Mississippi, Lentz (1931) stated that low- to moderate-intensity wildfires were frequent and that 80 to 90

percent of the Mississippi delta hardwood forest showed

evidence of damage. Gustafson (1946) and Toole (1959)

presented evidence of disastrous consequences to the

hardwoods from repeated fires. In Louisiana, Kaufert

(1933) dated fire scars on stumps back prior to the

Civil War, but based on the recollections of “oldtimers,” he did not think widespread burning of these

bottoms occurred until about 1890. Low-intensity fires

are the norm in these forests because fuel loadings are

generally light (except after damaging wind storms)

due to rapid decomposition on these moist, humid

sites. In the canebrakes, fire intensity was much

higher although fire severity was low except during

drought. Large fires occur only after extended drought,

usually a dry fall followed by a dry spring.

Fuels and Fire Behavior

Pines—The typical lowland wildfire in pitch pine

on the Pine Plains of New Jersey advances as a wall of

flame consuming overstory pine crowns and leaving a

stubble of shrub skeletons unless the underlying organic soil is also consumed, which occurs during severe drought fires (Little 1979). Fires in the Pine

Barrens tend to be of lower intensity and more like

fires elsewhere in pitch pine stands. Even on steep

slopes of the Southern Appalachians where pitch and

Table Mountain pine grow together, more than

20 percent of the overstory trees typically survive,

although more intense stand-replacement fires

occasionally occur. The short needles of Virginia

pine form a relatively compact forest floor, which

dries slowly and is conducive only to light surface fires

(Little 1974).

Fuels in pocosins occupied by pond pine comprise

varying proportions of shrubs, switchcane, and grasses.

High intensity fires can occur where high fuel loadings

accumulate, often including a rank growth of ericaceous shrubs. Typical live fuel and litter loadings are

6 to 8 tons/acre (13 to 18 t/ha) in low pocosins (about 4

feet high) (fig. 4-18), 8 to 10 tons/acre (18 to 22 t/ha) in

medium height pocosins (about 5 feet high), and upwards of 15 tons/acre (34 t/ha) in high pocosins (about

14 feet high) (Wendel and others 1962). Some pocosin

sites are depicted in a stereo photo series (Ottmar and

Vihnanek, in press). The probability of blowup fires

occurring in pocosins ranges from low in low pocosins

to high in high pocosins.

Northern Hardwoods—Hart and others (1962)

estimated average annual litterfall in a New Hampshire Northern hardwood stand at 1.4 tons/acre (3.16

t/ha). Leaf litter decomposition rates reported as halftimes (years required to lose one-half of the original

dry weight) ranged from 1.1 years for yellow birch to

2.5 years for beech, which are high (Gosz and others

1973). Accumulated duff is typically 2 to 3 inches (5-8

cm) in depth (Hart and others 1962).

Gore and Patterson (1986) sampled downed wood in

Northern hardwood stands including a large tract of

old-growth Northern hardwoods, and a recently clearcut stand in New Hampshire. Loadings of material <1

inch (2.5cm) in diameter were low, ranging from 0.4 to

2.7 tons/acre (1 to 6 t/ha) across all but the recently cut

stand (table 4-4). The total mass of downed wood

declined precipitously in the first 10 years following

cutting and stabilized at 18.7 tons/acre (42 t/ha) in the

old-growth stand (fig. 4-19). Patterson and others

(1983) reported similarly low total woody fuel loadings

for Northern hardwood stands burned in 1947 and

about 1880 in Acadia National Park, Maine. Total

dead fuel loadings averaged 10.0 tons/acre (22.4 t/ha)

and 11.8 tons/acre (26.5 t/ha), respectively, with duff

depths of 2.3 and 1.0 inches (5.8 and 2.5 cm).

Postfire Plant Communities

Pitch Pine

Vegetation Dynamics—Pitch pine is well adapted

to fire having thick bark, serotinous cones, and the

ability to refoliate from dormant buds located along

the stem and branches, or from the basal crook located

just below ground line (Little 1979). A majority of the

trees have serotinous cones especially where fires are

severe (Little 1974). Seed is produced at an early age,

3 to 4 years for sprouts and 10 years for seedlings
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 79

Chapter 4: Fire in Eastern Ecosystems Wade

(Little 1953). Pitch pine is much more resistant to fire

than its hardwood competitors. Fire every 4 to 6

decades will ensure a pitch pine component. As fire

frequency increases, the importance of pitch pine in

the stand also increases. Repeated intense fires at less

than 20 year intervals will eliminate even its fireadapted associates such as shortleaf pine, which take

longer to produce viable seed (Little 1974). Shortleaf

pine also produces basal sprouts when topkilled, but

loses this ability with age while pitch pine can do so

indefinitely (Little and Somes 1956). In the Pine Barrens, most associates are sprouters so that total cover

surpasses 100 percent the first year after wildfire,

severely restricting light and space for obligate seeders (Boerner 1981). The relationship between fire

frequency, intensity, and severity, and their effect on

postfire succession is discussed by Little (1979).

Common associates on upland sites in New Jersey

include shortleaf, Virginia, Table Mountain and eastern white pines; black, white, northern red, southern

red, chestnut, bear, post, scarlet, and blackjack oaks;

and various hickories (Little 1973,1979; Little and

Garrett 1990; Murphy and Nowacki 1997; Wright and

Bailey 1982). Elsewhere eastern white pine is a common associate and outcompetes pitch pine in the

absence of continued fire.

Successional trends following fire are toward domination by hardwood species. Without disturbance,

pitch pine declines (Smith 1991; Vose and others

1994). On upland sites, trees usually invade much

Figure 4-18—Low pocosins maintained by periodic wildfire, eastern North Carolina. Photo by Walter Hough, 1973.

Table 4-4—Average loading (tons/acre) of downed wood by diameter

class (inch) and time since cutting for New Hampshire

northern hardwood stands (Gore and Patterson 1986).

Years since stand was clearcut

Diameter 1 15 50 100 Uncut

0 to 0.25 3.4 1.1 1.5 0.5 0.9

0.25 to 1 6.5 2.2 2.4 2.7 2.4

1 to 3 17.8 1.7 2.9 1.0 2.1

>3 10.8 9.5 7.6 20.1 13.3

Total 38.5 14.5 14.3 24.3 18.7
80 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

faster than understory shrubs such as huckleberries

and blueberries (Little 1979; Little and Moore 1949).

On mesic to wet sites, fire exclusion leads to replacement by red maple, blackgum, sweetbay, American

holly, and gray birch (Little 1979). Shrub encroachment is also much faster, and dense understories of

sheep laurel, piedmont staggerbush, gallberry, and

leatherleaf often quickly develop.

Management Considerations—On xeric mixed

pine-hardwood ridges in the Southern Appalachians,

fire has been advocated to restore diversity and productivity (Swift and others 1993; Vose and others

1994, 1997). Where pitch pine occurs with oaks and

shortleaf pine in New Jersey, it is favored by winter

fires that produce good seedbeds. In the Pine Barrens,

a fire return interval of 12 to 16 years is used to

maintain pine-oak stands. However, low-intensity

winter prescribed burns have little effect on hardwoods (Little 1973). Boerner and others (1988) found

that hardwood growth was actually increased by winter prescribed burns, suggesting that these fires were

counterproductive. Competing hardwoods can be best

controlled by applying a winter fire that reduces surface fuels followed by a summer fire. Deep-burning

fires are needed on deep organic soils to prepare

seedbeds and kill competing hardwoods; however,

smoke management constraints generally severely

curtail opportunities for such burns. Several

underburns 8 to 16 years apart will reduce huckleberries while favoring an herbaceous ground cover including mosses and lichens (Buell and Cantlon 1953) that

benefit many wildlife species from butterflies to quail.

But according to Boerner (1981), species richness

peaks the first year postfire and declines precipitously

as resprouting heath cover closes toward the end of

that first growing season.

In the middle of the Pine Barrens, intense, frequent

fires (every 8 years or so) over a long period have

eliminated virtually all large trees. The result is a low

shrubby plant community consisting of a 3 to 7 foot (1 to

2 m) tall coppice growth of pitch pine, blackjack oak, bear

oak, and mountain-laurel (Little 1946; Little and Somes

1964; Lutz 1934; Windisch and Good 1991). Repeated

surface fires with moderate fire intensities can transform Pine Plains on xeric uplands into taller, less dense

stands due to selective survival and growth of taller

stems. Frequent crown fires (<20 year intervals) will

maintain the Plains type. To preserve this unique ecosystem, fire frequency and severity should be varied to

produce a diversity of stand ages, structures, and patch

types (Windisch and Good 1991). Without fire, scrub type

pitch pine eventually forms an overstory (Little 1998).

Figure 4-19—Downed wood in old growth Northern hardwoods on the Bowl Research Natural Area, White

Mountain National Forest.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 81

Chapter 4: Fire in Eastern Ecosystems Wade

Roughly 4,000 acres of relic pitch pine barren ecosystem known as the Albany Pine Bush occurs in upstate

New York. With periodic fire, pitch pine and scrub

oaks (bear and dwarf chinkapin) predominate; but in

the absence of fire, broadleaved hardwoods including

red and white oaks, red maple, and white ash become

established (Milne 1985). Prescribed fire is currently

used on a 10 year return interval to restore and

maintain this ecosystem, which provides habitat for

the Federally endangered Karner blue butterfly. See

Walker (1967) for discussion of management

recommedations pertaining to pitch, Virginia, and

pond pines.

Virginia Pine

Vegetation Dynamics—Virginia pine tends to

dominate only on nutrient-poor, xeric sites (Mattoon

1915; Williston and Balmer 1980) where other species

have trouble surviving. Virginia pine has only localized commercial importance. Although classed as a

southern yellow pine and often exhibiting more than

one flush during the growing season, it is much less

tolerant of fire than the major Southern pine species

because of its thin bark. Young trees sometimes produce basal sprouts when topkilled.

Management Considerations—Although polesized stands have been treated with low-intensity,

winter season prescribed fire without overstory mortality in New Jersey, fire use should be considered

experimental because of the likelihood of mortality.

Most wildfires kill a majority of the stand because the

thin, compact forest floor will only burn under relatively hazardous conditions. These fires, however, are

usually responsible for regenerating the species. Fire

is an effective tool for eliminating Virginia pine in

mixed pine stands (Slocum and Miller 1953). Prescribed fire was recommended for preparing a seedbed

for the next crop after harvest (Church 1955), and to

increase seedling vigor (Sucoff 1961).

Pond Pine

Vegetation Dynamics—Pond pine has semiserotinous cones, which are often produced by age 4 to

6 and open slowly over a period of years in the absence

of fire (Wenger 1958). Seeds released from cones opened

by fire almost invariably result in a blanket of reproduction, some of which survive the next fire if given 5

to 10 years to develop. Seedlings tend not to resprout,

although they can in some situations.

Common associates are loblolly and slash pines,

cabbage palmetto, Atlantic white-cedar, pond cypress,

bald cypress, swamp tupelo, sweetbay, loblolly-bay,

redbay, sweetgum, and red maple. Greenbrier is almost always a component of the understory along with

switchcane, gallberry, large gallberry, swamp cyrilla,

wax myrtle, saw palmetto, and sweetpepperbush

(Bramlett 1990; Wenger 1958).

Different successional pathways producing various

community types result from the interaction between

fire frequency, fire intensity, hydrology, and organic

soil depth (McKevlin 1996; Wharton 1977). The original pocosins once covered more than 2.5 million acres

(1 million ha) in North Carolina alone (Richardson

1981), but only a fraction of that remains because of

peat mining, drainage, and conversion to pine plantations or row crops. For overviews of this ecosystem see

Richardson and Gibbons (1993) and Stout and Marion

(1993).

Management Considerations—The prescription

fire window is narrow between conditions too wet to

carry fire and fires that sweep through the overstory

completely consuming many live understory stems

0.25 to 0.5 inch (0.6 to 1.3 cm) diameter. Among several

seedbed preparation techniques, prescribed fire was

judged the most risky, but it also produced the best

results (Crutchfield and Trew 1961). Where pond pine

is not commercially utilized, stands are often burned

every 10 to 20 years to regulate fuel buildup and

restore fire to the ecosystem.

One of the most important pond pine understory

communities is composed of switchcane. Cane also

occurs as open-grown thickets thought to have originated on abandoned Native American agricultural

fields and from Native American burning practices

(Platt and Brantley 1997). Early explorers often mentioned canebrakes because of their distinctive character; apparently they were once widespread ranging

from the valleys of the Appalachians to the Pocosins of

the coastal Plain. They have largely disappeared because of overgrazing, inappropriate fire management,

or deliberate type conversion. According to Wharton

(1977), river cane burns about every 5 years but

reaches its maximum fuel storage of 5 to 7 tons/acre

(11 to 16 t/ha) in 3 years. Regularly burned cane

provides some of the most nutritious native grazing in

the South (Biswell and others 1942; Hilmon and Hughes

1965b; Hughes 1966; Shepherd and others 1951). If

native range improvement is an objective, fires must

be frequent; otherwise shrubs will overtop the cane

within a decade. Continued exclusion results in one of

the most flammable fuel complexes in the South.

Guidelines for using prescribed fire in the Pine Barrens can be found in Little and Moore (1945).

Mixed Mesophytic Forest

Vegetation Dynamics—The hardwood forests of

the Appalachian Mountains, the Ozark Mountains,

and upland hardwoods of the Coastal Plain and the

Piedmont have been regularly grazed and burned

from the earliest settlement times (Komarek 1982;

Van Lear and Waldrop 1989). The vast Shenandoah
82 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

Valley was burned annually by Native Americans to

keep it from reverting to forest (Leyburn 1962). Foley

(1901) noted that fire along with logging and grazing

were major determinants of the species composition at

the turn of the century. In the absence of fire, a mixed

mesophytic forest develops. In the old-growth stage,

pine regeneration is precluded and the forest slowly

moves toward a hardwood climax (Cain and Shelton

1994). Literature reviews of the effects of fire on

Eastern hardwood forests are provided by Christianson

(1969) and Fennell and Hutnik (1970).

Management Considerations—The use of fire in

hardwood stands generally has not been recommended

because of the fear of damaging stem quality and

because of the danger of erosion, particularly on steep

slopes (Van Lear and Waldrop 1989). This recommendation is largely based on postburn observations of

wildfires, which often burn with higher intensity and

severity than prescribed fires. For example, in a survey of almost 6,000 harvested upland hardwood trees

nearly half had basal wounds, 97 percent of them

caused by fire (Hepting and Hedgcock 1937). The

incidence of decay originating in basal wounds was

greater for basswood and yellow poplar than oaks.

Seventy percent of the trees with basal wounds had

butt rot that resulted in an average cull of more than

15 percent. But the costs of decay are greater than just

reduced board feet, because fire-damaged trees take

up space that could be utilized by trees of superior form

(Gustafson 1946).

Reviews of fire research on Southern Appalachian

and Upper Piedmont sites showed that prescribed

fires had little negative impact on soil (Van Lear and

Johnson 1983; Van Lear and Waldrop 1989). Although

numerous questions about fire effects on soils remain

unanswered, generally fires that expose mineral soil

create the potential for erosion, while those that leave

a portion of the forest floor do not appear to have

deleterious soil or water consequences. Van Lear and

others (1985) found that Piedmont sites can be harvested following a series of low-intensity prescribed

burns with minimal soil loss and degradation of water

quality.

Augspurger and others (1987) and Waldrop and

others (1985) demonstrated that single fires have

little effect on the composition of young coppice stands.

Roth and Hepting (1943) and Roth and Sleeth (1939)

examined numerous hardwood stands of sprout origin

and found that sprouts on burned areas were forced to

develop at or below the ground line, which resulted in

well-anchored stems free from decay. Thor and Nichols

(1974) found that both the number of stems per sproutclump and the total number of clumps, especially oaks,

increased with annual and periodic burning in comparison to unburned stands.

Low-intensity prescribed fires have also been shown

to stimulate germination of yellow poplar seed (Little

1967), which can remain viable in the forest floor for

more than a decade, and produce more faster growing

seedlings than those on unburned sites (Shearin and

others 1972). Although fire has been demonstrated to

be useful in the regeneration of some mixed mesophytic forests, no references were found that advocate

underburning in the management of these forests.

However, research on the application of prescribed fire

in this type continues. For example, a moderately

intense prescribed burn was applied to a south-facing

slope in the Southern Appalachians to test its effectiveness for restoring a degraded pine/hardwood community and stimulating forage production after 70

years of fire exclusion (Elliott and others 1999).

Studies conducted in the Upper Piedmont and Southern Appalachians have shown that fire can be safely

used to dispose of logging debris and prepare seedbeds

(Swift and others 1993; Van Lear and Waldrop 1989).

For example, Sanders and others (1987) found that

low-intensity dormant-season fires had little adverse

effect on bole quality of mature hardwood stems. Van

Lear and Danielovich (1988) noted little visible evidence of erosion on mountain slopes up to 45 percent

following prescribed summer burning designed to reduce heavy logging debris and prepare the site for

planting. Sanders and Van Lear (1987) showed that

the judicious use of fire reduces the large amount of

highly flammable fine woody material present after

clearcutting by more than 90 percent. The fell-andburn technique that gained prominence in the late

1980s can regenerate mixed pine-hardwood stands

after clearcutting with minimal adverse site effects

(Abercrombie and Sims 1986; Danielovich and others

1987; Phillips and Abercrombie 1987).

Northern Hardwoods

Pre-1900 Succession—Paleoecological studies suggest that Northern hardwood species such as beech,

sugar maple, and birch decline following fire. Pollen

and charcoal samples from Lake Wood in Acadia

National Park, Maine, show that during the period

about 2,000 to 6,000 BP (before present time), Northern hardwoods and hemlock were dominant. During

that period, fires indicated by charcoal analysis occurred in conjunction with sharp declines in hemlock

probably as a result of an insect or disease outbreak

(Davis 1981). Declines of hemlock about 4,800 and

3,000 BP were followed by periods in which one or

more fires burned the watershed. With the rise in

importance of spruce and cedar about 2,000 BP, the

incidence of fire increased with return intervals of 200

to 400 years. The abundance of maple, beech, and

hemlock declined simultaneously. The watershed of

Lake Wood burned in a catastrophic fire in 1947; today
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 83

Chapter 4: Fire in Eastern Ecosystems Wade

it contains only one small stand (about 5 acres) of

hemlock and no Northern hardwood stands. The forest is

dominated by seral hardwoods (aspen, paper birch, and

gray birch), northern red oak, white pine, and red pine.

Northern hardwoods such as the Bigwoods of Minnesota generally are not very flammable (Grimm 1984);

fires burn as patchy, creeping ground fires. Grimm

(1984) noted that “the fire regimes of deciduous forests, such as the Bigwoods, are much different from

the commonly perceived model of a forest fire regime,

in which fuels and fire danger increase with time and

in which intense crown fires commonly cause great

destruction of forest.” This is consistent with our

observations in Northern hardwood stands in Maine

other than at Lake Wood (table 4-5). Stands burned in

1947 currently support forests dominated by sprouts

of beech and sugar maple rather than seral hardwoods

(Patterson and others 1983).

A vigorous debate exists about whether Native

American cutting and burning practices or climatic

cooling caused shifts from Northern hardwoods to oak

and pine at Crawford Lake, Ontario (Campbell and

McAndrews 1995; Clark 1995; Clark and Royal 1995;

McAndrews and Boyko-Diakonow 1989). Although

the relative importance of Native American burning

versus climate change as an influence on the larger

Northern hardwood region remains open, it seems

likely that changes evident in the Crawford Lake

pollen profiles were partly the product of human

manipulation of the forest. Clark (1995) concluded

that additional studies of Native American effects on

Northern hardwood forest composition are needed,

but there is little evidence that Native American

burning alone (without accompanying agricultural

activity) was as important in Northern hardwoods as

it apparently was in oak forests to the south (Abrams

1992).

Post-1900 Succession—Based on fire records from

1945 to 1976, Bormann and Likens (1979) concluded

that forests in the Green and White Mountains are

“among the least burnable in the ‘northern hardwood

region’.” On average, only 7 to 10 acres (3 to 4 ha) burn

annually per million acres (405,000 ha) on the Green

Mountain and White Mountain National Forests.

Fahey and Reiners (1981) calculated fire rotations in

Northern hardwoods of 910 years for Maine and 770

years for New Hampshire. Current work (Patterson

1999) documents the continued trend toward less area

burned (longer rotations) during the later half of the

20th century in New Hampshire and a low 20th

century fire occurrence in Vermont, which has the

largest representation of Northern hardwoods.

Stearns (1949), who examined a virgin Northern hardwood stand in northern Wisconsin, noted that although

hot slash fires “burned to the edge of the virgin stand

they did not penetrate into it more than a few rods.”

Although ecologists believe fire has been a relatively

unimportant ecological factor in Northern hardwoods

(Bormann and Likens 1979; Fahey and Reiners 1981),

they acknowledge the fact that Northern hardwoods

have burned in the past, especially when adjacent

stands were clearcut during the logging period and

when stands accumulated fuels from blowdown

(Lorimer 1977; Stearns 1949). Records suggest that

modern stands have been more influenced by fire

(chiefly as a result of anthropogenic fire during the

period 1850 to 1950) than stands will be in the future.

Even present Northern hardwood stands have been

influenced to a far smaller degree by fire than have

other vegetation types in the Northeast.

After the 1947 fire in Acadia National Park, beech

and sugar maple stands have returned to their original stand composition more rapidly than any other

Table 4-5—Average basal area (sq ft/acre) by species for Mt. Desert Island northern

hardwood stands burned in 1947 and before 1880 (Patterson 1999).

Burned in 1947 Burned before 1880

Sample year

Species 1980 1992 1980 1992

Red spruce — 0.1 0.6 1.6

Hemlock 0.1 0.1 0.4 —

Paper birch 4.8 4.3 4.0 4.2

Yellow birch 0.1 0.3 0.1 0.7

Red maple 0.3 1.2 0.4 0.9

Sugar maple 4.4 3.1 7.0 5.0

White ash 0.1 — 1.5 1.1

American beech 11.7 15.0 9.6 9.2

Striped maple 0.5 1.6 0.9 2.2

Bigtooth maple 0.7 1.5 0.2 0.4

Others 0.5 0.2 0.3 0.2

Total 23.2 27.4 25.0 25.5
84 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

forest types (table 4-5) (Patterson 1999; Patterson and

others 1983). Although Northern hardwood species

are widely viewed as having little resistance to fire,

maple and birch sprout vigorously from the stump;

beech suckers from the root system as vigorously as

aspen (Fowells 1965). This capacity for rapid vegetative reproduction appears to limit invasion of Northern hardwoods by seral aspen, paper birch, and gray

birch. These species are short lived and cannot persist

in competition with beech, maple, and yellow birch in

the absence of frequent, stand-replacing disturbances

(Patterson and others 1983). The present dominance

of white birch on some sites in the White Mountains is

probably more a reflection of increased incidence of

fire and logging in the 1800s than it is an indicator of

the long-term importance of fire on the landscape.

Management Considerations—As management

shifts toward longer harvest rotations and reduced

volume removal, Northern hardwoods will likely regain their historic position of importance on mesic,

fire-protected sites in the Northeast. Northern hardwoods are susceptible to fire (Swan 1970). Where

Northern hardwoods mix with conifers including hemlock, white pine, red spruce, and balsam fir, fires are

likely to be more common, especially in the wake of

catastrophic wind storms (Foster 1988; Lorimer 1977;

Stearns 1949). If climate warms and incidence of fire

is reduced, Northern hardwoods may return to some

sites at the present hardwood-boreal forest boundary

while giving way to transition hardwood-conifers to

the south. However, the increased presence of human

ignition sources may alter fire-vegetation relationships evident in presettlement forests.

Bottomland Hardwoods

Vegetation Dynamics—In young pole size stands,

fires often result in basal wounds. Although these

wounds often heal over, internal decay continues with

decay height closely related to time since fire. Kaufert

(1933) estimated that 90 to 95 percent of the decay in

merchantable Southern bottomland hardwood stands

was the result of past fires. When mature oaks die in

areas protected from fire, species such as red maple,

American elm, and green ash tend to replace them

(Abrams 1992). Lotan and others (1981) stated that

because the elm-ash forest is moderately fire-prone,

prescribed fire should be tested for its ability to control

insect and disease pests and unwanted understory.

Keep in mind that most bottomland hardwoods, even

large ones, are sensitive to fire. Low-intensity fires

appear benign at first glance, but the cambium has

been damaged and incipient decay begins even though

the bark remains intact for several years after fire.

Stand-Replacement Fire

Regimes _______________________

Major Vegetation Types

Vegetation types in the Eastern United States represented by stand-replacement fire regimes, where

fire typically kills more than 80 percent of the overstory, include prairie, wet grassland, and portions of

the oak-gum-cypress (bay forests) FRES ecosystem

types (table 4-1). The conifer cover types include sand

pine, Table Mountain pine, Atlantic white-cedar, and

spruce-fir. Pocosins without a significant component

of pond pine are also a stand-replacement fire regime

type.

Wet Grasslands

Kuchler (1964) recognized two major regions of herbaceous wetlands in the Eastern United States, exclusive of the Everglades of southern Florida (see chapter

7). These regions include the northern cordgrass prairie, which extends along the Atlantic coast from Maine

to southern Florida, and the southern cordgrass prairie, which spans the Gulf of Mexico from southern

Florida to southern Texas. Numerous marshes, some

quite extensive, occur in inland areas in the Eastern

United States, and many of these possess characteristics similar to freshwater coastal marshes.

It is convenient to distinguish salt and brackish

from oligohaline (tolerant of moderate salinities) and

fresh marshes because of consistent differences in

species composition and fire behavior. Along the Atlantic seaboard, salt marshes can be further subdivided into the New England group and the Coastal

Plain group (Mitsch and Gosselink 1993). Salt marshes

of the New England group extend from Maine to New

Jersey and are built mainly on marine sediments and

marsh peat, with relatively little sediment discharge

from distributaries. Salt marshes of the Coastal Plain

group extend southward from New Jersey to Florida,

where they are replaced by Mangrove forests at the

southern tip of Florida.

Salt and brackish marshes are largely dominated by

species of cordgrass and rush. The regularly flooded,

tidal salt marshes in the Eastern United States are

dominated almost entirely by smoothcordgrass. Regularly flooded areas typically are referred to as low salt

marshes to distinguish them from high salt marshes,

which occur inland from the low salt marshes, are less

frequently flooded, and often contain more stressful

soil conditions due to stagnation and evaporative

concentration of salts (Mitsch and Gosselink 1993).

Low salt marshes reach their greatest extent in South

Carolina, Georgia, and along the Gulf of Mexico (Teal
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 85

Chapter 4: Fire in Eastern Ecosystems Wade

1986). High salt marshes can be dominated by several

species such as smooth cordgrass, needlegrass rush,

pickleweed species, inland saltgrass, saltmeadow

cordgrass, and saltmeadow rush. Along the Texas

coast, high salt marsh can include extensive stands of

gulf cordgrass, which is also the characteristic dominant of salty prairie, an upland community type. In

brackish areas, salt marsh species yield dominance to

species of slightly less salt tolerance, and a greater

variety of both dominant and subordinate species can

be found (Gosselink 1984). In addition to a shift in the

herbaceous layer, brackish marshes often include woody

species, especially eastern baccharis and bigleaf

sumpweed. A more detailed discussion of the geographic variations in salt marsh geomorphology and

vegetation can be found in Mitsch and Gosselink (1993).

Coastal marshes that receive freshwater and are

removed from the direct influence of salt water form

the inner band of coastal marshes. These marshes

reach their greatest extent along the middle and

southern Atlantic Coast and along the northern Gulf

Coast. The Atlantic Coast freshwater marshes include

about 405,200 acres (164,000 ha), while those in the

northern Gulf of Mexico cover about 1,156,400 acres

(468,000 ha). Generally, fresh and oligohaline wetlands occur where salinities are less than 5 ppt, but

wetland types are more easily recognized by the known

salinity tolerances of the vegetation rather than the

average soil or water salinity (Brewer and Grace

1990). Many plant associations exist because of the

high diversity of species found in fresh and oligohaline

marshes. The lowest fresh marshes are characterized

by plants that root in relatively deep water, such as

species of pond-lily, waterlily, wildrice, and cutgrass.

Along the Atlantic Coast, tidal fresh marshes include

both annual streamside associations and perennial

associations of green arrow arum, pickerelweed, arrowhead species, and cattail species. In fresh marshes

of the Gulf Coast, bulltongue arrowhead, maidencane,

spikerush species, and numerous sedge and forb species are predominant.

Prairie

The prairie ecosystem in the United States, referred

to by many as the tallgrass prairie, forms a rough

triangle from the Minnesota and North Dakota border

south to the Texas Gulf Coast and eastward into

northern Indiana (Reichman 1987). It is dominated by

big bluestem, Indiangrass, and switchgrass. Drier

sites are dominated by little bluestem, and wet, lowland sites are often dominated by prairie cordgrass.

Within this grass matrix are more than 250 forb

species (Freeman 1998). The local species composition

at any one place is dependent on burn history, grazing

history, soils, aspect, and topographic position. In

Indiana, Illinois, and Wisconsin only 0.1 percent of

this ecosystem remains. The Flint Hills region of

Kansas has the most remaining prairie, 3 million

acres (1.2 million ha), which is only 17 percent of the

presettlement prairie in Kansas (Samson and Knopf

1994). The prairies that do remain outside of the Flint

Hills region are widely scattered and often smaller

than 1.2 acres (0.5 ha) (Betz and Lamp 1989). The

tallgrass prairie is one of the youngest ecosystems in

North America. Much of the region was glaciated only

10,000 years ago. This region has few endemic species,

so most plant and animal species have migrated into

this region from neighboring ecosystems (Risser and

others 1981).

Portions of the tallgrass prairie consist of oak savannas and glades or barrens. Herbaceous species found

in savannas are a mixture of forest and prairie species

(Curtis 1959); there are no endemic savanna species.

In wetter periods, or periods with reduced fire frequency, savannas can be converted to forests. In drier

years or with shortened fire intervals, savannas can be

converted to grasslands. Whether savannas are stable

ecosystems or simply an unstable continuum between

closed canopy forest and open grasslands is debatable

(Nuzzo 1986). Today, Nuzzo (1986) estimates that only

0.02 percent of presettlement savanna survives and

that the largest savannas are only about 50 acres (20

ha).

 Glades or barrens are patches of prairie within a

forest matrix. The barrens region stretches across

Missouri, Tennessee, Kentucky, southern Illinois, Indiana, and Ohio (Baskin and Baskin 1978). These

grasslands occur only on very dry, shallow soils, which

are usually found on south- or west-facing slopes.

Anderson and Schwegman (1971) stated that barrens

are “degraded forests that had been invaded by prairie

plants as a result of fire” and that in the absence of fire

the areas quickly revert to forest. However, Wade and

Menges (1987) stated that these glades don’t support

woody vegetation except for the shallow rooted eastern redcedar; thus, they are often termed cedar glades

(Baskin and Baskin 1978). If fire is excluded for long

periods, trees encroach around the edges of glades and

reduce their size.

Bay Forests

This ecosystem type occurs primarily in North and

South Carolina but can be found along the Atlantic

Coast from Virginia to Alabama. Carolina bays and

pocosins without an overstory of pine or Atlantic

white-cedar are the major vegetation types. Much of

the original type contained a merchantable overstory

that was harvested, thereby altering the fire cycle.

This type is characterized by a dense tangle of evergreen and deciduous shrubs and vines (Richardson
86 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

and Gibbons 1993). Carolina bays are swamps dominated by bay species. Numerous species of special

concern including several Federal and State listed

species occur in this vegetation type.

Conifers

Sand Pine—Two varieties of sand pine, Choctawhatchee and Ocala, are recognized here because

their fire ecology requires different management. The

natural range of Choctawhatchee sand pine is confined to the panhandle of Florida and Baldwin County,

Alabama (Brendemuehl 1990). It originally was restricted to the Gulf shoreline (islands and dunes)

where fire was infrequent (Outcalt 1997). The Ocala

variety has serotinous cones and is confined to the

central Florida ridge and on old sand dunes down both

coasts from central to southern Florida. It is easily

distinguished from the Choctawhatchee variety, which

generally lacks serotinous cones.

Table Mountain Pine—This serotinous cone species is endemic to the Appalachians from Pennsylvania to northeastern Georgia, with local populations in

New Jersey and Delaware (Little 1978). According to

Sternitzke and Nelson (1970), about 90 percent of the

standing inventory is in West Virginia, Virginia, and

North Carolina. Table mountain pine forms even-aged

pure stands or shares dominance with pitch pine.

Common hardwood associates include red maple,

blackgum, sourwood, chestnut oak, and scarlet oak

(Della-Bianca 1990), which eventually dominate these

sites in the absence of fire. A dense shrub layer of

mountain-laurel, which will burn within a week of

good drying conditions during the dormant season, is

often present along with other ericaceous shrubs such

as blueberries and huckleberries.

Spruce-Fir—This is a high-elevation forest type of

the Appalachians and is one of the rarest and most

threatened ecosystems in the South (White and others

1993). In the Southern and Central Appalachians,

stands comprise red spruce, the endemic Fraser fir,

and balsam fir. In the Northeast, red and white spruces

and balsam fir make up this forest type.

Atlantic White-cedar—This species tends to form

pure stands throughout a narrow coastal belt from

southern Maine to northern Florida (Little 1959) and

westward to the Mississippi (Little 1978). However,

its distribution is spotty because it avoids substrates

underlain with clay (Little 1950). It requires moist

sites with a long hydroperiod but without stagnant

water, generally in swamps with organic soils.

Fire Regime Characteristics

Wet Grasslands—Much of the coastal region of the

Southern United States, from Virginia to Texas, is

characterized by a presettlement fire frequency of 1 to

3 years (Frost 1995). Coastal marsh landscapes are

typically extensive, a factor that aids in the propagation of an individual fire. Natural barriers to fire

spread are relatively common and vary from wide

river channels to small stream channels and narrow

animal trails. Depending on the fuel and wind speeds,

fires may either bridge small to moderate-sized natural breaks or be stopped by them. Thus, the extent of

natural fires varies greatly as does the ease of accomplishing a prescribed burn. Lightning-strike fires are

thought to be common in coastal wetlands (Frost

1995), and often fire from the adjacent upland can

spread into the marsh. Spontaneous combustion has

been reported to occur in coastal marshes (Viosca

1931), though how frequently it happens is largely

unknown.

Away from the coastal influence of a persistently

high water table, peat fires can be common during

prolonged dry periods and can represent a substantially more severe fire, leading to loss of substrate and

protracted subterranean burning. While such fires

may not be the norm for coastal marshes, they can be

important under some circumstances (Hungerford and

others 1995). For all fires, coastal or inland, groundwater levels are important for both the behavior of the

fire and its effects on vegetation and soil (Bacchus

1995).

Prairie—Historically, Native Americans contributed to the creation and maintenance of the tallgrass

prairie ecosystem by frequently burning these ecosystems, which controlled woody vegetation and maintained dominance by herbaceous plants. In the Eastern tallgrass prairie, Native Americans were probably

a far more important source of ignition than lightning.

With grasses remaining green through late summer

and a low incidence of dry lightning storms, lightningcaused fires were probably relatively infrequent.

Few studies of the pre-Euro-American tallgrass prairie have been conducted. Most existing data are primarily anecdotal, based on widely scattered accounts

of early explorers. Existing data on burn frequency in

the tallgrass prairie come from studying tree rings on

the prairie-forest margin. In Missouri, Guyette and

McGinnes (1982) determined that fires occurred every

3.2 years prior to 1870. After 1870, the average fire

return-interval increased to 22 years. On the Mark

Twain National Forest, Guyette and Cutter (1991)

determined that from 1710 to 1810, fires occurred on

average every 4.3 years and that severe fires (fire scars

on three or more trees) occurred every 11 years. After

1810 fires occurred every 6.4 years, and no fire scarred

more than two trees. But fire scar studies are typically

conservative and tend to underestimate fire frequency,

which could be the case here because historical accounts reported annual burning in the tallgrass prairie
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 87

Chapter 4: Fire in Eastern Ecosystems Wade

(Pyne 1997). After Euro-American settlement, fires

were less frequent and burned smaller areas. The

reduction in fire frequency resulted in replacement of

large areas of grasslands by woodlands (Beilmann and

Brenner 1951; Muir 1965; Pyne 1997).

Some notion of the seasonality of historic fires can be

learned from early explorers and missionaries dating

back to the late 1600s. According to Pyne (1997), Wells

stated in 1819 that fires in western Pennsylvania were

set by Native Americans at the end of the Indian

summer, presumably October; Michaux observed in

1805 that fires occurred in March or April in Kentucky; and James reported in 1819 that fires in Missouri usually occurred in the fall. In Illinois, McClain

and Elzinga (1994) reported that historically almost

all fires were ignited during the Indian summer, late

October to early November. Euro-Americans continued the burning practices they learned from Native

Americans until early in the 20th century when these

historic fire patterns were altered through fire suppression, planting of cool season grasses, and the use

of prescribed burning concentrated in the spring to

improve forage for livestock. It appears from these

historical accounts that the season when fires were

common varied regionally and through time as a

result of the cultural practices of the people living in a

particular area.

In addition to fire, bison and other large grazers

influenced the tallgrass prairie. Bison were present

throughout the tallgrass region (McClain and Elzinga

1994); thus periodic grazing was a significant influence on plant communities (fig. 4-20). Bison prefer

burned to unburned grassland for grazing during the

growing season and can contribute to the pattern of

burning in prairie (Vinton and others 1993). The

variability of grazing by bison may have created variability in the patchiness of fuels and severity of subsequent fires. The effects of grazers in tallgrass prairie

plant communities are reviewed by Hartnett and

others (1996), Hartnett and Fay (1998), Howe (1999),

and Knapp and others (1999).

Bay Forests—This type now burns on about a 20 to

100 year cycle, but uncertainty exists about the historic fire frequency. Wharton (1977) stated that 50 to

150 years are required for mature bay forests to

develop. McKevlin (1996) believes fire frequency is

probably less now than it was 200 years ago in spite of

lowered water tables. The presence of charcoal lenses

at various depths in the underlying organic soil is

evidence of extensive fires in the past (Dolman and

Buol 1967).

Conifers—The fire cycle for Ocala sand pine corresponds roughly to stand longevity, which is 30 to 60

years (Christensen 1981). The historic fire frequency

for the Choctawhatchee variety is unknown but lightning fires were rare. This variety grows in pure stands,

directly inland from the beach, separated from more

pyrophilic vegetation types by wet intradune swales

and sparse dune vegetation.

Little is known about the historical fire frequency in

Table Mountain pine. Barden and Woods (1974)

found that fires in Table Mountain pine were more

frequent, more intense, and probably larger earlier

this century. Between about 1800 and 1944, the fire

interval in sampled stands averaged 10 to 12 years.

Now the fire return interval is between 7 and 70 years

with an average of 40 years based on USDA Forest

Service records. A commonly accepted hypothesis is

that the fire practices of the large Native American

population inhabiting the Southern Appalachians exposed both Table Mountain and pitch pine to frequent

understory burns keeping the stands open, thereby

ensuring successful regeneration after the occasional

stand-replacement fire. Zobel (1969) suggested an

evolutionary link between fire and Table Mountain

pine because of its fire adaptations.

Fires rarely occurred in spruce-fir forests prior to

Euro-American settlement. Two relatively recent instances in Maine were the 4 days in October 1947 when

about 7,000 acres (2,830 ha), including spruce-fir,

Figure 4-20—Grazers are attracted to freshly burned areas

because of increased forage quality, Konza Prairie Biological

Station, Riley County Kansas. Fire management can control the

distribution of grazers on the landscape. Photo by Greg Hoch.
88 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

burned, and the summer drought fire that involved

Mount Katahdin in the early 1990s. Before EuroAmericans harvested Atlantic white-cedar, this

prized species was mostly perpetuated by major disturbances, probably crown fire that occurred at 25 to

300 year intervals. Mature cones can develop by age

three in open stands, which supports the crown fire

hypothesis.

Fuels and Fire Behavior

Wet Grasslands—Coastal marshes typically support high fine fuel loadings. Estimates of standing

biomass range from 4 to 18 tons/acre (10 to 40 t/ha)

(Gosselink 1984; Odum and others 1984) and higher

(Gough and others 1994). Hackney and de la Cruz

(1978) have reported that for some species, full recovery

of standing biomass to preburn levels can occur in a

single growing season, while it may take several years

for other species. Informal observations of fire behavior

indicate that marsh fires are generally consistent with

other grassland fires in the Eastern United States.

Rates of lateral spread vary with wind speeds and can

exceed 3 feet/second (1 m/s). Flame heights can vary

from less than 3.3 feet (1 m) for short or sparse vegetation and up to 13 feet (4 m) for dense or tall vegetation

(fig. 4-21). In typical areas of fine fuels, marsh fires

produce fewer flying embers than shrubland or forest

fires and are generally less likely to produce spot fires.

Coastal marshes typically contain large quantities

of herbaceous vegetation and are considered highly

flammable. But not all marsh types and plant species

are equally conducive to propagating a fire. The vegetation of salt and brackish marshes is relatively

similar throughout the coastal region of the Eastern

United States, with species of cordgrass, rush, and

saltgrass as typical dominants. The cordgrass species,

which are generally the most widely represented group

throughout coastal salt and brackish marshes, tend to

be quite flammable. Species such as saltmeadow

cordgrass and gulf cordgrass can burn readily while

tissues are green and are capable of burning more

than once in a single growing season. All of the cordgrass

communities are capable of carrying fire over standing

water.

Flammability of vegetation throughout fresh and

oligohaline marshes can be highly variable due to the

considerable plant diversity (Gough and others 1994).

In general, grasses support much more intense and

continuous fires than do forb and sedge associations.

Some exceptions to this include dense stands of sawgrass

and cattail, which are capable of supporting extensive

fires. Common species that generally provide good fuels

Figure 4-21—Headfires in sawgrass behave similarly in other Eastern wet grasslands, Everglades National Park.

Photo by Wayne Adkins.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 89

Chapter 4: Fire in Eastern Ecosystems Wade

include reed, maidencane, and switchgrass. Dominant

species that represent poor fuels include bulltongue

arrowhead, spikerush species, alligatorweed,

hydrocotyle species, pickerelweed, and green arrow

arum. Generally, grass-dominated associations burn

more readily than those dominated by sedges or forbs.

Among grass-dominated associations, fresh and

oligohaline marshes (for example, panicums) tend not

to burn as reliably as those in the brackish and salt

marshes (Ford and Grace 1998a). Forb associations,

such as the extensive bulltongue arrowhead community of the lower Mississippi River wetlands, typically

provide poor fuel for fires during the dormant season

when plant tissues have decomposed and fuel is sparse.

In other forb-dominated plant associations, propagation of a fire may only be successful during the dormant season when aboveground tissues have senesced.

The presence of woody plants in coastal marshes can

alter fire behavior substantially in some cases. Where

the woody plant component is high, reduced airflow to

fine fuels may reduce the completeness of the burn and

the rate of fire spread. Nonetheless, fire generally kills

the aboveground portions of these woody species

(Scifres and Hamilton 1993). Succession to woody

associations in fresh marshes is common, particularly

where water depths are not excessive and substrates

are firm. A heavy dominance by species such as willow,

maple, cypress, tupelo, cedar, Chinese tallow, and

bayberry can substantially reduce herbaceous fuels.

These communities are likely to burn only under very

dry conditions.

Prairie—Early explorers and settlers described

abundant grasslands, prairies, savannas, and open

woodlands with grass and herbaceous understories

throughout the Eastern United States. The exact

characteristics of tallgrass prairie such as height,

loading, and percent cover are unknown but can be

estimated from historical accounts. Grasses were 3 to

6 feet (1 to 2 m) high; they were often described as tall

as a man on horseback. Percent cover was substantial

(>50 percent) over hundreds of square miles. From

these characteristics, loadings are conservatively estimated to range from 2 to 5 tons/acre (4.5 to 11.0 t/ha).

Fuel loadings ranged from 1.1 tons/acre (2.5 t/ha) for

sparse communities to 3.4 tons/acre (7.6 t/ha) for

abundant communities in fuel models developed for

fire behavior modeling (Reinhardt and others 1997).

These loadings were based on estimates of current

production (Garrison and others 1977) and accumulated thatch or litter. Flame lengths of 12 feet (3.6 m)

could be expected assuming these loadings, 6 percent

fuel moisture content, and windspeed of 20 mph (32

km/hr) (Rothermel 1983). Thus, historical prairie fires

probably burned much of the time with flame lengths

of 8 to 15 feet (2.4 to 3.6 m), too hot for direct frontal

attack with hand tools.

Bay Forests—Forest floor and downed woody fuels

averaged 10 tons/acre (22 t/ha) in the Piedmont of

North Carolina (Albrecht and Mattson 1977). In the

Great Dismal Swamp in Virginia, Day (1979) found

loadings of accumulated dead and live understory

fuels (< 1 inch diameter) in a mixed hardwood stand of

9.7 tons/acre (21.7 t/ha) and on a maple-gum site of

15.6 tons/acre (35.1 t/ha). Also, see pond pine fuels and

fire behavior in the Mixed Fire Regime section.

Conifers—Sand pine needles are short and form a

flat mat on the forest floor that burns poorly and

supports creeping fires. Needle-drape is not a flammability problem in either variety. Mature stands average about 10 to 15 tons/acre (22 to 34 t/ha) of available

fuel composed mainly of live understory biomass

(Custer and Thorsen 1996). Lightning fires are common, typically small, low-intensity creeping fires that

go out at night when the humidity rises. The primary

ground fuel is deer moss, which can produce flame

lengths of 2 to 3 feet (<1 m) when dry, but when

humidity is high, the moss absorbs moisture from the

air and will not even smolder. The Ocala variety often

supports a well-developed understory of scrub oaks,

ericaceous shrubs, and rosemary that will burn intensely when dry and pushed by a strong wind. This

variety also exhibits a “varnish stage” when the needles

unexplainably exude a sap-like substance that is exceedingly flammable (Hough 1973). This condition is

usually observed in the fall and occurrences can be

decades apart. The Ocala fuel complex typically burns

with intense fires that almost always enter the stand

from more pyrogenic adjacent communities (fig. 4-22).

The fastest spreading wildfire recorded in the United

States occurred in this fuel type; it covered 35,000 acres

(14,170 ha) in 4 hours with a spread rate of 6 mph.

Under Atlantic white-cedar stands in the Great

Dismal Swamp of Virginia, Day (1979) found average

loadings of 1.8 tons/acre (4.0 t/ha) for the litter layer,

1.2 tons/acre (2.7 t/ha) for woody fuels <0.75 inch (2 cm)

diameter, and 22.4 tons/acre (50 t/ha) for woody fuels

>0.75 inch (which included stumps as well as logs).

Postfire Plant Communities

Salt and Brackish Marshes

Pre- and Post-1900 Succession—Many coastal

marshes are dependent on fires that are lethal to

aboveground tissues and that reduce or eliminate

woody plants. Woody plants can, however, be excluded

from an area by excess salinity. Saline and hypersaline soil conditions generally preclude most native

woody species in both coastal and inland wetlands

throughout the temperate zone, with the exception of

salt-tolerant mangroves, which are restricted to subtropical-tropical latitudes in southern Florida (Mitsch
90 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

and Gosselink 1993). Marshes formed on unconsolidated or flotant substrates are also typically unsuitable for the long-term success of trees even in the

absence of fire (Doyle 1995). In some situations, woody

succession may be precluded by herbivores. For example, in much of southern Louisiana wild populations of nutria are believed to prevent successful

reestablishment of baldcypress (Connor and Toliver

1987). The herbaceous communities that develop in

marshes, as a result of fire, edaphic influences, or

herbivores, tend to be readily flammable and are well

adapted to frequent fires.

Despite the frequent presence of standing water, fire

is able to propagate in both low and high salt marsh

and in brackish vegetation. As a result, these systems

are frequently burned, both naturally and with prescription. Postfire succession patterns vary somewhat

with salinity and preexisting vegetation. In low salt

marshes, little species replacement occurs and smooth

cordgrass typically retains dominance. Thus, the main

effect on vegetation is a replacement of old tissues with

younger tissues that are more palatable to wildlife

species. In high salt marshes, fire causes few longterm shifts in vegetation, although it may provide a

brief period when dominance by cordgrass is reduced

and interstitial species increase. Successional patterns are more pronounced in brackish marshes following fire. In areas of the central and western Gulf

coast, succession from herbaceous dominance to dominance by eastern baccharis takes place over several

years. Frequent fires in these systems keep woody

species from dominating. On a shorter time span, fire

in brackish marshes reduces dominance by cordgrass

and rush species temporarily and favors earlier successional species such as chairmaker’s bullrush (Ford

and Grace 1998a).

Management Considerations—According to

Nyman and Chabreck (1995), the frequency of intentional fires increased around 1910 as burning in coastal

marshes became a more common practice for promoting wildlife populations and reducing the hazards of

wildfires. Today, prescribed fire is commonly and

frequently used in salt and brackish marshes to enhance productivity (Hackney and de la Cruz 1981),

manage food sources for wildlife and cattle, reduce

plant cover, reduce fuel loadings, and eliminate woody

species such as baccharis (Chabreck 1988).

In the absence of fire, succession to woody dominance can take place in only a few years. Where the

control of woody plant succession is of highest priority,

growing-season burns approximately every 3 years

are the most effective. Observations indicate that

baccharis-dominated systems remain flammable, in

contrast to those overtaken by the introduced exotic,

Chinese tallow. This species can greatly reduce fire

intensity and fire propagation in dense stands (Grace

1998). Only frequent fires are likely to effectively

prevent its invasion into wetlands. Once Chinese tallow exceeds a certain density, it typically becomes

nonflammable and acts as a firebreak. Once this

density threshold is reached, herbicides or mechanical

means will be required for its removal.

Waterfowl and mammals alike generally prefer early

successional plant species as well as the younger

tissues of resprouting plants. Some fire-promoted species such as chairmaker’s bullrush are considered of

exceptionally high value to muskrat (Ondatra

zibethicus), nutria (Myocaster coypus), and snow geese

(Chen caerulescens). In areas where burning is commonly used to promote wildlife populations, burns

may be conducted in the fall and winter to provide a

steady supply of young tissues throughout the winter.

For cattle grazing, burning is often recommended as a

method of temporarily increasing dietary crude protein and forage quality (Angell and others 1986; McAtee

and others 1979).

Figure 4-22—Stand replacement burn in Ocala variety sand

pine on the Ocala National Forest, Florida. Photo by George

Custer, 1993.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 91

Chapter 4: Fire in Eastern Ecosystems Wade

Marsh fires can be classified as cover burns, root

burns, and peat burns (Lynch 1941; O’Neil 1949).

Water levels control the depth of influence of fire.

Thus, proximity to the water table, tidal conditions,

and drought cycles can determine the severity of

impact to belowground plant parts and to substrate.

While more common in inland areas such as the

Everglades, peat burns have been reported in coastal

wetlands along the Gulf of Mexico (Hoffpauir 1968;

Lynch 1941). During dry periods it may be possible to

create fires that could substantially damage plant

roots. Documented success in using such burns to

control plant species, however, is lacking (Nyman and

Chabreck 1995).

Prescribed fire for promoting desired wildlife forage

species such as chairmaker’s bullrush appears to be

more successful when conducted during the fall or

winter. Spring burns are believed to damage the

regrowth of this species and lead to more persistent

dominance by saltmeadowcordgrass (Chabreck 1981).

Fall and winter burns may be used to avoid destroying

nests or killing young wildlife (Nyman and Chabreck

1995). Fires aimed at promoting nutria and many

other species are often limited in size to produce a

landscape mosaic of burned and unburned habitat

(Kinler and others 1987). Postburn water levels can

substantially influence the effects of fire on vegetation. When stubble is submersed for an extended

period, complete death can occur for many wetland

species (Herndon and others 1991; Sale and Wetzel

1983). See Chabreck (1988) for more information on

the use of fire to manage wildlife, and Kirby and others

(1988) for an extensive bibliography of literature dealing with fire effects on wildlife.

Avoiding use of fire to favor wildlife may be wise

under certain circumstances. For example, in the

Mississippi delta, geologic subsidence rates have contributed to extremely high rates of marsh loss because

coastal areas have subsided faster than accretion

occurs. In this situation, the feeding activities of nutria and other mammals can contribute to habitat loss

(Ford and Grace 1998b). Nyman and Chabreck (1995)

noted that the potential exists for deleterious effects of

marsh burning in this region. Hypothetically, fire

could accelerate rates of wetland loss through the

removal of organic matter that might otherwise contribute to sediment accretion and through the promotion of wildlife populations that lead to consumption of

vegetation. At present, experimental evaluation of

this hypothesis is lacking.

Another factor is the possibly deleterious grazing by

snow geese. Currently, excessive numbers of snow

geese are causing extensive damage to the northern

wetlands where they breed and they are also known to

cause extensive eat-outs of southern coastal wetlands.

There is no evidence at present that burning is significantly affecting their population. However, burning of

coastal wetlands can attract wintering snow geese to

recently burned areas and increase the potential for

localized damage. Finally, because of the important

role of marshes as sources of organic matter for estuarine food webs, high fire frequencies are not necessarily ideal for near-shore systems (Hackney and de la

Cruz 1978; Nyman and Chabreck 1995). A more complete discussion of salt marsh fire ecology can be found

in Lynch (1941), Bendell (1974), Daiber (1974), Frost

(1995), Nyman and Chabreck (1995).

Fresh and Oligohaline Marshes

Pre- and Post-1900 Succession—Various successional sequences are found from fresh marshes to

forested wetlands for the most inland of the fresh

marshes. Fresh and oligohaline wetlands may succeed

to dominance by baldcypress, swamp tupelo, water

tupelo, and red maple as well as to Chinese tallow

(Frost 1995). A number of factors may impede woody

plant development including unconsolidated substrate,

scouring by waves, and periodic fires. Fire-driven

successions in freshwater coastal marshes are poorly

documented. Evidence suggests that as with brackish

marshes, fire releases a diversity of early successional

species that are more palatable to wildlife (Ford and

Grace 1998a). Succession of the herbaceous communities in fresh and oligohaline marshes tends to be rapid

as it is in salt and brackish marshes. Preburn vegetation can regain its dominance in a few years. Van

Arman and Goodrick (1979) reported that 6 months

after prescribed burning a Florida freshwater marsh,

vegetative recovery was almost complete, and total

numbers of animal species and individuals were significantly higher in the burned area than in the

adjacent unburned marsh. Increases in these

macroinvertebrates and smaller fish populations at

the lower end of the food chain suggest that these

increases potentially could be passed on up the scale.

See Ewel (1995) and Frost (1995) for a more detailed

consideration of the role of fire in regulating succession in forested freshwater wetlands.

An increasingly common successional pattern in

fresh marshes is due to invasion of Chinese tallow,

which has limited tolerance to salinity and is largely

confined to fresh and oligohaline wetlands. Evidence

indicates that when Chinese tallow invades a wetland

or upland grassland it causes a shift from a grassdominated herbaceous layer to a sparse forb-dominated layer that is much less capable of carrying a fire.

As a result, stands of Chinese tallow act as firebreaks.

Below some minimum stand density, fire can be used

to effectively control Chinese tallow as long as adequate fuel remains (Grace 1998).
92 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

Management Considerations—The use of prescribed fire in coastal fresh marshes is much less

extensive than in salt and brackish marshes (Chabreck

1988). Fires in salt and brackish marshes, however,

often spread into fresh marsh areas, resulting in a

relatively frequent burn regime for associations that

will propagate fire. Prescribed burning in fresh marshes

is more likely to be used for fuel reduction and to

control woody plants such as wax myrtle, thinleaf

alder, and Chinese tallow than to promote wildlife

populations. Nonetheless, when wildlife or cattle production is the goal, the same management recommendations apply to fresh marshes as described previously

for brackish and salt marshes.

Prairie

Vegetation Dynamics—A primary effect of fire in

tallgrass prairie ecosystems is the control of invading

woody species (Anderson and Van Valkenburg 1977;

Wade and Menges 1987). The rapid conversion from

prairie to forest with the removal of fire was noted in

the early 1800s. Muir (1965) observed in Wisconsin

that as soon as sufficient firebreaks were created, a

thick oak forest invaded the prairie. In 1822, botanist

Edwin James reported, “Whenever the dominion of

man is sufficiently established in these vast plains, to

prevent the annual ravages of fire, trees will spring

up” (Pyne 1997). One of the most aggressive woody

species in the prairie is eastern redcedar. In the

absence of fire this species could quickly become the

dominant tree over much of the Ozark region (Beilmann

and Brenner 1951). Cedar forests now occupy 6.4

million acres (2.6 million ha) in five Midwestern States,

an increase of 113 percent during the last three decades (Schmidt and Leatherberry 1995). In as little as

30 years after fire, a treeless pasture can be converted

to a closed canopy cedar forest (Hoch and Briggs 1999).

In savannas, frequent fires tend to be of low intensity, do not kill overstory trees, and create an open

understory. Infrequent fires are more intense due to

litter accumulation, can kill overstory trees, and promote vigorous sprouting of woody species, often creating a thicket. Fires every 2 to 3 years held woody

canopy at a constant level (Faber-Langendon and

Davis 1995). Generally, more frequent fires reduced

tree canopy while less frequent fires increased tree

canopy. Faber-Langendon and Davis (1995) suggested

that a 4 year fire interval might be best for controlling

tree spread because at this interval fires would burn

more intensely than annual or biennial fires.

Fire frequency affects species differently. Generally,

big bluestem shows no response to time since fire,

while little bluestem, Indiangrass, and switchgrass all

decrease with time since fire (Collins and others 1995).

Gibson (1988) found that perennial forbs and coolseason grasses increased with time since fire, while

annual forbs and warm-season grasses decreased.

Annual or frequent burning tends to decrease herbaceous plant diversity in tallgrass prairie. Knapp and

others (1999) found that annually burned areas had

lower species richness than unburned areas and areas

burned every 4 years. Collins and others (1995) found

that species richness increases for 7 to 8 years after

burning, and that time since burning was the primary

agent in determining variation in species composition.

Gibson (1988) argued that burning every 4 years

might be the best strategy for maintaining maximum

diversity (fig. 4-23). In annually burned areas fewer

species can become established. In unburned areas

litter accumulation creates too much shade for many

species. Burning every 3 to 4 years ensures that most

of the species present will have at least one “optimal”

year for growth and reproduction. Collins (1987) showed

that a combination of burning and grazing resulted in

higher plant species diversity than burning or grazing

alone.

Generally, net primary production increases the

growing season following burning (fig. 4-24). If a prairie remains unburned, detritus accumulates that

shades the soil, especially at the beginning of the

growing season, and limits production (Knapp and

Seastedt 1986; Seastedt and Knapp 1993). However,

the decomposing detritus adds nitrogen to the soil, and

the detrital layer insulates the soil from drying. Researchers have theorized that long-term annual burning would reduce soil nitrogen levels and lead to

decreased productivity. However, even after 20 years

of annual burning at Konza Prairie, Kansas, productivity has not decreased (Blair and others 1998) probably because a large pool of soil nitrogen buffers the

system. Only now are declines in soil nitrogen becoming evident in these annually burned grasslands. The

influence of burning on net primary production also

depends on interactions with other factors such as

drought. In drought years, fire can decrease net primary production (Briggs and Knapp 1995). Following

fire, higher levels of net primary production will occur

on a long-term unburned area than on annually burned

areas, primarily due to the accumulation of soil

nitrogen.

Seasonality Influences—Seasonality of fire can

have a dramatic effect on species composition and

diversity (Platt and others 1988a). Henderson and

others (1983) found that early spring burning did not

affect cool season grasses while late spring burning

reduced flowering up to 70 percent. In contrast, burning in the fall or spring increased flowering in the

warm season grasses over flowering in unburned areas; late spring burning treatments showed the greatest flowering activity. Henderson (1992a) determined

that late spring burning significantly reduced diversity compared to early spring or fall burning due to
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 93

Chapter 4: Fire in Eastern Ecosystems Wade

Figure 4-24—Productivity of all plants, especially the grasses, increases following fire, Smith Pioneer

Cemetery, Vermilion County, Indiana. In the Eastern tallgrass region, dominant grasses such as big

bluestem pictured here, could reach heights of 11 feet (3.4 m) or greater. Photo by Greg Hoch.

Figure 4-23—Periodic fires increase the diversity of tallgrass prairie plants, Smith Pioneer Cemetery,

Vermilion County, Indiana. In its eastern range, the tallgrass prairie is found in isolated “remnant” prairies,

old cemeteries, and along railroads. Many are smaller than 1 or 2 acres (0.5 ha) and surrounded by

agricultural fields.
94 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

removal of cool season grasses and several mid- and

late-season flowering forb species. Prairie violet and

blue-eyed grass, early flowering species prone to damage by spring burning, were not lost. This study also

suggested that burning in the fall might benefit prairie violet, blue-eyed grass, and sky blue aster.

Henderson (1992b) found that pasque flower was

favored by early spring fires that occurred before bud

emergence. However, this pattern was reversed in

years having high rabbit herbivory or late frosts.

Areas burned early in the spring were preferentially

grazed by rabbits. In years with late frost the thatch

from the previous year seemed to insulate the emerging buds from late frosts.

Howe (1994, 1995) determined that species richness

was higher on plots burned in July than on plots

burned in March or unburned plots in a restored

prairie. Two species, blackeyed Susan and annual

fleabane, were found only in the summer burn plots.

Summer fires dramatically increased the seedling

establishment of forbs but had no effect on seedling

establishment of grasses.

Seasonality of burning can affect productivity, especially if the fire influences soil drying and moisture

levels (Adams and others 1982; Towne and Owensby

1984). Burning in winter or early spring removes the

insulation provided by accumulated litter and allows

the upper layers of the soil to dry. Burning late in the

spring (just after the warm season grasses emerge)

removes the thatch, warms the soil, and allows sufficient light to the soil surface; but there is insufficient

time for the soil to dry before the grasses begin growing. The effects of the seasonal timing of fire on the

tallgrass prairie can be complex depending on time

since last fire, latitude, and rainfall. In many cases,

timing of rainfall (Benning and Bragg 1993) and

microclimate (James 1985) have a greater influence

on productivity than season of burning.

Management Considerations—Most studies in

the tallgrass prairie show that prescribed fires should

be conducted in the fall, early spring, or summer to

maximize plant species diversity. However, burning

during these times may lead to direct mortality of

some animals or to indirect mortality from the removal of protective cover. If weight gain of livestock is

the primary goal, the area should be burned in late

spring to maximize production of warm season grasses.

Additionally, grazing can significantly influence ecosystem responses to fire. Management should strive to

preserve heterogeneity through maintenance or simulation of natural disturbance regimes. Where feasible,

mowing or grazing (Collins and others 1998) should be

considered in conjunction with burning. Steuter (1990)

provides an excellent example of the use of a “natural”

disturbance regime for grassland conservation and

management.

In remnant prairies, heterogeneity within remnants

and regionally between remnants is an important

management objective. Larger areas of prairie should

be managed as patchworks to promote diversity within

the area (Coppedge and Shaw 1998; Steinauer and

Collins 1996). Ideally, some element of randomness

can be incorporated in the fire management program.

This can be accomplished by burning some remnants

every 1 to 2 years, some every 4 to 5 years, and some

areas every 10 years at different times of year. Attempts should be made to burn only part of an area at

a given time. This allows refugia for animal species in

the unburned thatch. Many insects, which are important pollinators to prairie plants, winter aboveground.

If the entire area is burned, whole populations of these

species can be destroyed.

The tallgrass prairie is an extremely dynamic ecosystem. Grassland plants have shown genetic changes

as a result of management in as little as 60 years

(Painter and others 1989). Many species in remnant

prairies have become extinct within the last 60 years

probably due to fragmentation and certain management practices; more species probably became extinct

prior to this (Leach and Givnish 1996). Gibson (1988)

observed that even after several years of identical

treatment two areas may still have different plant

communities, probably a result of past land use. Thus,

it is difficult or impossible to answer, “What plant

species were originally found in this general area at

the time of settlement by Euro-Americans?” Climate

variability can have an even stronger influence on the

prairie than careful management. Henderson (1992a)

stated that “The vegetation of the prairie seemed to

change more from one year of severe drought than

from 10 years of frequent early spring, late spring, or

early fall burning.”

Bay Forests

Vegetation Dynamics—The overstory was historically dominated by one or more of the following:

Atlantic white-cedar, bald cypress, pond pine, slash

pine, swamp tupelo, and Blackgum. Other overstory

species include sweetbay, red bay, loblolly bay, red

maple, and sweetgum. The almost impenetrable understory includes lyonias, titi, swamp cyrilla,

gallberries, bays, blueberries, huckleberries, wax

myrtle, sweetpepperbush, hollies, Virginia willow,

various species of greenbrier, and sphagnum moss.

The vegetation of all communities is highly correlated

to time since the last fire (Christensen and others

1988). Fire is necessary to cycle nutrients, especially

on sites with deep organic soils; the absence of fire is

more of a disturbance in pocosins than intense fire

(Christensen 1981). During drought conditions, the

underlying organic soils burn creating substantial

smoke problems, often for weeks at a time. The various
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Chapter 4: Fire in Eastern Ecosystems Wade

plant associations that make up this ecosystem and

their response to fire are discussed by Wells (1928),

Richardson (1981), Christensen and others (1988),

and Richardson and Gibbons (1993).

Management Considerations—Although the

original overstory was usually harvested, few bays are

currently managed for forest products. In the decades

following World War II, management usually consisted of draining, bedding, and planting loblolly or

slash pine, or in some cases row crops. On the remaining bays, fire periodicity determines the successional

state of the site. If bay forests burn every 2 to 5

decades, they are called shrub bogs (Christensen 1977;

Wharton and others 1976), and if burned at least once

a decade, herb bogs. If the underlying organic soils are

completely consumed, both pocosins and bays will

revert to marsh (Richardson and Gibbons 1993). Herb

bogs must be burned at least once every decade or they

will succeed to shrub bogs (Wharton and others 1976).

Where the objective is to maintain herb bogs and their

suite of showy herbaceous species, many of special

concern, they should be burned on a 1 to 3 year rotation.

As succession proceeds, the fire prescription window

narrows and becomes less defined. Wind is the major

factor determining whether fire will carry through the

early successional stages, while drought becomes more

important as the flashy herbaceous groundcover is

shaded out by the developing shrub layer. Where shrub

composition is primarily wax myrtle, gallberry,

fetterbush, and other flammable species, such as in the

pocosins of North Carolina, these communities remain

receptive to wind-driven fires.

Sand Pine

Cone serotiny is weak in the Choctawhatchee variety and strong in the Ocala variety. Both species are

thin-barked and easily killed by fire. The Ocala variety

recaptures the site with seed from the freshly opened

serotinous cones. The Choctawhatchee variety seeds

in from adjacent unburned stands. Both varieties are

prolific seeders, producing viable seed by age 5 or 6

(Brendemuehl 1990). Two fires in quick succession

(less than about 6 years apart) will eradicate this

species from a site. Sand pine is somewhat shadetolerant and rapidly establishes in the direction of the

prevailing wind. It often grows in pure, even-aged

stands.

Choctawhatchee—This variety is the principal

overstory species. Common understory associates include occasional xeric oaks (turkey, bluejack, sand,

post) and prickly pear, with a sparse groundcover of

wiregrass and bluestems (Brendemuehl 1990).

Young sand pines are sensitive to fire, so when a fire

burns through a stand, it usually kills the overstory;

however, overstory species composition does not change

because of the copious amount of seed available for

regeneration. In fact this variety is often described as

a weed species because it tends to invade other vegetation types that are downwind. Because of its thin bark,

fire is often used to eradicate it when it seeds into other

stands.

Ocala—This variety occurs on nutrient-deficient

sands as the dominant overstory species in evenaged

stands. Common understory associates include evergreen shrubs such as myrtle, sand live oak, Chapman

oak, turkey oak, rusty staggerbush, rosemary, scrub

palmetto, and saw palmetto (Christensen 1981, Outcalt

1997). Little groundcover is found on these xeric sites

although cup lichen, or deer moss as it is locally known

(same species as in the Artic), is fairly common.

These stands usually burn with an intense, winddriven fire that generally consumes all live foliage,

kills thin-barked trees, and opens cones allowing the

stand to regenerate (Cooper 1951, 1965; Cooper and

others 1959).

Management Considerations—Industrial plantations of the Choctawhatchee variety have been

established in inland Florida and about 200 miles

north of its natural range on xeric sand-hill sites in

west-central Georgia. This variety can withstand light

surface fires once it reaches pole size; prescribed fire

under mild burning conditions is sometimes used for

hazard reduction in plantations. Care must be taken,

however, because the boles of even mature trees are

quite susceptible to severe surface fires. Because of the

lack of understory competition and slow buildup of

fuels, the use of fire in the management of this variety

is generally not a high priority. Prescribed fire, however, is advocated for regenerating the Ocala variety

(Cooper 1953,1973a; Price 1973). A tight prescription

and experienced crew are necessary to confine the fire

to its intended boundaries and to manage smoke from

these stand-replacement fires because of their proximity to urban areas (Custer and Thorsen 1996). See

Walker (1967) for a broader discussion of management

recommendations for both varieties.

Table Mountain Pine

Vegetation Dynamics—Table Mountain pine is

found on xeric, typically south- to west-facing sites

(Whittaker 1956) where it is perpetuated by fire,

although it will colonize more mesic sites following fire

(Williams and Johnson 1990). This species depends

upon fire to: (1) melt the wax seal on serotinous cones

to release seeds; (2) consume a large portion of the

accumulated forest floor to create a receptive seedbed;

and (3) reduce competition for sunlight, water, and the

pulse of mineralized nutrients important on sterile

soils (Zobel 1969). Groeschl and others (1993) stated

that fire reduces overall site quality, which results in
96 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Wade Chapter 4: Fire in Eastern Ecosystems

a more favorable environment for pine than for xericsite hardwoods.

After examining fire records from 1960 to 1971 in the

southern Appalachians, Barden and Woods (1974)

found no occurrence of lightning-caused crown fires,

perhaps because fine herbaceous fuels had vanished

after decades of fire exclusion. Of 85 lightning fire

records, only two fires killed a majority of the overstory and none resulted in more than token pine

reproduction (Barden and Woods 1976). Several human-caused crown fires during this period killed most

of the overstory and resulted in much better pine

recruitment. Sutherland and others (1995) found the

same situation upon examining two centuries of evidence in a southwestern Virginia Table Mountain pine

community. Most existing stands have resulted from

fires associated with logging early in the 20th century

(Williams 1998).

Management Considerations—The long-term

effects of fire exclusion in the Appalachians are becoming more apparent (Williams 1998). Most stands are

now degraded and succeeding toward hardwood dominance (Williams 1998; Williams and Johnson 1992).

The increased incidence of bark beetle attacks in these

stressed, aging stands is accelerating this successional trend. The Southern Appalachian Mountains

Assessment (SAMAB 1996) listed Table Mountain

pine as a rare community.

In the 1990s interest increased in restoring Table

Mountain pine communities (Waldrop and Brose 1999;

Welch and Waldrop, in press). Although other natural

events such as ice storms can create canopy gaps,

reduced duff depths are the overriding requirement

for seedling establishment (Williams and Johnson

1992; Zobel 1969); for this, periodic fire is generally

responsible (although see Barden 1977 and Williams

1998). Questions regarding the necessity of crown

fires are still unresolved (Waldrop and Brose 1999;

Whittaker 1956; Zobel 1969) but may be answered by

future research.

Spruce-Fir

Spruce-fir forests are a related variant of the extensive boreal forest biome described in chapter 3. Historically, the spruce-fir type was virtually fireproof

(Harmon and others 1983; Korstian 1937), but logging

followed by fire has devastated this forest type in the

Southern Appalachians (Korstian 1937). In many cases,

the deep duff layers have been consumed down to bare

rock, and species composition has shifted to yellow

birch, pin cherry, and mountain ash. According to

Minckler (1944), rehabilitation of these sites will take

500 to 1,000 years.

Mature spruce may initially survive low-intensity

fires, but Stickel and Marco (1936) found that over half

the survivors had been attacked by fungi, insects, or

both within 3 years postburn. Thus underburning

does not appear to be an appropriate practice in

management of this forest type. See Walker (1967) for

more discussion of management recommendations.

Atlantic White-Cedar

Vegetation Dynamics—Once trees reach pole size,

copious amounts of seed are produced (about 500,000

per pound) from several thousand cones per tree. The

seed is released every fall and stored in the forest floor

where it remains viable for about 3 years. Stands tend

to be exceedingly dense. Lower branches die at an

early age but persist for several decades before being

sloughed off. Under normal (wet) conditions, crown

fires destroy the aboveground vegetation. As droughts

get progressively worse, more of the forest floor and

stored seed are consumed. Two fires in close succession, before the seed bank is replenished, will produce

herb bog, shrub bog, or bay forest depending upon the

future fire return interval.

Reestablishment after fire depends upon fire severity

and age of the stand. When the water table is high and

fires just skim off the top of the forest floor, the replacement stand can be Atlantic white-cedar, pond pine, or

sprouting hardwood trees and shrubs. Succession depends upon the amount of stored cedar seed, preburn

species composition, and the composition of unburned

adjacent stands. Fires during severe drought, which

consume much of the organic soil, result in open water

and a dense cover of leatherleaf, hardwoods capable of

sprouting, and cedar, depending on the factors just

mentioned and postfire precipitation (Little 1959).

Generally, dense stands of Atlantic white-cedar are

formed after fire when the water table is neither much

above nor below the top of the peat. A high water table

allows sprouting hardwoods to gain a competitive

advantage before cedar seeds germinate. A low water

table allows most of the cedar seed to be consumed. In

both cases, regeneration may be insufficient to produce a monotypic cedar canopy.

Management Considerations—Current fire management of Atlantic white-cedar stands is to exclude

all fire until the stands are harvested. Then fire can be

used to dispose of logging debris and prepare the site

for the next crop. See Walker (1967) for more discussion of management recommendations.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 97

Stephen F. Arno

Chapter 5:

Fire in Western Forest

Ecosystems

Understory Fire Regimes _________

Major Vegetation Types

Major forest types that are characterized by nonlethal understory fire regimes include those where ponderosa pine or Jeffrey pine has been a major component either as a fire-maintained seral type or as the

self-perpetuating climax (table 5-1). This includes

extensive areas throughout the Western United States

from northern Mexico to southern British Columbia,

Canada (Little 1971). Also, sizeable areas of open

woodlands dominated by Oregon white oak, California

black oak, blue oak, or Digger pine were characterized

by frequent understory fires largely due to deliberate

burning by Native Americans (Boyd 1986; Lewis 1973).

These occurred in relatively dry areas west of the

Cascades and Sierra Nevada from the southwest corner of British Columbia to southern California. Recent

studies suggest that large areas of the redwood forest

in coastal northern California were characterized by

frequent understory fires resulting from burning by

Native Americans (Brown and Swetnam 1994; Duncan

1992; Finney and Martin 1989; Greenlee and

Langenheim 1990).

Additionally, portions of other forest types may also

have had understory fire regimes. For example, some

areas of interior Douglas-fir near the drought-caused

lower timberline in the higher valleys of the Rocky

Mountains may have been maintained in open condition in understory fire regimes (Arno and Gruell 1983;

Arno and Hammerly 1984). Nevertheless, most of this

type is best represented by the mixed regime.

Fire Regime Characteristics

Fires were frequent, with mean intervals between 5

and 30 years in most areas (Kilgore 1987; Martin

1982) in the ponderosa pine type and at similar intervals in the redwood type; fires occurred even more

frequently in some of the oak-prairie communities. At

one extreme, fire intervals averaged 1 to 2 years in an

area of northern Arizona and no more than 10 years

throughout the Southwestern ponderosa pine type

due to abundant lightning activity (Dieterich 1980).

Conversely, near the cold limits of the ponderosa pine

type in western Montana, mean fire intervals averaged between 25 and 50 years (Arno and others 1995b).

Relatively short mean intervals occurred where many

ignitions were made by Native Americans, while longer
98 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Arno Chapter 5: Fire in Western Forest Ecosystems Table 5-1

—Occurrence and frequency of presettlement fire regime types by Forest and Range Environmental Study (FRES) ecosystems, Kuchler potential natural vegetation

classes (1975 map codes), and Society of American Foresters (SAF) cover types. Occurrence is an approximation of the proportion of a vegetation class represented

by a fire regime type. Frequency is shown as fire interval classes defined by Hardy and others (1998) followed by a range in fire intervals where data are sufficient.

The range is based on study data with extreme values disregarded. The vegetation classifications are aligned to show equivalents; however, some corresponding

Kuchler and SAF types may not be shown.

Fire regime types

Understory Mixed Stand-replacement

FRES Kuchler SAF Occura Freqb Occur Freq Occur Freq Nonfire

Coastalc Cedar-hemlock-Douglas-fir Douglas-fir-w. hemlock 230 M 2: 40-150 M 3

Douglas-fir 20 K022

Mosaic of above and Oregon Pacific Douglas-fir 229 M 2: 40-150 M 3: 200-500

oak woods K028

Red alder 221

Calif. Mixed evergreen K029 Douglas-fir-tanoak-Pacific M 1

madrone 234

Redwood 27 Redwood K006 Redwood 23 M 1: 5-25 m 2

Hemlock- Spruce-cedar-hemlock K001 Sitka spruce 223 m2 M3 m

Sitka spruce 24 W. hemlock 224 M 3

W. hemlock-Sitka spruce 225 M 3

W. redcedar-w. hemlock 228 M 3

W. hardwoods 28 Oregon oakwoods K026 Oregon white oak 233 M 1

California oakwoods K030 Blue oak-digger pine 250 M 1

Canyon live oak 249 M 1,2

California coast live oak 255 M 1,2

Coastal c fir-spruce 23 Silver fir-Douglas-fir K003 True fir-hemlock 226 M3 m

Fir-hemlock K004 Mountain hemlock 205 M 2,3 m

Inland forests

Ponderosa pine 21 W. ponderosa pine K011 Pacific ponderosa pine 245 M 1: 5-30 m 2

Pine-Douglas-fir K018 Pacific ponderosa-Douglas-fir 244 M 1: 5-30 m 2

Mixed conifer K005 Sierra Nevada mixed conifer 243 M 1: 5-30 m 2

Jeffrey pine 247 M 1: 5-30 m 2

California black oak 246 M 1: 5-30 m 2

Arizona pine K019 Interior ponderosa pine 237 M 1: 1-25 m 2

E. ponderosa K016 Interior ponderosa pine 237 M 2

Black Hills pine K017 Interior ponderosa pine 237 m 1 M2 M2

Interiorc Douglas-fir 20 Douglas-fir K012 Interior Douglas-fir 210 m 1,2 M 2: 25-100

Larch 25 Grand fir-Douglas-fir K014 W. larch 212 M 2: 25-200 M 2,3

Grand fir 213 m 2 M 2,3

W. white pine 22 Cedar-hemlock-pine K013 W. white pine 215 M 2: 50-200 M 3: 130-300

(con.)
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 99

Chapter 5: Fire in Western Forest Ecosystems Arno Table 5-1—Con. Fire regime types Understory Mixed Stand-replacement FRES Kuchler SAF Occura Freqb Occur Freq Occur Freq Nonfire Lodgepole pine 26 Lodgepole pine-subalpine K008 California mixed subalpine 256 M 2 Rocky Mountain W. spruce-fir K015 Lodgepole pine 218 M 2: 25-75 M 2,3: 100-300 lodgepole pine c 26 W. spruce-fir K015 Whitebark pine 208 M 2: 50-200 M 3: 150-300 Interior c fir-spruce 23 W. spruce-fir K015 Engelmann spruce-subalpine fir 206 M 2,3: 100-400 m Spruce-fir-Douglas-fir K020 White fir 211 M 2 M 2,3 Blue spruce 216 M 2 M 2,3 W. aspen c 28 W. spruce-fir K015 Aspen 217 m 2 M 2 aM: major, occupies >25% of vegetation class; m: minor, occupies <25% of vegetation class. bClasses are 1: <35 year, 2: 35 to 200 years, 3: >200 years. cAdded subdivision of FRES.
100 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Arno Chapter 5: Fire in Western Forest Ecosystems

intervals occurred on similar forest sites that were

more remote from aboriginal occupation. For example,

in a western Montana study the mean fire interval

from 10 heavily used areas was 9 years while the mean

interval from 10 remote areas on similar sites was 18

years (Barrett and Arno 1982).

In the Southwestern ponderosa pine type, major fire

seasons occur after snow melt (April and May) and in

mid-summer just before the monsoon rains begin, and

a secondary season exists in the fall. In most other

areas the main lightning fire season is summer;

whereas Indian burning apparently occurred to some

extent in spring, summer, and fall (Barrett and Arno

1982; Gruell 1985a, 1985b). Low-intensity surface

fires were characteristic and may have been quite

large where dry forests and adjacent grasslands were

extensive—for example, on the gentle topography of

high plateaus in northern Arizona and New Mexico. In

contrast, in rugged mountainous topography, the understory fire regime was often confined to small areas

of dry sites on south-facing slopes (Arno 1980). The

adjacent moist sites supported other, denser forest

types, which burned less often and in mixed or standreplacement fire regimes. When stand-replacement

fires burned the adjacent moist types in 1889 and 1910

in the Northern Rocky Mountains, the dry forest types

still burned primarily in a nonlethal manner. See

chapter 6 for additional information about the Southwestern ponderosa pine type.

Fuels

During periods of high fire frequency, fuels were

primarily herbaceous material and forest floor litter.

After fire suppression became effective, forest floor

duff and live fuels such as shrubs and conifer regeneration accumulated. Measurements in recent decades (Brown 1970; Brown and Bevins 1986; Sackett

1979) show that litter typically ranges from 0.6 to 1.4

tons/acre (1.3 to 3.1 t/ha) and the entire forest floor of

litter and duff averages about 12 tons/acre (27 t/ha) in

both Arizona and Northern Rocky Mountain areas.

Forest floor quantities as high as 40 tons/acre (90 t/ha)

have been measured (Harrington 1987b). During periods of frequent fire, forest floor quantities would typically range from 1 to 4 tons/acre (2.2 to 9.0 t/ha).

Herbaceous fuels range from practically none in dense

stands to as much as 0.5 tons/acre (1.1 t/ha) in open

stands on productive sites. In the Black Hills of South

Dakota, herbaceous fuel quantities in open stands of

ponderosa pine averaged 440 lb/acre (490 kg/ha), which

was six times greater than in closed stands (Uresk and

Severson 1989). Herbaceous fuel quantities are typically about 400 lb/acre (448 kg/ha).

Frequent low-intensity surface fires perpetuated

open stands of trees whose lower branches were killed

by fire. With fire suppression, accumulated fuels

support higher intensity fire including torching and

crowning behavior and longer periods of burnout. The

increased burn severity results in greater mortality to

plants and soil organisms. Managers can easily overlook the significance of forest floor fuels; the upper layer

(litter) and part of the middle (fermentation) layer

provide the highly combustible surface fuel for flaming

combustion and extreme fire behavior during severe

fire weather. The lower part of the fermentation layer

and the humus layer make up the ground fuel that

generally burns as glowing combustion. A substantial

amount of forest floor material can remain after an area

is initially burned (Sackett and Haase 1996).

Postfire Plant Communities

Ponderosa Pine/Jeffrey Pine and Ponderosa PineMixed Conifer

Pre-1900 Succession—These semiarid forest types

are widespread in the inland portions of western

North America. They include pure pine climax types,

which are abundant in plateau areas of northern

Arizona and New Mexico, central Oregon, and eastern

Washington. They also encompass many sites in the

inland mountains where pines are seral to more shadetolerant conifers—interior Douglas-fir, white fir, grand

fir, or incense-cedar. Prior to 1900 these pine communities experienced frequent fires as a result of highly

combustible leaf litter, an abundance of cured herbaceous vegetation, and a long season of favorable burning weather. Stands had an open, parklike appearance, dominated by large old, fire-resistant trees.

Shrubs, understory trees, and downed logs were sparse,

as testified to by dozens of historical photographs and

narrative accounts (Cooper 1960; Leiberg 1899;

Wickman 1992). Travelers often rode horseback or

pulled wagons for miles through these forests without

the need of cutting trails.

Undergrowth was primarily of fire-resistant grasses

and forbs that resprouted after each burn. Shrubs

were suppressed by the frequent burning coupled with

overstory competition (Gruell and others 1982). In

most stands, duff depth probably averaged only about

half an inch (Keane and others 1990a). The majority of

overstory trees survived each fire, while many of the

understory trees were killed. The most fire-resistant

species—ponderosa pine, Jeffrey pine, and western

larch—were favored. In the large areas of this type

where ponderosa pine is seral, it maintained dominance only because of the frequent fires.

In much of the pure ponderosa pine type and in the

seral pine type on dry sites, pine regeneration occurred whenever overstory trees died, thereby creating small openings. These open microsites allowed a

few seedlings to grow fast enough to gain resistance to

survive the next fire (Cooper 1960; White 1985). Thus
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 101

Chapter 5: Fire in Western Forest Ecosystems Arno

stands tended to be uneven-aged and often contained

some 400 to 600 year old trees (Arno and others 1995b,

1997). Trees were often distributed in small even-aged

clumps. When small patches of overstory were killed

by fire or bark beetles, subsequent fires consumed

these fuel concentrations, locally reducing grass competition and creating mineral soil seedbeds. This favored establishment of ponderosa pine seedlings, allowing a new age class to develop in a micromosaic

pattern within a stand (Cooper 1960). These effects

helped create an uneven-age stand structure composed of small, relatively even-aged groups (Cooper

1960; Arno and others 1995b).

Mixed-severity fire regimes were characteristic on

some of the relatively moist sites and on steep slopes

throughout the ponderosa pine type. Variable and

mixed regimes were evidently widespread in ponderosa pine communities in the Front Range in Colorado

and in the Black Hills of South Dakota and Wyoming,

and stand-replacement regimes occurred in some

situations in the Black Hills. These are described

under the section “Mixed Fire Regimes,” later in this

chapter.

Post-1900 Succession—Important changes have

occurred in these forests since 1900 due to interruption of frequent burning. Nonlethal fire has decreased

while lethal fire has increased (fig. 5-1). Reduced fire

began in the late 1800s as a result of (1) relocation of

Native Americans and disruption of their traditional

burning practices; (2) fuel removal by heavy and extensive livestock grazing; (3) disruption of fuel continuity on the landscape due to irrigation, cultivation,

and development; and (4) adoption of “fire exclusion”

as a management policy. The general result has been

development of dense conifer understories, commonly

adding 200 to 2,000 small trees per acre beneath old

growth stands or thickets of 2,000 to 10,000 small

trees per acre where the overstory was removed.

Densely overstocked conditions have resulted in slow

growth and poor vigor of most trees in a large proportion of the ponderosa pine type where adequate thinning treatments have not been applied. Stand stagnation is accompanied by a sparse representation of

nonflowering herbs and shrubs, which reflects a loss of

natural biodiversity and of forage for wildlife (Arno

and others 1995a). Growth stagnation renders even

the dominant trees highly vulnerable to mortality in

epidemics of bark beetles, defoliating insects, diseases

such as dwarf mistletoe, and various root rots (Biondi

1996; Byler and Zimmer-Grove 1991; Cochran and

Barrett 1998). For example, in the 1980s about a

million acres of ponderosa pine-fir forests in the Blue

Mountains of eastern Oregon suffered heavy mortality from the above agents as a result of overstocking

and growth stagnation related to fire exclusion (Mutch

and others 1993).

Figure 5-1—Change in fire severity between historic (pre1900) and current conditions for Forest Service and Bureau of

Land Management administered forested potential vegetation

groups in the Interior Columbia Basin and portions of the

Klamath and Great Basins (Quigley and others 1996).

In many stands, duff mounds 6 to 24 inches deep

have accumulated under old growth trees, and burning these mounds may girdle and kill the trees

(fig. 5-2). Stands of small, slow-growing pines have

also colonized former grasslands (Gruell 1983). Overstory trees have been removed in more than a century

of logging, primarily partial cutting, and this has

aided development of thickets of small trees. On sites

where ponderosa pine is seral, there has been a compositional shift to the shade-tolerant species. These

successional changes have resulted in a buildup of

understory or ladder fuels that now allow wildfires to

burn as stand-replacing crown fires.

A combination of heavy forest floor fuels and dense

sapling thickets acting as ladder fuels, coupled with

the normally dry climate and frequent lightning- and

human-caused ignitions, has resulted in a dramatic

increase of severe wildfires in the ponderosa pine type

in recent decades (Arno 1996; Williams 1995). For

example, approximately 1 million acres (405,000 ha),

largely in this type, burned in severe wildfires in

central in Idaho between 1986 and 1996 (Barbouletos

and others 1998).

Management Considerations—Fires of the past

were important to the evolution of ponderosa and

Jeffrey pine forests (Keane and others 1990a; Mutch

1970; Weaver 1967). Today, prescribed fire and wildland fire are the obvious and most feasible substitutes

for filling the ecological role of historic fires in restoring these wildland ecosystems (fig. 5-3). Many alternatives exist for employing prescribed fire and fuels

management treatments to improve forest health and

reduce excessive ladder fuels in ponderosa pine and

pine-mixed conifer types (Arno 1988). Different prescriptions can be chosen to enhance critical resources

or values (Fiedler and others 1996), such as maintenance of old growth in a natural area, encouraging

Historic

Current

60

50

40

30

20

10

0

Lethal Mixed Nonlethal

Columbia River Basin Study Fire Regimes

Percent of area
102 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Arno Chapter 5: Fire in Western Forest Ecosystems

Figure 5-2—A prescribed burn during

October in central Idaho removed duff as

thick as the distance between the hands,

in this case with little damage to large

ponderosa pine trees.

Figure 5-3—Prescribed fire during May

is being used to reduce fuel loadings

after a retention shelterwood harvest in a

ponderosa pine/Douglas-fir forest.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 103

Chapter 5: Fire in Western Forest Ecosystems Arno

browse and cover on a big game winter range,

maintaining forest structure favored by neotropical

migrant bird species, northern goshawk, flammulated

owl, and other species of concern; protecting homes

and watersheds from severe wildfire while maintaining visual screening; or providing a continuous supply

of forest products without disruption of esthetics.

However, most long-term management goals for wildlands in these forest types can be enhanced by some

form of prescribed fire and fuels management.

Many contemporary stands have such an altered

stand structure and composition, including a buildup

of understory fuels, that it is difficult to restore forest

health with prescribed fire alone. Silvicultural cutting

and pile-burning or removal of excess small trees may

be necessary to allow successful application of prescribed fire and to return to more open structures

dominated by vigorous trees of seral species (Arno and

others 1995a). Failed attempts to restore more natural

stand conditions with prescribed burning alone may

result from inappropriate use of fire as a selective

thinning tool in dense fire-excluded stands, or from

burning too little or too much of the accumulated forest

floor fuels. A better approach to the latter problem

may be to apply two or even three burns to incrementally reduce loadings (Harrington and Sackett 1990).

Once a semblance of the desired stand and fuel conditions have been established, stands can thereafter be

maintained more routinely with periodic burning or a

combination of cutting and fire treatments. Prescribed

fire can be used in wildernesses and natural areas to

maintain natural processes.

The advantages of using fire and improvement cuttings to restore and maintain seral, fire-resistant

species include: (1) resistance to insect and disease

epidemics and severe wildfire; (2) providing continual

forest cover for esthetics and wildlife habitat;

(3) frequent harvests for timber products; (4) stimulation of forage species; and (5) moderate site disturbances that allow for tree regeneration (Mutch and

others 1993). Frequent prescribed fires will not produce heavy screening or hiding cover, which is not

sustainable over large areas in these forests. But such

fires can help maintain moderate cover and screening

indefinitely (Martin and others 1988). Management of

a large proportion of the forest in open conditions can

help ensure protection of strategically located patches

of heavier cover (Camp and others 1996). The frequent

disturbance cycles can also produce and maintain

large old trees characteristic of pre-1900 forests and of

high value for wildlife habitat, esthetics, and selective

harvesting for lumber. In other words, the management approach of using a modified selection system

and periodic burning can be used to maintain remnant

old growth stands and to create future old growth

(Fiedler 1996; Fiedler and Cully 1995).

Prescribed fire and wildland fire must be introduced

cautiously in stands where leave trees have poor vigor

or where tree roots are located in a deep duff layer

(Harrington and Sackett 1992). Burning thick forest

floor fuel layers can mortally injure roots and boles of

old pines that in past centuries survived many fires

(Sackett and Haase 1996). Exceptionally poor tree

vigor is reflected by growth stagnation over two or

more decades in the dominant trees in both second

growth and old growth stands. This widespread stand

stagnation has resulted from basal area stocking levels often two or more times those of historic conditions

(Arno and others 1995b; Biondi 1996; Cochran and

Barrett 1998; Habeck 1994). Fire can be highly stressful to trees that are suffering from growth stagnation,

and may, for instance, allow bark beetles to inflict

major damage to the weakened trees. If vigor of desired leave trees is poor, it may be wise to thin mechanically and allow the leave trees to recover or release

somewhat before applying fire, in perhaps 2 years

(Fiedler and others 1996).

Options for alleviating the condition of poor vigor or

deep duff are not ideal. Managers can simply accept a

20 to 50 percent loss of old growth in a single fuelreduction burn as being a cost of decades of fuel

buildup. Most old growth stands are now more heavily

stocked than in pre-1900 times, and some of the trees

would probably have been killed by natural fires had

suppression not intervened. Alternatively, fuels could

be manually removed or raked away and dispersed

from around the boles of old growth trees. The use of

a burn prescription that reliably removes a portion of

the fuel mound around a big tree has not been found.

If glowing combustion is able to establish in the deeper

mounds, it can continue even with high moisture

content of that material and result in total consumption and prolonged heat release.

Improvement cutting, thinning, and understory cutting with whole-tree removal or pile burning may be

necessary to achieve open stocking levels that will

sustain vigorous tree growth and to reduce ladder

fuels (Fiedler and others 1996) (fig. 5-4). Harvesting

and thinning should be designed to retain the most

vigorous trees. If stems are tall and slender, as in

dense second growth stands, it may be necessary to

leave clumps of three to five trees for mutual protection against breakage by wet snow and windstorms.

The restoration cutting process may require thinning

in two to three steps over 15 to 20 years. Spot planting

of the desired seral tree species in open burned

microsites can be used when shade-tolerant species

have taken over (Fiedler and others 1996).

Prescribed fire can be used in a variety of seasons to

meet management objectives (Harrington 1991;

Kalabokidis and Wakimoto 1992; Kauffman and

Martin 1990; Kilgore and Curtis 1987; Martin and Dell
104 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Arno Chapter 5: Fire in Western Forest Ecosystems

Figure 5-4—(A) An unwanted ponderosa pine-fir stand that resulted from partial cutting of large pines followed

by development of fir thickets. (B) Similar stand after the first restoration treatment consisting of thinning to

favor remaining pines and jackpot burning.

A.

B.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 105

Chapter 5: Fire in Western Forest Ecosystems Arno

1978; Weatherspoon 1990). This includes burning in

late winter following snowmelt, when understory Douglas-fir or true firs may burn readily and are easily

killed because of low foliar moisture content; spring;

cloudy-humid days in summer; or autumn. Trees are

less susceptible to fire damage when entering dormancy in late summer and fall. Potential for fire

damage can be reduced by thinning and whole-tree

removal, burning when slash is still green, successive

burns starting with damp fuels, and raking duff mounds

away from boles of old growth trees. Fire intensity can

be reduced by burning at dawn, late evening, or at

night. Conversely, to enhance burning in stands where

surface fuels are inadequate, such as in fir thickets,

thinning can be used to create sufficient loadings of

cured slash. Alternatively, waiting for grass fuels to

cure in the fall may be effective. If enclaves of dense

growth are desired, such as for wildlife habitat, these

can be protected from wildfire if buffered within a

matrix of treated stands that have light fuel loadings.

During the past decade tens of thousands of acres of

the ponderosa pine type have been treated with prescribed fire in central Oregon, northwestern Montana,

and parts of Arizona and New Mexico (Kilgore and

Curtis 1987; Losensky 1989; Simmerman and Fischer

1990). If these applications can be greatly expanded

they could correct some of the severe forest health and

wildfire problems that exist in the Inland West (Everett

1994; McLean 1992; Mutch and others 1993).

Redwood

Pre-1940 Succession—Forests where redwood was

the most abundant tree covered about 1 million acres

in a narrow strip along the coastal fog belt from the

extreme southwestern corner of Oregon to Monterey

County, California (Roy 1980). It has long been recognized that relatively frequent understory fires were

historically a common feature of the redwood forest

and that fires seldom killed many of the large redwood

trees (Fritz 1931). Redwood is even more fire resistant

than coast Douglas-fir, its primary associate, and

unlike most conifers in this region, redwood sprouts

vigorously from dormant buds when the crown is

heavily scorched. Pacific madrone and tanoak, the

most abundant hardwoods in the redwood type, are

fire susceptible. They resprout when top-killed by fire,

but burning probably taxed them physiologically when

they were growing beneath an overstory.

Recent advances in dating fire scars on redwood

stumps have shown that presettlement (pre-1850) fire

intervals averaged between about 5 and 25 years,

shorter than previously estimated (Brown and

Swetnam 1994; Finney and Martin 1989, 1992). The

pattern of frequent fires on fire-scarred tree stumps

was traced back to about 1300 A.D. in one study area

(Finney and Martin 1989) and to about 800 A.D. in

another (Fritz 1931). There is a convergence of evidence from historical journals and archeological and

anthropological discoveries that Indian burning was

primarily responsible for the frequent fires prior to the

mid-1800s (Duncan 1992; Greenlee and Langenheim

1990; Lewis 1973; Stuart 1987). Coincidentally, the

redwood forest is located close to interior coastal valleys where detailed journal accounts of Indian burning

have been assembled (Boyd 1986, 1999). In the presettlement period, before logging occurred, the frequent

fires probably kept understories open and reduced

forest floor fuels. Overstories of the tall trees with

branch-free lower boles were relatively dense because

of high site productivity and redwood’s fire resistance.

During the settlement period, in the later 1800s and

early 1900s, Anglo-Americans conducted logging that

created large quantities of slash. These inhabitants

continued a pattern of frequent burning in conjunction

with logging and for grazing or other purposes. They

also allowed accidental fires to spread.

Post-1940 Succession—Starting in the 1930s and

1940s, land use patterns changed with the implementation of the California Forest Practices Act and with

more vigorous fire suppression (Greenlee and

Langenheim 1990; Stuart 1987). The number of escaped fires from agricultural or logging activities was

reduced greatly; but the effects of this reduction of

fires have received little evaluation. It is generally

assumed that the long-lived, shade-tolerant redwoods

will continue to dominate regardless of removal of

underburning. Successional relationships in redwood

forest communities (Zinke 1977) suggest that with

suppression of underburning, shade-tolerant shrubs,

understory hardwoods, and western hemlock will increase as will forest fuels in general. In the extensive

parks and preserves where redwood forests are protected from logging, continued exclusion of fire will

probably allow fuels to accumulate to levels that support higher intensity wildfires, which can escape suppression. Thus, the removal of underburning may ultimately result in a mixed fire regime where wildfires kill

significant portions of the overstory and may induce soil

damage on the steep slopes associated with a large

portion of the redwood type (Agee 1993; Atzet and

Wheeler 1982). Additionally, removal of underburning

will probably reduce the abundance of early seral species (including chaparral species) and will allow trees to

colonize the small openings and prairies associated

with the redwood forest belt (Finney and Martin 1992;

Greenlee and Langenheim 1990; Zinke 1977).

Oregon Oak Woodlands

Pre-1850 Succession—Open woodlands dominated

by Oregon white oak once occupied the driest climatic

areas throughout the Puget Sound-Willamette Valley
106 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Arno Chapter 5: Fire in Western Forest Ecosystems

lowlands and southward in dry valleys behind the

coastal mountains to the California border. (Farther

south this woodland expands to cover large areas of

dry hilly terrain and it is dominated by several species

of oak in addition to other hardwood trees; those

communities are described under “Western Oaks” in

chapter 6.) Oregon white oak dominated in open woodlands and savannas associated with valley grasslands, and isolated small prairies that were surrounded

by the extensive coast Douglas-fir forest. Oak woodlands also were associated with droughty sites such as

bedrock with shallow soils on the southeast coast of

Vancouver Island and in the San Juan archipelago, in

the rain shadow of the Olympic Mountains.

Ample evidence from journal accounts and archeological sources shows the extensive oak woodlands of

the Willamette and other major valleys persisted as a

“fire climax” maintained by frequent aboriginal burning (Agee 1993; Boyd 1986, 1999; Habeck 1961). Prior

to the influx of Euro-American settlers in the mid1840s, the Kalapuyan Indians and other tribes typically set fire to large areas of the Oregon oak woodlands to aid hunting, food plant harvest, and for other

purposes (Boyd 1986). Firing was commonly done in

September, and many areas were burned at short

intervals, perhaps annually in some areas. Similar

patterns of frequent burning to maintain valley grasslands, isolated prairies, and open oak woodlands are

described from northwestern Washington southward

to central California (Lewis 1973; Boyd 1986, 1999),

but these practices are well documented only in the

Willamette Valley of northwestern Oregon.

Most of these fires must have been characterized by

short duration flaming as they quickly consumed

grass and litter that had accumulated since the previous burn. The thick-barked oaks survived, but regeneration of all shrubs and trees would have been heavily

thinned by frequent burning. Grass flourished. The

effect of frequent burning on oak regeneration from

sprouts and seedlings is not known, but this species

was more successful in establishment than Douglasfir or other competitors under a regime of frequent

burning. Results of experimental burning suggest

that heavily burned microsites are favorable for oak

seedling establishment (Agee 1993). Also, about half

of the 1 to 9 year old seedlings burned in a prescribed

fire resprouted and were alive 3 years later.

Post-1850 Succession and Management Considerations—In former oak savannas, fire exclusion

has led to an increased density of shrubs and oaks,

transforming them into woodlands (Agee 1993). In

former oak woodlands, shrubs, Douglas-fir, and other

tree species are replacing oaks. Livestock grazing and

fire exclusion have been major factors in the successional change that has occurred in Oregon oak woodlands since Euro-American settlement. Logging,

clearing, and firewood harvest also have changed

many woodlands. Additionally, large areas of oak

woodlands have been displaced by agricultural, commercial, industrial, and residential development.

Today, 150 years after Euro-American settlement

began, only a general idea of presettlement conditions

can be hypothesized for most stands. Much of the

remaining undeveloped area in this type will be replaced successionally by the year 2010 unless prescribed fire and other restoration treatments are conducted (Agee 1993). To complicate restoration of oak

communities, a variety of introduced herbaceous plants

and the shrub Scotch broom are now established.

Introduced plants can increase as a result of some

treatments. Nevertheless, some strategies of burning

carefully coordinated with cutting, seeding of native

plants, and other treatments hold some promise (Agee

1993; Sugihara and Reed 1987). Restoration of structural characteristics of the oak communities is an

attainable goal (see Agee 1993).

Mixed Fire Regimes _____________

Major Vegetation Types

Major forest types include coast Douglas-fir, redwood, California red fir, interior Douglas-fir, western

larch, lodgepole pine, whitebark pine, and ponderosa

pine types east of the continental divide in Montana,

South Dakota, Wyoming, and Colorado. (There is

evidence that a mixed regime may have been important for perpetuation of giant sequoia groves in the

Sierra Nevada; Stephenson and others 1991; Swetnam

1993.) These forests are widespread in the upper

Great Plains, the mountains of the Western United

States, and in the northwest coastal regions, extending into southern British Columbia (fig. 1-2). In large

portions of their distributions, some of these forest

types are also characterized by stand-replacement fire

regimes in large portions of their distributions, evidently as a result of differences in climate or topography. An example of this contrast is the larch-lodgepole

pine forest in Glacier National Park, Montana, where

the northern portion was under a mixed fire regime

while the southern portion had a stand-replacement

fire regime (Barrett and others 1991). A similar contrast is provided by lodgepole pine types in southwestern Alberta as described by Tande (1979) and Hawkes

(1980).

Fire Regime Characteristics

Fires were variable in frequency and severity, and

perhaps because these situations are difficult to characterize in simple terms, they have been largely overlooked in previous fire regime classifications (Brown

1995). Mean fire intervals were generally longer than
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Chapter 5: Fire in Western Forest Ecosystems Arno

those of understory fire regimes and shorter than

those in stand-replacement fire regimes. However,

some individual fire intervals were short (<30 years),

while the maximum intervals could be quite long

(>100 years). Our mixed category covers the spectrum

of fire regimes between nonlethal regimes and those

where stand-replacement fires were typical.

As described in chapter 1, mixed fire regimes may

consist of a combination of understory and stand-replacement fires. Examples are the seral ponderosa

pine-western larch forests in western Montana that

burned in replacement fires at long intervals (150 to

400+ years) with nonlethal underburns at short intervals (20 to 30 year averages) in between (Arno and

others 1995b, 1997).

Mixed severity fire regimes may also be characterized by fires that killed a large proportion of firesusceptible species in the overstory (such as western

hemlock, subalpine fir), but spared many of the fireresistant trees (such as redwood, Douglas-fir, larch,

ponderosa pine). These fires tended to burn in a finegrained pattern of different severities, including

patches where most of the moderately susceptible

trees (such as California red fir, white fir, lodgepole

pine) survived. Any given location within a mixed fire

regime could experience some stand-replacement fires

and some nonlethal fires along with a number of fires

that burned at mixed severities.

Evidence of mixed severity fire in ponderosa pine is

suggested in two landscape photocomparisons of central Montana from the 1880s to 1980 (Gruell 1983,

plates 32 and 43). Relatively long fire intervals and

mixed burning also occurred in ponderosa pine in the

Colorado Front Range (Laven and others 1980) and in

the Black Hills of South Dakota (Brown and Sieg 1996;

Shinneman and Baker 1997). The mixed fire regime is

reported to have been common in mixed-conifer forests of northern Idaho (Zack and Morgan 1994; Smith

and Fischer 1997) and western Montana (Arno 1980).

Pre-1900 fires often covered large areas. The uneven

burning pattern in mixed fire regimes was probably

enhanced by mosaic patterns of stand structure and

fuels resulting from previous mixed burning. Thus,

past burn mosaics tended to increase the probability

that subsequent fires would also burn in a mixed

pattern. Complex mountainous topography also contributed to variable fuels and burning conditions,

which favored nonuniform fire behavior.

Fuels

During the presettlement period fuels were probably quite variable spatially and temporally. At a

given time, some segments of the vegetative mosaic

would be patches of postfire regeneration that had

arisen where the last fire killed much of the overstory.

Fuel loadings in these patches might increase

dramatically as dead trees and limbs fell into a developing patch of saplings. If these regenerated patches

burned again, the resulting “double burn” might be an

area cleared of most living and dead fuel and thereafter more likely to support nonlethal underburning in

the next fire. In the presettlement period a given fire

could burn day and night for 2 to 3 months under a

great variety of weather conditions in a hodgepodge of

different vegetation and fuel structures. Reburning

might occur later in the same fire event. This could

result in an intricate pattern of different fire effects on

the landscape. Such complex burning patterns are

difficult to imagine in modern stands where 60 to 100

years of fire exclusion have allowed most of the landscape mosaic to age and advance successionally.

Patches of late-successional forests with accumulations of dead and living fuels have coalesced, increasing the likelihood of fires of unusual size and severity

(Barrett and others 1991).

The ranges in fuel loadings observed in vegetation

types characterized by mixed and stand-replacement

fire regimes exhibit considerable overlap. Fuel loadings vary widely within broad vegetation classes due

to stand history and site productivity. Dead woody

fuels accumulate on the ground often in a haphazard

manner due to irregular occurrence of natural mortality factors such as fire, insects, disease, tree suppression, and wind and snow damage. However, the greatest fuel loadings tend to occur on the most productive

sites, which are predominately stand-replacement fire

regimes. For example, in the Northern Rocky Mountains downed woody fuel loadings ranged from an

average of about 10 tons/acre (22 t/ha) on low productivity (30 cu ft/acre/yr) sites to about 30 tons/acre

(67 t/ha) on high productivity (90 cu ft/acre/yr) sites

(Brown and See 1981).

Average fuel loadings determined from extensive

forest surveys in the Northern Rocky Mountain National Forests (Brown and Bevins 1986; Brown and

See 1981) indicate that quantities of duff and downed

woody material differ between mixed and stand-replacement fire regimes (table 5-2). For example, total

woody fuel loadings for the spruce-fir and cedar-hemlock types, which are stand-replacement regimes, averaged about 30 tons/acre. It averaged about 17 tons/

acre in Douglas-fir and lodgepole pine types, which are

characterized by mixed and stand-replacement regimes. Variability within stands and cover types was

considerable. Fuel loading distributions for forest floor

and all classes of downed woody material were highly

skewed, with long right-handed tails. The ratios of

medians-to-means averaged close to 0.6 (Brown and

Bevins 1986). These statistics indicated that fuels

were not uniformly distributed but concentrated in

scattered patches.

Downed woody fuels greater than a 1 inch diameter

are considered coarse woody debris, which has
108 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Arno Chapter 5: Fire in Western Forest Ecosystems

important implications for managing biodiversity

and nutrient potentials. Based on a survey of coarse

woody debris knowledge (Harmon and others 1986)

quantities of downed woody material are considerably

higher in Cascade Mountains and coastal forests than

in the Northern Rocky Mountains. Loadings of coarse

woody debris ranged from 60 to 300 tons/acre (130 to

670 t/ha) in Sitka spruce and western hemlock of

coastal British Columbia and from 60 to 240 tons/acre

(130 to 540 t/ha) in the Douglas-fir type of the Cascade

Mountains.

Postfire Plant Communities

Coast Douglas-fir and Douglas-fir/Hardwoods

Pre-1900 Succession—These humid maritime forests are extensive at low and middle elevations west of

the crest of the Cascades and British Columbia Coast

Range from northern California to southern British

Columbia. Fire history studies (see Agee 1993) indicate that mixed fire regimes were common in the

Douglas-fir type south from west-central Oregon; drier

areas of the Douglas-fir type farther north; and in the

Douglas-fir/hardwood types of northwestern California and southwestern Oregon (Wills and Stuart 1994)

(fig. 5-5). Conversely, the cooler, wetter, more northerly portions of the coast Douglas-fir type tended to be

associated with stand-replacing fire regimes. Mixed

fire regimes were probably also associated with some

areas of the redwood type, perhaps on steep terrain and

in areas relatively remote from Native American use.

Mixed fire regimes favored development of stands

dominated by large old, fire-resistant trees, such as

coast Douglas-fir—and, where present, redwood. These

regimes were characterized by patchy nonuniform

burning (Morrison and Swanson 1990; Teensma

1987; Wills and Stuart 1994). Overall, most of the

Table 5-2—Average loadings of forest floor and downed woody fuel by diameter class for

randomly located sample points in Northern Rocky Mountain forest types:

Douglas-fir (DF), Lodgepole Pine (LP), Larch/grand fir (L/GF), Spruce/fir (S/F),

and Cedar/hemlock (C/H).

USDA Forest Service forest survey cover types

Fuel DF LP L/GF S/F C/H

 - - - - - - - - - - - - - - - - - - - - tons/acre - - - - - - - - - - - - - - - - - - - -

Litter 0.56 0.35 0.66 0.52 0.85

Duff 13.0 16.0 21.8 25.4 25.4

Woody

0 to 1⁄4 inch .18 .22 .22 .12 .30 1⁄4 to 1 inch 1.0 1.0 1.3 1.0 1.3

1 to 3 inches 1.8 2.1 2.3 1.9 2.7

3+ inches 12.9 14.4 17.7 23.8 29.4

Total woody 15.9 17.7 21.5 26.8 33.7

fire-susceptible trees (notably western hemlock) were

killed while many of the resistant trees survived.

Occasional nonlethal understory fires and stand-replacement fires also occurred. Effects of burning included removing understory conifers and ladder fuels,

preventing successional replacement by shade-tolerant trees, and creating openings of all sizes that

allowed regeneration of seral undergrowth (including

berry-producing shrubs), hardwood trees (such as red

alder, bigleaf maple, bitter cherry), Douglas-fir and a

few other seral conifers (such as western white pine

and shore pine).

In northwestern California and southwestern Oregon, dry Douglas-fir/hardwood types burned rather

frequently and supported a variety of seral shrubs,

including Ceanothus spp. as well as hardwood trees

that typically resprout after fire—including tan oak,

madrone, canyon live oak, and chinquapin (Agee 1993;

Husari and Hawk 1993; Wills and Stuart 1994). A

remarkable feature of the mixed fire regimes in dry

sites was the prevalence of large (>6 feet in diameter),

old Douglas-fir that had survived numerous fires.

Although forests associated with this fire regime type

are extensive and important for a variety of ecological

values, little is known about landscape patterns and

successional patterns associated with the presettlement fires. The existence of this mixed fire regime as

a widespread type was documented only recently (Agee

1993; Means 1980; Morrison and Swanson 1990).

Post-1900 Succession—Better knowledge of the

ecological role and importance of fire in these mixed

fire regimes is needed (Kauffman 1990). Although

major ecological changes in these forests due to

clearcutting and short-rotation forest management

have been recognized, other significant changes have

also occurred as a result of interruption of the mixed

fire regime (Agee 1990). Effects of fire exclusion are
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Chapter 5: Fire in Western Forest Ecosystems Arno

conspicuous in dry site coastal Douglas-fir communities

which burned rather frequently, for example, those in

southwestern Oregon (Agee 1991; Means 1982) and in

the Puget Sound lowlands (Agee and Dunwiddie 1984;

Boyd 1986). Such effects include loss of early seral

shrub species, advanced successional development, increased stand density, and increased mortality. In older

logged areas it is common to find dense second growth

Douglas-fir and hemlock that are stagnating and succumbing to root rot, whereas the original stands consisted of large Douglas-fir in moderately open stands

that had survived for several centuries in the presence

of repeated fires.

Management Considerations—These forest types

have high value for recreation, esthetics, endangered

species habitat, and timber production. It is generally

recognized that fire had an important role in creating

these forests. A century ago Gifford Pinchot, Forester

for the U.S. Department of Agriculture, argued for

research into the ecological role of fire in these forests,

but his call went largely unheeded (Pinchot 1899). Most

fire research in these vegetation types has been related

to burning of clearcuts (Kauffman 1990), and provides

knowledge that is of limited value for modern ecological

application. Other ecological aspects of these forests

have been studied in detail at the H. J. Andrews

Experimental Forest in west-central Oregon (Franklin

and Others 1981; Teensma 1987) and in some redwood

ecosystems (Zinke 1977). Major forest preservation

efforts have focused on these forests, but ironically have

generally accepted the continued exclusion of fire as

consistent with maintenance of ecological values.

California Red Fir and Sierra/Cascade Lodgepole Pine

Pre-1900 Succession—These are high elevation

forests characteristic of the southern Cascades, Klamath Mountains (in northwestern California), and the

Sierra Nevada. Fire history studies conducted thus far

suggest a variable or mixed fire regime.

The following summary of fire history in California

red fir forests is paraphrased from Husari and Hawk

(1993) with additional information from Taylor (1993).

Estimates of natural fire frequency in California red fir

range from 21 to 65 years (Taylor 1993; Taylor and

Halpern 1991). Analysis of lightning fire occurrence in

Yosemite National Park shows that, on a per acre basis,

the California red fir forest type there experiences more

lightning ignitions than any other vegetation type, but

the fires are mostly small (van Wagtendonk 1986). This

is probably attributable to lower overall productivity on

these sites, a compact duff layer, a short fire season, and

generally cool, moist conditions. The California red fir

forest shows structural evidence of a combination of

large intense fires, small fires and fires of mixed

Figure 5-5—Coastal Douglas-fir in a dry area of northwestern Washington, with scars from two different

historic fires.
110 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Arno Chapter 5: Fire in Western Forest Ecosystems

severity (Taylor 1993). This variety of fire characteristics has led to landscape diversity in California red fir

forests (Agee 1989).

Agee (1993) and Chappell and Agee (1996) describe

variable and mixed fire regimes in lodgepole pine forests of southern Oregon. Average fire intervals are

estimated to be 60 to 80 years. Mountain pine beetle

epidemics can result in an accumulation of heavy fuels,

which in turn supports stand-replacement fire. Stands

with moderate quantities of older downed logs and open

space between tree canopies often experience smoldering fires (“cigarette burns”) that spread primarily

through decayed logs on the forest floor. Southward, in

the California red fir-lodgepole pine-mixed conifer subalpine forest of California, small patchy fires are the

norm (Parsons 1980). This is due in part to an abundance of broken topography, rock, bare mineral soil,

and sparse fuels that hamper fire spread.

Post-1900 Succession—Suppression of the mixed,

patchy fires in these high-elevation forests may eventually result in a landscape mosaic consisting largely

of contiguous old stands with comparatively heavy

loadings of dead trees (standing and fallen) and canopy

fuels. This is probably the basis for Husari and Hawk’s

(1993) projection that in the future these forests will

be characterized by infrequent high-intensity, standreplacing fires.

Management Considerations—Intensive tree harvesting can be used to break up continuous heavy fuel

loadings that might result in large, severe fires. Laacke

and Fiske (1983) state that most silvicultural treatments have been designed to produce even-aged stands.

Clearcutting offers the opportunity to control the spread

of dwarf mistletoe and root disease, which are often a

concern in this type. Recent interest in perpetual retention of some tree cover on high-elevation forest sites and

in designing treatments consistent with historical disturbances will probably encourage attempts at unevenaged silviculture or retention shelterwood. California

red fir is a shade-tolerant species, which conceptually

makes it appropriate to consider for uneven-aged silvicultural systems. Concern about root disease and dwarf

mistletoe has worked against consideration of unevenaged systems, but use of underburning might serve

some function in controlling these pathogens (Koonce

and Roth 1980; Petersen and Mohr 1985). Fire history

and experience burning in white fir types (Petersen and

Mohr 1985; Weatherspoon 1990) suggest that understory burning might be useful in some California red fir/

lodgepole forests.

Interior Douglas-fir, Larch, Rocky Mountain

Lodgepole Pine

Pre-1900 Succession—A broad range of mid-elevation mountain forests dominated by interior

Douglas-fir, western larch, or Rocky Mountain

lodgepole pine were characterized by mixed fire regimes. These occurred from central British Columbia

(Strang and Parminter 1980) and Jasper National

Park, Alberta (Tande 1979), southward at least to

western Wyoming (Arno 1981; Loope and Gruell 1973).

They were abundant and diverse in western and

central Montana (Arno 1980; Barrett and others 1991,

Arno and Gruell 1983). Mixed fire regimes allowed an

open overstory of mature Douglas-fir and larch to

survive many fires. Small trees and associated less

fire-resistant species were heavily thinned by moderate-intensity burning. Additionally, some nonlethal

underburns occurred in lodgepole pine stands having

light fuels. Occasional stand-replacing fires were also

part of the mixture making up this fire regime.

Effects of these variable fires often included maintaining a fine grained forest community mosaic on

much of the landscape—as is illustrated by maps in

Tande (1979), Barrett and others (1991), and Arno and

others (1993) (fig. 5-6). Elements of this mosaic were

small stands dominated by various age structures of

seral coniferous species and seral hardwoods such as

Scouler willow and aspen. Some stands experienced

nonlethal underburns that maintained open understories by killing saplings and fire sensitive species.

Others experienced patchy fire mortality that gave

rise to patchy tree regeneration including seral species.

Occasional large stand-replacement fires may have

Figure 5-6—This mosaic of age-classes resulted from a mixedseverity fire regime in Glacier National Park, Montana. Dates

indicate fire years that resulted in establishment of seral western larch and lodgepole pine age classes. In addition, some

low-intensity underburns were detected in fire scars, but they

did not thin the stand enough to allow new age classes to

become established (from Barrett and others 1991).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 111

Chapter 5: Fire in Western Forest Ecosystems Arno

reduced the spatial diversity, but the varying distribution of seed sources and sprouting shrubs in the

preburn mosaic probably enhanced variability in

postburn vegetation. A fire effect near the lower forest

boundary was to maintain seral grasslands, shrublands, and aspen groves by periodically removing most

of the invading young Douglas-fir or lodgepole pine

(Arno and Gruell 1986; Patten 1963; Strang and

Parminter 1980).

Post-1900 Succession—With a reduction in fire

activity due to livestock grazing (removing fine fuels)

and fire exclusion policies, young conifer stands have

invaded former grasslands within or below the forest

zone. The trees are densely stocked and subject to

extreme drought stress. They often have poor vigor

and are susceptible to western spruce budworm or

other insect or disease attacks and to stand replacing

fires. Productivity of seral herbs, shrubs, and aspen

declines dramatically in the continuing absence of fire.

Stands within the forest zone may have undergone

significant changes in recent decades. As a result of

fire exclusion, the trees in most stands within the

landscape mosaic have become older, and often have a

buildup of down woody or ladder fuels. Recent wildfires have burned as larger stand-replacement fires

than those detected in fire history studies (Arno and

others 1993; Barrett and others 1991).

Management Considerations—Fire exclusion

may move these communities toward a long-interval

stand-replacement fire regime. This could decrease

vegetation diversity on the landscape and may reduce

values for wildlife habitat, watershed protection, and

esthetics. Numerous alternatives exist for simulating

more natural landscape structure and disturbance

regimes using silvicultural cuttings and prescribed

fire (Arno and Fischer 1995; Arno and others 2000;

Brown 1989; Gruell and others 1986).

Whitebark Pine

Pre-1900 Succession—The whitebark pine type

occurs near the highest elevations of forest growth

from southern British Columbia south to western

Wyoming in the Rocky Mountains and along the Cascades to northern California. In the drier mountain

ranges and on rocky sites with sparse fuel, whitebark

pine is characterized by a mixed fire regime, whereas

in moist and more productive areas it is perpetuated

by a stand-replacement fire regime. The mixed fire

regime has been noted in whitebark pine stands in the

Selway-Bitterroot Wilderness and nearby areas of

central Idaho and western Montana (Barrett and Arno

1991; Murray 1996). Prior to 1900 these forests experienced a range of fire intensities from nonlethal

underburns to large (but usually patchy) stand-replacement fires. Rugged terrain, including extensive

rocklands and cool-moist north slopes hampered the

spread of fires and resulted in a variable burn pattern.

Whitebark pine is a seral species on all but the

harshest sites and is replaced successionally by subalpine fir, Engelmann spruce, or mountain hemlock

(fig. 5-7). These species were readily killed in lowintensity fires, whereas whitebark pine often survived. Epidemics of mountain pine beetle periodically

killed many of the older whitebark pines. Fuels created by beetle kills and successional ladder fuels

contributed to patchy torching or stand-replacement

burning. Underburns had a thinning effect that removed many of the competing fir, while more intense

fires created open areas favorable for whitebark pine

regeneration. This species’ seeds are harvested and

cached often in burned areas by Clarks Nutcrackers

(Nucifraga columbiana). Many seed caches are not

retrieved, so pines regenerate (Tomback 1982).

Whitebark pine is hardier than its competitors and

functions as a pioneer species on high-elevation burns.

Post-1900 Succession—Fire exclusion may have

been particularly effective in many areas of the

whitebark pine type where rugged terrain provides

natural fire breaks. To cover large areas of the type,

which is restricted to high ridges, fires had to burn

across broad valleys. Many of the latter are now under

forest management, agriculture, suburban or other

development that prevents fire spread. Large wilderness areas and National Parks sometimes allow wildland fires with limited suppression as a substitute for

natural fires. However, even the most successful natural fire program in a large wilderness has had a lower

level of success in returning fire to whitebark pine

than to other forest types (Brown and others 1994,

1995). Major fires in the whitebark pine type are

confined largely to late summer in especially dry

years, when prescribed natural fires usually are not

allowed.

To complicate matters, white pine blister rust

(Cronartium ribicola), an introduced disease, is killing cone-bearing limbs and entire trees of whitebark

pine in about half of the species’ natural range (Arno

and Hoff 1990). A small percentage of the trees has

some resistance, but without ample sites upon which

to regenerate, resistant strains cannot develop and

multiply.

Management Considerations—Fires may be more

critical for whitebark pine’s survival now that blister

rust has inflicted heavy mortality across large areas of

the Northwestern United States. Evidence from mortality observations (Kendall and Arno 1990), permanent plots (Keane and Arno 1993), and ecological

process modeling (Keane and others 1990b) suggests

that unless active management is carried out on a

landscape scale, whitebark pine will continue to decline in a large part of its range and may virtually
112 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Arno Chapter 5: Fire in Western Forest Ecosystems

disappear in some areas. Use of wildland fires in

wilderness, prescribed fires, release cuttings to favor

the pine over its competitors, and aiding the propagation of natural rust resistance are the most obvious

alternatives. These measures are now being tested at

several sites in Idaho and Montana by Robert Keane,

Rocky Mountain Research Station, Missoula, Montana.

Ponderosa Pine

Pre-1900 Succession—Some ponderosa pine forests were historically characterized by mixed fire

regimes, although the extent and ecological relationships of these mixed regimes are yet to be determined.

It appears that mixed regimes were commonly associated with ponderosa pine growing east of the Continental Divide, and also with some forests west of the

Divide, especially those on steep slopes and on relatively moist sites. The most compelling evidence for a

large area of mixed fire regime comes from the Black

Hills of South Dakota (Brown and Sieg 1996; Gartner

and Thompson 1973; Shinneman and Baker 1997) and

the Front Range of the Rocky Mountains in Colorado

(Kaufmann 1998; Laven and others 1980). Many of the

ponderosa pine stands in the Black Hills and nearby

areas of northeastern Wyoming and southeastern

Montana develop dense patches of pine regeneration

after fire, which become thickets of small stagnant

trees, susceptible to stand-replacing fire. Intervening

areas with more open stocking presumably were more

likely to underburn in the frequent fires of the presettlement era. Factors contributing to a mixed fire

regime in ponderosa pine probably include relatively

moist sites that tended to produce pine thickets soon

after a fire, areas frequently exposed to high winds

during the burning season, steep topography, and

stands killed by bark beetle epidemics.

Post-1900 Succession—In the Black Hills, northeastern Wyoming, and southeastern and central Montana, fire exclusion coupled with livestock grazing has

resulted in an expansion of ponderosa pine from its

presettlement habitat on rocky ridges and other poor

sites into the adjacent grassy plains. This invasion

may involve over a million acres (Gartner and

Thompson 1973; Gruell 1983), but no data are available. The stands on the ridges have also thickened

with ingrowth of younger trees. Dense stands greatly

reduce production of native grasses and forbs and

diminished forage values for livestock and wildlife

(Gartner and Thompson 1973). Thousands of homes

are now located in these dense pine stands, and many

are threatened annually by wildfires. Nearly 50 homes

burned in a central Montana wildfire in 1984, and over

Figure 5-7—Whitebark pine regenerating on a 26-year-old burn (foreground). In contrast, the old, unburned

forest (background) consists of subalpine fir, Engelmann spruce, and dead whitebark pine snags.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 113

Chapter 5: Fire in Western Forest Ecosystems Arno

100 burned in 1991 in ponderosa pine woodlands near

Spokane, Washington.

Management Considerations—The dense stands

of small ponderosa pines that have expanded into

grasslands in the upper Great Plains are at high risk

of severe wildfire and have diminished forage values.

Breaking up the continuity of these stands using

silvicultural cuttings and prescribed fire treatments

would allow more effective control of fire. It may be

possible to thin some of these stands commercially

while retaining the largest and healthiest trees. Prescribed burning in conjunction with cutting could

reduce fuel loadings and stimulate forage (Gartner

and Thompson 1973; Kilgore and Curtis 1987). Initial

burning should be done with care not to injure leave

trees (see discussion of ponderosa pine in the section

“Understory Fire Regimes”).

There may be objections to converting dense stands

to open-growing ones, especially doing so on a landscape scale (Shinneman and Baker 1997). Patchy

even-aged cuttings and prescribed burns are one alternative that could be used to simulate historical mixed

regime patterns (Franklin and Forman 1987). Allowing dense forests on public lands to be killed by insect

epidemics that encourage severe wildfires may be the

eventual result of inaction. This latter alternative

could be made less hazardous to adjacent lands if fuels

management were carried out around the borders of

such areas (Scott 1998).

Stand-Replacement Fire

Regimes _______________________

Major Vegetation Types

Major forest types include coast Douglas-fir, true fir/

hemlock, interior true fir/Douglas-fir/larch, Rocky

Mountain lodgepole pine, white pine/western redcedar/hemlock, spruce/fir/whitebark pine, and aspen.

Several minor forest types are also characterized by

stand-replacement fires. These forests are widespread

in wetter forest regions of the Northwestern United

States and Western Canada and in subalpine forests

associated with the major mountain ranges. Some of

the same compositional types are also characterized

by mixed fire regimes in large portions of their distributions. Differences in regional climate, fuels, and

local topography can influence the resulting fire regime (for example, see Barrett and others 1991).

Fire Regime Characteristics

Stand-replacing fires kill most overstory trees, although the pattern of these fires on the landscape

varies with topography, fuels, and burning conditions.

Sometimes extensive areas burn uniformly in

stand-replacing fire events, especially in wind-driven

crown fires (Anderson 1968). However, a major proportion of stand-replacement is caused by lethal surface fire, as was the case with much of the 1988

Yellowstone Area fires. Lethal surface fire was responsible for about 60 percent (versus 40 percent in

crown fire) of the stand-replacement burning in the

Selway-Bitterroot Wilderness under the prescribed

natural fire program (Brown and others 1995). Under

different conditions, a complex landscape mosaic of

replacement burning from crown fire and lethal surface fire is interwoven with areas of lighter burning or

no burning. For instance, patchy burning patterns

may be accentuated by rugged mountainous topography containing contrasting site types, microclimates,

and vegetation. These mosaic elements represent diverse burning environments and the result is that

stand-replacement burning is restricted to certain

landscape elements. For example, in one area of northern Idaho, stand-replacement fires were associated

with a mid-slope “thermal belt” on southern and western exposures while other slopes tended to burn in

mixed severity fires (Arno and Davis 1980). Superimposed on the site mosaic is a fuels mosaic linked to the

pattern of past fires. On gentle topography and more

uniform landscapes, such as high plateaus, standreplacement fires tend to be more uniform or at least

to burn in large-scale patches.

Stand-replacement fires generally occur at long average intervals (table 5-1), ranging from about 70

years in some lower elevation Rocky Mountain lodgepole pine forests subject to extreme winds, to 300 to

400 years in some inland subalpine types, and over

500 years in some moist coastal mountain forests.

Often the range of actual intervals is broad since the

fires themselves depend on combinations of chance

factors such as drought, ignitions, and high winds.

Such combinations occur sporadically. In coastal forest types having long fire intervals, such as coast

Douglas-fir and true fir-mountain hemlock, it appears

that exceptional drought and perhaps an unusual

abundance of lightning ignitions were linked to major

fires (Agee 1993). The great length of intervals and

findings of substantial climatic changes during the

last few thousand years suggest that fire intervals

varied with climatic changes (Johnson and others 1994).

The irregular timing of stand-replacement fire is

heightened in several forest types by their propensity

to support double or triple burns in the aftermath of an

initial fire. For instance, the Yacolt fire (1902) of

southwestern Washington and the Tillamook fire

(1933) of northwestern Oregon reburned numerous

times (Gray and Franklin 1997; Pyne 1982). A history

of occasional replacement fires followed by severe

reburns also is common in the Clearwater drainage of

northern Idaho (Barrett 1982; Wellner 1970).
114 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Arno Chapter 5: Fire in Western Forest Ecosystems

In the Rocky Mountain lodgepole pine type, fuel

buildup is an important factor in length of fire intervals

(Brown 1975; Romme 1982). Mean fire intervals in

this type range from a low of about 70 years in productive lower elevation lodgepole pine forests in high wind

environments on the east slope of the Rockies—for

example, at Waterton Lakes National Park, Alberta;

Glacier National Park, Montana; and perhaps in the

Red Lodge area northeast of Yellowstone National

Park. At the other extreme, mean fire intervals on

unproductive sites such as the high-elevation rhyolite

plateaus in Yellowstone National Park are 300 to 400

years (Millspaugh and Whitlock 1995; Romme 1982).

The majority of studies in this type has found mean

fire intervals between 100 and 250 years (Agee 1993;

S. W. Barrett 1994; Hawks 1980; Kilgore 1987).

Fuels

Unlike understory and mixed fire regimes, fuels

play a critical role in limiting the spread of fire in

stand-replacement fire regimes. Accumulation of duff

and down woody fuels increases the persistence of

burning. This is important for keeping fire smoldering

on a site until a wind event occurs (Brown and See

1981). Typically a certain level of fuel is required to

allow fire to spread. This may be the result of dead and

down fuels—from insect epidemics, windstorms, or a

previous fire—or of extensive ladder fuels (fig. 5-8). In

contrast, stands with few down or ladder fuels often

fail to support fire (Brown 1975; Despain 1990). In

lodgepole pine, dead and down woody fuel loadings of

15 to 20 tons/acre (34 to 45 t/ha) are generally near the

lower threshold of what will support a stand-replacement through moderate-intensity surface fire (Fischer

1981). Ladder fuels and heavier loadings of down and

dead woody fuels contribute to torching, and with

winds a running crown fire may evolve.

In cover types supporting large trees such as Douglas-fir/hemlock and western white pine, large woody

fuel loadings typically are 40 to 50 tons/acre (90 to

110 t/ha) and duff about 30 tons/acre (67 t/ha) (Keane

and others 1997). In smaller tree cover types such as

lodgepole pine and spruce/fir, large woody fuels typically are 15 to 20 tons/acre and duff about 15 to 30 tons/

acre. However, the range in loadings may be considerably greater as reported for Kananaskis Provincial

Park, Alberta (Hawkes 1979), where downed woody

fuels ranged from 4 to 63 tons/acre (9 to 141 t/ha) and

duff from 8 to 58 tons/acre (18 to 130 t/ha).

Figure 5-8—Downfall of mountain pine beetle killed lodgepole pine results in an accumulation of large downed woody

fuels that increases the likelihood of stand-replacement fire.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 115

Chapter 5: Fire in Western Forest Ecosystems Arno

Postfire Plant Communities

Coast Douglas-fir

Pre-1900 Succession—These humid maritime forests are extensive at lower and middle elevations west

of the Cascades and British Columbia Coast Range.

The cooler, wetter, and more northerly portions of the

coastal Douglas-fir type (generally associated with the

mountains of western Washington and southwestern

British Columbia) burned in stand-replacement fires

at long intervals, averaging 200 to several hundred

years (Agee 1993). The range of pre-1900 fire intervals

on a given site is unknown because in most cases only

the most recent interval can be calculated due to decay

of the previous stand. Long and others (1998) described fire intervals over the last 9,000 years, and

Impara (1997) reports on the spatial patterns of historical fires in the Oregon Coast Range.

Western hemlock is the potential climax dominant

tree in most of this type, but seral Douglas-fir, which

arose after replacement fires during the last several

hundred years, is the actual dominant. The greater

size and longevity of Douglas-fir allows it to persist in

considerable quantities for 700 to 1,000 years between

major stand-opening disturbances such as fire or severe blowdowns (Agee 1993). Scattered individual

Douglas-fir survived fires and served as seed sources

in the burn. Seeds of this species may also survive and

mature in the crowns of some trees whose foliage was

killed (but not consumed) by a late-summer fire. The

seeds are also wind-dispersed from unburned stands.

Douglas-fir seedlings grow readily on burned seedbeds and outcompete other conifers in the postburn

environment.

Often red alder becomes abundant and temporarily

outgrows Douglas-fir in a recent burn. However, the

fir grows up beneath and displaces alder within a few

decades, benefitting from soil nitrogen fixed by symbiotic organisms associated with alder roots. Numerous

other seral conifers (western white pine, shore pine,

grand fir, and Sitka spruce) and hardwood species

(bigleaf maple, mountain ash, cascara, and others), as

well as seral shrubs (salmonberry, huckleberries) and

herbaceous plants, appear in the postburn environment, greatly enriching the biological diversity of

these forests (Fonda and Bliss 1969; Franklin and

Dyrness 1973; Hemstrom and Franklin 1982; Huff

1984; Yamaguchi 1986).

Post-1900 Succession—Due to the great length of

natural fire intervals it would seem unlikely that

significant successional changes have occurred in most

of these forests as a result of attempts to exclude fire

during this century. Large areas of these forests have

been clearcut in recent decades, sometimes followed

by broadcast burning. This has given rise to large

areas of early seral communities dominated by native

flora, often with planted Douglas-fir, which might

offset a shortage of early seral communities resulting

from natural fires. However, natural burns and

clearcuts differ ecologically, for example, in seedbed

preparation, in providing residual large woody debris,

and in having an overstory of dead trees (Kauffman

1990). Hansen and others (1991) point out that young

communities arising after fires are rich in structural

complexity and in species composition, but are the

rarest successional state, much rarer in today’s landscapes than is old growth. Although millions of acres

have been set aside as reserves for old growth Douglasfir, there are no measures for perpetuating these

communities through the use of prescribed fire, and if

present fire suppression policies succeed, young postfire

communities will continue to be rare.

Management Considerations—Until the 1990s

these forests were usually managed by clearcutting,

site preparation, and growing even-aged stands at

rotations (50 to 100 years) much shorter than those of

presettlement fire intervals. This management approach failed to consider perpetuation of many ecological functions in these forests (Kauffman 1990). The

need to develop more ecologically-based management

strategies gave rise to concepts of “new forestry” treatments (Franklin and others 1986; Gillis 1990; Hopwood

1991). Most of these leave patches of overstory and

understory trees after harvesting to increase structural diversity of the new stand and the “biological

legacies” of large woody debris. In this structural

respect, these treated stands are simulating the early

seral community following a natural fire. However,

many of the proposed treatments avoid actually utilizing fire due to a desire to limit smoke, increase woody

residues on the site, avoid the operational difficulties

in burning, and reduce treatment costs (Means and

others 1996). Douglas-fir can be regenerated in heavily

logged areas without burning. It is apparently assumed that the other effects of fire are expendable, for

example, in soil nutrition and maintenance of diverse

fire-dependent undergrowth species (Agee 1993;

Kauffman 1990). However, there is little scientific

basis for such an assumption. A more rigorous evaluation of the consequences of various management

alternatives, including use and avoidance of burning,

is needed for this ecological type.

Coastal True Fir-Mountain Hemlock

Pre-1900 Succession—These high-elevation maritime forests are found along and west of the crest of the

Cascades (north of Crater Lake, Oregon) and the

British Columbia Coast Range. Principal tree species

are Pacific silver fir, mountain hemlock, western hemlock, and noble fir. This type is a cooler and wetter
116 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Arno Chapter 5: Fire in Western Forest Ecosystems

environment than the coastal Douglas-fir type that

borders it at lower elevations and on warmer aspects.

Annual precipitation is commonly >100 inches and a

deep snowpack accumulates in winter and persists

into early summer.

The principal tree species are fire-sensitive and

seldom survive surface fires. Thus, fires were typically

of the stand-replacement type, although scattered

Douglas-fir found in this type often survived. Fire

return intervals were evidently between about 125

and 600 years (Agee 1993). Shorter intervals (<200

years) were associated with drier environments where

the type is confined to north-facing slopes and is

surrounded by drier types. Fires in this type usually

occur under conditions of severe summer drought

accompanied by strong east (foehn) winds. Major

blowdowns also initiate regeneration cycles and can

contribute large amounts of fuels in this type.

Postfire stands often go through a shrub-dominated

stage—commonly including early seral communities

of mountain huckleberry (Agee 1993; Franklin and

Dyrness 1973). A variety of early seral conifers becomes established in burns (noble fir, Douglas-fir,

subalpine fir, Alaska-cedar, lodgepole pine, and western white pine). Eventually the shade-tolerant Pacific

silver fir, mountain hemlock, and western hemlock

become dominant.

Post-1900 Succession—The long intervals between

stand-replacement fires and the remote location of

much of this forest type suggest that fire suppression

would have had a minor effect upon it, but no detailed

evaluation of this question has been made.

Management Considerations—Few fire management programs that plan for wildland fire use exist in

the National Parks and wilderness areas in this forest

type. Where fires occurred in the past, they often

resulted in shrubfield/open conifer stands. These persistent open, “old burn” communities have been an

important component of wildlife habitat and natural

diversity. Use of prescribed fire is limited in this type

because fire generally burns only under extreme

weather conditions. Because the principal tree species

are readily killed by fire, any burning in standing trees

increases loadings of dead and down fuels. Mechanical

fuel manipulation (tree harvesting or removal of dead

woody fuel) is necessary for creating fuel reduction

zones to contain wildfires (Agee 1993).

Interior True Fir–Douglas-fir–Western Larch

Pre-1900 Succession—This is a diverse group of

forests that have a stand-replacement fire regime and

are dispersed throughout much of the Interior West,

usually at middle elevations in the mountains. Principal tree species are white fir and grand fir (potential

climax), interior Douglas-fir (seral or climax), and

larch, lodgepole pine, and aspen (early seral associates). These forests commonly develop dense stands

with accumulations of ladder fuels and they often

occupy steep slopes on cool aspects. The forest floor

fuels are primarily a compact duff layer that does not

support low intensity surface fires. However, when

down woody or ladder fuels accumulate and severe

burning conditions arise, they can support a standreplacing surface or crown fire. Such fires occurred at

intervals averaging between 70 and 200 years. Similar

compositional types in other geographic areas or on

different topographic situations are associated with

mixed fire regimes. The relative amounts of these

types in mixed and stand-replacement fire regimes is

unknown (Brown and others 1994). Also, the factors

that determine whether one of these forests will have

a mixed or stand-replacement regime is not known,

but lack of receptiveness of surface fuels to burning,

characteristically dense stands, steep slopes, and frequent strong winds probably favor the stand-replacement fire regime. For example, one area in the standreplacement regime is on the eastern slope of the

Continental Divide in Montana where dense Douglasfir stands develop in an environment featuring severe

winds (Gruell 1983).

Relatively frequent stand-replacement fires kept

much of the landscape in open areas (seral grasslands

or shrublands) and favored seral shrub species (such

as serviceberry, willow, and bitterbrush) and aspen.

Such plant communities are important forage for

wildlife.

Post-1900 Succession—Photocomparison and fire

history studies suggest that fire exclusion has allowed

a greater proportion of these inland forests on the

landscape to develop as dense stands. The spatial

continuity of these stands may allow insect and disease epidemics and stand-replacement fires to become

larger than in the past (Arno and Brown 1991; Byler

and Zimmer-Grove 1991; Gruell 1983) (fig. 5-1). At the

same time seral grassland species, shrubs, aspen, and

seral conifers are being replaced by thickets of shadetolerant conifers.

Management Considerations—This major forest cover type is divided into mixed and stand-replacement fire regimes, but the environmental characteristics linked to each fire regime type are poorly known.

Knowledge of these characteristics would help land

managers determine where stand-replacement fire is

probable, which might help in setting priorities for

management of fuels to confine potentially severe

wildland fires. It should be possible to reduce frequency and extent of stand-replacement fires using a

variety of fuel-reduction treatments. Prescribed fire in

activity fuels (slash) can be useful in fuels reduction
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 117

Chapter 5: Fire in Western Forest Ecosystems Arno

and in obtaining other desirable fire effects such as

stimulation of wildlife forage. It may be possible to use

fire for fuels reduction and habitat improvement for

ungulates by felling a few trees per acre to create

enough fine slash fuels to allow a prescribed fire to

move through the stand. This is promising in Douglasfir stands because of this species resistance to low

intensity fire.

Rocky Mountain Lodgepole Pine

Pre-1900 Succession—This is a major type at

middle to high elevations in the more continental

mountain climatic areas of the Inland West, from the

Yukon Territory, Canada, to southern Colorado. In

parts of this geographic distribution, lodgepole pine

forests burned in a mixed fire regime, primarily where

fine surface fuels and dry climate allowed lower intensity fires to occur. Much of the lodgepole pine type,

however, is resistant to burning except when there is

an accumulation of down woody, ladder, and crown

fuels. When fuel loadings are sufficient to support fire,

it becomes a stand-replacing surface or crown fire.

Brown (1975) illustrated how fuel loadings are indirectly linked to stand age (fig. 5-9). Young dense

stands containing ladder fuels of associated spruce

and fir and accumulated downfall from a former,

beetle-killed or fire-killed overstory have high potential to support a stand-replacement fire. Conversely,

young pole-size stands of pure lodgepole pine (with

sparse lower limbs) arising after a burn that removed

most large fuels, have low potential to support fire.

When a lodgepole pine stand becomes mature or

overmature, tree growth and vigor declines markedly,

and likelihood of a mountain pine beetle epidemic

increases. Such epidemics kill many trees that begin

falling in a few years, and within 10 to 15 years large

amounts of dead woody fuels accumulate that greatly

adds to the potential of stand-replacement fire. Dwarf

mistletoe also builds up with stand age and adds

highly flammable witches-brooms to the tree crowns.

Stand-replacement fire destroys the dwarf mistletoe,

except in surviving patches, and removes potential for

mountain pine beetle until a mature stand has again

developed.

Post-1900 Succession—Some studies indicate that

attempts to exclude fire have had relatively little effect

in this fire regime (Barrett and others 1991; Johnson

and others 1990; Kilgore 1987). Certainly numerous

fires have been successfully suppressed while quite

small or while they were in adjacent, different fire

regime types; but this is presumably offset by additional ignitions from large numbers of human-caused

fires. The possibility exists that suppression could

have appreciable effects in some geographic areas,

especially where units of this type are small, isolated,

and surrounded by other kinds of forest or vegetation

where fires have been largely excluded. In the southern Canadian Rockies a decline in fire frequency in

largely stand-replacement fire regime types was attributed at least partially to prevention and suppression of human-caused fires (Achuff and others 1996).

However, in some geographic areas, the proportion of

area burned by lethal seveity has increased (fig. 5-1).

Management Considerations—As illustrated by

the political uproar regarding the Yellowstone Area

fires of 1988, this forest type represents a challenge for

fire management (Christensen and others 1989;

Wakimoto 1989). Ecologists, land managers, and many

environmental activists recognize the importance and

inevitability of large stand-replacement fires. In contrast, such fires are often viewed as an unnecessary

inconvenience or as a disaster by those who are unaware of the natural role of wildland fire.

Fires are critical to maintenance of biological diversity in this type. Many early seral species, including

herbs, shrubs, and aspen, depend on occasional fires to

remain as components of the lodgepole pine type

(Habeck and Mutch 1973; Kay 1993). Black-backed

Woodpeckers, many invertebrates, herbivores, small

mammals, birds, and even some aquatic organisms

depend upon fires for creation of seral communities,

snag patches, and beneficial nutrient cycling (Agee

1993; Despain 1990).

Stand-replacement fire regimes in lodgepole pine

forests can be influenced by management actions. For

example, fuel breaks can be developed near critical

property boundaries and to protect resorts and other

facilities (Anderson and Brown 1988; Kalabokidis and

Figure 5-9—Fuel cycles and fire intensity potential in a lodge- Omi 1998; Schmidt and Wakimoto 1988). Wildland pole pine stand-replacement fire regime (Brown 1975).

Young Immature Mature Overmature

Time since establishment

A

C

B

Fire intensity potential for fuels

Low High
118 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Arno Chapter 5: Fire in Western Forest Ecosystems

fire use programs coupled with prescribed stand-replacement fires could help develop landscape fuel

mosaics that limit the ultimate size of wildfires

(Weber and Taylor 1992; Zimmerman and others 1990).

Clearcutting and broadcast burning have long been

a common timber management treatment in this type.

If fine fuel consumption is complete, most of the seed

source in the slash will be destroyed. Seeds from openconed pines at the edge of units are often winddistributed about 200 feet into clearcuts in sufficient

quantities for forest regeneration (Lotan and

Critchfield 1990). Leaving some cone-bearing trees

standing throughout the burn can provide seed source

as well as light shade and snags for wildlife. In salvage

logging of fire-killed stands, it is important to leave

ample dead trees for a variety of wildlife species.

There is increasing interest in planning for patches

of trees to survive broadcast burning, to provide structural diversity in the post-treatment community (Arno

and Harrington 1998; Hardy and others 2000).

Underburning in a lodgepole pine shelterwood cut,

although difficult, may be possible if slash fuels are

light and moved away from the base of leave trees.

Western White Pine-Cedar-Hemlock

Pre-1900 Succession—This forest type is centered

in northern Idaho and the interior wet zone of southeastern British Columbia (Krajina 1965; Shiplett and

Neuenschwander 1994). It also extends into

northeastern Washington and northwestern Montana.

It occupies a “climatic peninsula” (Daubenmire 1969)

or inland extension of Pacific maritime climate between about 46 and 53 °N latitude. This is the only

area of the Rocky Mountain system where western

white pine, western redcedar, western hemlock, and

numerous other maritime associates (trees and undergrowth) are found. In this area the inland maritime

species form the dominant vegetation mixed with

Inland Rocky Mountain species such as ponderosa

pine, interior Douglas-fir, Engelmann spruce, western

larch, and whitebark pine.

Traditionally this type has been considered to mainly

represent a stand-replacement fire regime, but detailed fire history studies are few (Smith and Fischer

1997). The most characteristic, large, and influential

fires were stand-replacing, and because of scarcity of

information, the type is discussed only in this section.

Nevertheless two recent studies (Barrett 1993; Zack

and Morgan 1994) as well as earlier observations

(Arno 1980; Arno and Davis 1980; Marshall 1928)

indicate that substantial areas of the type were also

exposed to a mixed fire regime. Evidence of underburning is often found in valley bottoms, on gentle

slopes on dry aspects, and on ridgetops. Surviving

trees were primarily western white pine, western

larch, and large western redcedars. Underburns may

have been important locally in perpetuating a dominance by large trees of these species. Conditions associated with underburning have not been described,

but underburning has occurred in recent wildland

fires in these forests (Brown and others 1994).

The stand-replacement fires characteristic of the

western white pine-cedar-hemlock type occurred at

intervals of 130 to 300 years associated with severe

drought, which commonly occurred for a short period

during mid- or late-summer. Because of the productive growing conditions, ladder fuels can develop rapidly in young stands. Moreover, large woody fuels are

created in abundance by fires and by insect and disease epidemics. Two or three decades after a standreplacement fire, most of the dead trees have fallen

and become heavy down fuels. This, coupled with a

dense stand of small trees and tall shrubs, may constitute a fuelbed that allows severe burning in a second

fire, known as a “double burn” (Wellner 1970). When

fire occurs under conditions of extreme drought accompanied by strong winds, stand-replacement fires

are likely in many natural fuel configurations. If large

accumulations of down woody fuel (>25 tons/acre) are

present, stand-replacement fires can occur under extreme drought without strong winds or steep slopes

(Fischer 1981).

Past stand-replacement fires allowed seral fire-dependent species to dominate most pre-1900 stands.

The major early seral tree dominants were western

white pine, western larch, and lodgepole pine, but

were accompanied by lesser amounts of interior Douglas-fir, grand fir, Engelmann spruce, paper birch,

and ponderosa pine. Including the everpresent western redcedar and western hemlock, these disturbance

communities were diverse. Luxuriant seral shrub and

herbaceous plant growth added to this diversity, as

described by Larsen (1929), Daubenmire and

Daubenmire (1968), and Cooper and others (1991).

Occasional fires allowed a variety of seral shrubs to

thrive, including redstem and evergreen ceanothus,

currants, red elderberry, Scouler willow, serviceberry,

mountain maple, American mountain-ash, bittercherry, and chokecherry. After a double burn,

shrubfields would persist for decades (Barrett 1982;

Wellner 1970). Shiplett and Neuenschwander (1994)

classify five common successional scenarios in these

forests. These are (1) a relatively rapid succession to

the redcedar-hemlock climax, (2) a prolonged domination by a mix of seral tree species as a result of

disturbances, (3) a shrubfield resulting from multiple

burns, (4) a sere influenced by scattered larch relicts

that survived fires, and (5) lodgepole pine dominance

throughout fire cycles on less productive sites with

relatively frequent burns.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 119

Chapter 5: Fire in Western Forest Ecosystems Arno

Post-1900 Succession and Management Considerations—Clearcut logging has to a considerable

extent replaced fire as the principal stand-replacing

disturbance in this forest type. Often the logging is

followed by broadcast burning or dozer scarification

and piling of large woody residues, which are then

burned. Like stand-replacement fire, clearcutting favors establishment of early seral tree species. Unlike

fire, clearcutting does not leave a snag forest to moderate the microclimate and provide large quantities of

woody debris and future fuels. Recently, environmental concerns have been raised about cumulative impacts of road building and logging, including soil

disturbance, erosion, loss of water quality, aesthetic

impacts, loss of wildlife habitat, and smoke production

from prescribed fire. These concerns have encouraged

substitution of partial cutting or of no cutting at all

accompanied by fire suppression. These approaches

contrast strongly with natural fire in their effects on

vegetation and may result in epidemic forest mortality

resulting from root diseases and bark beetles (Byler

and Zimmer-Grove 1991).

A high percentage of western white pine has been

killed in the last 50 years as a result of white pine

blister rust, an introduced disease. White pine has a

low level of natural rust resistance, and resistant

genotypes are available for use in reforestation. However, this species requires fire or logging with site

preparation to make sites available for regeneration

and successful establishment.

Relatively rapid change is characteristic of the vegetation in this highly productive forest type. In the

past, fire was the principal agent initiating new cycles

of change. Heavy logging and site preparation was to

a limited extent a replacement for fire, but had some

undesirable impacts. In revising management to reduce those impacts it is important to consider strategies and treatments that provide the beneficial effects

associated with fire in these ecosystems (Smith and

Fischer 1997).

Spruce-Fir-Whitebark Pine

Pre-1900 Succession—This type makes up the

highest elevations of forest growth in the Rocky Mountains and other interior ranges of Western North

America from central British Columbia to central

Oregon and western Wyoming. Southward in the Rocky

Mountains to New Mexico, beyond the range of whitebark pine, the ecologically similar limber pine is often

associated with the spruce and fir. In northern British

Columbia, the high-elevation spruce-fir forest merges

with the white spruce and black spruce boreal types

discussed in chapter 3.

In drier regions of the Interior West and locally on

drier topographic sites is a mixed severity fire regime

characterized by an abundance of whitebark pine (see

the section “Mixed Fire Regimes” in this chapter). In

contrast, the spruce-fir-whitebark pine type has variable amounts of whitebark pine or none at all, but is

characterized by stand-replacement fires generally at

intervals of 100 to 400 years. For example, in the most

detailed study of this type so far, in the Bob Marshall

Wilderness Complex in Montana, Keane and others

(1994) found mainly stand-replacement fires at intervals of 54 to 400+ years in 110 sample stands distributed across a 1.5-million-acre (607,000 ha) area. In the

Southern Rocky Mountains, spruce is often the dominant subalpine forest cover and other major disturbances—spruce beetle epidemics, extensive snow avalanches, and areas of wind-thrown forest—interact

with stand-replacement fires in complex temporal and

spatial patterns (Baker and Veblen 1990; Veblen and

others 1994). In the wettest spruce-fir microsites, such

as naturally subirrigated basins, fire occurs rarely

and is not the prevalent factor controlling successional

cycles that it is in most of Western North America.

Pre-1900 fires added structural and compositional

diversity to the spruce-fir-whitebark pine forest.

Burned areas often remained unforested for extended

periods due to the harsh microclimate (Arno and

Hammerly 1984). In extreme cases regenerating conifers take on a shrublike (krummholz) form for 50 years

or longer. Often whitebark pine is able to become

established first in a high-elevation burn due to its

superior climatic hardiness and its advantage of having seeds planted in small caches by the Clark’s

Nutcracker (Arno and Hoff 1990).

Post-1900 Succession—Little is known about possible human-induced changes in successional patterns throughout this high-elevation type. Logging

has occurred in some sizeable areas of the type and has

to a limited extent been a substitute for stand-replacement fire. In other areas fire suppression may have

effectively reduced the landscape component made up

of young postfire communities. For example, Gruell

(1980) published many photographs taken at subalpine sites in northwestern Wyoming in the late 1800s

and early 1900s and compared them with modern

retakes. Most of these comparisons show that mature

forest is noticeably more extensive today. Presumably

the slow postfire recovery period resulted in large

areas being unforested at any given time.

Whitebark pine has suffered a major setback since

the early 1900s due to heavy mortality from mountain

pine beetle and blister rust (Keane and Arno 1993).

The introduced blister rust particularly reduces the

amount of whitebark pine seed source available for

regeneration. In some areas large outbreaks of spruce

bark beetle and root rot in subalpine fir have also

resulted in heavy loadings of large woody fuels, which

will support future stand-replacement fires (Veblen

and others 1994).
120 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Arno Chapter 5: Fire in Western Forest Ecosystems

Management Considerations—In smaller wilderness areas and parks it is hard to plan for standreplacement fires from lightning ignitions due to the

problem of confining fires within area boundaries.

Data presented by Brown and others (1994) suggest

that maintaining natural fire cycles in these highelevation forests is difficult because the forests only

burn when fire danger elsewhere is unacceptably high

as a result of extreme drought. Traditional timber

harvesting coupled with broadcast burning is now less

likely to occur because of environmental concerns

about road building, watershed impacts, and obtaining prompt tree regeneration. Concerns regarding fire

management options are generally similar to those

expressed in high-elevation lodgepole pine types. In

some cases prescribed fires might be used to maintain

natural fire cycles. Cutting to provide slash fuels could

allow prescribed burning to be done when wildfire

hazard is moderately low.

Aspen

For discussions about aspen, see chapter 3, “Fire in

Northern Ecosystems.”

Regimes Where Fire is Rare_______

In some forest types, which collectively cover only a

small fraction of the forested land in Western North

America, fire is so unusual or exceptionally infrequent

that it exerts little selective influence on vegetation

development. Certain stands and topographic situations within the subalpine spruce-fir type seldom burn

and their successional cycles are initiated primarily

by beetle epidemics or windthrow. In British Columbia, Hawkes (1997) reported fire cycles of 800 to 2,000

years on cool wet sites occupied by the spruce-fir type.

The only major forest type of Western North America

where fire is not a primary disturbance agent is Sitka

spruce-western hemlock. This type is generally confined to the wettest lowland sites along the immediate

Pacific Ocean coastal strip and in alluvial bottoms of

coastal valleys from southern Oregon to southern

British Columbia. Northward, in central British Columbia and southeastern and south-central Alaska, it

expands to cover most of the narrow, low elevation

coastal forest zone (Arno and Hammerly 1984). Fires

of appreciable extent are unusual in this type due to

the prevalence of moist conditions year-round (Agee

1993). Windfall is a more prevalent disturbance creating new stands. On the Alaskan coast most fires are

human-caused and occur during rare droughts (Noste

1969). Lightning fires are rare; nevertheless, the historic role of fire is not fully resolved. For example,

Harris and Farr (1974) report four episodes of extensive fires in southeastern Alaska between about 1660

and 1830.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 121

Timothy E. Paysen

R. James Ansley

James K. Brown

Gerald J. Gottfried

Sally M. Haase

Michael G. Harrington

Marcia G. Narog

Stephen S. Sackett

Ruth C. Wilson

Chapter 6:

Fire in Western Shrubland,

Woodland, and Grassland

Ecosystems

Western shrubland, woodland, and grassland ecosystems lie west of the eastern humid temperate zone,

which begins a short distance east of the 100th meridian. Shrublands include sagebrush, desert shrub,

southwestern shrub steppe, Texas savanna, and chaparral-mountain shrub ecosystem types. Woodlands

include southwestern ponderosa pine, pinyon-juniper,

and oak types that at times can be considered either

forests or woodlands. The woodland/forest dichotomy

can depend on phase of stand development and on the

realization of natural site conditions that can form

savannas with tree overstories. Grasslands include

plains, mountain, desert, and annual grassland ecosystems (table 6-1).

Understory Fire Regimes _________

Major Vegetation Types

Southwestern United States ponderosa pine consists of two varieties: (1) interior ponderosa pine found

over most of Arizona and New Mexico, and (2) Arizona

pine found in the mountains of extreme southwestern

New Mexico and southeastern Arizona, and extending

into northern Mexico (Little 1979) (fig. 6-1, 6-2). Based

on stand physiognomy (as in Paysen and others 1982),

many stands of this vegetation type can be considered

woodlands (relatively open grown), and many are

classical closed forests. Differences may be due to

inherent site conditions or to expressions of a developmental phase; fire frequency seems to play an important role as well.

Fire Regime Characteristics

Fires were frequent and of low intensity. Light

surface fires burned at intervals averaging less than

10 years and as often as every 2 years (Dieterich 1980;

Weaver 1951). The short fire-interval was caused by

warm, dry weather common to the Southwest in early

summer, the continuity of grass and pine needles, and

the high incidence of lightning. Two fire seasons usually

occurred each year, a major fire season after snow melt

and just before the monsoon season in midsummer
122 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Table 6-1

—Occurrence and frequency of presettlement fire regime types by Forest and Range Environmental Study (FRES) ecosystems, Kuchler potential natural vegetation

classes (1975 map codes), and Society of American Foresters (SAF) cover types. Occurrence is an approximation of the proportion of a vegetation class represented

by a fire regime type. Frequency is shown as fire interval classes defined by Hardy and others (1998) followed by a range in fire intervals where data are sufficient.

The range is based on study data with extreme values disregarded. The vegetation classifications are aligned to show equivalents; however, some corresponding

Kuchler and SAF types may not be shown.

Fire regime types

Understory Mixed Stand-replacement

FRES Kuchler SAF Occura Freqb Occur Freq Occur Freq Nonfire

Ponderosa pine 21 SW ponderosa pine c Interior ponderosa pine 237 M 1a:2-10 m 1

Arizona pine forest K019 M 1a:2-10 m 1

Pine-cypress forest K009 Arizona cypress 240 M 1,2 m 1

Pinyon-juniper 35 Juniper-pinyon K023 Rocky Mountain juniper 220 M 1

Juniper-steppe K024 Western juniper 238 M 1

Pinyon-juniper 239 M 1

Arizona cypress 240 M 1

Southwestern oaks d California oakwoods K030 Canyon live oak 249 M 1

California coast live oak 255 M 1

California black oak 246 M 1

Blue oak-digger pine 250 M1 M 1

Oak-juniper K031 Interior live oak 241 M 1

Shinnery 31 Shinnery K071 Mohrs oak 67 M 1

Texas savanna 32 Ceniza shrub K045 M 1

Mesquite savanna K060 Mesquite 68, 242 M 1

Mesquite-acacia savanna K061 M 1

Mesquite-live oak savanna K062 Western live oak 241 M 1

Juniper-oak savanna K086 Ashe juniper 66 M 1

Mesquite-oak savanna K087 M 1

Sagebrush 29 Sagebrush steppe K055 M 2a:20-70

Juniper steppe K024 Rocky Mountain juniper 220 M 2a

Great basin sagebrush K038 Western juniper 238 M 2a:20-70

Wheatgrass-needlegrass M 2a

shrubsteppe K056

Desert shrub 30 Mesquite bosques K027 Mesquite 68, 242 M 1,2a

Blackbrush K039 M 1,2a

Saltbrush-greasewood K040 M 1,2a

Creosotebush K041 M 1,2a

Creosotebush-bursage K042 M 1,2a

Paloverde-cactus shrub K043 M 1,2a

Cresotebush-tarbush K044 M 1,2a

SW shrubsteppe 33 Grama-tobosa K058 M 1,2a

Trans-pecos shrub savanna M 1,2a

K059

Chaparral-Mountain Oak-juniper woodland K031 M 1,2a

shrub 34

(con.)
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 123

Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen Table 6-1—Con. Fire regime types Understory Mixed Stand-replacement FRES Kuchler SAF Occura Freqb Occur Freq Occur Freq Nonfire Mountain mahogany-oak M 1,2a scrub K037 Transition of K031 & K037 M 1,2a Chaparral K033 M 1,2a Montane chaparral K034 M 1,2a Coastal sagebrush K035 M 1,2a Plains grasslands 38 Grama-needlegrass- M 1 wheatgrass K064 Grama-buffalograss K065 M 1 Wheatgrass-needlegrass K066 M 1 Wheatgrass-bluestem- M 1 needlegrass K067 Wheatgrass-grama- M 1 buffalograss K068 Bluestem-grama prairie K069 M 1 Mesquite-buffalograss K085 Mesquite 68, 242 M 1 Desert grasslands 40 Grama-galleta steppe K053 M 1,2a Grama-tobosa prairie K054 M 1,2a Galleta-three-awn shrubsteppe M 1,2a K057 Annual grasslands 42 California steppe K048 M 1,2a Mountain grasslands 36 Fescue-oatgrass K047 M 1 Fescue-wheatgrass K050 M 1 Wheatgrass-bluegrass K051 M 1 Foothills prairie K063 M 1a Cheatgrass c aM: major, occupies >25% of vegetation class; m: minor, occupies <25% of vegetation class bClasses in years are 1: <35, 1a: <10, 1b: 10 to <35, 2: 35 to 200, 2a: 35 to <100, 2b: 100 to 200, 3: >200. cThis type was not defined by Kuchler. dAdded subdivision of FRES.
124 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

Figure 6-1—Southwestern ponderosa pine distribution.

Figure 6-2—Typical Southwestern ponderosa pine fuels near Flagstaff, Arizona.

and a secondary season in the fall. Once a fire started,

the forest floor was generally consumed, but the damage

to trees was highly variable. Low intensity surface

fires predominated and were probably large where dry

forests and adjacent grasslands were extensive such

as on the gentle topography of high plateaus in Arizona

and New Mexico. Damage to trees was highly variable

but mortality to overstory trees was generally minor.

Fuels

The structural and compositional changes in Southwestern ponderosa pine over the past 100 years or

more have been repeatedly documented (Biswell and

others 1973; Brown and Davis 1973; Cooper 1960).

What was once an open, parklike ecosystem, maintained by frequent, low-intensity fires, is now a crowded,

stagnated forest. In addition to stand changes, general

fire absence has led to uncharacteristically large accumulations of surface and ground fuels (Kallander

1969).

The natural accumulation of pine needles and woody

fuels is exacerbated by the slow decomposition rates

characteristic of the dry, Southwestern climate

(Harrington and Sackett 1992). Decomposition rate

(k) (Jenny and others 1949) is the ratio of steady state

forest floor weight to the annual accumulation weight.

Harrington and Sackett (1992) determined k values of

0.074, 0.059, and 0.048 for sapling thickets, pole stands,

and mature old-growth groves, respectively. Decomposition rates this slow, which Olson (1963) considers

quite low, border on desertlike conditions. Humid,

tropic conditions would have k values approaching 1.0

where decomposition occurs in the same year as the

material is dropped on the ground.

Fuel loading estimates can be obtained from predictions based on timber sale surveys (Brown and others

1977; Wade 1969; Wendel 1960) and using Brown’s

(1974) planar intersect method for naturally accumulated downed woody material. Forest floor weights
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 125

Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen

have been studied extensively in Arizona and New

Mexico; results show high variability between sites.

Ffolliott and others (1968, 1976, 1977), Aldon (1968),

and Clary and Ffolliott (1969) studied forest floor

weights in conjunction with water retention on some

Arizona watersheds. These and other works included

prediction equations relating forest floor weight to

stand basal area (Ffolliott and others 1968, 1976,

1977), age (Aldon 1968), height and diameter (Sackett

and Haase 1991), and forest floor depth (Harrington

1986; Sackett 1985).

The forest floor consists of a litter (L) layer, recently

cast organic material; a fermentation (F) layer, material starting to discolor and break down because of

weather and microbial activity; and the humus (H)

layer, where decomposition has advanced. The loosely

packed L layer and upper portion of the F layer provide

the highly combustible surface fuel for flaming combustion and extreme fire behavior during fire weather

watches and red flag warnings (fig. 6-3). The lower,

more dense part of the F layer and the H layer make

up the ground fuel that generally burns as glowing

combustion.

Forest floor fuels (L, F, and H layers including woody

material ≤1 inch diameter) were sampled in 62 stands

in Arizona during the 1970s in Arizona and New

Mexico (Sackett 1979). Throughout the Southwest,

unmanaged stands of ponderosa pine had from 4.8

tons/acre (10.8 t/ha) in a stand on the Tonto National

Forest to more than 20 tons/acre (45 t/ha) in a stand on

the north rim of the Grand Canyon National Park.

The next two heaviest weights (18.3 and 18.0 tons/

acre) also occurred on the north rim of the Grand

Canyon. Mean forest floor loading for the entire 62

stands measured was 12.5 tons/acre (28.0 t/ha). When

woody material greater than 1 inch diameter was

added, the average increased to 21.7 tons/acre (48.6 t/

ha). The heavier material does not have much to do

with extreme fire behavior, except as a spotting potential; these fuels do contribute to localized severity

when burned. A range of forest floor fuel loadings is

summarized in table 6-2.

Of the 12.5 tons/acre (28.0 t/ha) average of forest

floor fuel load found in the Southwest, about 1.0 ton/

acre (2.2 t/ha) was L layer material, 3.8 tons/acre

(8.5 t/ha) was in the F layer, and 6.1 tons/acre (13.7 t/ha)

was H layer. Small diameter woody material and other

material comprised the remaining 1.8 tons/acre

(4.0 t/ha). The large woody material that accounted for

42 percent of the total fuel loading, consisted of

1.4 tons/acre (3.1 t/ha) of material 1 to 3 inches (2.5 to

7.6 cm) in diameter, 5.0 tons/acre (11.2 t/ha) of rotted

woody material 3+ inches in diameter, and 2.8 tons/acre

(6.3 t/ha) of sound wood 3+ inches in diameter. See

Sackett (1979) for complete summary.

Not only is there wide variation from site to site in

the Southwestern ponderosa pine ecosystem, but vast

differences exist within stands with respect to overstory characteristics (Sackett and Haase 1996). Experience indicates four separate conditions: sapling

(doghair) thickets, pole stands, mature old growth

(yellow pine) groves, and open areas in the groves

Figure 6-3—Section of ponderosa pine forest floor showing the fire intensity (FI) layer of fuel and fire severity (FS)

layer of fuel in relation to the L, F, and H layers of the forest floor.
126 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

Table 6-2—Average ponderosa pine surface fuel loadings (ton/acre) in the Southwestern United States by location

(Sackett 1997).

Number Forest floor and Woody fuel

Location of sites 0 to 1 inch diameter wood >1-inch diameter Total fuel

Kaibab NF 4 15.5 8.6 24.1

Grand Canyon NP 4 17.5 5.6 23.1

Coconino NF 4 14.7 19.8 34.5

Tonto NF 2 6.5 2.7 9.2

Apache-Sitgreave NF 14 11.3 11.2 22.5

San Carlos Apache IR 3 14.4 8.4 22.8

Fort Apache IR 2 15.1 20.5 35.6

Gila NF 10 11.2 7.3 18.5

Navajo IR 1 9.4 4.9 14.3

Cibola NF 3 8.8 8.8 17.6

Santa Fe NF 3 13.2 14.6 27.8

Carson NF 4 13.3 4.3 17.6

Bandalier NM 1 11.6 3.0 14.6

Lincoln NF 2 13.9 7.1 21.0

San Juan NF 5 11.9 4.8 16.7

without crowns overhead. Sapling thickets produce as

much as 1.1 tons/acre per year of litter and woody

fuels, pole stands 1.5 tons/acre per year, and mature,

old-growth groves as much as 2.1 tons/acre per year. A

substantial amount of forest floor material remains

after an area is initially burned (Sackett and Haase

1996). The amount remaining varies due to the original fuel’s configuration and the fire intensity and

behavior, which are affected by the overstory condition. This amount persists even with repeat applications of fire. The charred condition of the remaining

forest floor material resists re-ignition from the newly

cast needles that are consumed quickly.

Postfire Plant Communities

Southwestern Ponderosa Pine

Pre-1900 Succession—Chronicles from 19th century explorers, scientists, and soldiers described a

forest type quite different than what is seen today. The

open presettlement stands, characterized by wellspaced older trees and sparse pockets of younger trees,

had vigorous and abundant herbaceous vegetation

(Biswell and others 1973; Brown and Davis 1973;

Cooper 1960). Naturally ignited fires burning on a

frequent, regular basis in light surface fuels of grass

and pine needles maintained these forest conditions.

Light surface fuels built up sufficiently with the rapid

resprouting of grasses and the abundant annual pine

needle cast. Large woody fuels in the form of branches

or tree boles, which fall infrequently, rarely accumulated over a large area. When they were present,

subsequent fires generally consumed them, reducing

grass competition and creating mineral soil seedbeds,

which favored ponderosa pine seedling establishment

(Cooper 1960). These effects created an uneven-age

stand structure composed of small, relatively evenaged groups.

The decline of the natural fire regime in these

ecosystems started with extensive livestock grazing in

the late 19th century when fine surface grass fuels

were reduced (Faulk 1970). Subsequently, pine regeneration increased because of reduced understory competition, less fire mortality, and more mineral seedbeds (Cooper 1960).

Post-1900 Succession—In the early 1900s forest

practices, and reduced incidence of fire, led indirectly

to stagnation of naturally regenerated stands and

unprecedented fuel accumulation (Biswell and others

1973). Stand-stagnation exists on tens of thousands of

acres throughout the Southwest (Cooper 1960;

Schubert 1974) and still persists where natural or

artificial thinning has not taken place.

For several decades, trees of all sizes have been

showing signs of stress with generally poor vigor and

reduced growth rates (Cooper 1960; Weaver 1951).

This condition is likely due to reduced availability of

soil moisture caused by intense competition and by

moisture retention in the thick forest floor (Clary and

Ffolliott 1969). Thick forest floors also indicate that

soil nutrients, especially nitrogen, may be limiting

because they are bound in unavailable forms (Covington

and Sackett 1984, 1992).

A combination of heavy forest floor fuels and dense

sapling thickets acting as ladder fuels, coupled with

the normally dry climate and frequent lightning- and

human-caused ignitions, has resulted in a drastic

increase in high severity wildfires in recent decades
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 127

Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen

(Biswell and others 1973; Harrington 1982). Fire report

summaries (Sackett and others 1996) show a great

increase in the number of acres burned by wildfire

since 1970 (fig. 6-4). Of all the years since 1915 with

over 100,000 acres burned, almost 70 percent occurred

between 1970 and 1990.

Another characteristic of today’s Southwestern ponderosa pine stands is the sparseness of the understory

vegetation, including pine regeneration. The thick organic layers and dense pine canopies have suppressed

shrubby and herbaceous vegetation (Arnold 1950;

Biswell 1973; Clary and others 1968). Natural regeneration is also limited to areas where the forest floor

material has been removed either by fire or by mechanical means (Sackett 1984; Haase 1981). This condition

has reduced the wildlife, range, and timber values of

these forests and has generally minimized biodiversity.

Management Considerations—The need to alleviate the stagnated and hazardous forest conditions is

a primary consideration in the management of Southwestern ponderosa pine stands. The restoration of

forest health to the Southwest also needs to address

the following concerns:

• Dwarf mistletoe, once held in check by periodic

fires, is now a major cause of mortality in localized

areas.

• Bark beetle outbreaks are evident in overstocked

stands that are stressed from the high competition

Figure 6-4—The total number of acres burned by wildfires in Arizona and New Mexico from 1916 to 1990,

USDA Forest Service Smokey Bear fire summary reports. Heavy line represents 10-year average; light

line represents trends using the 4253H mathematical filter, used for smoothing noise in data.

for limited soil moisture, especially during drought

years.

• Some amount of fire injury to the overstory is

almost assured from the application of fire into an

area. This may be in the form of crown scorch to the

smaller trees and belowground injury to roots and

root collars of the larger trees.

• Fuel conditions that contribute to the elimination

of whole stands from wildfire need to be reduced.

These conditions include heavy forest floor accumulations and ladder fuel conditions created from

dense, stagnated sapling thickets.

Although the extent of these conditions will vary

throughout the region, the combination of any of these

situations on a particular forest creates a major concern and problem for the manager. Forest management objectives in the Southwest need to include the

maintenance or improvement of existing old-growth

stands and actions that promote the creation of future

old growth stands.

Because recurrent fire was a primary element in

sustaining presettlement forest health leading to the

establishment and maintenance of old-growth stands,

its use should be emphasized when restoring favorable conditions for ancient pine development. These

conditions include low levels of dead organic material

(fuels) to lessen the potential of high fire intensity and

severity, and open stand structure to reduce crown fire
128 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

potential and intraspecific competition. Fire can be

used to reduce fuel hazard, but its success is temporary. Failures denoted by too little or too much fuel

consumption generally result from improper burn

prescriptions and by attempting to correct long-term

fuel buildup with one treatment. Cooper (1960) questioned whether prescribed fire could be used in the

restoration of deteriorated forests. He concluded that

planned burning would be too conservative and accomplish little, or would destroy the stand. While this

observation has merit, with refined burning techniques as described in Harrington and Sackett (1990),

it appears that fire could be applied sequentially to

relieve the fuel and stand density condition. However,

it is apparent that considerable large tree mortality

could result. This seems to be an inescapable cost

dictated by years of forest degradation.

Because of these consequences, special attention

should be given to the excessive buildup of forest floor

fuels in present old-growth sites. Burning of these

deep forest floor layers can mortally injure the roots

and cambiums of old pines, which previously survived

many fires (Sackett and Haase 1996). Options for

alleviating this condition are not ideal. Managers

could simply accept a 20 to 50 percent loss of old

growth in a single fuel-reduction burn as being a cost

of decades of fuel buildup. Alternatively, the heavy

accumulation of fuels could be manually removed from

around the root-collar of the old-growth trees before

the fire is applied. Currently, methods are being investigated that will make this mitigation method a feasible option for managers. The use of a burn prescription that removes a portion of the fuel accumulation

has not been found for prescribed burning in the

Southwest. If glowing combustion is able to begin in

the deeper accumulations of material, high moisture

content of that material may not prevent total consumption of the forest floor. Nearly complete burnout

of duff has been observed in ponderosa pine forests at

moisture contents up to 90 percent (Harrington and

Sackett 1990) and in mixed conifers up to 218 percent

(Haase and Sackett 1998).

In forest regions where old-growth pine groups are

absent, designated areas based on site quality and

existing stand types should be selected for creating

future old growth. The best growing sites should be

chosen because old-growth characteristics would be

achieved more expeditiously than on poor sites. Moir

and Dieterich (1990) suggested that 150- to 200-yearold ponderosa pine (blackjack pine) in open stands

with no dwarf mistletoe be selected as the best stands

to begin developing old growth. Through sequential

silvicultural and fire treatments, the stands should be

relieved of wildfire hazards and competition, allowing

concentrated growth on a chosen group of trees. A

long-term commitment is necessary, because another

century may be needed before select old-growth pine is

represented (Moir and Dieterich 1990). If younger

stands are chosen for old-growth replacement, a greater

commitment of time is required for thinning, slash

disposal, commercial harvesting, and fire application.

Mixed Fire Regimes _____________

Major Vegetation Types

The pinyon-juniper woodlands (fig. 6-5) cover approximately 47 million acres (19 million ha) in the

Western United States (Evans 1988) and are characterized by a large number of diverse habitat types that

vary in tree and herbaceous species composition, and

stand densities. Climatic and physiographic conditions vary greatly within the range of this vegetation

type. Pinyon-juniper woodlands in the United States

are commonly divided into the Southwestern and the

Great Basin woodland ecosystems based on species

composition. True pinyon is common in the Southwest

and is usually associated with one or several species of

junipers, including one-seed, Utah, alligator, and Rocky

Mountain junipers. Singleleaf pinyon is identified

with the Great Basin and is generally associated with

Utah juniper. Other species of pinyon occur in southern California, Arizona, south of the Mogollon Rim,

along the United States-Mexico border, and in Texas

(Bailey and Hawksworth 1988). Several other species

of junipers also are found in the West; one of the more

Figure 6-5—Pinyon-juniper woodlands distribution.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 129

Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen

common is western juniper, which is found mainly in

Oregon and eastern California. Stand densities and

composition vary by elevation as it affects available

moisture; drier sites tend to be occupied by junipers

that are widely spaced and of low stature. Many of

these sites are often classified as savannas. Higher

elevation sites tend to be dominated by relatively

dense stands of pinyon trees of comparatively tall

stature and good form.

This report includes western oak species of obvious

concern to resource managers but it does not include

all oaks found in the Western United States (fig. 6-6,

6-7). Discussion concentrates on the important treeform deciduous and live oaks of California and of the

Southwestern United States (such as Gambel oak and

Arizona white oak). These are generally addressed as

a group. Little information has been documented for

these species (McPherson 1992), but their importance

to resource and fire management requires a beginning.

Shinnery, predominantly composed of sand shinnery

oak, is described as a separate ecosystem (fig. 6-8).

The Texas savanna (fig. 6-8) as a mapped ecosystem

occupies major portions of the Rio Grande Plains of

south Texas, the Edwards Plateau of south central

Texas and portions of the Rolling Plains, Grand Prairie,

North Central Prairies, Blackland Prairies, and Cross

Timbers. It corresponds roughly with Sections 315C,

D, and E of Bailey’s Ecoregions and Subregions of the

United States (Bailey and others 1994) and with major

Figure 6-6—Western oak distribution.

Figure 6-7—Western oak woodlands, Camp Roberts Military Training Reservation, Paso Robles,

California.
130 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

Figure 6-8—Distribution of shinnery and Texas savanna FRES

ecosystems.

Texas Savanna

Shinnery

portions of the Rio Grande Plain and Edwards Plateau

vegetation regions found in Box and Gould (1959). As

a plant community type, however, it has significant

representative elements that extend far north into the

southern portion of the Plains Grasslands ecosystem.

In fact, in the original Rainbow series volume, “Effects

of Fire on Flora” (Lotan and others 1981), the area

associated with the Texas savanna was lumped into

one huge “Prairie Grasslands” type (which also included the grasslands of the “Great Valley” in California). The vegetation of the “Texas savanna” can be

found in the northern portions of Texas in the southern Rolling Plains, the Grand Prairie, the North Central Prairies, the Blackland Prairies, and Cross Timbers, and extends to just south of the Texas Panhandle

area (Box and Gould 1959). These other areas will be

considered as part of the Texas savanna for the purposes of this publication. These areas receive 20 to 30

inches of precipitation annually—more than half of

which falls during the warmest months, and less than

a quarter during the period from December through

March (Garrison and others 1977).

The vegetation is a savanna with an overstory layer

of low trees and shrubs that varies from dense to open.

This overstory is of variable composition, having broadleaved and needle-leaved, deciduous and evergreen

species that predominate. These are mesquite, acacias, oaks, junipers, ceniza, and prickly pear species.

Honey mesquite is the most widespread woody plant

in the Texas savanna type and will receive the most

discussion. The grass of the Texas savanna varies from

short (<2 inches) to medium-height (2 to 12 inches), and

the herbaceous vegetation in general varies from dense

to open. These understory plants are mainly bluestems,

indiangrass, and switchgrass in the northeast, gramas, buffalograss, Texas wintergrass and Sporobolus

spp. in the south, central, and northwest, and curlymesquite and tobosagrass in the west and on the

Edwards Plateau. The particular mix of vegetation or

specific plant community that one might encounter

seems to be well correlated with soil orders, which are

variable in the Texas savanna system (Garrison and

others 1977).

Fire Regime Characteristics

Long-term fire frequencies for the pinyon-juniper

woodlands have not been clearly defined and are the

topic of continuing study and discussion. However,

there is agreement that fire was the most important

natural disturbance before the introduction of livestock, particularly the large herds in the 19th century

(Gottfried and others 1995). It is suspected that prior

to the introduction of heavy livestock use, large areas

of savanna and woodland periodically burned. These

fires could have occurred during dry years that followed wet years when substantial herbaceous growth

developed (Rogers and Vint 1987; Swetnam and Baisan

1996).

In the Intermountain West, presettlement mean

fire intervals of less than 15 years were documented in

the sagebrush steppe where western juniper now

dominates (Miller and Rose 1999). Other knowledge

that clearly documents the fire frequency, extent, and

seasonality of long-term fire regimes was developed

from a few studies at the upper limit of the pinyonjuniper type where it occurs with ponderosa pine. Fire

scars are rare in living pinyon pines due to the tree’s

susceptibility to damage by fire or to rot fungi that

enter resulting wounds. Fire scars have been noted on

junipers but most members of this genus are difficult

to age because of missing and false rings. Nonetheless,

some fire frequencies have been determined for the

Southwest. A sample of fire-scarred pinyon trees from

three locations in the Sacramento Mountains in New

Mexico indicated a mean fire interval of 28 years with

a range of 10 to 49 years (Wilkinson 1997). Despain

and Mosley (1990), working in the pinyon-juniper and

ponderosa pine ecotone at Walnut Canyon National

Monument in Arizona, reported a surface fire interval

of approximately 20 to 30 years. Other studies by C.

Allen and by T. Swetnam and his associates (Gottfried

and others 1995), on productive sites in New Mexico,

indicated that standwide fires, which covered more

than 25 acres, occurred at 15 to 20 year intervals.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 131

Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen

Dense pinyon-juniper stands (450 tree/acre or

greater) can burn in crown fires under extreme weather

conditions, generally low relative humidity and high

wind speeds. The key conditions are a closed canopy to

allow the spread of fire through the crowns and abundant dead material on the ground and as snags

(Gottfried and others 1995). It appears that presettlement fire regimes in dense stands were a mixture

of surface and crown fires, and that intensities and

frequencies varied depending on site productivity. The

Walnut Canyon site probably sustained patchy surface fires at intervals of 10 to 50 years and could carry

crown fires at intervals of 200 to 300 years or longer.

On less productive sites with discontinuous grass

cover, fires were probably infrequent and burns were

small and patchy. Fire frequencies were probably

greater than 100 years in these areas, but did occur

more frequently under extreme conditions (Gottfried

and others 1995). However, where grass cover was

more continuous, fire frequencies were probably more

frequent (10-year interval or less) and tended to maintain these sites as savannas or grasslands. Surface

fires would kill oneseed juniper trees less than 3 to 4

feet (1 m) tall (Johnsen 1962) but would have less of an

impact on older, larger trees that have thicker bark

and high crown base heights that exceed flame lengths.

This relationship between height and susceptibility to

fire also has been observed in western juniper stands

(Dealy 1990) and in Ashe juniper stands in Oklahoma

(Wink and Wright 1973). Fast moving surface fires in

the Southwest often do not burn near the trunks of

larger trees because the litter layer does not ignite.

In the Great Basin, fire susceptibility depends on

the stage of stand development (Meeuwig and others

1990). In young open stands, shrubs and herbaceous

cover may be sufficient to carry fire, but this cover

declines with time and eventually becomes too sparse

as the trees develop. The trees, however, may still be

too widely spaced to carry crown fires, except under

severe conditions.

In recent centuries, fire regimes in Western oak

forests were characterized by frequent, low intensity

fires. This was probably due to use of these types by

Native Americans, who probably carried out programs

of frequent underburning. Higher intensity fires at

long intervals have become more likely in the last half

of the 20th century.

Few data are available on fire frequencies within the

Texas savanna (Fuhlendorf and others 1996). With

understory fuels usually exceeding 2,240 lb/acre (2,000

kg/ha) each year under undisturbed conditions, it

is quite likely that fire frequencies were less than

10 years, and potentially more frequent in the northeast portion of the Texas savanna. Fires occur most

frequently during February and March when most

grasses are dormant and lightning strikes occur

commonly, and from July to September when grasses

are dry. Both winter and summer fires with ample fuel

loading in the grass understory can topkill trees resulting in major alteration of the woody physiognomy.

However, woody plant mortality and stand-replacements are rare. Winter fires that occur with low

understory fuel loadings can result in partial removal

of the overstory (Ansley and others 1995, 1996b).

Species such as mesquite, redberry juniper, and live

oak sprout if topkilled by fire and are rarely removed

from the vegetation complex by fire. However, Ashe (or

blueberry) juniper, which occurs in south-central Texas,

can be killed by fire and replaced by herbaceous

vegetation.

Fuels

Pinyon-Juniper—The main fuel consideration is

the amount of fine fuels, which varies with habitat

type, stand history, and climatic conditions. Fuel loading information for woody material is not readily

available; however, Perry (1993) measured an average

of 20 tons/acre (45 t/ha) after a pinyon-juniper clearcutting operation in Arizona; this stand produced

about six cords/acre of fuelwood. Fuel loadings of more

than 11 tons/acre are considered heavy. Slash left in

partially harvested woodlands may provide fuel ladders for ground fires to spread into the canopies. Grass

understory loadings can range from sparse to abundant (200 to 600 lb/acre). Typical crown fuels are 3.6

tons/acre (8.1 t/ha) for foliage and 1.8 tons/acre (4.0 t/ha)

for 0 to 0.25 inch branchwood (Reinhardt and others

1997).

Western Oaks—Fuels are quite variable between

stands, depending upon species, site, and stand condition. For example, a closed-canopy canyon live oak

forest may have little or no live understory. Surface

fuels will be made up of leaf and branch litter and the

amount will depend upon the time since last fire in the

stand. A more open stand may have an understory of

shrubs and nonwoody species. A closed forest of a

deciduous species, for example California black oak,

may well have an understory of annual grass; but a

more open woodland of the same species may have a

mix of grass and shrubs as an understory. In the latter

case, the combination of grass and shrubs can provide

a fuel ladder complex with associated erratic and

potentially dangerous fire behavior.

The aerial fuels in these oak stands are variable too.

Little information exists to characterize the deciduous

species; however, the live oaks can be thought of as

roughly comparable to chaparral in terms of crown

fuel character—both being sclerophyllous in nature.

The green material in these species will burn if fuel

moisture is low enough.
132 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

Texas Savanna—The predominant fuel that contributes to a fire’s propagation is the herbaceous understory. However, if the mesquite overstory has dead

stem material, it can be ignited and potentially kill the

plant. Britton and Wright (1971) found that up to 24

percent of mesquite that had been sprayed with a

topkilling herbicide were killed with fire that occurred

4 years after spraying. The standing dead stems burned

into live root crowns. When the overstory is dense—

either from a high density of individuals, or from dense

resprouted material—a crown fire can be sustained,

given the necessary wind and moisture conditions.

Such a high density overstory can be found as a phase

in Texas savanna stands. Mesquite crown fires would

only occur in summer months because the plant is

winter deciduous. However, other species of the savanna complex, such as junipers and live oak, could

carry crown fire any time of the year.

Herbage production, which indicates potential fine

fuel loading in the understory, was divided into four

major productivity classes (Garrison and others 1977):

Class Productivity (lb/acre)

1 2,250 to 3,000+

2 1,500 to 2,250

3 750 to 1500

4 0 to 750

Postfire Plant Communities

Pinyon-Juniper

Pre-1900 Succession—The pinyon-juniper woodlands are diverse, and successional pathways differ by

habitat type throughout the West. Traditional succession toward a “climax” vegetation considers the continuous replacement of one community by another.

The driving force in the successional process is competition among plant species of different genetically

controlled capabilities responding to changes in the

environment (Evans 1988). In the woodlands, succession involves the same species but in different amounts

and dominance over the landscape. Several successional seres following stand replacing fires have been

proposed for the Southwestern or Great Basin pinyonjuniper woodlands. Most of the successional projections are based on stands that had been grazed in the

past. Arnold and others (1964), working in northern

Arizona, developed one of the first models. A model for

southwestern Colorado (Erdman 1970), similar to that

of Arnold and others (1964), progresses from skeleton

forest and bare ground, to annual stage, to perennial

grass-forb stage, to shrub stage, to shrub-open tree

stage, to climax woodland. This pattern takes

approximately 300 years; however, new fires could set

back succession before the climax is achieved. Arnold

and others (1964) indicated that tree reoccupation

progressed from the unburned stand inward toward

the center of the burn. Barney and Frischknecht

(1974) reported a sere for a Utah juniper stand in westcentral Utah where pinyon was a minor component.

This ecosystem has had a long history of heavy grazing

since the late 19th century. The postfire progression

went from skeleton forest and bare ground, to annual

stage, to perennial grass-forb stage, to perennial grassforb-shrub stage, to perennial grass-forb-shrub-young

juniper stage to shrub-juniper stage, and to juniper

woodland. Junipers were well developed 85 to 90 years

after a fire. They indicated that the speed of tree recovery

would depend on the stage of tree maturity at the time of

the fire; older seed producing stands would recover more

rapidly than younger, immature stands. They noted the

importance of animal transport and storage of juniper

seeds in the speed of tree recovery. A new juniper could

start producing seed within about 33 years of establishment, hastening tree recovery.

Post 1900 Succession—Data on successional trends

apparent in the 1900s show that on similar sites

succession may follow several pathways (Everett 1987a;

Everett and Ward 1984). Shrubs, rather than annuals,

have been the initial vegetation on some burned sites

(Everett and Ward 1984), while the shrub stage may

be reduced or absent on some New Mexico sites (Pieper

and Wittie 1990). Predicting the course of succession

is difficult since it depends on a number of factors

(Everett 1987a). Specific successional pathways depend on fire severity and related damage to the original vegetation, area burned, available seed sources

either in the soil or from adjacent areas, species fire

resistance and ability to reproduce vegetatively, site

conditions, and climatic parameters throughout the

successional process. Everett and Ward (1984) indicated that the “initial floristic model” is appropriate

after a burn; initial species composition and density

may be as or more important than the progressive

succession. Most preburn species returned within 5

years of a prescribed burn in Nevada (Everett and

Ward 1984) and in southern Idaho (Bunting 1984).

The major human influence on the pinyon-juniper

woodlands and fire’s role in these ecosystems has been

ranching. Most of the Western rangelands were overgrazed, especially in the period following the 1880s.

Some areas around the Spanish controlled areas of

New Mexico have been heavily grazed since the 16th

century. Overgrazing has had an important effect on

the role of fire in the woodlands. The reduction of cover

of herbaceous species resulted in insufficient fuels for

fires to spread and to control tree establishment. Fires

ignited by lightning or humans tend to be restricted in

space. Fire suppression activities by land management

agencies also reduced the occurrence of fires. 

Woodland and savanna stand densities have increased throughout most of the West. Some people
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 133

Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen

believe that the woodlands have invaded true grasslands because of the lack of fire, but this is open to

debate (Gottfried and Severson 1993; Gottfried and

others 1995; Johnsen 1962; Wright and others 1979).

Climatic fluctuations, such as the drought in the

Southwest in the early 1950s, and global climate

change also have affected the distribution of woodlands

in the West. In the Intermountain West, Miller and

Rose (1999) quantitatively established that the cooccurrence of wet climatic conditions, introduction of

livestock, and reduced role of fire contributed to the

postsettlement expansion of western juniper. Prior to

1880, fire was probably the major limitation to juniper

encroachment. Other human influences related to the

harvesting of wood products by early American Indians (Gottfried and others 1995) and the harvesting of

large quantities of fuelwood to make charcoal for the

mines and domestic wood for supporting populations

in Nevada (Evans 1988) and near Tombstone in Arizona.

Management Considerations—During the 1950s

and 1960s, large operations were conducted to eliminate the pinyon-juniper cover in the hope of increasing

forage production for livestock (Gottfried and Severson

1993; Gottfried and others 1995). Other objectives

were to improve watershed condition and wildlife

habitat. Mechanical methods, such as chaining and

cabling, were used and resulting slash was piled and

burned. Burning these large fuel concentrations generated high heat levels that damaged soil and site

productivity (Tiedemann 1987). Many of these piled

areas were sterilized and remain free of vegetation

after over 20 years. Individual tree burning was used

on some woodland areas. Most of the control operations failed to meet their objectives. Many areas failed

to develop sufficient herbaceous cover to support renewed periodic surface fires.

A relatively undisturbed site with a rich variety of

understory species may recover differently than an

abused site with little understory development. Similarly, an older stand of junipers with a less diverse

population of perennial species will recover differently

than a younger stand (Bunting 1984). Burning in

stands with few desirable understory species may

worsen the ground cover situation, and depending on

the characteristics of the tree component, destroy a

valuable wood resource (Everett 1987b). A potential

problem exists if the preburn or adjacent vegetation

contains undesirable species, such as red brome. Very

hot fires can seriously slow initial succession of desirable species (Bunting 1984). Everett and Ward (1984)

indicated that relay floristics, the migration of species

into the site, is more important for the later stages of

development. Wink and Wright (1973) found that soil

moisture was important in determining rate of understory recovery; it is more rapid when soil moistures are

high. Dry conditions may increase drought stress of

surviving herbaceous plants (Wink and Wright 1973)

and retard seed germination. Aspect and elevation can

be used to predict some general successional trends

(Everett 1987a).

Currently, prescribed fire is used to reduce accumulations of slash from fuelwood harvesting or to reduce

or eliminate the tree cover in an attempt to increase

range productivity and biodiversity. In Arizona, slash

is usually left unpiled. Small piles are constructed

occasionally and are burned as conditions and crew

availability allows. There is increasing interest in

managing the pinyon-juniper woodlands for sustained

multi-resource benefits including, but not limited to,

tree products, forage, wildlife habitat, and watershed

protection (Gottfried and Severson 1993). This is particularly true for high site lands that have the ability

to produce wood products on a sustainable basis.

Prescribed burning to dispose of slash is less desirable

in partially harvested stands, where the selection or

shelterwood methods have been used to sustain tree

product production. Burning tends to damage residual

trees, especially where slash has accumulated at the

base, and advance regeneration. Established, smaller

trees are particularly important for the next rotation

because of the difficulty of achieving adequate regeneration of these relatively slow growing species. It may

be desirable to move slash away from areas of satisfactory regeneration prior to burning or to avoid burning

in them.

Several different slash disposal options may be applicable to any one management area (Gottfried and

Severson 1993). Burning of large piles is unacceptable

because of soil site degradation (Tiedemann 1987) and

no longer recommended in the Southwest (USDA

Forest Service 1993). However, small piles of slash

may be burned in low intensity fires to encourage

floristic richness or to promote temporary increases of

nutrient content in herbaceous vegetation. Piled or

unpiled slash can also be left unburned to provide

habitat for small mammals or to break up sight distances for wild ungulates. It also can be scattered to

provide protection for establishment of young trees

and herbaceous species, and to retard overland runoff

and sediment movement.

Mechanical methods of clearing pinyon-juniper are

increasingly expensive, but prescribed fire is an economical alternative. The method used in Arizona is to

ignite the crowns from prepared fuel ladders of cut

lower limbs that are piled around the base of the tree.

Ladders are ignited one season after the limbs are cut.

In denser stands, fire spreads into the crown layer and

through the stand from fuel ladders that are created

below strategically placed trees. A method used in

central Oregon on sites converted to juniper from

sagebrush/grass is to conduct prescribed fires several

years after harvesting trees. The increased production
134 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

of herbaceous vegetation following cutting provides

fuels to carry the fire, which reduces residual slash

and kills juniper seedlings.

Research in the Great Basin suggests that fire works

best on sites with scattered trees (9 to 23 percent cover)

where the trees begin to dominate the understory and

in dense stands (24 to 35 percent cover) (Bruner and

Klebenow 1979). Wright and others (1979) indicated

that prescribed spring burning was successful in sagebrush/pinyon-juniper communities. Bruner and

Klebenow (1979) recommended an index to determine

if a fire will be successful or if conditions are too

dangerous. This index is based on the addition of

maximum wind speed (mi/hr), shrub and tree cover

(percent), and air temperature (°F). Burning can be

successful if scores are between 110 and 130. Dense

stands where pinyon is more common than juniper are

easier to burn than pure juniper stands (Wright and

others 1979). Bunting (1984) indicated that burning of

western juniper stands in southwestern Idaho was

only successful during the mid-August to mid-September period; burning in the fall did not achieve desired

results because of low temperatures, low wind speeds,

and lack of fine fuels. Prescribed fire can be used in

previously treated areas to control new tree regeneration. This technique works best if the area is ungrazed

for one or two seasons prior to burning. Wink and

Wright (1973) reported that a minimum of 890 lb/acre

(1,000 kg/ha) of fine fuels is needed to burn and kill

Ashe juniper seedlings and to burn piled slash. Success where alligator juniper dominates has been limited because of the trees’ ability to sprout, so prescribed fire

is not recommended (USDA Forest Service 1993).

Ecosystem Management—Reintroducing low intensity fire into the pinyon-juniper woodlands could

help meet ecosystem management goals. For example,

prescribed fire could be used after harvesting to limit

tree regeneration and to maintain overstory stand

densities that would promote vigorous understory

vegetation for livestock and wildlife. Fire could be

used during the earlier part of the rotation period,

when crown cover is less, and modified later to protect

adequate tree regeneration. The prescription would

vary by the amount and condition of woody debris in

the stand so that stand replacing crown fires are

prevented. Pockets of regeneration could be protected.

Fire could also be used to maintain herbaceous cover

dominance in natural savannas and ecotonal grasslands. However, as indicated above, all surface fire

options would require that the land be rested from

grazing prior to treatment so that sufficient fuels can

develop to carry the fire. It usually requires 600 to 700

lb/acre (672 to 784 kg/ha) of fine fuel to carry a fire in

the Great Basin (Wright and others 1979).

Fire has also been used to create mosaics of woodland

and openings within some Southwestern landscapes.

Mosaics are beneficial to wildlife and livestock

(Gottfried and Severson 1993) and can create an

aesthetically pleasing landscape. Aerial and ground

firing techniques have resulted in mosaics on some

juniper/mesquite grasslands in southern Arizona.

Western Oaks

Pre-1900 Succession—There is little doubt that

western oak trees evolved over a time when climatic

change was occurring and when disturbance including

fire was common. The deciduous or evergreen habit

probably is related to environmental moisture—evergreen oaks belonging to more arid systems (Caprio and

Zwolinski 1992; Rundel 1987). Postfire succession

during pre-Euro-American settlement was probably

much like the dynamics that we see today, but there

were probably more oaks than we find today. Some

species were easily top-killed; many species sprouted

in response to fire.

Post-1900 Succession—The current reduction in

the occurrence of the oaks in many areas may be due

to a number of factors, including increased fire severity,

grazing, overt removal to provide more pasture land,

and urban encroachment. Fire is probably not the

primary factor, but it can kill a stand of oaks outright.

Some oaks are more easily top-killed than others,

which is generally a function of bark thickness. See the

categorization of oak sensitivity to fire by Plumb and

Gomez (1983). Almost all of the oak species sprout

after fire, if root crown or underground portions are

still alive (Plumb 1980).

Management Considerations—In some parts of

the West, oaks have become subjects of intense resource management interest. The ranges of some species have become severely reduced; some species do

not seem to be reproducing at a desired rate (Bartolome

and others 1992). Competition to seedlings from understory vegetation may be hampering seedling survival (Adams and others 1992); grazing may play a

part as well. Effective management of these species

has yet to be established. The use of prescribed fire as

a means of reducing competition and opening up

closed canopy stands is being attempted (Clary and

Tiedemann 1992). Although results are not definitive

yet, it shows promise. For now, the use of prescribed

fire in western oaks should be approached with caution and patience. Some species are sensitive to fire

(table 2-1) but may survive under certain conditions

(Paysen and Narog 1993). Many oaks seem to be prone

to disease, such as heart rot. Injury from fire or other

treatment may not kill a tree, but might conceivably

inflict damage that could provide a port of entry for

disease. Much research remains to be done on these

species. For now, management treatments should be

carried out carefully.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 135

Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen

Texas Savanna

Pre-1900 Succession—Historical accounts differ

as to original density and distribution of mesquite in

Texas. Bartlett (1854) described much of Texas rangeland as open grasslands with scattered large mesquite

(a mesquite savanna). Marcy (1866) described some

upland areas of central Texas as “covered with groves

of mesquite trees,” and an area in the lower Texas

Panhandle as “one continuous mesquite flat, dotted

here and there with small patches of open prairie.”

These observations suggest that honey mesquite was

a natural part of the northern Texas vegetation complex prior to Euro-American settlement and, apparently in some instances, occurred as dense stands.

There is no indication as to the growth form of mesquite trees prior to Euro-American settlement. Fire

was a part of the environment when these explorers

traveled through Texas (Wright and Bailey 1980), but

the specific role it played in shaping the scenes they

observed is difficult to know. However, biological agents

and fire are credited with having limited mesquite

densities on rangelands before Euro-American settlement in the Southwest (Jacoby and Ansley 1991).

Post-1900 Succession—Honey mesquite density

increased in the Southwest during the 20th century. It

is likely that most of the multistemmed thickets that

occur in Texas today have greater stem and foliage

density because of increased anthropogenic disturbance of the canopy (including use of fire to topkill

shrubs, which induces sprouting) than would have

occurred naturally. Individual shrub densities have

increased since the late 19th century as well. This has

also been attributed to human influence—either

through suppression of natural fires, or dissemination

of mesquite seed by the herding and migration of

domestic livestock (Brown and Archer 1989).

Much of the vegetation in the Southwest is in a state of

flux and may have been changing for centuries in many

areas. This seems to be particularly true for the Texas

savanna type. Its dynamics, however, may have been

accelerated by the influence of recent human activities.

Intensive animal grazing coupled with extremes of

climate may be instrumental in causing active fluctuation of vegetation composition and physiognomy.

Domestic livestock have played a major role in dissemination of mesquite seed into mesquite-free areas

(Archer 1995; Brown and Archer 1989). Observations

of recently seeded Conservation Reserve Program

(CRP) stands on cropland near mesquite stands indicates that in the absence of cattle grazing, mesquite

seeds were probably deposited by wildlife (coyotes,

hogs, birds). However, this appears to be restricted to

the margins of already existing mesquite stands. Early

settlers accelerated dissemination into mesquite-free

areas first via the cattle drives that occurred about 1900,

and second with continuous grazing within fenced areas.

Current landscape patterns may reflect a trend that

has been ongoing for centuries, or phases in a pulse

equilibrium that may exist in much of the Southwest.

The current pattern may depend upon recent combinations of weather and human activity. Mesquite

encroachment, or encroachment of other woody species would probably occur in the absence of domestic

livestock grazing, but such grazing has probably accelerated this process.

Management Considerations—Historically, the

Texas savanna has provided a home to an abundance

of wildlife. But, in recent times, land clearing for

agricultural purposes has reduced the habitat for

some of these species (Garrison and others 1977).

Livestock grazing has been a predominant factor in

managing this vegetation type. The woody overstory

plants of the savanna, especially mesquite, have been

viewed as pests by most landowners. Mesquite’s thorny

branches, increasing density on rangeland, and perceived competition with forage grasses have made it

the target of eradication efforts over recent years.

Chemical and mechanical controls have been the primary agents used in this effort (Fisher 1977). More

recently, fire has gained increased acceptance as a

management tool (Wright and Bailey 1982).

Mesquite now has an emerging image as a resource

that should be managed rather than eradicated (Ansley

and others 1996a; Fulbright 1996; Jacoby and Ansley

1991). Unfortunately, decades of control attempts

have destroyed many mature stands of mesquite that

contained single to few-stemmed trees. These trees

were desirable in that they occupied far less surface

area than multistemmed growth forms that resulted

from destruction of aerial tissue and subsequent

resprouting. Complete elimination of mesquite has

been a goal that few landowners have achieved, and

the concept of complete removal is questionable, both

economically and environmentally (Fisher 1977).

Mesquite has many potential benefits to the ecosystem when maintained at controlled densities such as

in a savanna. Such benefits include nitrogen fixation,

livestock shade, habitat for nesting birds, and the

potential as firewood or wood products. Mesquite has

the potential to produce commercial hardwood in some

regions with higher rainfall (Felker and others 1990).

In lower rainfall areas, shrubby growth forms of mesquite can have other benefits, such as wildlife habitat.

A mesquite savanna offers a pleasant landscape and

may improve the value of a property over either an

unmanaged woodland or a treeless grassland.

Recent research suggests that mesquite savannas

can be sustained as long as the herbaceous understory

is maintained at sufficient densities to out compete

mesquite seedlings (Archer 1989; Brown and Archer

1989; Bush and Van Auken 1990). A savanna of this

nature can be created and maintained in large part by
136 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

using prescribed fire—one of the more environmentally acceptable and most economically sustainable

options for managing woody plants (Ansley and others

1996a). In the initial stages of stand treatment, herbicides may be a useful supplement to the use of prescribed fire. However development and maintenance

of the desired savanna growth form can often rely on

the use of low-intensity fire, which can be achieved by

burning under certain fuel loadings, humidities, and

air temperatures. Creating a savanna from thickets

using low-intensity fires will take time and should be

part of a long-term management plan.

Response of honey mesquite to fire is highly variable

and is a function of fine understory fuel loading and

condition and of season of the year (Ansley and others

1995; Lotan and others 1981; Wright and others 1976).

Abundant fine fuels tend to produce hotter fires and

result in more topkill of the woody plants than lighter

loadings (Wright and Bailey 1982). Summer fires will

produce more topkill than winter fires (Ansley and

others 1998). Fine herbaceous fuel loading and season

of the year can work in various combinations to produce partially defoliated mesquite, or completely

topkilled mesquite that quite often produces abundant sprouts from the root crown. Mesquite age also

affects survival of individual plants after fire. Individual trees 1.5 years of age or less are easily killed by

a fire when the soil surface temperatures are above

500 °F (260 °C) (Wright and others 1976). At 2.5 years

of age, they can be severely harmed, and if older than

3.5 years, they are seemingly fire resistant at these

soil temperatures.

Stand-Replacement Fire

Regimes _______________________

Major Vegetation Types

The major vegetation types within this fire regime

type are varied. Broadly, they include grassland and

shrubland vegetation types (fig. 1-2).

Grasslands

The grassland types (fig. 6-9) include:

• The plains grasslands, which range from Canada

south to northern Texas in a broad swath that

covers much of the Mid-Western United States.

• The mountain grasslands, which consist of open,

untimbered mountainous areas from Canada south

through the Northern and Central Rocky Mountains and the Coastal Range.

• The desert grasslands, which occur in the Southwestern States and in the Great Basin.

• The annual grasslands, which are concentrated

for the most part in the valleys and foothills of

California and along the Pacific coast.

• Cheatgrass (fig. 6-10), which has invaded and

gained dominance in many plant communities in

the Intermountain and Columbia Basin regions

(Monsen 1994).

Figure 6-9—Distribution of plains, mountain, desert, and annual grassland FRES ecosystems.

Plains Grasslands

Annual Grasslands

Mountain

Grasslands

Desert

Grasslands
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 137

Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen

Shrublands

Shrublands are described here as desert shrubland

types and the chaparral-mountain shrub type. Desert

shrublands transcend North America’s four major

deserts—Mojave, Sonoran, Chihuahuan, and Great

Basin (fig. 6-11, table 6-3). These deserts encompass

about 500,000 square miles (1,717,000 km2) within

the physiographic Basin and Range Province, surrounded by the Rocky Mountains and Sierra Nevada

in the United States, and the Sierra Madre Occidental

and Sierra Madre Oriental in Mexico (MacMahon

1988; MacMahon and Wagner 1985). They are characterized by low but highly variable rainfall, 10 inches/

year (25 cm/year), and high evapotranspiration. Each

desert differs in precipitation patterns, temperature

variables, and vegetation structure (Burk 1977;

Crosswhite and Crosswhite 1984; MacMahon 1988;

MacMahon and Wagner 1985; Turner and Brown

1982; Turner and others 1995).

Bailey’s (1978) Desert Division includes Mojave,

Sonoran, and Chihuanhuan Deserts, considered warm

deserts because their precipitation is mostly rain. The

Mojave receives winter rainfall, the Chihuahuan summer rainfall, and the Sonoran both. Winter rainfall

tends to be of long duration, low intensity, and covers

large areas, whereas summer rainfall is of short duration, high intensity, and covers limited areas

(MacMahon 1988). The Mojave Desert has greater

elevation and temperature variations than the Sonoran

Figure 6-10—Cheatgrass.

Figure 6-11—Distribution of desert shrub FRES ecosystems.
138 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

Table 6-3—Physiognomic fuel types for desert shrublandsa associated with the four North American deserts.

North American deserts

Desert shrublands Chihuahuan Sonoran Mojave Great Basin

Sagebrush F-29

Great Basin sagebrush K-38 X X

Desert shrub F-30

Blackbrush K-39 X X

Saltbush/greasewood K-40 XXXX

Creosotebush K-41 X X X

Creosotebush/bursage K-42 X X

Mesquite bosques K-27 X X X

Paloverde/cactus shrub K-43 X X

Southwestern shrubsteppe F-33

Grama/tobosa shrubsteppe K-58 X X

Trans-Pecos shrub savanna K-59 X X

a

FRES (F) shrubland ecosystems and the Kuchler Potential Vegetation System (K) equivalents (Garrison and others 1977).

Desert, which is lower, flatter, and warmer. Although

the Chihuahuan Desert lies south of the Sonoran, it

varies more in elevation and has colder winters. The

Mojave Desert is considered transitional between the

Sonoran and Great Basin Deserts, respectively, sharing components of each at its extreme southern and

northern ends. The Great Basin desert is considered a

cold desert because its precipitation is primarily snow

(MacMahon 1988).

Vegetation in these regions varies from predominantly shortgrass prairie, consisting of sparsely distributed bunch grasses, to predominantly shrubs, sometimes with scattered small trees, and often with exposed

areas of soil (fig. 6-12). Desert and desert shrubland

vegetation has been classified in numerous ways

(Shreve and Wiggins 1964; Turner 1982; Turner and

Brown 1982; Vasek and Barbour 1977). We focused on

desert shrublands within the United States (table 6-1).

Figure 6-12—Bare soil, evident between shrubs and small trees, is a common characteristic

of North American deserts as seen in the Mojave Desert, California.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 139

Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen

Although these shrublands are distributed as a

continuum of natural ecosystems, the use of vegetation classification systems gives us a convenient functional format for making fire management decisions.

For our purposes, desert vegetation will be subdivided

according to the FRES ecosystems as organized in

table 6-3. We included the FRES sagebrush and Southwestern shrubsteppe types in our description of desert

shrublands based on their similar fuels types, geographical proximity, and species integration.

Great Basin Sagebrush—This type characterized

by sagebrush species (fig. 6-13) covers plateaus and

vast plains at elevations ranging between 1,600 and

11,000 feet (490 and 3,500 m) with varied soils derived

from lava flows, ancient lake beds, and alluvium

(Garrison and others 1977). The Great Basin sagebrush, the largest range ecosystem in the Western

United States, covers about 247 million acres (100

million ha) of arid lands (Blaisdell and others 1982).

Sagebrush and associates are valuable for soil stabilization, wildlife habitat, animal feed, and ecosystem

stability. There are about 22 species and subspecies;

some have been studied extensively (Harniss and

others 1981; Koehler 1975; Monsen and Kitchen 1994;

Roundy and others 1995; Tisdale and Hironaka 1981).

Sagebrush, composed of dwarf and tall sagebrush

species, range between 1 and 7 feet (0.3 and 2 m) tall

Figure 6-13—Distributiuon of Great Basin sagebrush FRES

ecosystems.

and grow in dense clumps or scattered plants.

Shadscale, spiny hopsage, Mormon tea, and milkvetch

are important co-dominants in this vegetation type.

Understory grasses such as wheatgrass, brome, fescue, and bluegrass, and variable forbs form discontinuous patches with bare soil.

Blackbrush—This type is composed of dense to

scattered low stature shrubs and dense to open grass

at elevations below 6,550 feet (2,000 m) (fig. 6-14).

Blackbrush is one of the least studied landscape

dominant shrubs in the United States. It prefers level

topography and is not common on slopes or in drainages (Lei and Walker 1995). It maintains the highest

cover of any desert shrub community. This transitional community between the Great Basin and the

Mojave Desert occurs where annual precipitation is

about 7 inches (18 cm) (MacMahon 1992). Moisture

may limit its range. Blackbrush usually occurs in

almost pure stands, although it intergrades with

creosotebush and bursage at lower ecotones and

sagebrush/juniper ecotones at higher elevations (Lei

and Walker 1995).

Saltbush-Greasewood—This shrubland is

characterized by halophytes and succulent subshrubs.

Vegetation dominants include shadscale, black greasewood, and saltbush with saltgrass, winterfat, and

sagebrush also present. This shrubland is common to

all four deserts (table 6-3) and occurs on approximately

42 million acres (17 million ha) on heavy depauperate

soil, often with underlying hardpan or alkaline flats. It

is found below the sagebrush zone, generally below

elevations of 6,900 feet (2,100 m). Saltbush and black

greasewood are dominant and co-dominant species

throughout much of their range from Canada to northern Mexico, eastern California to Colorado and northeast Montana.

Creosotebush—This vegetation consists of low to

medium-tall, typically open shrubs (fig. 6-15) that

grow on bajadas, valley floors, gentle slopes, sand

dunes, and in arroyos below 5,000 feet (1,500 m) in the

Mojave, Sonoran, and Chihuahuan Deserts.

Creosotebush is a widespread dominant or co-dominant that also forms transitional vegetation between

the three warm deserts. Creosotebush occurs in mixed

to pure stands of open, low but variable diversity plant

communities on about 46 million acres (18.4 million

ha) (Cable 1973).

Joshua Tree—In parts of the Mojave Desert,

creosotebush is associated with the Joshua tree woodland (fig. 6-16). Joshua trees can resprout after fire,

develop fire-resistant bark on trunks, have protected

apical meristems usually high above surrounding fuels,

and reseed from offsite sources. Resource managers at

the Joshua Tree National Monument in California are

testing prescribed burning as a tool to create fuel
140 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

Figure 6-15—Creosotebush shown growing

on the Mojave Desert valley floor may resprout

(inset) after fire.

Figure 6-14—Prescribed burning to reduce blackbrush fuels at the urban wildland interface,

Carson City, Nevada.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 141

Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen

breaks to reduce large-scale destruction of this unique

resource by wildfires (fig. 6-17).

Creosotebush-Bursage—This is a transitional

plant association found below 5,250 feet (1,610 m)

elevation. It merges with the paloverde-cactus shrub

association found in the Sonoran Desert. In this region

creosotebush-bursage has higher species diversity including a larger tree component (table 6-4).

Paloverde-Cactus Shrub—This type is characterized by open-to-dense stands of low-to-medium tall

shrubs, small trees, cacti, and succulents (fig. 6-18).

Paloverde, pricklypear, cholla, saguaro, and bursage

are dominant species in this vegetation type. These

communities are a diverse mosaic of mixed vegetation

that occur in the Sonoran Desert at elevations generally below 4,000 feet (1,200 m) (table 6-4).

Southwestern Shrubsteppe—This shrub type or

the semidesert grass-shrub type (called desert grasslands in the FRES system) is composed of gently

sloping desert plains found below the Rocky Mountains and between the low mountain ranges of the

Sonoran Desert, Mexican Highland, and Sacramento

section in Arizona, New Mexico, and Texas (fig. 6-19).

Annual precipitation in this ecosystem varies from 10

inches (25 cm) in western areas to 18 inches (46 cm) to

the east. Despite the fact that half of the rainfall occurs

during warm months (frost free periods occur 180 days

or more of the year), evapotranspiration is between 80

and 90 inches (203 to 229 cm) per year and may exceed

the precipitation by a factor of 10.

Vegetation is composed of short grasses and shrubs

of variable composition. Grasses inhabit the more

developed Aridisol and Mollisol soils. Shrubs inhabit

the shallow soils. Junipers occur exclusively on Entisols,

which are predominantly found in the South. Yucca,

mesquite, creosotebush, and tarbush are the dominant woody plants, while black grama, tobosa, and

threeawn are the dominant herbaceous plants. Curlymesquite and other grama species also contribute

significantly to the biomass of these shrubsteppe communities, which are used mainly as rangeland.

Two shrubsteppe types are recognized. The Gramatobosa shrubsteppe occupies areas at elevations

between 1,610 and 7,045 feet (488 to 2,135 m) and

includes the more shrub dominated communities of

the shrubsteppe. Black grama, sideoats, and tobosa

are climax indicators occupying arid grassland communities throughout the Southwest. Black grama

prefers more gravelly upland sites; sideoats is less

selective, while tobosa prefers heavier clay lowland

soils. The Trans-Pecos shrub savanna is found on

Figure 6-16—Joshua tree clones provide clusters of fuel in otherwise sparse desert shrublands, Mojave

Desert, California.
142 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

Figure 6-17—Prescribed burning in a Joshua tree forest to reduce fuel loading at the urban/wildland

interface, Covington Flats, Joshua Tree National Park, California.

the Stockton Plateau and southwestern portion of

Edwards Plateau. It has a higher average elevation

(4,000 to 6,000 feet; 1,220 to 1829 m) and greater

rainfall than the grama-tobosa shrubsteppe. This is a

shrub dominated type characterized by grasses and

the common occurrence of junipers (fig. 6-20). Junipers occupy more than 6 million acres (2.4 million ha)

of rangeland in dense to open communities with oaks,

Texas persimmon, and mesquite.

Chaparral-Mountain Shrub—This ecosystem

type (fig. 6-19, 6-21) occupies lower and middle elevation mountain areas in the Pacific States, the Southwestern States, and the Rocky Mountains. The

vegetation consists of dense to open shrubs or low trees

with deciduous, semideciduous, and evergreen species

represented. Some of the types are so dense that

understory vegetation is practically eliminated, while

other types support a highly productive understory.

Fire Regime Characteristics

Fire frequency was variable in the stand-replacement fire regime types and depended upon ignition

sources and plant community development. In the

grassland types, fires could occur in any given year,

provided the grass was cured and dry enough to burn.

Although fire frequencies could not be measured precisely, mean fire intervals probably ranged from about

4 to 20 years depending on climate and ignition sources

(Gruell and others 1985a). In the plains and grasslands, Native Americans ignited fires for a wide variety of cultural reasons. This was the predominant

source of ignition in heavy use areas particularly at

lower and middle elevations. But, an ever-present

ignition source was lightning, which was probably

more important in valleys surrounded by forests than

in plains grasslands due to differences in efficiency of

lightning (Gruell and others 1985b). Grasslands, occupying flat to gently rolling terrain, would burn over

large areas until a break in terrain or a change in

weather stopped the fires. Fires swept over extensive

areas sometimes covering several hundred square miles.

Desert shrublands have been influenced over the

last 12,000 years by climatic shifts, varying soils, and

fire. Prior to Euro-American settlement, fires in these

desert shrublands were set by lightning and Native

Americans (Humphrey 1974; Komerek 1969). Wyoming big sagebrush experienced fire intervals ranging

from 10 to 70 years (Vincent 1992; Young and Evans

1991). Arid land fire history studies report fire intervals between 5 and 100 years (Wright 1986). Griffiths

(1910) and Leopold (1924) reported that before 1880
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Table 6-4—Physiognomic and taxonomic descriptions of vegetation types modified from Kuchler (1964) showing habitat typea fuel,

and forage associated with each. Note: Although numerous grass species are not listed for each vegetation type, they

have become cosmopolitan throughout each type as a result of anthropogenic disturbance. Their impact on the fire

dynamics of these desert ecosystems should be considered in making fire management decisions.

Vegetationa

-Fuels (Fu) Dominant species

-Forage (Fo) •Associated genera Treeb Shrub Herb Cactus

Great Basin sagebrushc Artemisia tridentata S

Dense to open low to medium • Artemisia, Atriplex, Chrysothamnus, Coleogyne S

shrubs • Ephedra, Eriogonum, Tetradymia s

Fu-0 to 2,000 lb/acre • Astragalus, Lupinus, Phacelia H

Fo-0 to 700 lb/acre • Agropyron G

Blackbrush S

Dense to open broadleaf • Artemisia, Gutierrezia, Haplopappus S

evergreen shrubs • Ephedra s

+ herbaceous understory • Hilaria G

Fo-250-500 lb/acre

Saltbush/black greasewood Atriplex confertifolia/Sarcobatus vermiculatus S/s

Open small shrubs • Lycium, Artemisia, Atriplex, Grayia, Krascheninnikoviad S

Fu-250 to 750 lb/acre • Allenrolfea, Menodora, Suaeda ss

Fo-50 to 200 lb/acre • Kochia H

• Distichlis G

Creosotebush Larrea divaricata S

Open dwarf to medium shrubs • Yucca brevifoliae T

Fu-40 to 100 lb/acre • Lycium, Baccharis S

Fo-12 to 40 lb/acre • Encelia, Franseria, Sphaeralcea s

Creosotebush/Bursage Larrea divaricata/Ambrosia dumosa S/s

Open dwarf to medium shrubs • Cercidium, Dalea, Prosopis, Olneya T

Fu-40 to 100 lb/acre • Lycium, Acacia, Fouquieria S

Fo-12 to 40 lb/acre • Encelia, Franseria s

• Hilaria G

• Opuntia, Ferocactus c

Mesquite Bosques Prosopis glandulosa; P. velutina T

Open to dense forest low • Cercidium, Olneya, Prosopis, Populus, Dalea, Salix T

broadleaf deciduous trees • Acacia, Baccharis, Lycium S

Fu-250 to 1000 lb/acre

Fo-0 to 500 lb/acre

Paloverde/Cactus Shrub Cercidium microphyllum/Opuntia spp. T c

Open to dense low trees, • Cercidium, Olneya, Prosopis T

shrubs, and succulents • Jatropha, Larrea, Lycium, Simmondsia, S

Fu-100 to 250 lb/acre Acacia, Condalia, Fouquiera, Celtis

Fo-30 to 100 lb/acre • Calliandra, Ephedra, Franseria, Janusia s

• Carnegiea c

• Ferocactus, Echinocereus, Opuntia c

Grama-tobosa shrubsteppe Hilaria spp., Bouteloua spp. G

short grass with shrubs Larrea S

Fo-0-600 lb/acre Yucca spp. ss

- - - - - - -

Trans-pecos shrub savanna Juniperus spp. T

shrubs with short grass Hilaria spp., Bouteloua spp., Muhlenbergia spp. G

Fo-0-600 lb/acre

- - - - - - - -

aBased on Kuchler’s classification system bT = tree; S = shrub; s = subshrub; ss = succulent shrub; H = herbaceous; G = grass; c = cactus

cGreat Basin sage is broken into four productivity classes (Garrison and others 1977)

dEurotia lanata (Pursh) Moq. = Krascheninnikovia lanata (Pursh) A. D. J. Meeuse & Smit, (Jepson 1993) eYucca brevifolia (Joshua trees) become a significant tree component in parts of the Mojave Desert and grama-tobosa shrubsteppe
144 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

Figure 6-20—Mixed fuels found in juniper shrub savanna (New

York Mountains, California).

Figure 6-18—Mixed vegetation of the paloverde/cactus shrub in the Sonoran desert near Four Peaks,

Maricopa County, Arizona.

Figure 6-19—Distribution of Southwestern shrubsteppe and

chaparral-mountain shrub FRES ecosystems.

Chaparral and Mountain Shrub

Southwestern Shrubsteppe
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 145

Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen

desert grasslands produced more grass and fires recurred at approximately 10-year intervals. Before

settlement deserts were characterized by sparse vegetation, broken by barren soil, and were not expected

to burn except under unusual circumstances. But

when fire occurs in warm desert shrub habitat, a long

recovery is expected. This recovery depends on geographical location, species composition, and climatology after the burn. Recovery is more rapid in areas

receiving higher precipitation. The various desert

shrublands vary in wildfire risk ranging from nonexistent risk of the sparsely vegetated saltflats to high

risk associated with heavy fuel loadings often found in

the mesquite type. Postfire survival by desert plants

may depend on genetic variation (Munda and Smith

1995), resprouting capability, resistant seeds, and

delayed mortality.

In California chaparral, fire intervals for large fires

(more than 5,000 acres) typically ranged from 20 to 40

years (Wright and Bailey 1982). But at higher elevations and north aspects fire return intervals were

longer, perhaps as infrequent as 50 to 100 years.

Young stands of chaparral whose canopy has not

closed and stands that have not restocked well after

disturbance often have a grass component that can

burn on any given year, as is the case with the grasslands. These fires may or may not be stand-replacement fires, depending upon the amount of heat transferred from the grass component to the sparse shrub

overstory. Fully developed chaparral stands can be

difficult to ignite unless there is some component of

dead material and good fuel continuity. However,

given an ignition and some wind, they will propagate

a moving fire even when virtually no dead material

exists in them. Because these are crown fires, they are

almost always stand-replacement fires. With both the

grasslands and chaparral, all or most of the

aboveground portion of the plants are killed. Most of

the perennial grasses have a perennating bud at or

near ground level, often protected by bunched stems

that act as insulators; often, tufts of these stems

remain after fire. Chaparral shrubs are often killed

down to the root collar; sometimes the entire individual is killed outright.

Figure 6-21—Typical chaparral vegetation (Arctostaphylos, Ceanothus), Mill Creek, San Benardino National Forest, California.
146 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

Fuels

Grassland Fuels—When cured and dry, grassland

fuels are ideally suited for burning. For the most part,

they fall into the fine fuel category; however, the

compact arrangement of stems in the “tufts” of bunchgrasses makes these portions of the plant difficult to

ignite regardless of their dryness. Once ignited, however, they can smolder for long periods if enough old

stem material has accumulated.

Plant density is also a critical factor in a grassland’s

ability to propagate fire. Heat output is relatively low

from grass fuels, so fairly continuous fuels are necessary for fire spread to occur. Light winds can sometimes compensate for moderately sparse fuels by providing required flame bathing. The amount of fuel can

vary with site condition, precipitation, and disturbance history. Typical annual productivity in desert

grasslands can vary from next to nothing upwards to

1,000 lb/acre (1,120 kg/ha); in plains and mountain

grasslands, productivity can be as high as 2,000 lb/

acre (2,240 kg/ha) (table 6-5).

The character of a grassland fire is also affected by

the overall geometry of the stand, which changes

throughout the life cycle of the plants in the stand. The

most dramatic example of this can be seen in annual

grasslands where the plants germinate, seed, and die

in a single season. A stand of recently cured annual

grass can be quite dense and tall (up to 6 or 7 feet); its

bulk density can be optimum for propagating a fast

moving fire. In a relatively short period, a process of

stand collapse begins and the bulk density of the stand

becomes steadily modified. By the end of the season,

the biomass is in a dense thatch on the ground and will

begin decomposing—in some localities, fairly completely. Fire can still propagate during these later

stages, as long as not too much moisture has accumulated in the thatch, but spread rates will not be as great.

Cheatgrass is a highly flammable fuel because of its

finely divided plant structure, long period in a cured

condition, rapid response to drying, and a tendency to

accumulate litter (Bradley 1986a). Cheatgrass dries 4

to 6 weeks earlier than perennials and can be susceptible to fire 1 to 2 months longer in the fall. It produces

large quantities of seed that usually develop into

dense stands providing ideal fuel continuity for fast

spreading fires. It grows well in areas of low precipitation that frequently undergo severe fire seasons.

Desert Shrublands—Fuels include cacti and other

succulents, grasses, shrubs, small trees, and mixtures

of these. Fuels occur in discontinuous patches to areas

where trees, shrubs, and grasses are contiguous. Fuel

loadings may reach 2,000 lb/acre (2,240 kg/ha) (fig. 6-22,

6-23). See table 6-4 for fuel loading and forage production for each associated shrub community.

Fuel loading in sagebrush varies depending on the

site and species. Based on shrub height and percent

cover, big sagebrush varies from 0.26 to 4.6 tons/acre

(0.55 to 10.2 t/ha). For a stand 2.5 feet in height and 20

percent cover, conditions typically found, sagebrush

foliage and stemwood averages 1.5 tons/acre (Brown

1982). Herbage production for this vegetation type can

vary from about 200 lb/acre (224 kg/ha) under poor

growing conditions (Brown 1982) to 1 ton/acre (2.2 t/ha)

under favorable conditions (Garrison and others 1977).

Forage production generally is one-fifth of the annual

herbage production. Humphrey (1974) noted that sagebrush was more subject to burning than any other

desert type.

Dwarf sagebrush (14 habitat types) is usually relegated to shallow soils and is not considered a fire

management problem because fuel continuity is poor

and it generally cannot carry fire. Tall sagebrush (29

habitat types) occurs on deeper soils, often has a

substantial grass component, and burns readily

(Blaisdell and others 1982). The presence of a herbaceous understory increases the potential for big sagebrush to carry a fire. Threetip, basin, Wyoming, and

mountain big sagebrush occupy about 60 percent of

the total sagebrush area. This sagebrush association

is practical to burn (Blaisdell and others 1982). Techniques for managing sagebrush/grass ecosystems with

fire and other means are discussed by Blaisdell and

others (1982), Bushey and Kilgore (1984), McGee

(1976, 1977), and Onsager (1987) (fig. 6-24). Fuel and

fire behavior models were developed by Brown (1982),

Frandsen (1981), Reinhardt and others (1997), and

Tausch (1989) for burning in Great Basin sagebrush.

Fire behavior studies in big sagebrush show that fire

intensity and rate-of-spread can be two to three times

greater when sagebrush foliage is cured, yet the proportion dead has little effect on predicted fire behavior

(Brown 1982).

In blackbrush fuel production ranges from 0 to 500

lb/acre (0 to 560 kg/ha), and forage production ranges

from 0 to 150 lb/acre (0 to 168 kg/ha). Blackbrush is

negatively associated with fine fuels of litter and

grasses. In saltbush-greasewood fuels production

varies from year to year, depending on the amount of

Table 6-5—Fuel loadings (lb/acre) from FOFEM fuel models

(Reinhardt and others 1997) for FRES grassland

ecosystem types based on annual productivities.

Fuel class Desert Plains Mountain

Sparse 300 600 900

Typical 600 1,250 1,900

Abundant 900 1,900 2,800
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Figure 6-22—During wet years, a herbaceous layer develops in the bare spaces between the dense

thorn-shrub of the Sonoran desert, Maricopa County, Arizona, increasing the potential for major fires.

Figure 6-23—A wildfire burned 10,000 acres of this Sonoran desert thorn-shrub in Four Peaks, Tonto

National Forest, Arizona.
148 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

precipitation. Production is also related to soil salinity

and texture (West 1994). Herbage production is generally 0 to 500 lb/acre (0 to 560 kg/ha).

Creosotebush has low leaf to stem biomass, yet its

standing dry biomass may reach about 3.8 ton/acre

(8.5 t/ha) and produce about 892 lb/acre (1,000 kg/ha)

per annum of new fuels (Chew and Chew 1965). The

resinous foliage is flammable, but fire generally will

not carry well in this community because the plants

are usually surrounded by bare soil. Herbage production ranges from 40 to 100 lb/acre (44 to 112 kg/ha),

about one-third of which is considered forage. High

species diversity within the creosotebush/bursage

shrub type produces diverse fuels. In some areas dense

stands with herbaceous understory supply contiguous

fuels for fire.

Mesquite bosques (fig. 6-25), characterized by low

deciduous mesquite trees, are typically found in high

moisture areas, and may produce up to 2,000 lb/acre

(2,240 kg/ha) of herbage, particularly in areas that

flood periodically and where the mesquite has been

artificially reduced. Fuels are highly concentrated in

mesquite bosques. Herbage production is commonly

between 750 and 1,000 lb/acre (840 and 1,120 kg/ha)

with forage production from 0 to 500 lb/acre (0 to 560

kg/ha) (Garrison and others 1977). Higher fuel loading

on a site will increase the fire mortality of mesquite.

Areas with 2.25 ton/acre (5.06 t/ha) of fine fuel sustain

up to 25 percent mortality, but only 8 percent mortality for 1.1 ton/acre (2.47 t/ha) (Wright 1980). Dunes

may form in association with mesquite thickets.

In paloverde-cactus shrub fuels production ranges

from 100 to 250 lb/acre (112 to 280 kg/ha); about 35

percent of this vegetation has forage value. Fuels in

the Southwestern shrubsteppe are mixed grassshrublands. The Grama-tobosa region has a higher

grass component while the Trans-Pecos shrub savanna has a higher shrub component. The variable

fuels in the Trans-Pecos shrub savanna produce up to

450 lb/acre (505 kg/ha) of forage. Creosotebush and

yucca are present, but grama and tobosa primarily

contribute to the maximum 1,500 lb/acre (1,680 kg/ha)

herbage production in this type.

Chaparral—Generally, fuels are not as easily ignited as grass fuels, but once ignited will burn readily

if conditions are right. Plant density can vary with

site, and sometimes with species. This is but one factor

that affects fuel continuity in a stand. Another factor

is the basic within-plant geometry that varies by

species. Geometry and arrangement of the woody fuel

portion and the leaves of chaparral plants are key to

understanding the ability of chaparral stands to propagate fire. The woody fuel inside a given shrub varies in

size class ranging from fine fuel (<0.12 inch diameter)

Figure 6-24—Fire is used as a range management tool for sagebrush found on the Great Basin plains.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 149

Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen

Figure 6-25—Mesquite thickets form highly concentrated fuels in desert washes, Mojave Desert, California.

to heavy fuel (3 or 4 inches and larger in the case of

some manzanita species). The arrangement and distribution of these size classes within a shrub varies by

species. Two extremes illustrate this: the arrangement of the woody portions of chamise and manzanita

species (fig. 6-26). The smaller woody size classes are

quite dominant in chamise and tend to be in proximity

throughout the crown; the opposite is true for manzanita. The leaves of chamise are small and needlelike

and are often relatively dense on a given twig.

Manzanita is a broadleaf shrub. The leaves of some

species are relatively sparse—being held distant from

each other by the woody structure of the shrub. Other

species of manzanita have dense clusters of leaves—so

dense that their thick sclerophyllous structures act

like an insulator. Other chaparral species, some members of the Ceanothus genus for example, have only a

moderate amount of fine woody material and have

small broad leaves that are sparsely distributed

throughout the shrub crown. In general, the geometry

of chaparral shrubs is not well suited to the spread of

fire. Chamise is an exception, especially in dense

stands with overlapping crowns. The maintenance of

crown fires in chaparral almost always requires dry,

windy conditions, which commonly occur in this vegetation type. Figure 6-26—Arrangement, distribution, and size of woody

fuels can vary by species.

With few exceptions, fully developed stands of chaparral have no understory layer of vegetation, and

therefore no potential for the “ladder effect” to propagate fire. However, when a litter layer exists, which

occurs under gentle slope conditions, it can significantly aid fire spread under marginal burning conditions. In this situation, fuel moisture content becomes

an important factor.
150 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

The dynamics of dead fuel production in chaparral

remain a mystery. Some suggest that dead fuel production increases with stand age (Rothermel and

Philpot 1973). While this is undoubtedly true, age is

not the only factor involved (Paysen and Cohen 1990).

Complexities of onsite growing conditions and periodic events seem to be important. For example, the

authors suspect that an unusual drought can produce

fine dead material in chaparral stands that may

be only present onsite for a year or so—making the

assessment of dead fuel dynamics unclear. Considerable down and dead material can be found in old

chaparral stands. The concept of “old,” unfortunately,

has to remain a relative one for now. The age at which

significant amounts of dead material are produced in

a given stand of a given species composition cannot be

predicted yet.

Postfire Plant Communities

Plains Grasslands

Pre-1900 Succession—The literature on plains

grasslands communities is rife with contradictory

interpretations of grassland dynamics. A few facts

seem to be agreed upon. First of all, pollen records and

rat middens indicate that most of the Central Plains

was covered with boreal forest dominated by spruce,

while much of the Northern Plains was glaciated

during the Pleistocene. There are indications that the

Southern Plains and the arid grasslands of the Southwest were also dominated by various conifer and

broadleaf trees. The climate change that brought about

the end of the glacial period ushered in the retreat of

the boreal forest and its replacement by grasslands—

a kind of vegetation able to cope with the drier climate

and soil conditions that predominated.

Fire was not a predominant force in delimiting the

extent of the plains grasslands. But given their existence and their flammability characteristics, the presence of fire had to have an impact on the character of

the grasslands, their species composition, and the

distribution of dominance. Modifications of climate

and soil development led to invasion of some grassland

areas by woody species. Under these circumstances,

fire probably had a distinct role to play in the maintenance, or loss, of these grassland areas. Working in

concert with grazing animals, fire could check the

advance of more fire-sensitive, woody species, providing enough grass fuel was available. It could also

encourage the advance of woody species that were

adapted to disturbance and harsh climate conditions.

Where invasion by woody species was not an issue,

fires could maintain a highly productive mode in some

grasslands, and in others cause shifts in grassland

species composition; under conditions of drought, it

could result in severe site damage.

Clearly, fire was a common element in presettlement times, and there is some conjecture that its

frequency might have increased with the arrival of

Euro-American settlers (Jackson 1965). For years,

attempts to suppress fires in the plains were either

nonexistent or not effective. As late as the 1890s,

from the Dakotas to the Texas Panhandle, fires would

run unchecked for days. During this period, fire,

drought, and grazing played a role in maintaining, and

at times debilitating, the grassland character. When

fire, or any other phenomenon that reduced the vegetative cover, occurred during periods of serious

drought, wind erosion often retarded the processes of

succession.

Post-1900 Succession—The general set of natural

forces affecting succession just prior to 1900 has not

really changed in principle. Land use has alternately

intensified, and disappeared, and returned again in

some cases. Some of the plains grasslands have been

converted to agricultural use—producing corn, wheat,

barley, and various legumes; some have been put to

intensive grazing use—successfully in some instances,

and in others with disastrous results. In the Southern

Plains, the conjunction of inappropriate farming practices and a devastating drought in the 1930s brought

about a perceived ecological disaster and social phenomenon, called the “dust bowl,” that shook the fabric

of Southwestern culture. In retrospect, no surprises

should have existed.

The semiarid climate of the plains grassland area,

the ever-present potential for drought, yearly temperature extremes, and the potential for high winds

exist today, as they have existed for centuries. They

were operative in forming the plains grasslands and

continue to drive the processes of succession. The

factors relevant today are the firmly entrenched agricultural practices and the use of the grasslands as

pasturage for grazing animals. Land use patterns

such as these, once terminated, will drive the processes of succession in various directions—dominated

by the presence of the existing natural factors.

Deviations from successional patterns of past centuries

are difficult to predict other than on a case-by-case

basis.

Management Considerations—Management of

plains grasslands should be undertaken with a view

toward maintaining stability under local climate and

soil conditions. In the Northern Plains, a temperature

range of more than 130 °F between yearly maximum

and minimum temperatures can occur (a range of

174 °F has been recorded in one place). The average

growing season can range from 116 days in the northernmost portion to 160 days in the southern part

(Rogler and Hurt 1948). Native grasses tend to be

hardy and drought resistant—such species as blue

grama, buffalograss, western wheatgrass, and
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Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen

needlegrass. If the native grasses are to be used as

livestock forage, then overutilization should be guarded

against. Native range utilization by livestock should

be supplemented by locally produced forage and seed

crops whenever needed to protect native species.

The Southern Plains are also characterized by temperature extremes and a highly variable climate. Precipitation is comparatively light and infrequent; a

major proportion of it falls during the active growing

season, from April through September (Savage and

Costello 1948). Humidity is low, winds are high, and

evaporation is rapid. Hot temperatures and high winds

often reduce the effectiveness of precipitation that

does occur. Overutilization of rangelands during

drought always has to be guarded against.

Fire can be either a disaster or a useful element in

the plains grasslands, depending on its timing and

severity. A range fire that denudes a large area preceding a drought can set the stage for severe soil movement in many areas of the Great Plains—the high

winds and frequently aolean soils indicating the process. When good recovery is favored by adequate

precipitation, fire can improve productivity for a while.

The effectiveness of fire, both good and bad, can be

mitigated by current levels of productivity and by

intensity of utilization. Recently grazed grassland, or

a year of low productivity, can reduce the impact of fire

by minimizing fuel consumption, fireline intensity,

and general extent of burning.

The use of fire as a management tool can improve

productivity if it is applied in a manner consistent with

the grassland’s productivity, given climate and soil

character. Kucera (1981) contrasted the application of

prescribed fire between the more moist, highly productive grasslands and those of lower moisture availability and less productivity (fig. 6-27). Timing of the

application centers on the development of thatch. In

the higher productivity grasslands, the buildup of

thatch tends to suppress productivity after a few

years. In the lower productivity grasslands, the development of thatch provides a means of storing moisture

and thus increases productivity—at least over a period

of a few years. Thus, relatively frequent application of

prescribed fire in the high productivity grasslands can

be beneficial by removing thatch that has accumulated beyond desirable levels.

Mountain Grasslands

Pre-1900 Succession—Although bunchgrass species vary in their individual susceptibility to fire damage, repeated fires at intervals of about 5 to 40 years

(Gruell and others 1986) maintained the bunchgrass

community. The abundance of individual species no

doubt varied not only by site conditions but by the

actual frequency and seasonal timing of fire. A successional process of major importance was the continual

checking and reduction of woody plant encroachment.

Mountain grasslands were intertwined with forests

and shrublands ranging from rose and aspen in Alberta

to conifers and sagebrush of Rocky Mountain foothills.

Encroachment into grasslands by woody species was

an ongoing process kept in check by repeated fires.

Post-1900 Succession—Grazing by livestock, elimination of Native American ignitions, and fire control

efforts greatly reduced the amount of fire in these

grasslands. As a result tree species such as ponderosa

pine, Douglas-fir, and lodgepole pine, and sagebrush

have increased substantially along ecotonal boundaries. In some areas dense Douglas-fir forests now

dominate sites to such an extent that evidence of

former grasslands is lost except by soil analysis

(Bakeman and Nimlos 1985 ). Elimination of periodic

burning has apparently reduced diversity of herbaceous species in some areas (Wright and Bailey 1982).

In a study of fire regimes in the Interior Columbia

River Basin involving grasslands and other vegetation

types, Morgan and others (1994) suggested that human influences have had a variable effect on the

nature of fire regimes. Fires tended to be less frequent

but not always more severe. For example, where exotic

annuals have invaded sagebrush steppe vegetation,

fires have become so frequent that sagebrush does not

have time to reestablish, and the annuals return

quickly. Changes in fire regimes can move in one

direction as a result of active fire suppression that

results in a buildup of fuel, or in another direction as

a result of livestock grazing and other activities that

break up fuel continuity. No single successional formula can be offered for grasslands in general.

Management Considerations—Prescribed fire

can be effectively used to hold back woody plant

encroachment and maintain high levels of productivity in mountain grasslands. The complexity of mountain

Figure 6-27—The productivity of humid grasslands versus dry

grasslands after fire.
152 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

grasslands, however, requires careful consideration of

species composition and site dryness to design prescriptions for successful prescribed fire (Wright and

Bailey 1982). For this, knowledge of species response

can be helpful.

Idaho fescue is sensitive to fire partly because it is

susceptible to smoldering in the clump that can kill

plants or reduce basal area. It tends to recover slowly

from fire; however, on some sites it can withstand

burning (Bradley 1986b). Burning when soils are moist,

such as in the spring, helps to minimize damage.

Needlegrasses can also be severely damaged depending on severity of fire. Damage from wildfires can be

minimized by the grazing of livestock to reduce fuels.

Needle-and-thread grass reproduces by seed and can

increase markedly in 2 to 4 years after fire (Gruell and

others 1986). Bluebunch wheatgrass and Sandberg

bluegrass recover quickly from fire (Bradley 1986c;

Howard 1997). Rough fescue generally responds favorably to fire even after an initial reduction in basal

area. Preburn coverages can be attained in 2 to 3 years

(McMurray 1987).

Cheatgrass

Succession—Cheatgrass was accidentally introduced into the United States sometime around the

turn of the 20th Century, supposedly through contaminated grain (Pyke and Novak 1994). Cheatgrass

did not emerge as a noteworthy element in the Great

Basin environment until the period between 1907 to

1930 (Morrow and Stahiman 1984). By 1930, it had

achieved its current distribution (Pyke and Novak

1994). In the early 1900s, it had been noted in isolated

places—notably embankments, railroads, and highways. During the next 3 decades, it spread rapidly into

overgrazed sagebrush rangeland (Billings 1994).

Following disturbance by fire in areas where

cheatgrass is present, it reestablishes from abundant

seed. Even if fire destroys 90 percent or more of its

seed, it can reestablish and compete significantly with

native perennials (Bradley 1986a; Monsen 1992). Over

a period of years, cheatgrass gains dominance over

perennials and increases the flammability of the site

(Peters and Bunting 1994). Repeated fire will diminish the perennial seed bank and allow cheatgrass to

increase its dominance. Once cheatgrass becomes abundant enough to increase the likelihood of fire, repeated

fires may occur frequently enough to eliminate shrubs

such as sagebrush and native perennials. As wildfires

become more common cheatgrass can essentially dominate a site (Monsen 1994).

Management Considerations—Native species

can occupy sites that were dominated by cheatgrass,

but this is not a common occurrence. Use of mechanical tillage, herbicides, and properly timed fire can be

effective in reducing cheatgrass cover if other species

that germinate under cool conditions can be introduced. Prompt rehabilitation of burned areas by seeding accompanied by livestock restrictions is important.

Fire usually gives cheatgrass a competitive advantage. However, prescribed fire can be used to reduce

cheatgrass and to allow seeded species a chance to

establish. The narrow prescription window during

which substantial seed can be destroyed is from the

time cheatgrass becomes flammable, when it leaves

the purple stage, until seed falls a short time later.

During the 1990s a greenstripping program gained

favor. The objective was to reduce wildfire frequency

and size by establishing strips of fire-resistant vegetation, such as forage kochia, at strategic locations on

the landscape to slow or stop wildfires (Pellant 1994).

Greenstripping is aimed at effectively disrupting fuel

continuity, reducing fuel accumulations and volatility

on areas with a high density shrub cover such as

sagebrush, and increasing the density of plants that

retain higher moisture contents.

Annual Grasslands

Pre-1900 Succession—In California where this

type prevails, the Spanish settlers kept poor records,

so knowledge of native vegetation types is poor. Many

believe that the prehistoric vegetation was perennial

(Garrison and others 1977), but meager evidence is

available to support this belief. However, evidence

from the early 1800s indicates dominance by annual

grasses.

Post-1900 Succession—Intensive agricultural

development has taken over much of the original

annual grasslands. At the lower elevations of the

ecosystem, cultivated lands make up one of the richest

agricultural areas in the world (Garrison and others

1977). Remnants lie at upper elevations in the Sierra

foothills, and many are components of a hardwood

savanna or shrub savanna that are quite common in

these foothills. The annual grasslands are quite responsive to rainfall, and productivity and species dominance both vary accordingly. Fire is very much a part

of the ecosystem and does not seem to have detrimental effects. In fact, it is being used by ranchers to

eliminate woody overstory species and enhance productivity of the grasses.

Management Considerations—The most productive portions of the ecosystem are not producing annual grasslands, but rather agricultural crops. Clearly,

as long as this activity can be sustained, it will remain

the primary management activity in the “bottomland”

portions of this system. In the upland portions, grazing

and fire can be achieved to attain various management

goals. However, they are both system disturbances and

must be used judiciously. Annual rainfall is probably
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the most important consideration in applying management treatments in a manner consistent with ecosystem viability. Drought years should probably not be

accompanied by intensive disturbance activities.

Desert Shrublands

During the era of Euro-American settlement, fire

frequencies initially increased. Newspaper records

between 1859 and 1890 report that settlers engaged in

active fire suppression, including deliberate overgrazing of rangeland to reduce fuels. Woody species were

favored by the reduction of grass and forb competition

caused by overgrazing (Wright 1986). Grazing altered

the role of fire in those desert areas once dominated by

grasses. The consequent reduction of major fires was

followed by shrub invasion into desert grasslands

(Bahre 1985). Early 1900s wildland management policies continued to promote historical fire suppression

and rangeland use in desert landscapes. A new management strategy was initiated when desert managers recognized that continued shrub encroachment

was associated with overgrazing and fire reduction

(Komerek 1969; Leopold 1924). Shifts in land management resulted in reduced grazing, increased fuels and,

thus, changed the fire dynamics. Currently, burning of

thousands of acres is becoming more common, and fire

has become a serious management issue in some

shrubland areas (Blaisdell and others 1982; Bunting

and others 1987; Narog and others 1995; Schmid and

Rogers 1988; Wilson and others 1995a).

Desert shrubland management traditionally focused

on shrub eradication in favor of grasses. The objective

was to improve forage for livestock and increase efficient management of range by increasing livestock

and wildlife visibility. Fire, disking, herbicides, and

heavy grazing were all commonly used. Often, the end

result of this heavy range management was to decrease the amount of annual biomass and actually

reduce the productivity of these ranges.

The use of fire in desert shrublands is controversial.

Experts do not agree on historical fire cycles or what

the land-use goal must be. Presently, desert range

management practices rely on generalized studies

made on limited areas. Anthropogenic influence has

changed the vegetation and its dynamics in these dry

sensitive areas. High fuel loading, from multiple

branching shrubs, and contiguous herbaceous fuels

are now common in many of these deserts. Fire can be

used to achieve desired objectives in many of these

desert shrubland communities (Bunting and others

1987; Lotan and others 1981; McGee 1977; Wright

1990). Fire also may contribute to the loss of desirable

fire intolerant species that are sometimes replaced by

less desirable fire tolerant species. The present resource management challenge is to determine which

species to maintain and what management priorities

are suitable for each specific area.

Sagebrush

Pre-1900 Succession—Historical accounts of sagebrush habitat are sketchy, but fires in big sagebrush

were set both by lightning and humans. The many

species and subspecies of sagebrush are quite susceptible to fire. Typical succession after fire would begin

with a grass/forb dominance, and eventually lead to

sagebrush recovery in 30 or more years.

Until the mid-1800s, the American bison was the

primary herbivore impacting the fuels of sagebrush/

grasslands (Young and others 1979). In the late 1800s,

overstocked free ranging cattle led to a depletion of

perennial grasses and other palatable forage. The

subsequent introduction and spread of cheatgrass in

the early 1900s corresponds with increased fire frequency and the reduction of big sagebrush. This, in

turn, increased erosion and further damaged perennial native grass and forb components (MacMahon

1992).

Post-1900 Succession—Since 1900 the cultivation

and abandonment of marginal land, abusive grazing,

and widespread recurrent prescribed burning of sagebrush resulted in an imbalance between the numbers

and sizes of shrubs, and associated native grasses and

forbs (Blaisdell and others 1982). Thus, much of the

resource potential of the sagebrush range was depleted. By 1936, 85 percent of sagebrush lands were

considered depleted (Tisdale and others 1969). Prescribed fire was used to remove shrubs and replace

them with native perennial grass forage (Cornelius

and Talbot 1955; Pechanec and others 1954; Pechanec

and Stewart 1944; Reynolds and others 1968). This

ecosystem readily burns, particularly where there is a

contiguous understory of grasses. Habitat changes

coincident with increased fire have included plant

community composition changes (Blaisdell 1949;

Hassan and West 1986), altered soil seed banks (Blank

and others 1995), and increased soil repellency (Salih

and others 1973). The absence of sagebrush is often an

indicator of past burns (Humphrey 1974). Secondary

consequences of wildfires in sagebrush can include

range deterioration, flooding, erosion, lowered grazing capacity, and reductions in the amount and quality

of wildlife habitat. Extensive research has focused on

rangeland degradation (Young and others 1979) and

loss of productivity (Beetle 1960; Harniss and others

1981).

Management Considerations—Sagebrush land

managers are now confronted with recovering its productivity. Sagebrush production loss continues even

with recent improvements in management. Currently,
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Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

the value of the sagebrush rangelands is being reevaluated. Multiple factors need to be incorporated

into resource management plans. Big sagebrush can

gain dominance over the herbaceous layer in 5 to 30

years after a burn. Season of burn modifies species

dominance (White and Currie 1983) and affects postfire

sagebrush productivity (Mueggler and Blaisdell 1958).

For example, silver sagebrush mortality is higher and

regrowth is less after a dry fall burn (White and Currie

1983). After fires, sagebrush mortality is proportional

to fuel reduction. Although many sagebrush species

are readily killed by fire, at least three species (threetip

sagebrush, silver sagebrush, and California sagebrush)

are known to resprout (Malanson and O’Leary 1985;

Tisdale and Hironaka 1981). Most sagebrush species

reseed after fire, but may require fire intervals of up to

50 years to regain their dominance (Bunting and

others 1987). Frequent fires can cause type conversion

from sagebrush species to rabbitbrush, horsebrush,

and snakeweed. Where wheatgrass occurs, the burn

season is extended and wildfires are reported to consume more area per burn.

Introduced cheatgrass can outcompete indigenous

herbaceous species. This brome is undependable forage because of its large fluctuations in yield from year

to year. After two to three reburns, sagebrush sites can

be converted to stable cheatgrass; fire return intervals

of 5.5 years maintain cheatgrass dominance.

Cheatgrass is often accompanied by other invasive,

noxious, and undesirable species. Together these pose

a serious fire hazard, particularly following wet springs.

Planning prescribed fires in sagebrush should include specific objectives and consider many factors

such as species and subspecies of sagebrush, soils

(Salih and others 1973; Simanton and others 1990),

fuel loading, fuel moisture content, and windspeed

(Britton and Ralphs 1979; Brown 1982). Early spring

or late summer burns can be used to promote native

perennial grasses. There is little postfire recruitment

for 3 to 5 years following a fire in perennial grasses, yet

surviving grasses and accompanying forbs increase

biomass production. Often forbs will dominate an area

for several years postburn. Harniss and Murray (1973)

found increases in herbage production for 20 years

after a burn.

Attempts at restoring sagebrush rangeland to achieve

higher biomass yields are being investigated (Downs

and others 1995). In general, shrublands that have

been converted to grasses by large wildfires are difficult to restore. Fire negatively impacts soil seedbeds

important for sagebrush regeneration (Blank and others 1995). Sagebrush seed can be viable up to 4 years.

Sagebrush can be restored through reseeding.

Cheatgrass seed banks present on sagebrush sites

may negatively influence reestablishment of native

bunch grasses and shrubs (Hassan and West 1986). If

sagebrush is in good “natural” condition an initial

postfire influx of cheatgrasss will occur. Given adequate precipitation, perennial native grasses and

shrubs can outcompete cheatgrass by the second year

(West and Hassan 1985). Postfire rehabilitation efforts can be unsuccessful if other measures such as

grazing are not incorporated (Evans and Young 1978).

Species and associations of the sagebrush-grass type

are influenced by edaphics and microclimate (Meyer

1994). Restoration efforts are complicated by the level

of site disturbance and ecosystem variability and

specificity (Blaisdell and others 1982; Blank and others 1995). Wildfire in cheatgrass dominated sites may

afford managers an opportunity to reseed with perennial grasses and reduce the cheatgrass to lengthen the

fire return interval. Presence of woody fuels may

provide a hotter fire that can kill more cheatgrass

seeds. Herbicide applications may facilitate native

shrub and grass reestablishment (Downs and others

1995).

Wildlife such as pronghorn, deer, elk, coyotes, rabbits, rodents, and an endangered prairie dog reside in

sagebrush rangelands. Abundant avifauna (over 50

species) that nest and feed in sagebrush include eagles,

hawks, owls, doves, chukar, and sage grouse. Wild

ungulates and domestic sheep may benefit from the

maintenance of high quality sagebrush browse

(Rodriguez and Welch 1989). Wildfires have removed

large areas of sagebrush and may have destroyed a

significant amount of sage grouse habitat (Downs and

others 1985). Short- and long-term effects of fire on

wildlife in this habitat need further evaluation (Gates

and Eng 1984).

Blackbrush

Succession—Historical documentation of blackbrush fire cycles is limited. As late as 1981 (Lotan and

others 1981; Martin 1975), land managers did not

perceive desert fires as a serious land management

problem because of small fire size and minimal damage to resources. Current data refute this perspective

(Narog and others 1995; Wilson and others 1995a,

1995b). Cyclic desert precipitation above 10 to 14

inches (25 to 36 cm) may increase biomass and fuel

continuity enough to increase fire behavior potential.

Since 1900, it appears that neither fire nor exotic

annuals have altered soil microflora apparently required for blackbrush survival or reestablishment.

However, burning has promoted succession to grassland by destroying the cryptogamic crust that stabilized the soil. Frequent large fires have eliminated

blackbrush from some areas. Some sites show no

recovery after almost 4 decades (Wright and Bailey

1982). Currently, burning is not a recommended practice for range enhancement purposes in this shrub
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 155

Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen

type (Callison and others 1985) because blackbrush is

often replaced by species of similar forage potential.

Management Considerations—Fire has been

used for range improvement by reducing the shrub to

grass ratio in areas where shrubs are gaining dominance. Land managers must also focus on protecting

cacti and succulents, which will complicate fire management because of their various responses to fire

(Thomas 1991). Fire may continue to be a necessary

tool to modify fuel buildup. Currently, increases in

desert shrubland fires and fire size have become a

serious concern particularly with the recent increase

in urban encroachment and resource degradation issues on these lands.

Research is needed to develop management and

restoration recommendations for blackbrush

(Pendleton and others 1995). Fire destroys the shortlived blackbrush seedbanks (produced by masting)

necessary for it to reestablish. High temperatures,

wind, and low humidity are usually required to propagate fire in blackbrush. If blackbrush becomes decadent or in some way presents a wildfire hazard, removal by burning may be appropriate. In some cases

mature shrubs may survive low intensity fires; however, fire generally kills both seeds and mature shrubs.

Although blackbrush is somewhat effective for erosion

control, it may take more than 60 years to reestablish

after a disturbance such as fire (Bowns and West

1976).

Wildlife such as deer, elk, desert bighorn, pronghorn, squirrels, rabbits, and game and nongame birds

use blackbrush for cover, browse, and seeds. Livestock

are more limited: sheep and goats browse blackbrush,

but its low palatability and nutritional value make it

unsuitable for cattle and horses.

Saltbush-Greasewood

Succession—Little is written regarding historic fire

patterns in the saltbush-greasewood type. In some

areas little change has occurred since 1900 in black

greasewood dominated vegetation, while in others both

saltbush and black greasewood have expanded into

areas previously dominated by sagebrush (Sparks and

others 1990). Rangeland seeding and invasion of grasses

forming a highly flammable understory have increased

the fire frequency in the saltbush-greasewood type.

Postfire recovery is often rapid due to postfire resprouting

and vigorous reseeding strategies used by the various

shrub species in this vegetation type.

Management Considerations—In the past fire

management was not a concern in saltbush-greasewood vegetation because sparse understory, bare soil

frequently found in intershrub spaces, and the low

volatilization of many saltbush species made this

vegetation type resistant to fire (Tirmenstein 1986).

These communities may burn only during high fire

hazard conditions. In wet years brought by El Nino,

such as 1983 to 1985, fine fuels may become contiguous across otherwise gravelly soils. Recently these fine

fuels have become a fire hazard problem (West 1994).

Grazing and other disturbance can encourage increases

in biomass production, especially in the spring

(Sanderson and Stutz 1994). Introduced cheatgrass

has increased the fire risk, particularly when the area

is ungrazed (West 1994). Disturbance may also allow

this vegetation type to increase its range. Many species in this type resprout (West 1994). Black greasewood vigorously resprouts after fire or other disturbance. Season of burn, fire intensity, and fuel loading

may be important factors to consider when using fire

to regenerate or increase the productivity of this vegetation type (Harper and others 1990). Intense fall

fires may increase plant mortality in spite of a species’

resprouting potential. Some Atriplex species resprout

and others produce abundant seeds. Thus postfire

reestablishment from onsite and offsite seed sources is

possible.

Saltbush-greasewood vegetation provides valuable

forage for livestock and wildlife, particularly during

spring and summer before the hardening of spiny

twigs. It supplies browse, seeds, and cover for birds,

small mammals, rabbits, deer, and pronghorn. Saltbush and black greasewood can be used to revegetate

mine spoils and stabilize soils. Saltbush concentrates

salts in leaf tissue and may be used to reduce soil salts

and reclaim degraded land for agriculture. Outplanting

methods are being developed for saltbush restoration

projects (Watson and others 1995).

Creosotebush

Succession—Historically, creosotebush was restricted to well-drained knolls and foothills. However,

by 1858 it had begun to invade the grama grasslands

and by the early 1900s creosotebush had encroached

into areas dominated by grasslands (Valentine and

Gerard 1968). Overgrazing and drought contributed

to the expansion of creosotebush range (Buffington

and Herbel 1965). Fire suppression may be contributing to this expansion.

Management Considerations—Creosotebush

invades desert grasslands. Although creosotebush

may suffer up to 80 percent dieback during drought,

it still resprouts (Humphrey 1974). On the other

hand, it is sensitive to fire, especially in spring (Brown

and Minnich 1986; McLaughlin and Bowers 1982).

Fire and herbicides have been used to control

creosotebush. High fuel loading and spring and summer

burning will lead to higher creosotebush mortality

from fire (Martin 1966). This indicates that wildfires

could have kept it from invading grasslands before
156 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

Euro-American settlement (Wright and Bailey 1982).

Selective thinning of creosotebush by fire suggests

that this ecosystem is not resilient to burning and

creosotebush may be replaced by other species, particularly with recurrent fires (Cable 1973). For example, bush muhly growing under creosotebush canopies may out-compete smaller shrubs and become the

dominant after fire. Following heavy precipitation,

herbaceous fuel increases and may increase fire potential in the creosotebush vegetation type (Brown and

Minnich 1986). Creosotebush can withstand some fire

exposure (O’Leary and Minnich 1981). Brown and

Minnich (1986) report slow recovery for creosotebush

after low-severity fire, and limited sprouting and germination were observed after fire in most of the species in the creosotebush associations.

Sheep will use creosotebush for cover, but

creosotebush is unpalatable browse for livestock and

most wildlife. However, pronghorn, bighorn sheep,

mountain goats, game and nongame birds, fox, small

mammals, and many reptiles and amphibians are

some of the wildlife that use creosotebush for cover

and its seed for food. Interestingly, the protected

desert tortoise (Gopherus agassizi) typically burrows

in soil stabilized by this plant (Baxter 1988).

Creosotebush can be outplanted to facilitate rehabilitation of disturbed desert areas where it improves

microsites for other plants and for fauna.

Creosotebush-Bursage

Fire use prior to 1900 may have limited the range of

creosotebush-bursage and kept it from invading desert

grasslands (Humphrey 1974). Since the early 1900s

white bursage has become dominant to creosotebush

on disturbed sites. McAuliffe (1988) reports that

creosotebush may use white bursage as a nurse plant.

Bursage species are easily topkilled but can resprout.

Following a fire, cover of creosotebush and bursage is

reduced but then increases over time (Marshall 1994).

Because fuel loading can vary seasonally and annually, fire management considerations in the creosotebursage type requires a site-specific analysis of plant

cover, fuel loading, and fuel continuity.

Mesquite

Succession—Mesquite density and distribution

increased prior to 1900 with fire suppression and seed

dispersal by livestock. After 1900 mesquite continued

to increase even though numerous eradication practices such as biological control, herbicides, mechanical

removal, and prescribed burning were used to limit its

density and spread—with mixed results (Glendening

1952; Jacoby and Ansley 1991; Wright 1990; Wright

and Bailey 1982).

Management Considerations—Fire as a management tool for controlling mesquite has its limitations. Mesquite may become more prevalent 5 years

following a burn than it was before fire (Martin 1983).

Mesquite can root sprout; top-killed individuals may

resprout from dormant buds found in upper branches

or from the base of the trunk below the ground surface.

Mesquite seedlings can survive fire (Cable 1961), but

on a burned site mesquite is sometimes reduced (Wright

1980). Fire may kill a good proportion of mature

mesquite, particularly the smaller trees (<2 inch diameter) (Cable 1949, 1973). It is most susceptible to

fire during the hottest and driest part of the year

(Cable 1973). Drought years may increase mortality of

mesquite if eradication is attempted. If managers

wish to open dense mesquite stands, then roots must

be killed, not just aboveground biomass. Fire can be

used to reduce the density of young mesquite populations, particularly during dry seasons that follow 1 to

2 years of above normal summer precipitation (Wright

1980). Adequate precipitation, no grazing, and using

fire about every 10 years allow grasses to successfully

compete with mesquite (Wright 1980). Rehabilitation

of mesquite-invaded grasslands requires removal of

livestock before burning, otherwise the shrubs

outcompete the grasses (Cox and others 1990). Shrub

reinvasion depends on grazing management combined

with continued use of fire at the desired frequency

(Wright 1986).

In managing for mesquite savanna (Ansley and

others 1995, 1996b, in Press; James and others 1991),

shaded rangeland may be a preferred condition rather

than attempting to completely eradicate mesquite (see

the Texas Savanna section). Low-intensity fire may

allow mesquite to retain apical dominance on upper

branches while reducing overall foliage. Season, air

temperature, relative humidity, and duration and

temperature of fire were factors reported to affect

mesquite response to fire (Ansley and others in press).

Mesquite topkill is related to heat in the canopy, not at

the stem bases. Single and repeated summer burns

kill mesquite aboveground, but do not kill roots (Ansley

and others 1995). Prescribed burning may be used to

kill mesquite seedlings while leaving tree sized and

shaped older individuals (James and others 1991).

Paloverde-Cactus Shrub

Succession—Prior to 1900, fires in paloverde-cactus shrub were not considered to be important and

occurred mainly in the restricted desert grasslands

(Humphrey 1963). Conversion of desert shrubland to

grassland to enhance forage for livestock and wildlife

had been the primary land-use goal during the 1800s

(Martin and Turner 1977; Phillips 1962). The high
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 157

Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen

shrub component in this desert is attributed to historic

overgrazing and overburning.

Since 1900, increases in ignitions and fire size are

evidence of changing land management practices in

the paloverde-cactus shrub. Exotic grass invasion now

supplies a contiguous fuel source in many areas so that

the historical small and infrequent fires were replaced

by more frequent and larger fires (Narog and others

1995). Rogers (1986) speculated that finer fuels and

higher rates of spread may allow desert fires to become

larger than nondesert fires before being controlled.

Although many of the species in this vegetation type

can resprout (Wilson and others 1995b), postfire communities generally experience changes in species composition, particularly with an increase in the grass

component, at the expense of cacti and succulents

(Cave and Patten 1984; McLaughlin and Bowers 1982;

Rogers and Steele 1980).

Management Considerations—Current management policy for some of the paloverde-cactus shrub

vegetation now includes multiple interests with an

increasing emphasis on recreation and tourism. This

new policy involving reduced grazing, an increasing

number of ignitions, and a greater herbaceous component is altering the fire regime (Robinett 1995). Fire

dynamics information is required to effectively manage these changing needs. The increase in fire frequency and size may have serious consequences particularly for plant and wildlife species of special interest

such as the giant saguaro (Thomas 1991; Wilson and

others in press) and the desert tortoise; both may be

fire intolerant. Little information exists on maintaining desert species in the presence of fire. Restoration

in the paloverde-cactus shrub type needs to be addressed if the thousands of acres recently burned are

to be rehabilitated.

Southwestern Shrubsteppe

Succession—Historically, fire suppression and seed

dispersal by herbivores have allowed grama-tobosa

range to become dominated by creosotebush, tarbush,

and mesquite. Tobosa is an early postfire seral component. Since the 1900s fire has been used to regenerate

decadent stands of tobosa. Fire may stimulate or

damage grama depending on climatic conditions, season, and fire severity. Reestablishment after fire is

generally through stolons. Grama species can regenerate by seed, stolons, rhizomes, or tillering; tobosa

mainly regenerates by rhizomes.

Management Considerations—Tobosa can be

managed with prescribed fire, which causes low mortality, improves palatability, and increases biomass

production. Tobosa is one of the few native grasses

that have competed well with nonnative grasses. Spring

burns produce the best results when precipitation is

adequate. Litter of up to 3.0 tons/acre (6.7 t/ha) easily

carries fire and is completely consumed. Broomweeds,

snakeweeds, and firewhirls are prescribed burning

hazards in tobosa. For optimum forage production

prescribed burns should be conducted every 5 to 8

years on tobosa stands. Nonbunchgrass species of

grama may take 2 to 3 years to recover following fire.

Grama and tobosa supply abundant forage for livestock and wildlife. Grama is palatable all year, but

tobosa is poor forage in winter months. Black grama is

drought adapted and can be used for restoration to

prevent soil erosion.

Trans-Pecos Shrubsteppe

Succession—Historically, junipers were relegated

to rock outcrops and upland limestone sites, preferring shallow limestone soils. Fire suppression and

overgrazing have allowed the woody species to expand

from their historically more limited range onto the

mixed prairie, sometimes in dense stands (Sparks and

others 1990). Dense juniper stands are highly competitive and reduce understory grassy forage. Junipers

dominate over oaks on drier sites, are shade intolerant, and may be succeeded by pinyon pine. Junipers

are facultative seral trees with extensive lateral roots

that effectively compete for surface moisture in xeric

environments. They may or may not root sprout depending upon species. Chemical control, mechanical

control, and prescribed burning have been used to

reduce juniper density to improve rangeland forage

productivity.

Management Considerations—Management techniques to reduce juniper and shrub density to improve

rangeland for livestock are employed in many areas.

Prescribed burning in junipers is recommended to open

dense stands; however, ground fuels are not always

adequate to carry fire. Ahlstrand (1982) found that

plant response to fire in this community is predominantly by vegetative means. He suggests that prescribed burning can be used to improve the grass

component at the expense of the shrubs. Pretreatment

with chemicals or mechanical methods is also recommended. A minimum of 1,000 lb/acre (1,120 kg/ha) of

continuous fine fuels is needed for prescribed burns

(Rasmussen and others 1986). Fire history studies

suggest that fire-free intervals of less than 50 years

restrict the expansion of junipers, and that nested fire

cycles have actually driven the juniper’s range (Bunting

1994). Fire rotations of 10 to 40 years are recommended

to control junipers. Reburn intervals between 20 to 40

years or when junipers reach 4 feet tall are recommended to maintain converted grasslands (Wright and

others 1979). Variable fire effects in this type can be

obtained (Tausch and others 1995). For specific fire

prescriptions refer to Wright and others (1979).
158 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Paysen Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems

Mechanical treatment followed 5 years later by

burning to kill saplings is recommended to maintain a

landscape mosaic of open stands and grassland. Mature junipers in moderate to dense stands are resistant to fire, yet may suffer some mortality. Small

stemmed individuals are easily killed by fire. Rapidly

burning grass fires occurring at intervals of 10 years or

more are adequate to allow juniper saplings to reach

sufficient heights 3 to 6 feet (1 to 2 m) to withstand fire

injury. Burned areas may be invaded through seed

dispersal, and establishment can occur within 10 to 40

years (Rasmussen and others 1986). Dead junipers are

volatile fuels, and spot fires from firebrands can be a

problem.

Fauna in the Trans-Pecos shrubsteppe ecosystem

are similar to the species found in desert grasslands.

Pronghorn and deer are widely distributed across the

shrubsteppe range as are dove, quail, rabbits, and

small rodents. Javelina are common in the south.

Common carnivores include coyote, bobcat, eagle, owl,

and hawk. Juniper berries and acorns are a favored

food by many species.

Chaparral—Mountain Shrub

Pre-1900 Succession—The species that we refer to

collectively as chaparral evolved as a component of the

understory of Laurentian forest types. They were

adapted to harsh conditions and could withstand disturbance. Chaparral development had no particular

relation to fire (Axelrod 1989). With warming and

drying trends, chaparral species became more opportunistic and were able to fill niches once occupied by

species less able to compete under these conditions. A

disturbance that chaparral was able to cope with was

fire—an element whose presence was probably important in providing opportunity for chaparral to attain

status as a recognizable vegetation type. By the end of

the 19th century, newspaper accounts of fires burning

through this type for days and weeks in southern

California became common. By accounts of historic

fires, maps of vegetation from the first third of the 20th

century, local lore, and by remnants of previous vegetation, a picture of chaparral’s ecological amplitude

begins to emerge with fire as an important environmental component. This logic continues into the 21st

century.

Post-1900 Succession—The benefit of more or less

real-time observation allows us the opportunity to fine

tune our view of chaparral’s role in succession. The

dynamics of chaparral’s environment have made it

difficult to definitively document chaparral’s role in

the successional process. Several salient points can be

made, however, with little fear of argument:

• Chaparral succeeds many forest types after a major

disturbance—whether from fire or logging. It is

often seral—especially at elevations where we currently consider chaparral as a montane understory

type. Given a reasonable number of disturbancefree years, the forest type will regain dominance.

• Chaparral often succeeds chaparral after fire, especially at elevations where we consider chaparral

as the dominant vegetation type. Species composition can shift drastically, probably depending on

whether the fire occurred before or after seed set

for a given species. The concept of an infinite store

of chaparral seed in the soil is becoming more and

more questionable due to seed predation by rodents, ants, and birds (Quinn 1980).

• The concept of chaparral being a fire climax refers

to a delicate balance between characteristics of the

chaparral species on a site and the fire regime.

Fire frequency and timing can tip the balance so

that chaparral can be overtaken by herbaceous

vegetation types, such as annual grasses, and in

southern California by an allied “soft chaparral”

type—a highly volatile semiwoody group of shrubs.

But the present fire regime appears to be about the

same as during the presettlement period. Conard

and Weise (1998) presented evidence that fire

suppression has offset increased human ignitions

during the past century, thus preventing fire frequency from increasing to the point of degrading

the ecosystem. Area burned per year, size of large

fires, and seasonality of fires in chaparral changed

little during the past century.

Management Considerations—Management of

chaparral has been directed primarily at concerns

about fuel hazard, wildlife habitat, and as a cover type

that plays an active role in maintaining slope stability

and watershed capability. Some would prefer to manage chaparral as a problem because it occupies potential rangeland that could be used for livestock grazing.

All can, in their place, be perfectly good reasons for

managing chaparral, but you have to pick one—or

maybe two.

Prescribed fire can be used to remove dead fuel for

hazard reduction, increase structural diversity for

wildlife habitat purposes, and increase the proportion

of young biomass in a stand—for both hazard reduction and wildlife habitat improvement. In some areas,

but not all, prescribed fire can be used to maintain

stands of chaparral in their current state (that is, to

maintain a fire climax). For prescribed fire to be

successful, species that reproduce only from seed, the

presence of seed must be assured. Some chaparral

seeds need scarification, which fire often provides.

Besides heat-shock scarification, smoke-induced
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 159

Chapter 6: Fire in Western Shrubland, Woodland, and Grassland Ecosystems Paysen

germination is important to many chaparral species

(Keeley and Fotheringham 1998). Seeds of chamise

can be destroyed if directly exposed to fire. Many

chaparral species sprout after fire; reproduction from

seed is not as much of an issue for these species as long

as individual plants are not killed. However, next to

nothing is known about the effects of physiological age

on sprouting ability of chaparral species.

Individual shrubs can be killed outright by fire.

Shrubs lacking in vigor will probably not respond to

fire in their normal fashion. Thus, stresses such as

protracted drought might cause an unexpected effect

if fire were to be introduced. An extremely severe fire

can result in little reproduction from either sprouting

or seed germination. A series of fires with short return

intervals may result in reduced chaparral shrub density if shrubs burn before they reach seed-bearing age,

or young shrubs developing from sprouts are physiologically unable to respond.

Extremely old chaparral stands can be found with

little or no dead material in them, and others can be

found with a significant down and dead component

(Conard and Weise 1998). The difference can be dictated by species composition, site conditions, and history of the site. Management of stands with a lot of

dead material in them has to be taken on a case by case

basis. From a fuels standpoint, these stands do have

an elevated hazard level. Whether or not they present

a serious threat should be evaluated in light of their

juxtaposition to other resources and the condition of

the other resources. Old chaparral stands should not

automatically be considered as “decadent.”

Conversion of chaparral to rangeland has to be

undertaken with caution. Soil and slope conditions

should be evaluated to avoid loss of soil. For this

reason, steep slopes and easily eroded soils should be

avoided in conversion projects. In all cases, chaparral

management should be undertaken with a clear view

of species present, site conditions, stand history, fuel

situation, and successional potential.
160 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Notes

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USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 161

Ronald L. Myers

Chapter 7:

Fire in Tropical and

Subtropical Ecosystems

Apart from savannas and grasslands, wildland fire

in tropical environments has received scientific scrutiny only within the past few decades. It is now widely

recognized that the vast majority of wildland fires

occur in the tropics and subtropics (Goldammer 1993).

Because a global treatment of the effects of fire in

tropical and subtropical ecosystems is beyond the

scope of this volume, the reader is referred to J. G.

Goldammer (1990), wherein Mueller-Dombois and

Goldammer (1990) outlined generalized tropical and

subtropical fire regimes.

In this chapter, we focus only on fire effects in

subtropical Florida, Puerto Rico, the United States

Virgin Islands, and Hawaii by drawing on appropriate

literature from southern Florida, the Caribbean,

Mexico, and the Pacific Islands.

Understory Fire Regimes _________

Major Vegetation Types

In the regions covered here, major vegetation types

having woody dominants exposed to surface fires

that are generally nonlethal to the overstory occur

only in subtropical Florida. They include (1) pinelands

and savannas dominated by slash pine (var. densa),

(2) wetland savannas and woodlands dominated by

pondcypress, (3) cabbage palmetto forests and savannas, and (4) southern live oak hammocks (a local term

for groves or forests dominated by hardwoods). Close

analogs of these vegetation types are the Caribbean

pine forests, woodlands, and savannas of the Bahamas, Cuba, and coastal Central America from Belize to

Nicaragua; some of the other pine types in Cuba; the

mountain pinelands of Hispaniola; and some of the

seasonally inundated palm swamps and savannas in

the tropical and subtropical Americas. There does not

appear to be an obvious analog of the pondcypress

type: a subtropical, fire-tolerant, wetland conifer forest or savanna, which occurs in seasonally flooded

depressions throughout the Southeastern Coastal Plain

of the United States, but reaches its greatest coverage

in subtropical Florida. For a synopsis of the role and

effects of fire in southern Florida ecosystems see Wade

and others (1980) and Myers and Ewel (1990).

All of southern Florida’s vegetation types characterized by understory fire regimes are found on low, flat,

poorly drained substrates. The landscape consists of a

vegetation mosaic where hydrology exercises considerable influence over the availability of fuels and
162 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Myers Chapter 7: Fire in Tropical and Subtropical Ecosystems Table 7-1—Occurrence and frequency of presettlement fire regime types by Forest and Range Environmental Study (FRES) ecosystems, Kuchler potential natural vegetation classes (1975 map codes), and Society of American Foresters (SAF) cover types. Occurrence is an approximation of the proportion of a vegetation class represented by a fire regime type. Frequency is shown as fire interval classes defined by Hardy and others (1998) followed by a range in fire intervals where data are sufficient. The range is based on study data with extreme values disregarded. The vegetation classifications are aligned to show equivalents; however, some corresponding Kuchler and SAF types may not be shown. Fire regime types Understory Mixed Stand-replacement FRES Kuchler SAF Occura Freqb Occur Freq Occur Freq Nonfire Longleaf-slash pine 12 Subtropical pine forest S. Florida slash pine 111 M 1a: 1-5 m 1b K116 Cabbage palmetto-slash pine 86 M 1a m 1b Caribbean pinec M 1a m 1b Palmetto prairie K079 M 1a Oak-gum-cypress 16 S. floodplain forest K113 Bald cypress 101 m 2a M 2b,3 Cypress savanna K091 Pondcypress 100 M 1 Mangrove K105 M 2 Live oak 89 M 1b m 2a Wet grasslands 41 Everglades K092 Marshes c M 1a Tropical hardwoods d M 2,3 Melaleuca d M 1 M 1,2 aM: major, occupies >25% of vegetation class; m: minor, occupies <25% of vegetation class. bClasses in years are 1: <35, 1a: <10, 1b: 10 to <35, 2: 35 to 200, 2a: 35 to <100, 2b: 100 to 200, 3: >200. cThis vegetation type fits as a subdivision of FRES or Kuchler types, but is not an SAF type. dSpecial type for chapter 7, not a FRES, Kuchler, or SAF type.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 163

Chapter 7: Fire in Tropical and Subtropical Ecosystems Myers

interacts with fire in determining the dominance of

species. Along a hydroperiod (period the water table is

above the soil surface) gradient, pondcypress occurs on

the wettest sites. Relatively higher sites support pine,

palm, or oak. The juxtaposition of the pine, palm, and

live oak-dominated vegetation is controlled to a large

degree by landscape features such as ponds, lakes, and

drainages that create fire shadows where the intensity and incidence of fire is reduced. Twentieth century

anthropogenic alterations of both fire and hydrologic

regimes, coupled with fire exclusion in some areas and

the introduction of non-native species, has markedly

altered some of southern Florida’s historic vegetation

patterns and fire regimes.

Fire Regime Characteristics

Fire is essential for existence of the southern Florida

woody vegetation types having understory fire regimes. These types are subject to nonlethal, relatively

frequent, low-intensity surface fires ignited by either

lightning or humans, although lethal long-returninterval ground fires may be important in the development of certain types of pondcypress communities.

The shortest fire return intervals of 1 to 5 years are

found in the pinelands and savannas. The longest fire

intervals measured in centuries are found in bald

cypress forests. If fire is removed as a process, or if the

normal range of variability of the fire regime is altered, the character of the ecosystem changes.

Lightning is frequently cited as the primary ignition

source responsible for southern Florida’s fire-maintained vegetation. Florida has the reputation of being

the “lightning capitol” of the United States, if not the

world, with frequent, intense electrical storms generated during the summer by a double sea breeze system

operating from both the east and west coasts of the

Florida peninsula. Lightning fires occur during every

month of the year, but with peaks in ignitions and area

burned concentrated during the summer months

(Snyder and others 1990). This period coincides with

the season of locally generated convectional storms

(fig. 7-1). The greatest area burned from lightningignited fires usually occurs at the onset of this season

when fuels are dry and wetlands exhibit their lowest

water levels.

We have little historical documentation of the burning activities of pre-Columbian Native Americans, so

it is difficult to assess their influences on southern

Florida fire regimes. The Calusa Indians disappeared

shortly after contact due to European diseases and

Spanish slave raiding parties from Cuba and the

Greater Antilles. Because there were no Spanish settlements in, nor expeditions through, southern Florida,

there are no known written records of Calusa burning

practices. However, the conventional wisdom that

Figure 7-1—Monthly distribution of total number of fires and

area burned from 1948 through 1997 in Everglades National

Park, Florida, by ignition source. Courtesy of John Segar

(Everglades National Park) and James Snyder (Big Cypress

National Preserve).

lightning ignitions were the sole determinant of southern Florida’s fire regimes has been challenged (Myers

and Peroni 1983; Robbins and Myers 1992; J. R. Snyder

1991). Extensive, ecologically significant Native American burning in southern Florida can be inferred from

archeological surveys, especially the number and distribution of occupation sites; from documented burning practices by the post-Columbian Seminoles, who

were forced into southern Florida in the 1700s; by

extrapolating from general principles of aboriginal

burning elsewhere; and most significantly, from the

aseasonal flammability of southern Florida’s fuels.

The latter means that the landscape could burn at any

time of the year, and extensive areas possessed available fuels outside of the summer lightning season.

The most compelling evidence suggesting an important aboriginal influence on fire regimes in subtropical Florida is provided in a vegetation and fire history

of Andros Island in the Bahamas (Kjellmark 1995,

1996). Alternating periods of human occupation and

depopulation over the past several thousand years are

well documented on Andros Island. Pollen analyses

indicated that during periods of human occupation,

pinelands of Caribbean pine expanded at the expense
164 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Myers Chapter 7: Fire in Tropical and Subtropical Ecosystems

of tropical hardwoods. During known periods of depopulation, the pinelands contracted considerably,

persisting only on the western side of the island where

lightning-caused fire may have been more prevalent.

The Bahamian pinelands are nearly identical in structure, function, and species composition to some of

southern Florida’s pinelands.

Florida’s present-day human-caused fires are concentrated in January through May with the greatest

area burned in March and April. This predictable dry

period, occurring before the onset of the lightning

season, coupled with the year-round availability of

fuels in some vegetation types, suggest that Native

American burning likely shifted the seasonal component of fire regimes, increased the incidence of ignitions, and ignited fires at times when wetland vegetation types were likely to burn extensively. This suggests

that reliance solely on lightning ignitions in the Everglades might lead to a landscape that may never have

previously existed.

Fuels

The fuels controlling Florida’s understory fire regimes consist of ground cover of perennial grasses, low

woody shrubs, saw palmetto, and surface litter. Due to

the presence of volatile oils and waxes, live surface

fuels can readily carry fire, contributing to Florida’s

year-round incidence of fire. Postburn accumulation of

fuel is rapid as most grasses, shrubs, and palmetto

resprout within a week of the burn regardless of the

season. In denser pine stands, needle drop from crown

scorched trees can form a continuous litter fuel bed

within weeks of a burn. This rapid accumulation of

fuel allows for low-intensity reburns on some sites

within a year. Sites without pine, or with open stands,

may require up to 3 years of fuel accumulation before

an effective burn can recur. In circumstances where

fuels accumulate for a decade or more, the probability

of lethal effects increase, either due to crown scorch

and consumption, or through smoldering ground fires

in the accumulated duff at the base of trees.

Although many of southern Florida’s fuels are available throughout the year, a predictable dry period

begins in the winter months and continues until the

onset of the summer convectional storms. At times,

this dry period is protracted leading to severe drought

in March, April, and May. Because water levels have

receded and fuel moistures are low, more fuels are

exposed, increasing loadings on local areas and on a

total area basis. These drought conditions may be

exacerbated further if the winter includes widespread

frosts or freezes, which top-kill many plant species of

tropical origin.

Pinelands—There are two types of subtropical

pinelands—flatwoods on sandy soils and pine rockland

on limestone. Ground cover in the pine flatwoods

consists of a diverse array of shrubs, grasses, forbs,

and saw palmetto. The palm fronds, grasses, and

pinestraw compose the fuel that carries the fires.

Volatile compounds in the leaves of some of the grasses,

particularly wiregrasses, the ericaceous coastalplain

staggerbush and the aquifoliaceous gallberry and

dahoon holly, wax myrtle, and the fronds of saw

palmetto, allow for vigorous burning even at high dead

and live fuel moistures. For example, the moisture of

extinction of the NFFL Southern Rough Fuel Model

(Albini 1976), which is representative of these pineland

fuels, is a high 40 percent.

Pine rockland fuels comprise grasses, shrubs, and

pine needles, which also are represented by the NFFL

Southern Rough Fuel Model. In contrast to flatwoods,

the shrub layer is composed primarily of species of

West Indian origin including white bully, varnish leaf,

cocoplum, and myrsine. Some of these shrubs will

reach tree stature in the absence of fire. Of about 100

shrub species found in the pine rocklands, only seven

come from the pine flatwoods flora (Snyder and others

1990). The herbaceous layer is a diverse mixture of

herbs of both tropical and temperate origin, with a

high percentage of endemic species.

Postfire Plant Communities

Pine Flatwoods and Pine Rocklands

Vegetation Dynamics—Fire is not a successioninitiating process because the pinelands are fire-maintained vegetation types. Fire is as vital as rainfall in

maintaining the vegetation. Postfire species composition is virtually identical to the prefire vegetation

composition. Some mortality of the overstory pine may

occur in any or all age classes; the soil surface may be

exposed, but only for a few weeks; the fuel biomass is

reduced and nutrients are released. Released nutrients coupled with exposed soil interact to stimulate

flowering in a number of the ground cover species.

Some species, particularly the grasses, will rarely if

ever flower without fire. The season when a fire occurs

can have a strong inluence on the flowering response

of some species (Robbins and Myers 1992).

The pinelands of southern Florida fall into two broad

categories based on soil substrate and composition of

the ground cover: subtropical pine flatwoods and subtropical pine rocklands. Pine flatwoods occur on flat,

poorly drained acid sands that were deposited on

ancient marine terraces. Across Florida, pine flatwoods

compose the most extensive ecosystem type and form

the fuel matrix in which many other vegetation types

are embedded. Fires originating in the flatwoods have

strongly influenced the structure, composition, and

juxtaposition of other fire-maintained and fire-influenced habitats, particularly cypress swamps, bays,
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Chapter 7: Fire in Tropical and Subtropical Ecosystems Myers

marshes, and hammocks. See chapter 4 for additional

discussion of pine flatwoods and other southern pine

forest types.

Similar fire-maintained pine forests and savannas

that occur on poorly drained acid sand substrates are

found in western Cuba and on the Isle of Pines south

of Cuba (Borhidi 1996). The pine species in Cuba are

Caribbean pine and Pinus tropicalis, which appear to

form vegetation complexes maintained by fire regimes

identical to the regimes of slash pine and longleaf pine

flatwoods, respectively. For example, like longleaf

pine, Pinus tropicalis has a definite grass stage, increasing the probability that young individuals can

survive frequent fires.

The pine rocklands, in contrast to the acid sands of

the flatwoods, occur on alkaline limestone bedrock

that forms a ridge running from north of Miami south

into Everglades National Park. Pine rocklands occur

elsewhere in southern Florida on outcrops of limestone, particularly in the Big Cypress National Preserve and on some of the Lower Florida Keys: Big Pine,

Little Pine, No Name, Cudjoe, and Sugarloaf Keys. In

some areas of the pine rocklands, exposed rock makes

up 70 percent or more of the surface (Snyder and

others 1990) (fig. 7-2). Although somewhat elevated

from surrounding wetlands, the water table in the

pine rocklands, like in the pine flatwoods, is at or near

the surface during the wet season.

The pine species in both subtropical flatwoods and

rocklands is the southern variety of slash pine (var.

densa), which might be somewhat more fire tolerant

than its northern Florida counterpart (var. elliottii).

South Florida slash pine is described as having a firetolerant grass stage, similar to longleaf pine, that

provides young individuals some protection from lowintensity surface fires (Little and Dorman 1954). The

trait, however, is not nearly as pronounced as in

longleaf pine. Seedlings of south Florida slash pine

over a year old do develop a thicker stem and bark than

the northern variety. This may offer south Florida

slash pine relatively greater protection, but they possess nothing like the true grass stage of longleaf pine.

The other pine species typical of Florida’s flatwoods,

longleaf pine and pond pine, do not extend into the

subtropical zone.

As a mature tree, south Florida slash pine has thick

protective bark that insulates the tree and high open

branches that facilitate heat dispersal. Portions of the

outer layer of the platelike bark also peel off as they

are heated by a passing flame front, serving to dissipate heat. Like all of the fire-tolerant southern pines,

south Florida slash pine continuously flushes new

needles throughout the growing season. This gives

trees the capacity to survive 100 percent crown scorch

from burns occurring during the growing season, provided the buds are not killed. In contrast, equivalent

levels of scorching from fires during the dormant season

are more likely to lead to death of the tree, either

directly or through stress-induced beetle infestation.

Other than the marked floristic differences between

pine flatwoods and pine rocklands, they are quite

similar: an open overstory of pines and a continuous

ground cover of herbaceous and low shrubby fuels.

Fire behavior and the historical range of variability of

their fire regimes are nearly identical: frequent (every

1 to 5 years), low-intensity surface fire that can occur

at any time of the year. Large severe fires are associated with predictable drought, primarily in March,

April, May, and June, but can also occur at other times

of the year. Smaller, and perhaps patchier, fires occur

at more humid and flooded times of the year. Locally

intense fires may be associated with fuels accumulated from tropical storm damage.

Figure 7-2—Pine rockland vegetation on Big Pine Key, Key

Deer National Wildlife Refuge, Florida, consists of an overstory

of south Florida slash pine, a diverse underwood of tropical

hardwoods and palms, and a ground cover of grasses and

forbs. Photo by Ronald Myers.
166 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Myers Chapter 7: Fire in Tropical and Subtropical Ecosystems

In general, the majority of the larger pines survive

the fires, with mortality occurring in clumps possibly

associated with areas of high fuel accumulation or

greater intensity generated by interacting flaming

fronts. Intensity being equal, pine mortality is greatest after fall and winter burns. Likewise, seedlings

and saplings survive fires in isolated clumps that may

be associated with gaps created when adult trees were

killed by previous fires.

In spite of similar fuel characteristics and dominance by the same pine species, the developmental

trajectories of pine flatwoods and pine rocklands in the

absence of fire are markedly different. On betterdrained sites, long-unburned flatwoods develop into

evergreen hardwood hammocks dominated by oaks,

particularly live oak. On poorly drained sites they

develop into evergreen bays (red bay, sweet bay, loblolly

bay, and other genera). The changes are relatively

slow because propagules must arrive from pockets of

hammock or bay forests persisting in fire shadows,

drains, or hollows. In contrast, pine rocklands, in the

absence of fire, take on the character of tropical hardwood forests (hammocks) within a few fire-free decades as the species in the extant shrub layer reach

tree stature, and as other tropical hardwoods invade

from tropical hammocks scattered through the

rocklands. Once a significant midstory of hardwood

vegetation forms in both pine flatwoods and rocklands,

a fire can become lethal and severely damage the pine

overstory. In such cases, postfire recovery of the pines

depends on seed sources, environmental conditions for

regeneration, and subsequent return of frequent fires.

In the continued absence of fire, such a site rapidly

becomes a dense hardwood forest dominated by either

temperate or tropical species.

Unlike many Western forests maintained by understory fire regimes, southern Florida’s pinelands did

not experience broadscale changes due to 20th century

fire suppression efforts. The loss of flatwoods and

rocklands habitats has been largely due to urban

development and associated fragmentation rather than

successional changes brought on by successful fire

suppression. Although fire prevention efforts and suppression activities were implemented in southern

Florida, woodsburning has a long tradition throughout Florida and the Southeastern United States. Where

fuels are uninterrupted across broad landscapes, fires

are common. Fires are set by ranchers, hunters, arsonists, and by accident. The exclusion of fire has been

more pronounced in fragmented, developed landscapes

along the coasts; in relatively small parks, preserves,

and refuges; and in areas where suppression equipment had ready access.

Management Considerations—Both the pine

flatwoods and pine rocklands in southern Florida have

been impacted by the interaction of fire and the spread

of non-native species. Remnant rocklands that burn

infrequently have been invaded by Brazilian pepper, a

subtropical shrub that, once established, will persist

in the shrub layer even if fire is reintroduced. Many

pine rocklands are also threatened by Burma reed, an

exotic grass whose spread is facilitated by frequent

burning. The acid sandy soils of wetter pine flatwoods

are susceptible to invasion and complete dominance

by melaleuca. Melaleuca invasion can lead to a mixed

fire regime consisting of some understory burns and

some crown fires that are nonlethal to melaleuca but

lethal to the pines. Both pineland types are threatened

by cogongrass, a pantropical grass known for its close

association with fire. In stands of pine that have

escaped fire for several cycles and in other woodland

vegetation, Old World climbing fern is becoming a

problem by creating ladder fuels that carry fire into

tree canopies. Fires in the pinelands that normally

would stop at the margins of flooded hardwood and

cypress swamps can burn into these vegetation types

when their canopies are covered with the fern. Burning mats of the light-weight fern break free and are

kited away by convection columns, igniting spot fires

well downwind from the main fire.

Because of a long history of burning, first indiscriminate then prescribed, fire has been much more readily

accepted as an essential land management activity in

Florida than it has been elsewhere. Florida law allows

the State to conduct hazard reduction burns on unoccupied private wildlands, and it recognizes prescribed

burning as a right of the landowner. Prescribed hazard

reduction burns and ecological management burns

began in the Everglades in the 1960s. Today, the

National Park Service’s Big Cypress National Preserve burns more acreage annually than any other

National Park Service unit in the United States, and

more than any other publicly administered wildland

in Florida.

Failure to burn either pineland type at frequent

intervals (2 to 7 years) leads to rapid fuel buildup,

changes in the vegetation structure, changes in species composition, and eventual habitat loss. Selection

of an appropriate fire intensity depends on ecological

and management objectives. Burns can be conducted

at any time of the year. With fires of equal intensity,

the pines are more susceptible to stress from scorching, other damage, and death from fall burns. Dry

season burns are more effective in removing duff and

exposing the mineral substrate, but in the pine

rocklands, deep burns may occur in organic matter

that has accumulated within cracks in the rock. These

may result in root damage and death of the pines, but

also they may effectively reduce the density of shrubs

on long unburned sites.

Many of the understory species respond favorably to

growing season burns, particularly when occurring at
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 167

Chapter 7: Fire in Tropical and Subtropical Ecosystems Myers

the transition between dry and wet seasons. Generally, managers favor winter burns for fuel reduction

and growing season burns for ecosystem maintenance.

Pondcypress Wetlands

Vegetation Dynamics—Cypress-dominated vegetation types in southern Florida cover roughly 800

square miles (2,000 sq. km) (Wade and others 1980).

Two variants of cypress occur in southern Florida:

baldcypresss and pondcypress (var. nutans). Baldcypress grows in floodplain forests, around shores of

large lakes, and in the interior of large cypress strands

(broad vegetated drainage depressions), all of which

burn at intervals of centuries. Pondcyress grows in

frequently burned savannas, shallow depression ponds

called cypress domes or cypress heads, and on the

periphery of cypress strands (fig. 7-3).

Although the role of fire in the dynamics of cypressdominated wetlands is poorly understood (Ewel 1995),

it is known that nonlethal, understory fire regimes

prevail in the pondcypress savannas and woodlands

rather than the baldcypress forests. The state-transition model (fig. 7-4) illustrates that fire regimes of

various types play a role in the dynamics of cypressdominated wetlands, with understory fire regimes

responsible for monospecific pondcypress forests and

savannas, including what are known locally as cypress

domes, strands, and dwarf forests (Wade and others

1980).

Cypress domes occur in circular depressions or ponds;

cypress strands form in elongated shallow drainage

channels. Dwarf cypress forests occur on shallow soils

over limestone bedrock and may include domes,

strands, and savannas. In both domes and strands,

trees increase in stature from the periphery to the

center or midline of the depression, giving the vegetation a domed or ridged appearance.

Numerous theories involving the interplay of soil

depth, hydroperiod, water depth, and fire have been

postulated to explain the cause of the domed and

ridged appearance. Fires are generally more frequent

around or along the periphery of domes and strands,

and one study concluded that this “differential marginal fire theory” accounts for the doming and ridging

(McJunkin 1977). As one enters a dome or strand, tree

height increases, crown closure becomes more complete, and fine herbaceous fuels are either no longer

continuous or are submerged. This, coupled with longer

hydroperiods toward the interior, limits the influence

of fires to those times when protracted drought coincides with fires originating in the surrounding vegetation, such as pine flatwoods, wet prairie, or cypress

savanna or woodland. The longer interval between

Figure 7-3—Cypress-dominated landscape with cypress domes, woodlands, and savannas, Big Cypress

National Preserve, Florida. Photo by Ronald Myers.
168 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Myers Chapter 7: Fire in Tropical and Subtropical Ecosystems

fires coupled with increased flooding allows the accumulation of organic soils.

During those rare instances when fire penetrates to

the interior of domes and strands, the organic matter

may be consumed and the fires become lethal to the

cypress causing stand-replacement, even though there

is no flaming front. Such fires usually recur on the

order of centuries. When they occur more frequently,

open water or herbaceous marsh vegetation is present

rather than forest. Some domes have an open water

center giving them an appearance of a doughnut from

the air. Similarly, the midline of strands may possess

a series of ponds. Rare, severe fire events that consume organic soil probably play a role in the creation

of these ponds.

Dwarf cypress forests and savannas occur on shallow sandy or muck soils overlying limestone bedrock.

Density of stunted cypress trees varies from a few to

nearly 700 trees per acre, with a ground cover of

grasses, sedges, and forbs serving as the principal

carrier of the surface fires (Wade and others 1980).

Dwarf cypress forests form a continuum grading into

cypress strands and domes as soils deepen,

hydroperiods lengthen, and maximum water depth

increases. Productivity and hydroperiod mediate the

accumulation and density of the herbaceous fuel bed,

which to a large degree controls the recurrence of fires.

Return intervals may range from 5 to 20 years in dwarf

cypress forests.

Little is known about the fire tolerance of pondcypress, but its bark has good insulation properties

relative to other native wetland species, and the trees

seem quite tolerant of low-intensity surface fires (Hare

1965). Limited studies have shown that pondcypress

is much more likely to survive surface fires than

competing hardwoods and shrubs (Ewel and Mitsch

1978), and years of observing wildfires and prescribed

burning in the Big Cypress National Preserve attest to

pondcypress’ ability to withstand fire. Some trees,

particularly those in dwarf cypress forests, possess the

ability to resprout from their swollen buttresses if topkilled. Observations also point to the importance of

fire in checking the encroachment of swamp hardwoods and permitting the establishment of cypress

seedlings (Gunderson 1977).

After a long fire-free interval, pondcypress domes

and strands develop a hardwood understory. In the

continued absence of fire, the pondcypress is replaced

by a hardwood-dominated swamp or bay. Once the

shrub layer forms a continuous fuel bed under the

cypress, conditions for lethal fires exist with the potential to eliminate the cypress overstory. It has been

postulated that many of the bayheads (clumps of

evergreen hardwood shrubs and trees) in the Everglades represent former cypress domes that were

burned out during the droughts of the late 1930s

(Wade and others 1980). If such fires include ground

fuels, all woody vegetation would be removed; depending on the depth of the burn this would create open

water, marsh, or willow, or permit re-establishment of

cypress, provided there was a seed source. Without a

ground fire, the hardwoods would resprout creating

either a deciduous hardwood swamp dominated by red

maple and Carolina ash, or a bay dominated by red

bay, sweetbay, and loblolly-bay.

Shallower pondcypress wetlands are susceptible to

invasion by melaleuca following fire. There are a

number of examples where a single high-intensity fire

in pondcypress that included scattered mature

melaleuca led to the replacement of pondcypress by

melaleuca.

Management Considerations—Most prescribed

burning in cypress is limited to the savannas and

woodlands that have herbaceous fuel beds. This effectively reduces the hazard of fire getting into deeper

domes and strands during droughts. Burning during

exceedingly dry periods is usually avoided to prevent

excessive damage to trees on organic soils and to avoid

extended smoke problems. However, preventing or

avoiding “muck” fires will eventually lead to the loss of

Figure 7-4—Generalized succession scheme for showing the

role of different fire regimes in cypress wetlands and related

vegetation types (Gunderson 1984).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 169

Chapter 7: Fire in Tropical and Subtropical Ecosystems Myers

deeper water cypress forests and their habitats as they

succeed to hardwood swamp forests.

Cabbage Palmetto Savannas and Forests

Cabbage palmetto-dominated savannas and forests

are most prevalent on soils having a calcareous layer

(shell, limestone) near the soil surface. The role of fire

in creating and maintaining palm-dominated communities is poorly understood, but cabbage palmetto is

tolerant of fire (Myers 1990a). Its single terminal bud,

even as a seedling, is well protected from fire. The

absence of a fire-sensitive surficial cambial layer in

monocot stems allows palms to survive considerable

charring, even some consumption of the stem. The

palms appear able to withstand higher intensity fires

than slash pine. Cabbage palmetto is frequently a

codominant with live oak in hardwood hammocks.

Where cabbage palmetto is the sole dominant, it may

indicate that either the pine forest or live oak forest

had been eliminated in a severe fire or series of fires,

or perhaps that the pines had been harvested. In the

absence of fire, oaks invade palm forests, eventually

becoming the dominants.

Live Oak Forests

Live oak is generally considered a non-fire type or

climax of pine and palmlands in the absence of fire.

Live oak forests, however, do burn. Saplings and small

trees have the ability to both resprout and to send

runners forming clones. Large trees are protected by a

relatively thick bark. Live oak litter produces a particularly compact fuel that limits flame length. It

rarely burns without the momentum of a heading fire.

The litter also holds moisture making it fire resistant

except during periods of extreme drought. When it

burns, flame lengths are short (<2 feet). Intense fire

with some crowning may occur where the oak forest

has expanded over a continuous saw palmetto fuel bed

such as in long unburned pine flatwoods or dry palmetto prairies (Huffman and Blanchard 1991). Live

oak forests are commonly embedded in flatwoods,

prairies, and palmlands on sites somewhat protected

from the frequent fires that normally spread through

the matrix vegetation. For example, oak forests are

frequently found on the leeward side of wetlands,

ponds, lakes, and wet drainages where they would

rarely be exposed to direct heading fires. The importance of these “fire shadows” was first pointed out by

Harper (1911). Live oak forests that have gone for

extended periods without being exposed to fire may

exhibit a higher diversity of broadleaved trees and

shrubs including other oaks and red maple. At more

southerly and coastal locations in the Florida peninsula, this component takes on a more tropical character. Where oaks have expanded into other vegetation

types due to fire exclusion, repeated growing season

burning, undertaken before the summer rainy season,

is most effective in killing oaks and restoring historic

landscape patterns (Huffman and Blanchard 1991).

Mixed Fire Regimes _____________

Major Vegetation Types

Melaleuca forests, woodlands, and savannas are an

unusual wooded vegetation type dominated by an

introduced Australian tree, melaleuca. It occurs in

both Florida and Hawaii. In Florida, melaleuca is an

aggressive invader of fire-prone wet prairies, shallow

marshes, wet pine flatwoods, and cypress swamps

(Meskimen 1962; Myers 1983); in Hawaii, it is a

component of upland plantations composed of a variety of other introduced species. Although the Hawaiian melaleuca plantations are subject to anthropogenic ignitions creating significant fire control

problems, it is in Florida where melaleuca seems most

at home and appears to be gaining control of the fire

environment (fig. 7-5). By the mid-1990s nearly 500,000

acres (202,400 ha) were infested with melaleuca, and

of that, 40,000 to 50,000 acres were considered “pure”

monospecific forests (Ferriter 1999).

Fire Regime Characteristics and Vegetation

Response

The fire regime mediating Florida’s melaleuca forests varies from one characterized by low-intensity

surface fires in savannas, with some torching of individual trees, to high-intensity crowning fires in denser

stands. Regardless of the fire intensity, little or no

mortality occurs to any of the trees beyond the seedling stage (Myers and Belles 1995). The bole of

melaleuca is protected by thick spongy bark. The outer

bark layers carry fire into the crown, consuming

branches and leaves and releasing seed from serotinous capsules. New branches sprout from the barkprotected larger stems. Fire in stands of melaleuca

containing any mature capsule-laden individuals leads

to the spread of the melaleuca forest into susceptible

habitats nearby, resulting in a shift from a fire regime

controlled by surface fuels to one dominated by aerial

fuels. This is a fire regime heretofore unknown in the

Florida environment and is likely to result in significant changes to wetland habitats, especially the species composition. Once melaleuca gets a foothold in a

pine or cypress dominated habitat, the shift from a

low-intensity to high-intensity fire regime results in

the mortality of the native pine and cypress and

subsequent conversion to melaleuca. This regime of

varying fire intensity makes it difficult to categorize

the melaleuca fire regime in table 7-1. Understory

burns occur in melaleuca savannas. In mixed stands of
170 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Myers Chapter 7: Fire in Tropical and Subtropical Ecosystems

melaleuca and cypress or pine, the fires are lethal to

cypress or pine but not to melaleuca. In pure melaleuca

forest, high-intensity crowning fires are not lethal to

the main stem of the trees. The combination of limited

stem mortality and high-intensity fire is unusual in

North American ecosystems. Placing melaleuca forest

in the mixed fire regime is a compromise between low

mortality and high intensity.

Management Considerations—Controlling the

spread of melaleuca is a major concern and challenge

on public lands in southern Florida. Successful control

involves a strategy of first targeting outlying individuals and populations, then treating mature trees individually with herbicide, followed by prescribed burning after released seeds have germinated. Seedlings

and saplings less than 3 feet tall are generally killed in

these burns. Larger individuals will resprout (Myers

and Belles 1995).

Stand-Replacement Fire

Regimes _______________________

Major Vegetation Types

Vegetation types ranging from hardwood forests to

grasslands in subtropical Florida, Hawaii, Puerto Rico,

and the Virgin Islands are characterized by standreplacement fire regimes.

Grassland vegetation types are considered to be

stand-replacement regimes because the dominant

aboveground vegetation is burned and replaced. When

grasslands possess an overstory of trees—that is, a

savanna type—the fire regime becomes an understory

fire regime if repeated fires are primarily nonlethal to

the overstory (see chapter 1). In Florida, the prairie

types, known as dry and wet prairies, have ground

cover and surface fuels that are identical to that found

in pine flatwoods or pondcypress savanna without the

overstory trees. The lack of a tree overstory, which

determines whether a fire regime type is stand-replacement or understory, may be related to a history

of frequent (nearly annual) burning or other aspects of

fire and land use. Whether open pine, cypress or

prairie, the fuel characteristics, burn conditions, and

fire behavior are nearly identical, and the ground

cover vegetation responses are the same.

In many tropical environments, including Hawaii,

Puerto Rico, and the Virgin Islands, there is an interplay between lethal and nonlethal stand-replacement

fire regimes, one fueled by grasses, the other by forest

fuels. Fires originating in agricultural fuels, usually

non-native range grasses, burn up to and penetrate

forest edges killing trees and allowing grasses to

encroach at the expense of forest. Once within the

forest, these fires create feedbacks in future fire susceptibility, fuel loading, and fire intensity that favor

grass fuels—a process, which if left unchecked, has the

potential to transform large areas of tropical forest

into shrubland, savanna, or grassland (D’Antonio and

Vitousek 1992; Koonce and Conzales-Caban 1990).

The stand-replacement fire regime type also characterizes the fires that occasionally occur in mangrove

vegetation, along with frequent fires in Florida’s salt

and freshwater marshlands, and fires that occur in a

number of successional stage communities between

marshes and swamp forest. These successional communities are dominated by willows, ashes, and bays.

Grasslands and herbaceous wetlands are common

fire-maintained vegetation types in both Florida and

Hawaii, with notable differences between the two

locations. In Hawaii virtually all of the grasslands are

dominated by introduced exotics, such as thatching

grass, Natal redtop, molasses grass, broomsedge, fountain grass (fig. 7-6), and Columbian bluestem. These

fuels have created altered fire regimes to which the

Figure 7-5—Torching melaleuca during prescribed research

burn in a melaleuca-invested wet prairie, Big Cypress National

Preserve, Florida. Photo by Holly Belles.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 171

Chapter 7: Fire in Tropical and Subtropical Ecosystems Myers

native Hawaiian flora are not adapted (Smith and

Tunison 1992). In some cases, the nonnative grasses

form the understory of introduced pine and eucalyptus

plantations. In Florida a few introduced grasses are

causing problems in pinelands, but Florida’s extensive

wet and dry prairies are native fire-maintained ecosystems.

In subtropical Florida, most native herbaceous vegetation types are associated with wetlands. They include salt marsh, sawgrass marsh, wet prairie, and

miscellaneous broadleaved herbaceous marshes. The

vegetation type known as dry prairie or palmetto

prairie represents the ground cover vegetation of pine

flatwoods or palm savanna without the trees. It is

more common in the transition zone between temperate and subtropical vegetation in Florida and is on the

periphery of the region discussed in this chapter.

Fire Regime Characteristics and Vegetation

Response

Florida Freshwater Marshes and Wet Prairie—

Freshwater marsh and wet prairie vegetation types

include dense and sparse sawgrass marshes, wet prairie, marl prairie, spikerush flag marsh, beakrush flag

marsh, and American white waterlily marsh (Kushlan

1990). Fire in these marsh types affects species composition and may limit or reduce peat accumulation. The

sharp demarcation between different marsh types

frequently indicates the boundary of a past burn

(Kushlan 1990). The fire regime and fire effects in

these herbaceous wetlands result from the interplay of

hydroperiod and fuels, which together determine

whether the fires are lethal or nonlethal to the dominant species, and whether or not fire is a successioninitiating process. In most cases, the aboveground

vegetation is consumed, but the fires are usually not

lethal to the dominant species that make up the fuel.

These species simply resprout from underground buds,

tubers, or rhizomes.

Two conditions can create lethal fires and initiate a

vegetation change (Herndon and others 1991). One is

when fire coincides with severe drought and consumes

some, or all, of the organic substrate, destroying root

systems and underground regenerative organs of the

dominant species. The second occurs when water levels rise faster than vegetative regrowth after a burn

and the site remains flooded long enough to cause the

death of the vegetation. In either case, vegetation is

replaced primarily by species present in the seed bank.

Figure 7-6—The introduced fountain grass retains dead leaves due to its bunchy structure and is highly

flammable. It is an aggressive invader that readily replaces native plants especially where it carries fire into

less flammable native vegetation. It invades lava beds creating continuous fuels where fuel breaks formerly

existed. Photo by Jim Brown.
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Myers Chapter 7: Fire in Tropical and Subtropical Ecosystems

Sawgrass, once covering several million acres, is the

dominant marsh vegetation type in the Everglades.

Dense sawgrass marsh occurs on organic soils, while

sparse sawgrass marsh is found on marl soils. Fuel

loadings in dense sawgrass are sufficient to allow fires

to burn over standing water, and lightning ignitions

are common. These wet season fires go out when they

burn into other marsh types. During the dry season,

fires may burn through sparse sawgrass stands and

wet praires with low fuel loads by smoldering through

a dry algal mat called periphyton, composed mostly of

filamentous blue-green alge, that forms over the substrate surface when the marsh is flooded. During

severe droughts, organic soils may be consumed in

both dense sawgrass and in many of the deeper water

marsh sites. Sawgrass marshes and wet prairies can

burn every 3 to 5 years (Wade and others 1980). In the

absence of fire, sawgrass will succeed through a willow

stage to hardwood bay vegetation.

Wet prairies are the least flooded of Florida’s marsh

vegetation types. The dominant species are identical

to those in pine flatwoods and cypress savannas. Soils

vary from periphyton-derived marls to sands. Dominants include maidencane, cordgrass, beakrush, and

hairawn muhly. The fire frequency is 2 to 5 years.

Flag marshes are named after herbaceous broadleaved marsh species that have a flag appearance.

They have a long hydroperiod and only burn during

severe droughts. These fires may consume organic

soils.

Except for the deepest water marshes, all types are

susceptible to invasion by melaleuca. Dense sawgrass

marshes and wet prairies are particularly susceptible.

Florida Salt Marsh and Mangrove—Although

salt marshes occur in both the temperate zone and the

tropics, fire mediates the tension zone between mangrove and salt marsh only in the tropics. In subtropical

Florida, salt marsh is wedged between mangrove on

the seaward side and freshwater marsh on its inland

edge. Freezing temperatures probably have some influence on the juxtaposition of mangrove and salt

marsh, but fires originating either in the salt marsh or

further inland in freshwater marsh, control the inland

advance of mangrove. When intense fires are stopped

by mangrove, the outer fringe of trees is killed and the

marsh expands (Wade and others 1980). Under moderate burn conditions the mangrove acts as a firebreak. Similar fire dynamics probably occur between

marshes and mangrove in Cuba and the Bahamas.

Florida’s salt marshes are dominated by black rush,

gulf cordgrass, sand cordgrass, and inland saltgrass,

mixed with a number of species found in fresh water

marshes, notably sawgrass and cattail. Because salt

marsh is under tidal influence and relatively isolated

from human activities, a large proportion of ignitions

are lightning caused. Fires supported by high fuel

loadings frequently burn over standing water. Fire

behavior varies considerably depending on the dominant species contributing to the fuel. Cordgrass fuel

loadings are frequently as high as 22 tons/acre (49 t/ha)

with a fuel bed depth of 8 feet. Beakrush fuel loadings

and fuel bed depths are half that amount. Saltgrass fuel

loadings vary considerably, and the fuel bed depth is

only 1 to 2 feet (Everglades NP FMP 1991).

Mangroves rarely burn, but they are influenced by

fire in seasonal environments such as in Florida.

Mangroves are a tropical and subtropical forest type

growing in brackish to high salinity coastal sites that

have weak wave action. Four species of mangrove are

found in Florida: red mangrove, black mangrove, white

mangrove, and buttonwood. Each tends to be indicative of different zones of salinity or tidal influence.

Lightning may be an important factor in the structure

and dynamics of mangroves by creating numerous

circular holes of dead and dying trees that may develop

into patches of more flammable herbaceous vegetation. Fire is responsible for checking the encroachment of mangrove into salt marsh, and it is not

uncommon to find red or white mangrove scattered

through long unburned fresh water marshes. They

have even been observed in the understory of cypress

domes.

Florida Coastal Prairies—Coastal prairies are

closely related to salt marsh; they occur in southern

Florida, the Bahamas, and Cuba. They are less frequently inundated than salt marshes, but are sometimes flooded with salt or brackish water. Species

composition includes saltgrass, seaside tansy, and

batis. Coastal prairies are maintained by a combination of fire and hurricanes (wind damage and storm

surge). If fire is absent for several decades, coastal

prairies develop into buttonwood forests (Craighead

1971). The extent of coastal prairies may be a function

of past clearing for charcoal production followed by

cattle grazing and associated frequent burning. Coastal

prairies on the Zapata Peninsula in Cuba have been

contracting at the expense of buttonwood since the

area was made a national park and the burning

associated with cattle operations was curtailed (Myers

1999). See chapter 4 for additional discussion of freshwater and salt marshes.

Florida Tropical Hardwood Forests—Hardwood

forests in southern Florida are usually islands imbedded in a matrix of marsh, prairie, or savanna. Fires

burning in tropical hardwood forests (called hammocks in Florida), hardwood swamps, and bays or

bayheads likely originate in the more easily ignited

matrix fuels. The forest islands usually serve as effective firebreaks with fires burning only at their

periphery. Fires have the opportunity to enter these

ecosystems during extreme droughts, where they may

cause conversion to earlier successional stages or
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 173

Chapter 7: Fire in Tropical and Subtropical Ecosystems Myers

shifts to marsh vegetation. Depending on fuel and

weather conditions, vegetation structure, and type of

substrate, the fires may be low-intensity surface fires,

ground fires that burn out organic soils, or crowning

fires moving through dense low shrubs, palmettos, or

trees. Many of the tropical hardwoods have the ability

to resprout if top-killed, but fires are lethal if the

organic substrate is consumed. It is not uncommon for

fires burning through grass fuels to go out at night as

humidity recovers, but will hold over in hardwood

forests, igniting the grass fuels the next day.

Hawaiian Forests and Grasslands—Fire regimes

in Hawaiian native forest were probably always standreplacement. Lava flows are the most common cause of

natural ignitions, but resulting fires probably have

not been an important evolutionary force shaping the

characteristics of the vegetation. Some tree species

such as koa and a’ali’i show some tolerance to fire, but

these may represent preadaptations from their

noninsular evolutionary environment (Smith and

Tunison 1992). Most of Hawaii’s native vegetation is

extremely sensitive to fire. Historically, lightning may

have caused some ignitions; today, lightning ignitions

may be more common due to the prevalence of exotic

grass fuels. Human sources, however, are by far the

most frequent cause of fires.

The litter fuels in native Hawaiian forests, which

range from dry forest to montane rainforest, are generally not sufficient to carry fire. The woody vegetation

itself is not flammable. The presence of flammable

grasses is essential for fires to spread. The grass fuels

tend to be available throughout the year. Fire may

have played a natural role only in the seasonal montane

environment where, in places, native grasses and shrubs

form a continuous fuel bed (Mueller-Dombois 1981).

Vegetation Dynamics—Although the vegetation, and

thus fire regimes, of Hawaii changed dramatically in

the later part of this century, some changes actually

began over 1,600 years ago when the Polynesians

colonized the Islands. The Polynesians encountered

vegetation that was not fire prone, but, like indigenous

peoples elsewhere, they used fire to manipulate the

vegetation to plant crops, to facilitate travel, and to

stimulate native grasses used for thatch such as pili

grass. Early Euro-American visitors to the Islands

reported encountering open savannas and grasslands

at lower elevations and observed Polynesian burning

practices (Kirch 1982).

Fire frequency and fire size greatly increased in the

late 1900s when non-native grasses were introduced

and spread (fig. 7-6). The invasion of fire-prone alien

grasses coincided with the removal of feral goats from

areas like Hawaii Volcanoes National Park. The grasses

provided the fuel that carried fire into nonfire-adapted

vegetation leading to the death of the native vegetation and its replacement by the grasses. Many fires are

confined to mesic and dry forest habitats; however,

grasses are also encroaching into some rainforest

habitats and subalpine ecosystems. The prevalence of

fire in these areas has been increasing (MuellerDombosis 1973).

Management Considerations—Managing fires in

Hawaii is problematic. Fires of all intensities, timing,

and sources are destructive to the Islands’ native

ecosystems, and most fires should be aggressively

suppressed even if they result from natural ignitions.

Prescribed fire has potential as a tool to reduce alien

grass fuels and create firebreaks to prevent fires from

entering sensitive native vegetation. Evidence indicates that prescribed fire may have a limited application in the restoration of a’ali’i shrublands and koa

forests or woodlands because fire stimulates resprouting in both of these species. Prescribed fire

might be useful in managing habitat for the endangered Hawaiian goose or nene (Nesochen sandwicensis)

(Smith and Tunison 1992).

Forests of Puerto Rico and the Virgin Islands—

Unlike Cuba and Hispaniola, which have native firemaintained pine and palm forests and savannas and

fire-maintained herbaceous marshes and wet prairies, Puerto Rico and the U.S. Virgin Islands do not

possess any significant fire-adapted native vegetation

types. Native forest types of Puerto Rico and the

Virgin Islands include mesic forests on the windward

sides that grade into rainforest or montane cloud

forest with increasing elevation. The leeward lowlands support dry forest types. Fires occur in the dry

forests, but the sparse accumulation of litter fuels only

supports low intensity surface fires that generally go

out at night as the humidity rises. If the thin-barked

trees are top-killed, many have the ability to resprout,

probably from an adaptation to drought, not fire.

Where disturbances have created grassy openings,

higher intensity fires may cause significant damage to

the forest cover particularly at the forest edge. Repeated burning favors the grasses at the expense of

forest. Similar fire damage occurs on more mesic sites

where agricultural fires may escape and encroach into

the forest during dry periods. This pattern of burning

is more prevalent and a greater problem in other

tropical areas where slash and burn (swidden) agricultural practices are common.

Severe fires have occurred in hurricane-damaged

tropical dry and seasonal forests on the Yucatan Peninsula in Mexico (Whigham and others 1991). The

potential for similar damage and subsequent fires

exists in all of the tropical regions covered in this

chapter.
174 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Notes

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USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 175

Kevin C. Ryan

Chapter 8:

Global Change and

Wildland Fire

Global change, the combined effect of human activity on atmospheric and landscape processes (Vitousek

1994), affects all aspects of fire management. Scientists have documented changes in the global carbon

cycle due to increases in atmospheric carbon dioxide

(CO2), changes in biogeochemical cycling due to increased nutrient deposition (for example, nitrogen),

and changes in land use and cover. These changes are

expected to continue for the foreseeable future (IPCC

1996a,b).

Changes in the global atmospheric chemistry are

attributed to biomass burning and industrial processes. These alterations in the chemical makeup of

the atmosphere are predicted to have a significant

impact on biogeochemical processes and Earth’s radiation balance, the so called “greenhouse effect.”

These changes in the chemical composition of the

atmosphere and Earth’s energy balance can be expected to modify precipitation, temperature, humidity, and vegetation development—all affecting fire

management. In addition, historic changes in patterns of land use—roads, subdivisions, timber harvesting, farming, and ranching—have altered vegetation and fuels, affecting the potential ignition, spread,

and severity of wildland fires. Continued migration of

people into wildlands further complicates prescribed

fire management and wildfire suppression.

Because of the complex interactions of all these

processes, it is difficult to make definitive estimates

about the rate and, in some cases, even the direction of

change. However, given current knowledge, the anticipated changes can be expected to increase the pressure

on fire management organizations. This chapter examines the complexity of global change and the possible

influences on vegetation and fire management.

Changes Over Time______________

Vegetation and fire regimes have been in a constant

state of flux throughout geologic time. Climate has

changed throughout the millennia (Bradley 1999).

New species have evolved as others became extinct,

with climate and herbivory as the dominant influences. In the current geological epoch, the Holocene (0

to 10,000 years before the present), activities of humans have increasingly influenced vegetation and

fire. Numerous climate fluctuations during this time

include the Medieval Warm Period, AD 900 to 1350,

and the Little Ice Age, 1450 to1900. Average temperatures have varied by as much as 5.4 °F (3 °C) over
176 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Ryan Chapter 8: Global Change and Wildland Fire

periods as short as a few decades (Bradley 1999; Mann

and others 1999). Human societies (Ahlstrom and

others 1995; Lipe 1995) and the prevalence of fire

have been significantly affected by these changes

(Bonnicksen 1999; Clark 1990; Swetnam 1993).

Prior to the arrival of Europeans in North America,

native people routinely used fire to drive game animals and manage vegetation near encampments

(Barrett and Arno 1982; Bonnicksen 1999; Boyd 1999;

Clark and Royall 1995, 1996; Pyne 1982). In some

regions Native Americans developed large agrarian

communities where vegetation was extensively altered. Although the degree to which fire was used to

initiate and maintain agriculture is uncertain, agriculture and harvesting of biomass for energy did lead

to substantial change in fire regimes and vegetation in

some areas. Across much of the landscape, lightning

was the primary source of ignition. And fire spread

was only hindered by the availability of fuels, weather,

and natural barriers. Forest and rangeland sites developed with characteristic fire regimes and vegetation. Landscapes developed with a characteristic mosaic of stands of varying age, structure, and species

composition. Fauna developed life cycle and behavior

patterns tuned to these landscape patterns. While

early Euro-American settlers may have seen many

desirable features in existing patterns of vegetation,

these features were not static but represented only a

point in time in the development of North American

vegetation (Betancourt and others 1990; Bradley 1999;

Bonnicksen 1999; Delcourt and Delcourt 1987; Prentice

and others 1991; Woolfenden 1996).

Since Euro-American settlement across the continent, fragmentation of the landscape resulting from

agriculture, mining, and urbanization (Bahre 1991;

Baker 1992; Veblen and Lorenz 1991) has significantly

altered the fire potential of many ecosystems. This

transformation of the vegetation prevented fires that

formerly swept across prairies and steppes into adjacent forests (Gruell 1985). Domestic livestock have

reduced the availability of fine fuels for fire spread.

Grazing and fire exclusion together have led to the

replacement of grasslands by shrublands in some areas

(Wright and Bailey 1982). The introduction of exotic

species has led to substantial changes in the species

composition and fire potential of many ecosystems,

particularly in arid and semiarid areas (Billings 1990).

Timber harvesting has led to unnatural patterns of

vegetation, modified fuel beds, and altered fire severity.

One of the most significant changes in land use in

the 1900s was the suppression of wildfires. Fire suppression has led to changes in species composition and

vegetation structure, and it has led to a significant

buildup of fuels (Arno and Brown 1989) and increased

forest health problems (Mutch 1994). The shift from

ranching to ranchettes and urban encroachment on

wildlands (Riebsame and others 1997) is also leading

to a buildup of fuels. The result is that fires, though

apparently less frequent than in the 1800s, are now

often larger and more severe (Agee 1993; DeBano and

others 1998; Sampson 1997) than formerly. When fires

do occur they can result in serious threat to life and

property. Larger, more severe fires also have greater

potential to adversely affect postfire vegetation composition and structure, as well as soils and water, cultural

resources, and air quality, as described in other volumes of the Rainbow Series (“Effects of Fire on Soil and

Water,” “Effects of Fire on Cultural Resources and

Archaeology,” and “Effects of Fire on Air”).

Emissions from industrial processes, burning of

fossil fuels, and slash-and-burn agriculture in the

tropics have increased the concentrations of greenhouse gasses (GHG) in the atmosphere particularly in

recent decades (Tett and others 1999). Chief among

these is carbon dioxide (CO2), but water vapor (H2O),

ozone (O3), methane, nitrous oxides (NOx), and various chlorofluorocarbons (CFCs) are also important

(IPCC 1996a). Carbon dioxide has risen from approximately 270 ppm in the preindustrial atmosphere to

around 365 ppm. While there is not universal agreement that increases in GHG have caused temperatures to rise, the observed 20th century warming

reversed a millennial cooling trend. The 1990s were

the hottest decade in the millennium, and 1998 was

the hottest year (Mann and others 1999). There is

growing scientific consensus that we are experiencing

a greenhouse warming effect (IPCC 1996a).

Changes in the atmosphere due to GHG are expected to alter global weather patterns and significantly change regional climate. Due to the complexity

of general circulation models used to predict climate

change, there is much uncertainty as to the magnitude

of effects of increased GHG on regional climate. Globally, average annual temperatures are expected to

increase on the order of 2 to 8 °F (1.1 to 4.5 °C),

depending on location. At this time estimates of regional climate changes are more tentative than estimates of global change, but increases are expected to

be greater at high latitudes, in mid-continent regions,

and in fall and winter (IPCC 1996a). The growing

season may be extended by 1 to 2 months depending on

latitude and altitude. Average annual precipitation

may increase as much as 20 percent, but little summer

rain is expected in much of North America’s interior.

Maritime climates may be wetter than today, but it is

uncertain if increases in precipitation will be adequate

to compensate for higher temperatures (Franklin and

others 1991). Because continents are expected to warm

up more rapidly than oceans, the interiors of the

continents are expected to experience major drought

by the middle of the next century (IPCC 1996a; Rind

and others 1990).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 177

Chapter 8: Global Change and Wildland Fire Ryan

Climate is generally defined as the 30-year average

weather for a location. While temperature change (that

is, global warming) is the major focus, the atmosphere,

hydrosphere (water), cryosphere (ice), and biosphere

(flora and fauna) are mechanistically coupled, and their

interactions affect the relations between climate and

wildland fire and between vegetation and fire regimes

(K. Ryan 1991) (fig. 8-1). If the expected global warming

occurs, the increase in CO2 and changes in precipitation

will alter growth and competitive interactions of plant

communities. This will result in changes in ecosystem

structure and species composition.

Climate change, therefore, will directly affect the

frequency and severity of weather favorable to fire

spread. Changes in plant communities will affect fire

regimes indirectly by altering the physical and chemical properties of fuels. Fire is a major source of mortality in many communities, but in turn it creates gaps for

new species. As a result, changes in fire regimes will

modify the rate at which communities respond to

climatic change.

Climate, Weather, and Fire

Interactions ____________________

Regional fire potential is strongly coupled to regional climate. This connection between weather and

fire potential is a fundamental part of all models and

training in fire behavior, so clearly any significant

climate change will affect the frequency and severity

of conditions suitable for the ignition and spread of

fires. In addition, each type of fuel has characteristic

physical and chemical properties that affect flammability, and these properties vary with climate and

weather. On the other hand, predicted temperature

increases are not expected to have large direct effects

on fire potential (fig. 8-2) but are expected to be

correlated with a number of other changes that will

affect fire potential (table 8-1). One of these changes is

drying rate. Each type of fuel has its own drying rate,

and its moisture content varies according to site moisture history. Decreased relative humidity can be expected to result in lower fine fuel moisture (fig. 8-2)

and more rapid fire spread, whereas higher humidity

can be expected to result in higher fuel moisture and

less rapid spread. Therefore, if humidity patterns are

altered, the moisture and combustion properties of

fuels will change correspondingly.

Increases in the frequency and severity of drought

and an extension of the length of the fire season (Rind

and others 1990; Wotton and Flannigan 1993) will

result in more severe fires and increased consumption

of long time-lag fuels (logs, duff, and organic soils).

Wind is also a major factor in determining how fast

fires spread. If the frequency of high winds changes,

the potential for large fires will also change. Lightning

is a major source of ignitions, and increased temperature, precipitation, and evaporation will change thunderstorm patterns. As a result, the frequency of lightning-caused fires is expected to increase some 30 to 70

percent depending on location (Price and Rind 1994).

Further, much of the increase is expected to come

during periods of moisture deficit.

Based therefore on climate projections, increases

are expected in the length of the fire season (Wotton

and Flannigan 1993), the frequency of lightning fires

(Price and Rind 1994), the frequency of drought (Rind

and others 1990), and area burned (Flannigan and

Van Wagner 1991; Stocks 1993) in much of Canada

and the United States; but some regions are expected

to experience a decrease in fire activity (Bergeron and

Flannigan 1995). In addition to these direct influences

Figure 8-1—Climate, vegetation, and fire are dynamically

coupled. Any change in one factor will lead to changes in the

other two.

Figure 8-2—Fine fuel equilibrium moisture

content as a function of relative humidity and

temperature.
178 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Ryan Chapter 8: Global Change and Wildland Fire

of weather on fire, storms are predicted to be more

severe in the altered climate (IPCC 1996a). If so,

increased wind damage to forests could greatly increase available fuels.

Complete combustion of biomass results in the production of CO2 and H2O, but combustion is rarely

complete, and a variety of other chemical species are

produced (table 8-1) (Crutzen and Goldammer 1993;

Goode and others 1999; Hao and others 1996; Levine

1996; Ward and Hardy 1991). Globally, biomass burning is a major source of several chemical species in the

atmosphere. Many of the compounds released by burning are greenhouse gasses. Particulate matter (for

example, Pm 2.5, Pm 10) can produce a local shortterm cooling effect by reducing solar heating. Particulates can also result in reduced precipitation (Rosenfeld

1999).

Biomass burning contributes to the overall problem

society faces in managing greenhouse gasses and

providing for clean air. See the volume “Effects of Fire

on Air” in the Rainbow series for a state-of-the-knowledge review of the fire management issues associated

with fire’s impact on local and regional atmospheric

conditions.

Climate and Vegetation

Interactions ____________________

Climate is considered the principal determinant of

vegetation distribution throughout the world

(COHMAP Members 1988; Neilson 1995; Woodward

1987). Solar radiation, temperature, humidity, precipitation, and wind all affect the physiological ecology

of plants (Bazzaz 1996), thereby affecting their ability

to complete life cycles and sustain populations

(table 8-2). Vegetation, therefore, is governed by the

cumulative history of climate, vegetation, and disturbance processes, and as climate changes, the distribution of the world’s vegetation will change. Moreover,

the pattern and severity of disturbance, especially

fire, will also change (Overpeck and others 1990, 1991;

K. Ryan 1991).

Climate-controlled relationships between vegetation structure and species composition occur similarly

along both altitudinal and latitudinal gradients.

Changes occur along these gradients such that every

1,640 foot (500 m) increase in altitude is roughly

proportional to a 171 mile (275 km) increase in latitude (Hopkins bioclimatic law) (McArthur 1972). Given

Table 8-1—Fire interactions with the climate/weather system.

Climate/weather influences on fire Fire influences on climate/weather

Relative humidity Carbon dioxide (CO2)

Wind (speed, persistence, extremes) Carbon monoxide (CO)

Drought (frequency, persistence) Methane (CH4)

Length of fire season Water vapor (H2O)

Lightning (dry vs. wet) Particulates (Pm 2.5, Pm 10)

Dry cold fronts (frequency) Nitrous oxides (NOx)

Blocking high pressure (persistence) Ammonium (NH4)

Trace hydrocarbons

Trace gasses (including VOC)

Table 8-2—Climate/weather system interactions with vegetation.

Climate/weather influences Vegetation influences on

on vegetation climate/weather

Solar energy Albedo

Temperature Evapotranspiration (H2O)

Relative humidity Photosynthesis (O2)

Precipitation (timing, amount) Respiration (CO2, H2O)

Atmospheric chemistry Methane (CH4)

Wind (direction, speed, extremes) Convection

Advection

Desertification
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 179

Chapter 8: Global Change and Wildland Fire Ryan

time to establish equilibrium following an average

annual temperature increase of 6.3 °F (3.5 °C), vegetation zones in the Rocky Mountains can be expected to

shift approximately 2,167 feet (630 m) up mountain

slopes (fig. 8-3), or 213 miles (350 km) farther north.

The rate of vegetation movement associated with the

shift in isotherms is several times faster than known

species migration rates (Davis 1990; Gates 1990).

Massive shifts in biome boundaries should be expected (IPCC 1996a; King and Neilson 1992; Neilson

1993; Overpeck and others 1991). In many cases,

species will not be able to migrate and populations

will become fragmented (Peters 1990).

Greenhouse changes will affect numerous biochemical processes that will alter ecological relationships

(Joyce and Birdsey 2000; Schimel and others 1999).

Photosynthesis, respiration, decomposition, and nutrient cycling will all be affected (Agren and others

1991; Bazzaz 1996; Long and Hutchin 1991; Mooney

1991). Responses of these four major physiological

functions are interdependent and vary with temperature and atmospheric chemistry. However, they each

exhibit different responses to changes in temperature.

For example, the temperature response curves for

photosynthesis and respiration differ, and an increase

in CO2 will have a major impact on photosynthetic

rates. So the current balances within individual plants

or communities cannot be projected into the future.

Substantial increases in water use efficiency (ratio of

the amount of CO2 assimilated during photosynthesis

to the amount of H2O transpired) may result from

increased atmospheric CO2 (Bazzaz 1996; Houghton

and others 1996; Mooney and others 1991; Strain

1987). As a result plant growth may accelerate greatly.

However, all living cells respire, and the respiration

rate increases with temperature (M. Ryan 1991).

Because woody plants have more nonphotosynthetically active living tissue (for example, large root

systems and sapwood) than herbaceous plants, temperature-influenced changes in forests and woodlands

are expected to be relatively large compared to grasslands (IPCC 1996b), and mature forests are more

likely to be severely affected than young forests (M.

Ryan 1991; Waring and Running 1998). The loss of

carbon during respiration will increase with temperature, thereby potentially reducing the effect of increased water use efficiency. If increased CO2 alters

carbon to nitrogen ratios of plants, then decomposition, nutrient cycling, and insect and disease resistance will be altered (Vitousek 1994).

Regional climate dominates the zonation of vegetation, but microclimate, soils, life cycle processes (for

example, germination and growth), and ecological

interactions such as competition, herbivory, and fire

strongly affect the external morphology and physiological ecology of communities within vegetation

zones. All of these can be expected to change in response to changes induced by greenhouse gasses.

Studies have not addressed interspecific and intraspecific interactions that affect growth rates, allocation (that is, how a plant’s growth is allocated

between leaves, roots, fruiting, and so forth), and

community relationships (Joyce and Birdsey 2000;

Mooney 1991). However, given the complexity of species traits, it is unreasonable to expect current community relationships to remain unchanged in the

future (Delcourt and Delcourt 1987; Foster and others

1990; IPCC 1996a,b). For example, temperature, moisture, and photoperiod exert strong controls over phenology and growth.

Species are adapted to a range of seasonal patterns.

Significant changes in these seasonal patterns can

lead to asynchronous development, which can lead to

reproductive failure and growth loss (Grace 1987).

Also, height growth and foliage biomass have been

Figure 8-3—Current vegetation zones in the Bitterroot Mountains of Montana and Idaho as a function of elevation (Arno

1980) (A). Projected vegetation zone changes associated with

warmer annual temperatures associated with a doubling of CO2

in the atmosphere (B). This simple one-dimensional projection

does not take into account the many dynamic interactions but

does illustrate the relative magnitude of possible shifts in

vegetation zones.
180 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Ryan Chapter 8: Global Change and Wildland Fire

shown to increase at elevated CO2 levels (Kramer and

Sionit 1987). If canopy structure changes, the light

available for understory plants will be altered, and

elevated CO2 partially compensates for low light (Cure

and Acock 1986). Because not all species sustain

enhanced growth for long periods of time, the effect on

the competitive relationships between overstory and

understory species is uncertain. Water use efficiency

varies by species, and some species respond to elevated CO2 by increasing root to shoot ratios. Thus,

competition for water and nutrients will change. Species with the C3 photosynthetic pathway (for example,

woody plants and “cool season” grasses) show greater

increases in growth at elevated CO2 than plants with

C4 pathway (for example, “warm season” grasses)

(Houghton and others 1996; Smith and others 1987).

Cheatgrass, an exotic C3 grass in the Western United

States, is especially responsive to elevated CO2 (Smith

and others 1987). Climate change is expected to favor

early successional species assemblages over later ones

(Bazzaz 1996).

Climate not only affects regional vegetation, but

vegetation in turn affects both regional climate and

microclimate (fig. 8-1, table 8-2). The character of

surface vegetation affects the amount of solar energy

absorbed versus reflected. Evapotranspiration from

actively photosynthesizing foliage contributes substantial amounts of water vapor to the atmosphere,

potentially affecting local precipitation. Both living

and dead vegetation produce CO2 during respiration

and release a variety of other compounds to the atmosphere. Some of these compounds are greenhouse

gasses (for example, CH4), and some contribute to air

quality problems such as regional haze and smog (for

example, trace hydrocarbons).

Without question, global change has affected interspecific relationships and will continue to do so, likely

at an accelerated rate. The effects will likely cascade

throughout ecosystems. For example, increased water

use efficiency of upland plants can be expected to

reduce stream flows (Running and Nemani 1991),

thereby affecting aquatic systems.

Interspecific relationships are too complex and poorly

understood to be predicted. Given this complexity of

the interactions, managers and policy makers are not

likely to have significantly improved scientific bases

for their actions, and many changes are likely to go

undetected until major shifts occur.

Fire and Vegetation

Interactions ____________________

The species composition of a community is determined by the successive birth and death of the individual community members. How fast vegetation responds to changing climate depends on species life

histories, migration rates, and rates with which suitable regeneration gaps are created. Fire has played

and will continue to play a significant role in determining vegetation physiognomy, structure, and species

composition in the world’s temperate and boreal ecosystems (Agee 1993; Crutzen and Goldammer 1993;

DeBano and others 1998; Rundel 1982; Wright and

Bailey 1982).

Fire is a major cause of plant mortality. For example, fire preferentially kills trees of short stature or

thin bark (fig. 8-4). Likewise, fire creates gaps that

new individuals colonize. Thus, changes in fire may

greatly accelerate vegetation’s response to changing

climate. Such interaction of climate, vegetation, and

fire has influenced the presence and rate of most

ecosystem processes in forest and rangeland settings

(Heinselman 1981). For example, fire return intervals

influence the distribution of life forms and regeneration modes present on a site (Noble and Slayter 1980).

The composition and structural integrity of some ecosystems are so strongly influenced by the fire regime

that they are considered to be “fire dependent” (DeBano

and others 1998; Habeck and Mutch 1973; Turner and

Romme 1994; Wright and Bailey 1982). The severity of

fire, which depends on the amount and type of biomass

present and weather conditions at the time of the fire,

exerts a strong influence on plant survivorship and

regeneration (fig. 8-5). Therefore, altered fire regimes

under a future of global climate change can be expected

to accelerate vegetation changes on the landscape

(King and Neilson 1992; Overpeck and others 1991;

Figure 8-4—Variation in predicted stem mortality as a

function of bark thickness and fire duration (adapted from

Peterson and Ryan 1986). Area above a bark thickness

curve implies cambium death, area below implies survival.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 181

Chapter 8: Global Change and Wildland Fire Ryan

Romme and Turner 1991; K. Ryan 1991; Weber and

Flannigan 1997).

Fire exerts selective pressure both at the individual

plant and community level (Noble and Slatyer 1980;

Rowe 1983). Short fire cycles favor species that endure

fire by juvenile sprouting or evade fire by storing seed

in the soil, invading from offsite, and having short life

cycles. Intermediate fire cycles favor species that resist fires when mature or evade fire by storing seed in

the canopy, but sprouting and invasion by offsite

colonizers also occur. Long fire cycles favor species

that typically avoid fire. Such species exhibit low

resistance to fire injury and regenerate predominantly

by seed. If the fire return interval is reduced to a period

less than the time to sexual maturity, then a species

Figure 8-5—Postfire vegetation recovery varies with fire severity (adapted from

Ryan and Noste 1985). In this concept of fire severity, the y-axis represents heat

pulse above the fire and the x-axis represents the heat pulse down into the soil.

Table 8-3—Fire interactions with vegetation.

Vegetation influences on fire Fire influences on Vegetation

Biomass (“loading” mass/area) Survival (resistance to fire injury)

Bulk Density (mass/volume) Regeneration (seeding vs. sprouting)

Size Distribution (surface area/volume) Injury (stress and loss of vigor)

Chemistry (volatiles vs. nonvolatiles) Competition (light, water, nutrients)

Live vs. dead ratio Community dynamics

Shading/exposure Structural composition

Strata (surface vs. overstory)

Continuity (horizontal and vertical)

will no longer be able to complete its life cycle on the

site and could be lost from the site. The ensuing rate of

reseeding will depend largely on the size of the area

burned and the mobility of the seed.

The quantity, chemistry, and size distribution of

fuels will change as species, growth patterns, and

decomposition change (table 8-3). For example, high

temperatures, drought, and nutrient shortages may

lead to stress-induced mortality (King and Neilson

1992; Waring 1987) and early leaf senescence, thereby

accelerating fuel accumulation. Also, an increase in

carbon to nitrogen ratios will reduce decomposition

(Agren and others 1991) and modify the role of fire as

an agent of decomposition and nutrient cycling (Gosz

1981; Rundel 1982).
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Ryan Chapter 8: Global Change and Wildland Fire

Uncertainty of Interactions: Can We

Predict the Future? ______________

Interactions between climate, vegetation, and fire

are complex and uncertain; thus, expectations for fire

management are general and tentative. We can hypothesize how change in one factor will change another, but in actuality several climatic forcing factors

will change simultaneously and initiate many internal adjustments within individual plants and communities. The relative abundance of species may shift

because some are less adapted to the climate-altered

site. Some species may regenerate but will be unable

to successfully complete their life cycle given new

climate and fire regimes. For example, redstem

ceanothus, a valuable wildlife forage species in the

Northern Rocky Mountains, and similar species that

rely heavily on seed stored in the soil, sometimes for

centuries, could be eliminated from sites by regeneration failure resulting from new climatic extremes,

particularly early season drought and severe fire. If

changes in climate and fire regimes lead to extensive

species losses on a site, then migration of species from

offsite will be accelerated. Species with wide ecological

amplitude should be favored over those with narrow,

specific habitat requirements. Regeneration strategies best suited to unstable conditions should also be

favored. The additional environmental stresses and

the increased frequency and severity of disturbance

will likely favor the expansion of exotic and invasive

species (Baskin 1998; Hogenbirk and Wein 1991).

Given changes in climate, soils, nutrients, and fire,

many endemic populations will not be able to compete

and successfully complete their life cycles on their

current sites. They will become locally rare or extinct

unless they are able to colonize new areas. Some

species, particularly those that predominantly reproduce vegetatively or from seeds stored in the soil, are

not highly mobile. While they may regenerate prolifically following site disturbance, they are less likely to

take advantage of climatic-induced disturbance off

site. These species should be slow to migrate to new

areas that are within their ecological amplitude.

Given the altitudinal shift in life zones, numerous

alpine species will become locally rare or endangered

because there is no higher zone into which they can

migrate (Franklin and others 1991; Peters 1990; Romme

and Turner 1991). Similarly, subalpine species such as

whitebark pine will be lost from all but the higher

mountain ranges. Poor soil development will retard the

migration of subalpine species into the former alpine

zone, but montane species should migrate freely to

higher elevations. If high temperatures and moisture

stress severely limit productivity, they could threaten

the continued existence of low elevation forests (IPCC

1996b). The advance of dry woodland and steppe species

into these forests may be slowed by their lack of shade

tolerance, but they should invade readily on sites disturbed by fire. Increased temperature and drought can

be expected to increase the decomposition of peat soils

and increase their susceptibility to fire (Hungerford

and others 1995). The current boundaries between

temperate and boreal forests, and between boreal

forests and tundra, are expected to shift northward

but not necessarily at the same rate.

In general, climatic change may be expected to

result in improved habitat conditions at the coolerwetter limits of a species’ range and poorer conditions

at its warmer-drier limits. However, many communities exist as “habitat islands” isolated by ridges or

valleys or surrounded by cultivation and urban areas

(Peters 1990; Peters and Darling 1985). These form

effective barriers against species migrations. The rate

of climatic change may be much more rapid than

species’ ability to migrate (Davis 1990; Gates 1990).

Understanding the potential impacts of climate

change on vegetation and fire will require a level of

integration not previously attempted in ecosystem

studies (Mooney and others 1991). Several authors

(Agren and others 1991; Franklin and others 1991;

Keane and others 1997; Neilson 1993; Overpeck and

others 1991) have attempted to understand the complex interactions by using process-based computer

models to simulate long-term ecosystem changes in

response to changes in climate. Keane and others

(1997) provided the most comprehensive treatment of

fire and climatic interactions on biogeochemical cycling. They used the Fire-BGC and FARSITE (Finney

1998) models to simulate changes in stand structure,

species composition, and water and gas exchange over

a 250-year time span in Glacier National Park, Montana. For model comparisons they simulated four fire

management scenarios: (1) current existing climate

and complete fire exclusion, (2) current existing climate and recent historical fire frequencies, (3) future

expected climate and complete fire exclusion, and (4)

future expected climate and expected future fire frequencies. Their results indicate that, because fire

tends to maintain younger forests and younger forests

have lower respiration, the Glacier National Park

landscape respires less carbon to the atmosphere with

periodic fires, even after taking fire emissions into

account (table 8-4). Smoke emissions nearly doubled

in the future climate/fire scenario (4), but these fluxes

were small relative to those from autotrophic and

heterotrophic respiration in unburned forests (scenario

3). Future climate was predicted to result in more

frequent and severe fires.

These results are for one ecosystem, and results are

likely different for other ecosystems, especially where

fire has not played such a strong historic role in

vegetation development. However, the prediction

of reduced atmospheric flux of greenhouse gasses
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 183

Chapter 8: Global Change and Wildland Fire Ryan

Table 8-4—Annual carbon flux (thousand tons C/year) on the McDonald and St. Mary drainages, Glacier National

Park landscape averaged across the 250 year simulation period. Table adapted from Keane and others

(1997).

Historical Future

No fires, fires, No fires, fires,

current current future future

climate climate climate climate

Carbon sourcesa (1) (2) (3) (4)

Heterotrophic respiration (HR) 820 768 942 810

Autotrophic respiration (AR) 1,168 1,087 1,466 1,128

Total respiration (TR=HR+AR) 1,989 1,855 2,409 1,938

Total fire emissions 0 15 0 24

Total carbon emissions 1,989 1,871 2,409 1,962

a

Carbon, expressed in units of 1,000 tons/year, can be converted to Gg/year if multiplied by 0.9072. Gg is a gigagram (109 grams).

associated with periodic fire illustrates that, because

of the complex interactions among ecosystem functions, ecosystem responses may be counterintuitive.

Considerable uncertainty still exists as to how far and

how fast climate will change. The autecology of many

species is poorly known so it is not possible to make

quantitative determinations of how they will respond.

Because future climate and vegetation are uncertain, it

is not possible to quantify changes in fire potential.

Considerable research is needed before we can confidently predict the magnitude of climate change, its

effects on vegetation and fire, and feedbacks to the

climate system. Given the complexity of the problem,

it is unreasonable to expect significantly better

information in the near future. Given the magnitude of

potential implications to fire management, long-range

planning should recognize the need for greater resources

in fire management (K. Ryan 1991; Stocks 1993).

Global change is a fundamental fact that natural

resource managers must face. The direction and magnitude of climate change over the next few generations

are uncertain, particularly at the regional level. But

the continued changes in land use are likely to affect

fire management regardless of the degree of climate

change. Given that weather patterns and atmospheric

chemistry are likely to change, and given the introduction of exotic species, management activities based on

the goal of restoring the historic range of variation

may not succeed (Millar 1997). Active manipulation of

wildlands and their disturbance regimes may be necessary to try to maintain the continued presence of

numerous species (Peters 1990; Sampson 1997).
184 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Notes

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USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 185

James K. Brown

Chapter 9:

Ecological Principles,

Shifting Fire Regimes and

Management Considerations

This chapter presents a broader, more fundamental

view of the ecological principles and shifting fire regimes described in the previous chapters that have

important implications for ecosystem management.

Also included are strategies and approaches for managing fire in an ecosystem management context and

sources of technical knowledge that can assist in this

process. Research needs are also described. The ecological fundamentals that underlie the effects of fire

on flora and fuels can be described under four broad

principles:

1. Fire will occur with irregular pattern depending

on climate.

2. Diversity of species and vegetation pattern depends on fire diversity.

3. Fire initiates and influences ecological processes

such as regeneration, growth and mortality, decomposition, nutrient fluxes, hydrology, and wildlife activity.

4. Humans exert a commanding influence on ecosystems by igniting and suppressing fire.

Ecological Principles ____________

Fire Recurrence

Fire as a disturbance process on wildlands has

occurred as long as vegetation has been present on

earth. The history of fire can be traced through charcoal fragments back to the Paleozoic Era, several

hundred million years ago (Agee 1993). Lightning that

can start fires occurs at a mind boggling rate. Approximately 8 million strikes per day occur globally (Pyne

1982). Human ignitions were common historically and

continue to be common today. Wildland fires will

continue to happen; the important questions about

fire occurrence are when, where, and of what severity?

The frequency of historical fire varied widely across

North America depending on climate. Fire return

intervals typically ranged from 2 to 5 years in ecosystems supporting abundant cured or dead fine fuels

such as the Southern pines, Southwestern ponderosa

pine, and oak savanna. They ranged from 5 to 35 years
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Brown Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations

for dry site conifers, shrublands including California

chaparral, and most grasslands; 35 to 200 years for

mesic site Western and Northern conifers; 200 to 500

years for some Eastern hardwoods and wetter site

conifers; and 500 to 1,000 years for extremely cold or

wet ecosystems such as alpine tundra and Northwestern coastal spruce-hemlock forests.

Our knowledge of fire frequency is largely based on

tree ring analyses and postfire stand ages, which only

allow a glimpse of fire history over the past several

hundred years—a rather short climatic period. Nonetheless, it provides a basis for understanding the

recurrence of fire that can be useful in planning. Keep

in mind that climate could indeed change and in turn

influence the occurrence of fire and the nature of

vegetation response.

Historically, fires have occurred at irregular intervals, largely determined by climate. Dendroclimatological studies in western Canada (Johnson and

Larsen 1991) and the United States (Swetnam 1993)

have shown that climatic cycles within cycles sometimes influence fire frequency. For example, in giant

sequoia forests, precipitation was the most important

influence on fire occurrence over periods of years such

as the recurrent episodes of the climatic phenomena El

Nino and La Nina (Swetnam and Betancourt 1990).

However, temperature was the most important influence on fire frequency over periods of decades to

centuries. In both cases fuel moisture content was

probably the important fuel property most influenced

by climatic trends in precipitation and temperature. A

study of presettlement fire frequency regimes of the

United States (Frost 1998) suggests that patterns of

fire recurrence, termed “fire periodicity,” can be considered as regular or irregular. For fire regimes having high fire frequencies (average fire-return intervals

of 0 to 10 years), individual fire occurrences were

considered nonrandom because they clustered around

a mean fire frequency. For fire frequencies greater

than 10 years, individual fires occurred irregularly or

in a random pattern.

Biodiversity

Biodiversity is broadly defined as the variety of life

and associated ecological processes that occur in an

area. This variety is sometimes broken down into

genetic, species, and ecosystem components (Salwasser

1990). In dealing with vegetation, it is convenient to

think of the spectrum of components as being plant,

community, and landscape. The landscape can be

viewed as a mosaic of patches, which are plant communities typically described as vegetation types, successional stages, stands, and age classes.

Fire regime types influence biodiversity in various

ways (Duchesne 1994). In forest ecosystems, understory fire regimes have the greatest influence on

biodiversity within plant communities because the

understory vegetation is more affected by fire than the

overstory. Stand-replacement fire regimes substantially influence biodiversity across the landscape by

affecting the size, shape, and distribution of patches.

Mixed fire regimes probably have the most influence

on biodiversity within plant communities, but also

affect patch characteristics or between community

diversity. In grassland ecosystems, fire frequency and

seasonal timing largely determine biodiversity.

Biodiversity can be increased by fire in many ecosystems and reduced by eliminating fire (Keane and

others, in press). Variability of fire regimes in time and

space creates the most diverse complexes of species.

Thus, landscapes having fires with high variability in

timing, intensity, pattern, and frequency tend to have

the greatest diversity in ecosystem components

(Swanson and others 1990). The phrase “pyrodiversity

promotes biodiversity” coined by Martin and Sapsis

(1992) aptly summarizes this concept. However, biodiversity can be reduced when fires occur much more

frequently than happened under the historical fire

regime. An understanding of the underlying relationships provides a basis for managing fire to meet

conservation of biodiversity goals.

Plant Response to Fire

Chapter 2 explained the many adaptive traits that

allow plant species to survive fire. In fact, many

species depend on fire to continue their existence.

Traits such as thick bark, fire resistant foliage, and

adventitious buds allow plants to survive low to moderate intensity fires of relatively short duration. Traits

such as fire stimulated germination, belowground

sprouting parts, and serotinous cones allow plants to

reproduce following high severity fires. For any particular plant to survive and persist, its adaptive traits

must be compatible with characteristics of the fire and

the timing of its occurrence. Fires can vary in intensity, duration, severity, seasonal timing, and frequency.

Other factors, especially weather and animal impacts,

can greatly affect whether a species can reproduce and

continue its existence following fire. Grazing by ungulates can influence postfire successional patterns and

flammability of future fires (Smith 2000).

Fire severity and intensity have a large influence on

composition and structure of the initial plant community following fire. Fire intensity mostly influences

survival of aboveground vegetation. Fire severity accounts for both upward and downward heat fluxes;

thus, it is a better indicator of initial postfire flora and

other fire effects. For example, when moisture contents of the forest floor fuels are high, a surface fire

may burn at high intensity yet not damage sprouting

tissues in the duff layer and mineral soil. Conversely,

under low forest floor moisture contents, a surface fire
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Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations Brown

may burn at low to moderate intensity yet consume

the forest floor and damage many sprouting tissues.

As a general rule, burned areas tend to return to the

same flora that was there before fire (Christensen

1985; Lyon and Stickney 1976). However, fires of high

severity create opportunities for new plants to establish from offsite seed. Large, high severity burns can

be slow to recover depending on available seed sources.

Fires of low severity are followed by a strong sprouting

response except where annuals are the dominant

vegetation.

The timing of fire including both seasonality and

frequency is crucial to managing for conservation of

biodiversity. This aspect of fire management can be

easily overlooked because of emphasis on controlling

fire and meeting air quality constraints. Seasonal

timing of fire is important because it largely determines fire severity and related mortality. It particularly affects reproduction of herbaceous plants and

shrubs. For example, in some ecosystems spring and

summer fire may produce abundant postfire flowering

while late summer and fall fires may produce little.

Perennials in Texas survive spring fire, but annuals

are harmed if fire occurs before seed is produced

(Chandler and others 1983). Evidence suggests that to

maintain long-term (decades) diversity in a tall grass

ecosystem, fire should be applied at different times of

the year to achieve successful seedling establishment

and productivity for a variety of plants (Bragg 1991).

Fire frequency is a particularly important consideration in short fire return-interval regime types because a period of several years to perhaps a decade can

be critical for survival of some species. Frequent fire

regimes that allow control of shrubs are critical to

maintaining grassland ecosystems (Wright and Bailey

1982). Many rare and threatened species have declined with reduction of fire frequency (see Greenlee

1997). Some fire dependent species in the Southeastern United States seem to require a 1 to 3 year fire

return-interval (Frost 1995). In contrast, local species

extinctions can occur with fires that occur too frequently, although it is generally accepted that locally

rare plants have greater chances of surviving on landscapes having diverse vegetation communities and

structure created by diverse disturbance histories

(Gill and Bradstock 1995). A problem today is that

plants adapted to short fire return-intervals can be

harmed by fires burning with high intensity and

severity in accumulated fuels that resulted from long

fire-free periods (Sheppard and Farnsworth 1997).

Community and Landscape Responses to Fire

Species diversity within a vegetation community

such as a stand or a patch depends on the collection of

species in the community, their adaptive traits, the

timing of fire, and the nature of fire as it moves

through the community. The spatial arrangement of

fuels and individual plants can be important to survival, particularly where fuels are unevenly distributed. Variable fire weather can also influence survival. Concentrations of live or dead fuels can generate

much greater fire intensities and severities on relatively small sites. This could enhance or reduce diversity depending on the community. For example, in a

Douglas-fir forest, localized fuel concentrations may

result in fire-created gaps or holes in the canopy. This

would create structural diversity and stimulate understory vegetation, a typical response to fire in a

mixed fire regime (fig. 9-1). However, in a ponderosa

pine forest, excessive mortality to highly valued old

growth trees could be a consequence.

Ecosystems and plant communities are considered

to be fire dependent when their continued existence

depends on recurrent fire. Where fires occur regularly

and frequently, such as in African savannas, open pine

communities, and Mediterranean shrublands, they

may remain stable for millennia (Chandler and others

1983). Repeated fires in fire-dependent communities

maintain a dynamic process that creates diversity

across the landscape, but if fire is excluded, biodiversity would probably diminish (Chang 1996). It has

been argued that fire-dependent communities have

evolved flammable characteristics that help ensure

repeated fires and the cycle of renewal (Mutch 1970).

However, the evolutionary argument remains unsettled (Chang 1996, Christensen 1993b).

Stand-replacement and to some extent mixed regime fires create patches on the landscape of differing

dominant vegetation and stand structures (fig. 9-2).

Patches can vary greatly in size and shape depending

on the biophysical features of the landscape and fire

behavior. Winds of variable speed and direction can

cause fire behavior to create a variety of fire shapes.

Terrain and landforms, rather than other fire influences, primarily determine patch dynamics in heavily

dissected landscapes (Keane and others, in press). For

example, fires in the nonmountainous boreal forests

were typically large (often well over 10,000 acres) but

medium to large (100 to 10,000 acres) in conifer forests

of western mountains (Heinselman 1981). Even in

large fires in mountainous terrain, fire severity can

vary considerably within the burn, leaving a patchy

distribution of fire effects (Turner and Romme 1994).

Generally, on landscapes characterized by large standreplacement fires, the pattern is naturally coarse

grained. On landscapes supporting smaller standreplacement fires, the pattern is finer grained. On

landscapes having understory fire regimes, occasional

trees are killed, creating gaps. This leaves a fine

grained pattern in the overstory such that the notion

of patches is not as helpful for describing landscape
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Brown Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations

Figure 9-1—A mixed severity fire burned through this Douglas-fir stand in Yellowstone National Park killing

about half of the trees leaving gaps and large openings in the canopy.

Figure 9-2—Stand-replacement fire sustained during low wind speeds by burning in heavy accumulations of

dead surface fuels, Yellowstone National Park.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 189

Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations Brown

diversity. In these fire regimes, considerable structural diversity can exist within communities.

As time since last fire increases, succession advances all stands to similar communities gradually

reducing structural diversity (Keane and others, in

press). Extending fire-free periods also increases the

likelihood of larger fires, hence larger patches and less

patch diversity (Bonnickson and Stone 1982;

Heinselman 1981; Swetnam 1993). In whitebark pine

forests, Murray (1996) found that lack of fire created

high elevation landscapes with high mean patch size

and low diversity. Romme (1982) found that fire control policies tended to reduce landscape richness and

patchiness and increase evenness in Yellowstone National Park, although in some situations, exclusion of

fire actually increased landscape diversity. Knowledge of fire regimes can help managers choose alternative land practices involving fire that favor landscape

diversity compatible with natural ecosystems.

Ecological Processes

Fire is an ecological process that triggers an amazing network of other processes and associated conditions. To explain this network, it can be helpful to

categorize fire effects into first and second orders.

First order effects are the immediate actions of fire

and include plant mortality, consumption of organic

material, creation of smoke, and changes to the physical-chemical environment. Second order effects are

many and depend on the nature of first order effects

and the postfire environment, especially soil, weather,

and animal activity. For example, here is a partial list

of second order effects:

1. Change in microclimate

2. Increase in range of soil temperatures

3. Change in soil nutrients and microbial activity

4. Regeneration of vegetation

5. Succession and new vegetation patterns

6. Change in plant growth rates and competitive

interactions

7. Altered wildlife habitat and activity of invertebrates and vertebrates

8. Changed water storage capacity and pattern of

runoff

Plant mortality, regeneration, and growth are fire

effects of obvious importance to land managers because they determine the characteristics of flora and

fuel that are readily observable as succession proceeds. Less apparent but nonetheless important, especially to the pattern of fuel change, is the decomposition process that involves fire, insects, and pathogens

in varying roles.

Successional Pathways

The classical concept of succession was based on the

perception that plant communities evolved over time

toward a final climax state that remained stable indefinitely. However, modern ecologists have rejected

this concept and now view succession as a dynamic

process that can move in alternative directions under

the influence of periodic disturbance and never reach

a stable end point (Christensen 1988). A useful method

of portraying succession utilizes the multiple pathways approach (Connell and Slayter 1977; Kessell and

Fischer 1981) where successional classes or stages are

linked along pathways converging to one or several

somewhat stable late-successional community types.

Successional classes are described by vegetation type

and structural stage. The number of succession classes,

pathways, and time steps between classes can vary

depending upon knowledge and the application. This

approach allows fire of varying severity and other

disturbances such as grazing and silvicultural cuttings to be incorporated in the conceptualization of

successional processes.

Time is a key element in understanding succession

(Wright and Heinselman 1973) and explaining it to

others. Some plant communities such as mesic and

wet site grasslands regain their former composition

and structure within only 1 or 2 years after disturbance (fig. 9-3). For other ecosystems, some compositional change may continue to occur long into the

future. Forest and shrubland communities vary greatly

in the time necessary for recovery to a mature condition. In understory fire regimes, vegetation usually

recovers rapidly. Structural changes are small or finegrained and may not be readily apparent. In standreplacement fire regimes, a young forest condition

may appear within 20 or so years. But it could take

several times longer in large severe burns where tree

seed sources are limited.

Decomposition

Fire, insects, and pathogens are responsible for the

decomposition of dead organic matter and the recycling of nutrients (Olson 1963; Stoszek 1988). Fire

directly recycles the carbon of living and dead vegetation. The relative importance of fire and biological

decomposition depends on site and climate (Harvey

1994). In cold or dry environments biological decay is

limited, which allows accumulation of plant debris.

Fire plays a major role in recycling organic matter in

these environments. Without fire in these ecosystems,

nutrients are tied up in dead woody vegetation. In

forests, tree density and understories thicken causing

increased competition and moisture stress. In turn,
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Brown Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations

this increases the likelihood of mortality from insects

and diseases leading to increased dead fuels, higher

intensity fires, and possibly volatilization of more

nutrients. In grassland ecosystems where both fire

and grazing are excluded, thatch or dead herbaceous

litter accumulates, which depresses herbage yields

and the number of plant species (Wright and Bailey

1982). Fire can help control encroaching shrubs and

trees; increase herbage yield, utilization of coarse

grasses, and availability of forage; and improve habitat for some wildlife species.

Fire both creates and consumes fuel. It increases

available fuel by killing shrubs and trees, which leads

to falldown of dead material into the surface fuel complex. Moisture contents of dead fuels average much

lower than live fuels, which also increases fuel availability. Insects and diseases perform similar roles.

They both kill vegetation, which creates available fuel,

and decompose organic matter. Fire in some circumstances enhances the opportunity for insect and disease

attack. For example, bark beetles may overwhelm fireinjured conifers, and wood rotting organisms may invade fire-scarred deciduous trees. A complex interaction that is not well understood exists between insects

and disease organisms, fire, and the environment.

However, we do know that fire, insects, and pathogens

evolved together as vital components of ecosystems.

Fuel Accumulation

Fuel accumulation is a term often used loosely to

indicate an increasing potential for fire to start, spread,

and intensify as the time since the last fire increases.

Generally, in ecosystems where annual biomass increment exceeds decay, total vegetative biomass increases

steadily with time because photosynthesis is an ongoing process. Fuels accumulate but not necessarily in a

steady fashion (Brown 1985a). On forested sites much

of the annual biomass increment is tied up in live tree

boles where it is unavailable for combustion. In grasslands and forests having short fire intervals, fuels

increase regularly over time as biomass increases.

However, in medium to long fire interval conifer forests, available fuel, and fire potential may decrease as

a postfire stand develops, then increase as the stand

becomes old and overmature (Brown and See 1981).

Fuel accumulation and associated fire potential depend on fuel quantity as well as other important fuel

properties such as compactness and continuity (vertical and horizontal). To be useful for estimating fire

behavior, fuel quantity must be expressed by size

classes for live and dead components. In a given

vegetation type, fuel quantity, size distribution, deadto-live ratio, and continuity are the important properties that change as succession progresses. Generally,

Figure 9-3—One year after a prescribed fire in a mountain big sagebrush community, this mesic site recovered

to domination by perennial grasses and forbs, Caribou National Forest, Idaho.
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Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations Brown

fuel quantities accumulate to greater levels on the

more productive sites in grassland, shrubland, and

forest ecosystems (Brown and See 1981; Wright and

Bailey 1982). In forest ecosystems much of the dead

fuel exists as coarse woody debris, which includes

pieces larger than 3 inches in diameter and sometimes

larger than 1 inch diameter (Harmon and others

1986). The more productive sites grow larger trees,

which eventually become coarse woody debris. An

important consideration in management of temperate

ecosystems is that coarse woody debris be recognized

for the many roles it plays. It contributes to biodiversity by being part of the life cycle of macroinvertebrates,

soil mites, insects, reptiles, amphibians, birds, and

mammals (McMinn and Crossley 1996). It is a source

of nutrients, habitat for terrestrial and aquatic life,

and fuel for wildfire (Harmon and others 1986). As a

fuel its most significant feature is that it becomes

rotten wood, which prolongs burnout and allows fire to

persist on site for long periods. Historically, large fires

occurred because fire remained smoldering in rotten

wood and duff for extended periods until low fuel

moistures combined with high wind speeds to support

intense, fast spreading fires.

Flammability increases as dead-to-live ratios increase.

As fuels accumulate through growth and mortality of

plants, flammability thresholds may be reached that

allow fires to increase greatly in intensity. Surface fires

become crown fires in conifer forests, and shrub communities burn intensely as a single fuel complex.

Fuel continuity is important because it partly controls where a fire can go and how fast it travels. In

grasslands and open shrublands, heavily grazed areas

and areas of low productivity form discontinuous fuels

that limit spread of fire, which can be a critical obstacle

to use of prescribed fire. In forests, existence of ladder

fuels from understory vegetation allows surface fires

to reach into the crown canopy. If the canopy is mostly

closed, crown fire can readily develop under adequate

wind speeds. Open canopies do not support crown

fires. Increased fuel continuity can account for changes

in fire severity from understory to mixed and from

mixed to stand-replacement. Many options are available to land managers for altering fuel continuity

through manipulation of vegetation.

Effects of fire on fuel arise basically two ways: first,

reducing fuel through consumption, and second, increasing fuel by killing vegetation. Both processes

affect several properties of fuel and fire potential.

Initially dead surface fuel loadings are reduced, also

lowering the dead-to-live ratio. If substantial amounts

of shrubs, small conifers, and limbs and foliage of larger

conifers are killed by fire but not consumed, they will

contribute to surface fuels in the years ahead as they

accumulate on the ground. Fire greatly influences fuel

continuity by creating vertical and horizontal gaps

within and between surface fuels and crown fuels.

Accumulation in Forests—Live and dead fuels, as

well as small and large diameter fuels, can follow

different patterns of accumulation. Typically, live herbaceous and shrub fuels increase following fire during

early stages of stand development. Then as tree canopies close, live herbaceous and shrub fuel quantities

tend to decrease on mesic sites ( Habeck 1976; Lyon and

Stickney 1976). However, a decrease in biomass may

not occur where understories contain shade tolerant

species. Fine dead fuels from foliage, bark flakes, twigs,

and cured herbaceous vegetation become incorporated

in the forest floor. Once crown canopies close, the

amount of litter fuel remains fairly constant as newly

fallen litter is offset by older litter moving into the duff

layer. Duff quantities continue to increase for some

time until equilibrium with decay is reached. This

period varies widely from approximately 5 years in

Southeastern United States (McNab and others 1978)

to well over a hundred years in some boreal ecosystems.

Dead branches and tree boles accumulate on the

ground in response to natural mortality and factors

causing downfall (Brown 1975). Mortality factors such

as fire, insects, disease, canopy suppression, and wind

and snow damage impact stands in a rather haphazard manner. Thus, accumulation of downed dead fuel

often occurs in an irregular pattern that is correlated

poorly with stand age (Brown and See 1981).

Conifer crown fuels increase regularly; however,

likelihood of crown fire may increase then decrease as

the lower canopy level grows further above surface

fuels. Eventually, crown fire potential increases again

when surface fuels increase and understory conifers

become ladder fuels. Shade tolerant species tend to

have more foliar biomass than intolerant species due

to their longer needle retention and higher crown

densities (Brown 1978; Keane and others 1999). Because of their shade tolerance they can fill in crown

canopy gaps and develop into understory ladder fuels.

Fuels critical to fire spread differ considerably between short and long fire interval fire regime types

(Brown 1985a). In short fire interval forests, fine fuels

such as grass, live shrubs, and needles create flammable understory fuels even in forests with vastly

different decomposition rates such as in longleaf pine

and ponderosa pine. The substantial quantity of fine

fuels coupled with long periods of suitable burning

conditions largely account for the understory fire regime. In long fire interval forests the forest floor and

accumulated coarse woody debris are critical fuels.

They burn with considerable heat release over a relatively long duration resulting in extensive mortality to

overstory trees. They ignite other surface and aerial

fuels and serve as excellent receptors of spotting embers that often allow fire to move in a leap frog fashion.

Fire intervals and environments differ considerably

between long fire interval types such as cedar-hemlock forests on warm moist sites and subalpine and
192 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Brown Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations

boreal forests on cold, dry sites. Nevertheless, in both

cases accumulated forest floor and downed woody

fuels support stand-replacement fire particularly during extended dry periods (Romme and Despain 1989).

Accumulation in Shrublands and Grasslands—

On many grasslands, grazing eliminates most of the

annual production so fuel accumulation is inconsequential. In the absence of grazing, fuel quantities

depend primarily on annual production, which varies

substantially by site potential and annual precipitation (Wright and Bailey 1982). Fuel loading may

increase for several years after a fire as some slow

responding grassland communities recover. Frequently, however, productivity is increased within 1 or

2 years following fire (Wright and Bailey 1982). Herbaceous litter accumulates in some grassland ecosystems but only marginally in others. Ratios of accumulated litter-to-current production typically range from

0.25 to 0.50 (Reinhardt and others 1997).

In shrub and shrub/grass ecosystems young communities generally have a low dead-to-live ratio. Flammability depends largely on grass and sedge fuels. As

shrubs become senescent or undergo mortality, dead

stemwood accumulates, which significantly increases

potential flammability. Dead fuel quantities tend to

increase with time since last fire or with age of plant

community as suggested for chaparral, however, not

in a uniform nor readily predictable fashion (Paysen

and Cohen 1996). Besides age, other factors such as

drought, winter kill, insects, and disease can cause

periodic dieback that creates substantial dead fuel

quantities. As cover and height of shrubs such as

sagebrush increase, fire intensity and rate of spread

potential increase markedly (Brown 1982).

Human Influences

People are part of ecosystems and certainly have

exerted a major, far reaching influence on fire across

the landscape. Indian burning was common throughout the United States and Canada. Pyne (1982) quotes

Henry Lewis as saying, “To simply note that all Indians used fire to modify their environments is no more

an ecological generalization than to note that all

farmers used plows.” The extent of Indian burning

varied considerably, however, depending on locale and

population movements (Boyd 1999; Pyne 1982). Indian burning greatly extended grasslands especially

in the Eastern and Midwestern United States. Most of

the coastal plain from Massachusetts to Florida to

Texas was savanna. Western valleys and foothills

were maintained as grasslands and open forests (Gruell

1985).

Considerable debate exists about the relative importance of Native Americans and lightning in maintaining historical fire regimes ( Barrett and Arno 1982;

Frost 1998; Keane and others 1999). The relative

importance of Native American fires was probably

greater in topographically complex areas where fire

compartments were smaller and where lighting ignitions were infrequent (Frost 1998). Also debated is

whether anthropogenic burning should be considered

part of the native or natural fire regime (Arno 1985;

Kilgore 1985). Fires set by Indians were often of

different seasonality, frequency, and landscape pattern than those started by lightning (Frost 1998; Kay

1995). Indian and lightning-caused fire existed for

thousands of years, a short evolutionary period but a

long time for plant communities to adjust to fire

disturbance. This long period of fire on the landscape

argues strongly for accepting both sources of ignition

in considerations of Euro-American presettlement fire

history used to guide management of ecosystems.

Efforts to suppress fires were modest at first relying

on wet blankets and buckets around dwellings and

campsites (Pyne 1982). Modern suppression capabilities relying on sophisticated communications, rapid

attack, specialized equipment, and many fire fighters

are a far cry from the early 1900s. Fire protection has

succeeded in reducing the extent of fire and increasing

fire intervals. Chandler and others (1983) suggested

that as protection succeeds, fire intervals become

greater and flammability increases. Then, more protection is needed to keep burned acreage down. A given

protection effort and annual burned area will eventually reach equilibrium. Since the 1980s, the costs of

protection and greater understanding of the role of fire

have led to more hazard reduction and ecosystem

maintenance rather than just protection.

For the past 100 years or so, human use of fire—

earlier termed controlled burning and now prescribed

fire and wildland fire use—has met with considerable

controversy politically and within land management

organizations. “Light burning” (understory fire) was

once widely applied in the southern pines and ponderosa pine type especially in California. However, the

perceived threat to effective organized fire control

largely curtailed the program on publicly owned lands

(Pyne 1982). Some benefits of controlled burning were

still recognized, especially hazard reduction and preparation of seed beds for regeneration. In the West

justification for prescribed fire was fuel reduction,

namely slash burning. This single purpose use of

prescribed fire resulted in short-term successes but

long-term failure to optimize societal objectives for

forests (Agee 1993).

More recently, the concept of ecosystem management has led to a much wider understanding of the

ecological role of fire and its importance in the functioning of ecosystems. Concerns over air quality, control of fire, and costs, however, remain as major constraints on the application of prescribed fire and
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 193

Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations Brown

wildland fire use. The responsibility to see that fire is

properly managed as a component of the ecosystem is

now greater than ever because land managers have

the power to delay and exclude fire as well as an

understanding of fire’s important ecological role.

Shifting Fire Regimes ____________

Chapters 3 through 7 clearly show that fire regimes

have shifted from what they were historically across

most of the United States and southern Canada. In a

comprehensive assessment of burning in the contiguous United States, Leenhouts (1998) estimated that

approximately 10 times more area must be burned

than at present to restore historical fire regimes to

nonurban and nonagricultural lands. The greatest

departure from historical fire regimes is in the Rocky

Mountains where only a small fraction of the pre-1900

annual average fire acreage is being burned today

(Barrett and others 1997). Kilgore and Heinselman

(1990) estimated that the greatest detrimental effects

of fire exclusion were in short interval fire regimes of

the Rocky Mountains. In contrast, in long fire regimes,

the effects of fire protection have not had a significant

influence. In the Canadian and Alaskan boreal forest

limited protection due to remoteness has maintained

fire regimes essentially as they were historically.

Extensive grazing by domestic stock that reduces

fuels, and fragmentation by agriculture and human

developments, have also contributed to shifting fire

regimes. Lengthened fire return intervals have resulted in changes of minor to major consequence to

vegetation and fuels by increasing wildfire severity

and decreasing species and structural diversity. A

comparison of historical and current fire regimes in

the Interior Columbia River Basin of about 200 million

acres showed that fires have become more severe on 24

percent of the area (Morgan and others 1998) (see fig.

5-1 in chapter 5 of this volume). Fire severity was

unchanged on 61 percent of the area. Fires were less

frequent on 57 percent of the area, unchanged on 33

percent, and more frequent on 10 percent of the land

area. Fire protection, reduced fine fuels from grazing,

decreased fuel continuity from human development,

and in some cases exotic plants are the most probable

causes (Chang 1996; Keane and others 1999). Further

analyses of changes in fire regimes and condition

classes of vegetation are currently under way for the

United States (Hardy 1999).

Forests and Woodlands

Changes in forest composition and structure due to

shifting fire regimes have been widely documented.

Generally, shade-intolerant species are being replaced

with shade-tolerant species. Stand densities are

increasing with development of multiple layer canopies. Outbreaks of insects and occurrence of root diseases appear to be worsening (Stewart 1988). The

greatest impacts have occurred in the understory fire

regime types typified by ponderosa pine and longleaf

pine ecosystems (fig. 9-4). Although these two ecosystems experience widely different climates, they share

the same end results of fire exclusion made worse in

some locations by selective harvesting of old growth

trees. Where fire regimes have shifted, growth and

vigor of trees is reduced, insect and disease mortality

is increased, and understory fuel loadings and continuity increased so that wildfires tend to be of high

intensity, killing most or all of the overstory pine.

Diversity of understory herbs and shrubs is decreased.

The loss or depletion of the pyrophytic herb layer is

considered to be one of the unrecognized ecological

catastrophes of landscape history (Frost 1998). The

extent of the problem is greater in ponderosa pine

where relatively little prescribed fire has been applied. Although prescribed fire is widely applied in the

South it has largely been used only for rough (accumulated understory fuels) reduction during the dormant

season. Thus, lack of seasonal fire diversity in the

southern pine types has limited plant diversity.

In mixed fire regime types such as coastal and

inland Douglas-fir, whitebark pine, red pine, and

pinyon-juniper, the results of fire exclusion have created the same problems as found in understory fire

regimes. Mixed fire regimes are experiencing considerably less nonlethal understory fire than in the past

(Brown and others 1994). The mixed fire regime is

shifting toward a stand-replacement fire regime that

favors more shade tolerant species and less landscape

diversity.

In stand-replacement fire regimes, fire intervals

have generally lengthened; however, the effects of this

vary widely depending largely on presettlement fire

return intervals and accessibility for fire suppression

efforts. For example, in the lodgepole pine/subalpine

fir type, which dominates the Selway-Bitterroot Wilderness, presettlement stand-replacement fire was

1.5 times more prevalent than during the recent period (Brown and others 1994). The presettlement fire

return-interval was approximately 100 years. In the

same type in Yellowstone National Park, characterized by a fire return-interval of about 300 years, the

area burned probably has not differed between presettlement and recent periods (Romme and Despain

1989).

The age distribution of marginally commercial and

noncommercial forests such as those in wilderness

areas and parks is shifting to an abundance of older

stands (Brown and Arno 1991). Succession is increasing the shade tolerant component of stands, making

a major species shift likely if fire continues to be
194 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Brown Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations

excluded. In the case of western aspen more than half

of the type has been lost (Bartos 1998), much of it due

to successional replacement by conifers (Bartos and

others 1983). Fire protection policies have resulted in

the fire cycle in aspen shifting from about 100 years to

11,000 years; thus, if this degree of fire exclusion

continues, the loss in biodiversity will be considerable.

In jack pine forests the more shade tolerant balsam fir

is gradually assuming dominance aided by natural

deterioration and harvesting of jack pine.

Fuel accumulation patterns vary widely in coniferous stand-replacement fire regime types. Mature forests may support abundant or relatively little available fuel. However, as fire intervals are allowed to

increase and stands become over mature, downed

dead woody fuels and live ladder fuels from shade

tolerant understory conifers can be expected to dramatically increase. The result will still be standreplacement fire but at higher intensities, which will

tend to propagate larger fires in spite of suppression

efforts. This trend could lead to fewer but larger fires

burning during severe fire weather years, causing less

diversity in patch size and age (Keane and others

1999).

Grasslands and Shrublands

Grassland fire regimes have shifted dramatically

from the presettlement period. Many ecologists consider the reduced frequency and extent of fires on

rangelands due to fire protection to be among the most

pervasive influences in the United States by nonNative Americans (Pieper 1994). The shift to woody

plant domination has been substantial during the past

hundred years. Grazing and possibly climate changes

have acted with reduced fire to give a competitive

advantage to woody plant species. Some woody plants

such as honey mesquite become resistant to fire, develop fuel discontinuities, and reduce spread of fire. In

time, recovery following fire favors shrubs over perennials (Archer 1994). This can alter the composition of

ecosystems to the point that a return to the grassland

type becomes nearly impossible or impractical (Brown

1995).

Historically, fires were more frequent in Eastern

than in Western grasslands. High productivity of

biomass was maintained in the tallgrass prairie by

frequently occurring fire that recycled accumulated

thatch. A diverse composition was probably favored by

Figure 9-4—A stand-replacement fire supported by accumulated dead surface fuels and live ladder fuels from dense

understory trees occurred in this understory fire regime type killing the old growth ponderosa pine, Yosemite National Park.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 195

Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations Brown

variable frequency and seasonality of fires (Abrams

and Gibson 1991; Bragg 1991). Western grasslands

appear to have generally experienced fire less frequently (Gruell 1985; Wright and Bailey 1982) but still

frequently enough to hold back invasion of woody

plants.

Fire regimes have shifted to too much fire in the

drier portions of the sagebrush-steppe ecosystem that

occupies over 100 million acres in Western United

States. Fire frequency has increased in many areas

due to invasion of cheatgrass and medusahead, introduced annuals that cure early and remain flammable

during a long fire season. Increased fire frequency

exerts strong selective pressure against many native

plants (Keane and others 1999). A contrasting situation exists for the more mesic mountain big sagebrush

type where decreased fire frequency and encroachment by conifers is causing a reduction in herbaceous

and shrub vegetation (fig. 9-5).

Managing Fire __________________

Fire is an integral component of ecosystems that can

affect all aspects of ecosystem management. Fire regimes have shifted as a result of human influences and

may continue to shift with clearly detrimental results

in some ecosystems. Land managers need to know how

to plan and carry out fire management strategies that

successfully incorporate the ecological role of fire.

Constraints on managing prescribed fire and smoke

make it difficult to achieve resource goals, while protection against wildland fires allows development of

undesirable ecological consequences (Brown and Arno

1991). Overcoming this predicament requires that

land managers and the public alike recognize the role

of fire in the functioning of ecosystems and in meeting

varied resource objectives.

Strategies and Approaches

Vegetation and fire management objectives should

be derived from broader ecosystem management goals

to achieve desirable fire effects. Determining objectives, and the strategies and approaches for achieving

them, can be simple to complex depending on land

ownership and direction provided by the owners. For

example, a small woodlot owner may simply want to

reduce fire hazard, in which case fuel reduction objectives can be clearly stated and, if appropriate, a prescribed fire conducted to reduce the unwanted fuel.

Where the direction is ecosystem management, a goal

recently adopted on many Federal and some State

lands (Salwasser 1994), a more elaborate process may

be required to determine objectives and strategies.

Figure 9-5—Without disturbance, this sagebrush/grass community being encroached by Douglas-fir will

eventually become a closed canopy forest with sparse understory vegetation, Deerlodge National Forest,

Montana.
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Brown Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations

To steer this process, a guiding principle or goal for

ecosystem management is to provide for conservation

of biodiversity and sustainability of ecosystem composition, structure, and processes (Kaufmann and others

1994). This involves molding a management plan

based on an understanding of ecosystem processes. An

element missing or minimally considered from many

past planning efforts was the landscape of varying

scales. For this a perspective is needed that involves

consideration of ecological processes across a hierarchy of land units (Hann and others 1993).

The setting of goals and objectives starts out broadly

with a goal specifying the future condition of the

ecosystem or a particular tract of land. This desired

future condition is a vision for the future and not an

objective for management action (Kaufmann and others 1994). An assessment of the ecosystem, resource

potentials, and needs of people is a prerequisite for

setting the desired future condition. From this, more

specific objectives can be derived for managing fire.

They should be specified in terms that can be monitored. Different approaches may be appropriate for

doing an assessment and setting the desired future

condition and the ensuing management objectives.

Consider the planning task by three types of land

use zones (Arno and Brown 1989):

• Zone I – wilderness and natural areas objectives call for allowing fire to play its natural role to

the greatest extent possible. Fire objectives may

vary depending on whether it is a wilderness or

natural area intended to preserve a particular

condition or process.

• Zone II – general forest and range management, where the need to provide resource values

means a wide range of vegetation and fire objectives will be appropriate.

• Zone III – residential wildlands, where the

natural role of fire will be constrained considerably and fuel management is the primary objective.

Two occasionally troublesome facets of setting goals

and objectives in Zones I and II that rely on knowledge

about the ecological role of fire involve the “historical

range of variability” and the goal orientation of “process versus structure.”

Historical Range of Variability

The historical range of variability (also called natural range of variability) in ecosystem components can

be used to help set desired future conditions and fire

management objectives. It can serve as a basis for

designing disturbance prescriptions at varying spatial scales and help establish reference points for

evaluating ecosystem management (Morgan and others 1994). Reference points to past functioning of

ecosystems can be interpreted from various sources

such as historical records, palynology, natural areas,

archival literature and photographs, GIS data layers,

and predictive models (Kaufmann and others 1994;

Morgan and others 1994). Historical fire regimes of

forest ecosystems are often characterized by determining age distribution and areal extent of seral

classes across a large landscape and dating fire scars

to determine fire return intervals. These techniques

provide a snapshot of ecosystem conditions that covers

the past 100 to 400 years. Pollen analysis can extend

this period but with less precision about disturbance

events (Swanson and others 1993). Estimation of historic fire frequencies in grasslands and shrublands is

more problematical because of a lack of fire scars and

easily determined age classes. It relies largely on

historical accounts of human activities.

To what extent should knowledge of the historical

range in variability be relied upon to help establish

goals and objectives? This depends largely on soundness of the ecological knowledge and other ecosystem

issues such as human needs and threatened and

endangered species (Myers 1997). A strong argument

can be made that knowledge of historical fire should be

used as a guide for understanding landscape patterns,

conditions, and dynamics, but not necessarily for creating historical landscapes. Knowledge of historical

variability provides a basis for bringing the range of

existing conditions in a landscape within the historical range (Swanson and others 1993).

A scientifically based rationale underlies the use of

historical variability as a guide for managing biodiversity. Native species evolved and adapted to natural

disturbance events over at least the past 10,000 years.

Numerous ecological studies emphasize the close dependence of species on disturbance regimes (Swanson

and others 1993). Genetic diversity (Frankel and Soule

1981) as well as landscape diversity are maintained

through disturbance regimes. Where fire regimes have

shifted markedly, species and landscape diversity

have declined.

Concerns and limitations to using historical variability as a guide to managing ecosystems (Morgan

and others 1994; Swanson and others 1993) are:

1. Difficulty interpreting past variability due to

insufficient data.

2. Degree to which past and future environmental

conditions may fall outside the established range

of historical conditions. For example, the possibility of future climate change due to global

warming is a significant concern.

3. Extent to which the range of ecosystem conditions desired by society differs from historical

variability.

The natural range of variability can be determined

and applied with reasonable confidence in high
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 197

Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations Brown

frequency fire regimes of forests. In understory fire

regimes, considerable data on fire frequency often can

be obtained by consulting published accounts or conducting studies of fire intervals on fire scarred trees.

Variability of fire-return intervals can be quantified

and compared with recent fire history to determine

whether a significant departure has occurred (Brown

1993). In long interval stand-replacement fire regimes

of some forests and tundra, estimates of the historical

range of variability are more difficult to establish with

certainty because of the limited number of disturbance events that can be studied. Perhaps the best

technique for measuring fire regime characteristics in

this situation utilizes satellite and GIS technologies to

map vegetation pattern (Morgan and others 1994), an

approach requiring considerable resources.

A question that often arises in interpreting fire

history especially concerning wilderness and other

natural areas is how Indian ignitions should be treated

(see Lotan and others 1985). The prevailing thought

seems to be that because Indian burning occurred over

a long period, ecosystems were adjusted to fire effects

from human and lightning ignitions combined and

this reflects historical fire regimes. Disturbance history can only be readily and reliably measured for the

past 200 to 400 years. Variability in climate, vegetation composition, and disturbance patterns has been

substantially greater over the past several thousand

years than over just the last 400 years. But land

managers need a consistent basis on which to plan,

and using measurable fire history is a practical approach. The concept of the historical range of variability can be valuable in understanding and illustrating

the dynamic nature of ecosystems and in evaluating

current ecosystem health.

Process Versus Structure Goals

Process and structural goal setting approaches are

important to management of Zone I lands. These

concepts originated with establishment of wildernesses

and natural areas where the goal was to manage for

naturalness. The proper role of fire in wilderness and

natural areas has been characterized in terms of

process-oriented and structure-oriented goals (Agee

and Huff 1986). Expressed simply, do we want a

natural fire regime (process) or rather the vegetation

that a natural regime would have created (structure)

(Van Wagner 1985)? The answer to this may always

involve some degree of debate because of philosophical

differences over the concept of natural (Kilgore 1985).

In practice, both approaches or a mixture of the two

may be appropriate depending on circumstances. Practical aspects such as costs, fire safety considerations,

and size and boundaries of the ecosystem will often

determine the most appropriate approach.

A strictly process-oriented goal is probably only

appropriate in large wilderness areas. The process

goal approach modified by practical considerations

will usually be necessary.

In understory fire regimes where surface fuels have

accumulated to the point that high intensity fire is

likely, a structure-oriented goal is the best approach to

ultimately achieve natural conditions. After fuels have

been reduced using a prescription for low severity fire

to avoid killing the overstory, a process goal of allowing natural ignitions can be followed if it will maintain

the understory fire regime (Bonnickson and Stone

1985). Structural goals will continue to find application in understory fire regime types to restore and

even maintain the natural role of fire. The structural

goal approach is probably the best for management of

threatened and endangered species. It may also be

more efficient and esthetically pleasing (Agee and

Huff 1986).

Mixed fire regime types in wilderness areas present

variable, complex landscape patterns that can make

structural goals difficult to achieve. Fire frequencies

in the mixed type typically range from 35 to 100 years.

In some localities fire has been absent long enough

that fuels and stand structures appear to be falling

outside the range of historical variability (Arno and

others 2000). In such cases, where accumulated surface fuels and naturally occurring ignitions would

favor stand-replacement fire, structural goals aimed

at retaining a portion of the overstory may be appropriate to restore the mixed fire regime. If excessive

fuels have not accumulated, process goals seem to be

the most reasonable.

Another consideration in wilderness areas, regardless of whether structural or process goals are chosen,

is when and where to use prescribed fire to meet

wilderness objectives. In the contiguous United States

75 percent of Congressionally classified Wilderness

areas, which occupy half of the classified wilderness land

area, are too small to maintain natural fire regimes by

relying strictly on natural ignitions (Brown 1993). Constraints such as concern over escape fire, lack of lightning-caused fires, conflicting wilderness goals, and air

quality regulations will require prescribed fire to restore

fire and mimic natural processes. Decisions to use prescribed fire must be ecologically based, but also with the

realization that exacting solutions to mimicking natural

fire processes are probably not feasible. Neither the

determination of fire history nor applications of prescribed fire are precise undertakings.

For residential and commercially zoned lands (Zones

I and II), structural goals are the most appropriate.

Clearly definable and measurable end points are being sought. For example, specific conditions such as

tree species and size, stand age distribution, patch

size, stimulation of shrubs, increased forage production,

and reduced fuel quantities may be desirable objectives.
198 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Brown Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations

Landscape Assessment

Managing biodiversity and for sustainability of ecosystem components and processes requires a landscape perspective. Small ecosystems are found within

larger ecosystems, individuals occur within communities, and short-term processes are nested within longer

term processes (Kaufmann and others 1994). The

various scales fit into a hierarchical structure that

determines patterns of diversity for an area (Bourgeron

and Jensen 1993). A major challenge to setting vegetation and fire objectives in the context of ecosystem

management is evaluating and interpreting the ecological significance of multiple scales. Vegetation scales

range from individual plants, communities, seral

stages, potential vegetation types, to the biome level.

Species and individual plant communities are dealt

with using a fine filter approach. Traditionally, assessments of fire effects and other environmental

impacts have been done on a project basis using fine

and mid scale evaluations. The coarse scale aspects of

ecosystems have been largely neglected. The coarse

filter approach, which deals with higher scale levels

such as aggregations of communities, can operate with

relatively little information, yet be an efficient way to

meet biodiversity goals ( Bourgeron and Jensen 1993;

Hunter 1990; Kaufmann and others 1994). A single

ecosystem can be too small to hold viable populations

of all its species, especially large predators. Thus, the

coarse filter approach is best used on assemblages of

ecosystems such as watersheds and mountain ranges.

Both approaches are necessary to evaluate all facets of

an ecosystem and meet the goals of ecosystem management (Hann and others 1993a).

Assessment of landscape and ecosystem properties

can be undertaken with varying degrees of sophistication and effort. Some of these planning efforts, which

are evolving through trial and error, are mentioned as

examples. During the past decade agencies such as the

U.S. Forest Service and Bureau of Land Management

have undertaken landscape analyses on extremely

large areas such as the 200 million acre Upper Columbia River Basin (Keane and others 1996) and smaller

areas such as the Pike and San Isabel National Forests and Cimarron and Comanche National Grasslands in Colorado (U.S. Forest Service 1997) and the

130,000 acre Elkhorn Mountains and 46,000 acre

North Flint Creek Range in Montana (O’Hara and

others 1993). Details of these landscape evaluations

varied but they followed three general steps (Hann

and others 1993b):

1. Characterize the general composition, structure,

and processes of the ecosystems and landscapes

within the designated analysis area.

2. Analyze data to assess changes in structure and

composition and relate the changes to previous

management treatments.

3. Examine the ecosystem processes important for

the area and their effects on ecosystem and

landscape composition, structure, and rate of

change.

Succession Modeling—Simulation of succession

provides a means of predicting the long-term interaction of processes such as fire, insects, disease, and

cutting of vegetation on landscapes of varying scale.

Simulation can be helpful to managers and the public

by helping them understand how ecosystems function

and for evaluating different management alternatives. The wider availability of powerful computer

capabilities has led to an increase in succession modeling efforts particularly for landscape applications.

Manager-oriented computer models that simulate

successional processes across large landscapes are

faced with a tradeoff between realistic portrayal of

ecological processes and utility of the model. Some

models are too complicated to use without special

training or assistance. Nonetheless, managers are

increasingly using succession models in their planning while models are continually evolving and computer capabilities growing.

In choosing a model for a particular application, it is

important for the temporal and spatial scales of the

model to match the intended use (Reinhardt and

others, in press). Models that operate over a period of

decades are useful for scheduling treatments. For

example, the Fire and Fuels Extension to Forest Vegetation Simulator (Beukema and others 1997) simulates fuel quantities, tree characteristics, and tree

mortality in the event of a fire for single stands.

Managers can use the model to help schedule thinnings

and fuel treatments when potential fire behavior and

fire effects on an area are deemed unacceptable

(Reinhardt and others, in press).

Models that simulate fire effects over centuries are

useful for providing targets for managers, for estimating the historic range of conditions, for evaluating

implications of climate change, and for understanding

possible long-term consequences of management actions. For example, CRBSUM was used to simulate

landscape changes for different management scenarios

in the Columbia River Basin (Keane and others 1996).

Some of the current models that have been applied to

assist land managers are summarized in appendix B.

Restoration of Fire

Restoration of fire is needed to varying extents in

most ecosystems of North America to meet the holistic

goals of ecosystem management. The need for restoration is most evident in high fire frequency regimes

such as understory fire regime types and some grasslands and shrublands where fire has been excluded for

several times longer than the average fire return

interval. Although considerable knowledge supports
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 199

Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations Brown

the need for restoration of fire into wildland ecosystems, constraints and obstacles confront land managers (Brown and Arno 1991; Mutch 1994). Limited

funding, air quality restrictions, concerns over escape

fire, and inadequate public support can pose difficulties. Some breakthroughs in managing emissions and

obtaining support have provided more latitude for

prescribed fire programs (Mutch and Cook 1996).

Successful restoration involves clearly stated objectives, plans based on scientific knowledge of fire’s role

in the ecosystem, and adaptive learning from prescribed fire efforts. Adaptive learning is important

because prescribed burning usually improves with

experience. Prescription conditions and firing techniques may need to be modified to achieve objectives

such as a given level of fuel reduction or to meet

constraints such as holding overstory mortality to

certain limits. Fire may not spread adequately under

an initial prescription, thus requiring lower fuel moisture contents or higher wind speeds to be successful.

Restoration of fire can be undertaken on an entire

ecosystem or on an individual plant community basis.

Ideally, restoration of individual plant communities

would be based on ecological considerations of the

broader ecosystem of which they are a part. The extent

of ecosystem assessment that is appropriate for planning restoration will depend largely on land ownership and direction given to management. For large

land ownerships, restoration of entire ecosystems or

large landscape areas is the soundest approach to

manage landscape pattern and meet biodiversity goals.

It also allows for effective placement of fuel treatments

designed to disrupt fuel continuity and reduce threat

of large fire occurrences. The steps undertaken by

Keane and Arno (1996) to restore fire in the whitebark

pine ecosystem may be useful in other situations

including grasslands and shrublands. They recommend first, an inventory of landscape and stand characteristics at multiple scales; then, writing descriptions of the important processes of the landscape and

stands. Landscapes and stands can then be prioritized

for restoration treatment and selected based on inventory, description, priority, and feasibility. Treatments

should be designed for each selected stand or landscape based on inventory and description information

and implemented as efficiently as possible. Finally,

treatments should be monitored to evaluate restoration success.

Restoration of fire in grasslands, shrub steppe, and

savannas requires careful consideration of seasonal

timing and frequency to assure that prescribed fires

will spread at appropriate severities. Once woody

plants have encroached to a point of dominating a site,

it becomes difficult to get fire to spread with sufficient

heat to kill aboveground stems such as oak in savannas (Huffman and Blanchard 1991) and juniper in

sagebrush/grass communities. Perhaps the greatest

obstacle to success lies with areas that have successionally lost the native mix of species and lack sufficient grass fuel to carry fire. Seeding of native species

following fire may be necessary to restore a resemblance of former plant composition. Where conifers

invade grasslands such as pinyon-juniper and inland

Douglas-fir (Gruell and others 1986), successful spread

of surface fire may require fuel enhancement work

such as cutting numerous trees to create adequate

surface fuels. Otherwise, crown fire may be required,

which will necessitate a more flammable, narrow fire

prescription that can limit burning opportunities.

Prescribed Fire and Silviculture

Prescribed fire and silviculture can go hand in hand

for restoration of forest stands and ecosystems. Some

consider prescribed fire to be a silvicultural technique

even though it goes far beyond the usual goals of

silviculture that are oriented to producing tree products and desirable forest stand structures. One debatable point is the extent to which it is desirable to have

management mimic the kinds of stands and landscape

structures that typified presettlement fire regimes.

However, an understanding of similarities between

characteristics of fire regime types and silvicultural

stand structures can be helpful for integrating fire

with silviculture to restore fire as a process and meet

ecosystem management goals. The following description of stand structure and silvicultural practices

based on a discussion by Weatherspoon (1996) applies

to individual stands. Stands can be treated differently

to manage landscape-level vegetation.

Even-Aged Stands—These stands originated naturally mostly from high-severity, stand-replacement

fires that killed most of the trees. Silvicultural methods that produce even-aged stands include clear-cutting, seed tree, and shelterwood cutting. Shelterwood

or seed trees are typically removed after regeneration

is secured. Pile burning or broadcast burning is commonly used to reduce fuels and prepare sites for

regeneration. Leaving snags, large downed woody

material, and untreated patches in larger treatment

units is important for meeting biodiversity goals.

Two-Storied Stands—These stands were associated with moderate to high severity fire typical of the

mixed fire regime type. Retention shelterwood (also

called irregular shelterwood or shelterwood without

removal) is the silvicultural method for treating the

stand. Prescribed underburning can often be practiced

to manage fuels and create within-stand diversity.

Once created, the stand would never be devoid of large

trees because each regeneration cutting would be

accompanied by retention of some overstory trees.

Snags could be readily created.
200 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Brown Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations

Uneven-Aged Stands with Even-Aged or EvenSized Groups—These were associated with low to

moderate severity fires associated with the understory fire regime type and perhaps to some extent with

the low severity end of the mixed fire regime type.

Silviculturally this stand structure is mimicked with

the group selection cutting method. Skillful prescribed

underburning is required to apply the proper severity

for maintaining this structure. Jackpot burning and

two-stage burning under different prescription conditions may be appropriate.

Uneven-Aged Stands with Fine Tree Mosaic—

These stands are characterized by three or more sizes

and ages of all tree species distributed rather uniformly throughout the stand. This stand type is thought

to have developed primarily with shade-tolerant conifers over long periods following stand-replacement

fire. It is incompatible with frequent fires. The individual tree selection method is used to maintain this

structure. This stand structure could be considered to

represent open stands of ponderosa pine and longleaf

pine. Ecologically, however, they fit better with the

previous category of even-aged groups.

Understory Fire Regime Type

Restoration of the understory fire regime type requires application of frequent, low intensity fire, which

has been excluded for excessive periods of time. Restoration approaches can vary considerably depending on

stand and fuel conditions. The objective generally is to

create more open stand structures consistent with

historical disturbance regimes. A wide range of stand

densities can be appropriate depending on site potential and silvicultural objectives. Various even-aged

and uneven-aged stand structures can be utilized.

Favoring the long needle pine component through

regeneration and retention of old growth trees is

frequently a high priority need. Often the major problem to overcome is excessive understory fuel accumulations particularly live ladder fuels, and buildup of

duff around the base of desirable leave trees. Another

consideration is burning to encourage the historical

understory vegetation diversity. This requires burning during the growing season, which is a departure

from the traditional application of prescribed fire

during the spring, fall, or winter dormant seasons.

Conducting the first prescribed fire after a prolonged period of no fire must be done cautiously to

avoid flare-ups in sapling thickets or rough that might

kill desirable trees. For ponderosa pine, thinning of

dense understories and piling and burning slash before conducting a prescribed underburn may be necessary to reduce flammability and remove competitor

species that might survive most prescribed fires (Fiedler

and others 1996). However, too much caution where

the understory consists of thick patches of fir will

result in inadequate fire. Some fuel augmentation by

cutting small fir can help carry the fire with adequate

intensity to kill the fir. A series of prescribed fires

aimed at gradually reducing the accumulated live and

dead fuels may be necessary to return stands to where

maintenance underburning is easily manageable

(Sackett and others 1996). The best approach to restoration must be determined on a case by base basis, but

it will usually require a combination of mechanical

treatments and prescribed fire repeated over a period

of years.

Mixed and Stand-Replacement Regimes

The mixed fire regime includes a wide range of stand

structures and landscape patterns that result from

highly variable fire severities. Individual fires may be

of either nonlethal understory or stand-replacement

severity, or a combination of both severities. Thus,

managers have considerable latitude in designing

prescribed fire and silvicultural activities (fig. 9-6).

Although little guidance based on past restoration

efforts exists, the best way to determine restoration

objectives is on a large landscape basis because of the

wide latitude in individual stand structures. The challenge is to provide a diversity of stand structures with

retention of snags and some coarse woody debris in

forest ecosystems and unburned patches in grasslands and shrublands. In wilderness and natural area

management where fires have not been previously

allowed, avoiding excessive stand-replacement due to

accumulated fuels may be important.

Stand-replacement fire severities can be created

from either severe surface fire or crown fire. Wildfires

over prolonged burning periods can leave large proportions of both severities as observed in lodgepole

pine (Brown and others 1994). High severity surface

fires may be more readily prescribed and achieved

than crown fires due to the higher risk and fewer

burning opportunities for prescribed crown fires. Ecological effects of severe surface fire and crown fire

differ. Crown fire consumes foliage that otherwise

would fall and protect the soil. It can kill seeds in

cones, redistribute nutrients in ash, and provide more

chance for regeneration by offsite colonizers. Where

silvicultural objectives are being pursued, an important consideration is avoidance of excessive fragmentation caused by intensive small-scale cutting and

prescribed fire activities. Provision for snags and coarse

woody debris is also important.

Grazing and Exotic Plants

Introduced exotic species and grazing are two major

problems that can seriously interfere with efforts to

restore fire as an ecosystem process. Well-intentioned
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 201

Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations Brown

prescribed fire, and silvicultural and rangeland enhancement activities, can fail drastically unless grazing and exotic plants are anticipated and managed

properly.

Grazing—Excessive grazing can be the biggest hindrance to successful use of prescribed fire where grass

vegetation is a major component, particularly in western grasslands and shrub/grass vegetation types

(Wright and Bailey 1982). It is more of a problem for

bunchgrasses than rhizomatous grasses (Mack and

Thompson 1982). Overgrazing in the absence of fire as

well as following fire can reduce plant diversity. Grazing too soon following fire can eliminate or greatly

reduce desirable vegetation. In grassland areas woody

plants are competitively favored, which could defeat

the purpose of burning to halt woody plant

encroachment.

Depending on site potential and grazing pressure,

grazing should be deferred 1 to 2 years following fire in

ecosystems such as sagebrush/grass and semidesert

shrub (Wright and Bailey 1982). In forests such as the

aspen type, intensive grazing of sprouting plants by

livestock and wild ungulates, especially elk, following

prescribed fires can greatly retard plant recovery.

Small prescribed burns are particularly vulnerable to

overutilization because of concentrated grazing (Bartos

and others 1991).

Grazing prior to a prescribed burn can easily reduce

fine fuels to a point where fire will not spread successfully nor have sufficient heat to ignite or kill woody

plants. At least 600 lb/acre of herbaceous fuel is needed

for successful prescribed fire in grassland and grass/

shrub vegetation (Wright and Bailey 1982).

Exotic Plants—Fire can create favorable sites for

nonindigenous plant species to become established

and flourish. If exotic plants already grow in or near

areas that are candidates for prescribed fire, a potential problem exists. Aggressive exotic species can competitively exclude native vegetation. Severe fires that

expose large areas of mineral soil are most apt to be

invaded by exotic plants; if exotics are already established, their dominance may be accelerated. Lower

severity burns are more resistant to proliferation of

exotics because many native species sprout and quickly

occupy the site.

Cheatgrass, a nonindigenous annual that dominates millions of acres, is an extreme example of a

species favored by fire. Its invasion of the sagebrushsteppe vegetation type has led to increased frequency

of wildfire due to abundant, early curing fine fuels. The

result is permanent conversion to annual grassland

and disruption of the historic fire regime (Whisenant

1990). Another problem with nonindigenous plants

can occur from seeding nonnative grasses such as

Figure 9-6—Aspen is being successionally replaced by fir, Bridger-Teton National Forest,

Wyoming. Restoration will require a stand-replacement disturbance, which could be facilitated by

cutting some of the conifers.
202 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Brown Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations

annual ryegrass on severely burned sites as part of

wildfire rehabilitation efforts. This practice, which is

intended to stabilize soils, can delay reestablishment

of native species and possibly alter long-term community composition (Conard and others 1991).

A far different problem is caused by exotics such as

Chinese tallow, which has invaded coastal marshes of

the Southeast. Its invasion causes a shift from grassdominated communities to a sparse forb-dominated

community that is much less flammable and acts as a

fire break. Consequently, once Chinese tallow gains

dominance on a site, prescribed fire cannot be effectively used to control the exotic and encroaching woody

plants. Thus, the grass-dominated marsh communities are reduced.

Fire Prescriptions

Ecosystem management has brought new challenges

to the application of prescribed fire primarily due to

the increased scale and complexity of some prescribed

burning (Zimmerman and Bunnell 1998). Traditionally, prescribed fire was applied on small, relatively

homogeneous units of a single land ownership. Prescribed fire will continue to be important for smallscale operations. But to meet some ecosystem goals,

prescribed fire needs to be applied over extensive

areas that contain a variety of vegetation communities

and fuel conditions.

In designing fire prescriptions, a strong, clear connection is needed between ecosystem goals, resource objectives, and fire objectives. This helps assure that prescribed fire will accomplish the desired effects. It can

also help in choosing proper technical aids for determining the prescription and in assuring fires are cost

effective and safely conducted. Designing prescriptions

through a visible, logical process can also demonstrate

professional competence and promote credibility of those

in charge of the prescribed fire activities.

Defining fire objectives boils down to specifying first

order fire effects that describe what the burning should

immediately accomplish (Brown 1985b). Treatment

objectives need to specify: (1) how much of what kind

of organic matter should be consumed, (2) what vegetation should be killed, and (3) what the size of

burned and unburned patches should be. Constraints

on achieving the treatment objectives must also be

considered. These can be thought of as the fire effects

that should be avoided. Controlling fire, managing

smoke, and avoiding overstory mortality are the common constraints. Specifying objectives and constraints

is a matter of declaring what the fire should accomplish and avoid. Both are fire objectives of sorts, so why

regard them differently? One reason is that it helps in

demonstrating an awareness of beneficial and undesirable aspects of fire and in explaining the prescribed

fire plans to others.

Depending on resource objectives, the fire objectives

may call for a wide or narrow prescription window. For

example, the resource objective to restore fire as a

process in a nonlethal understory fire regime type may

only require that prescribed fire be able to spread with

minimal mortality to the overstory, an objective that

could be accomplished with a wide prescription window. The specific resource objective of attaining natural regeneration while retaining some large downed

woody material may call for a fire objective that

specifies exposure of 20 to 30 percent mineral soil

without consuming more than half of the large downed

woody material. This would require a narrow prescription window.

Occasionally, conflicts may arise between fire objectives and constraints. A common example is between

the objective to reduce fuels by burning at low fuel

moistures and the constraint to control smoke production. Conflict can arise between different objectives;

for example, to expose a high percentage of mineral

soil and to leave large downed woody material for

other ecosystem benefits. When conflicts arise, compromise may prevent the fire from achieving the resource objectives. It is important to recognize those

situations so a potentially unsuccessful prescribed fire

can be avoided.

Many technical aids are available to assist in preparing fire prescriptions. Most involve prediction of

information such as weather probabilities, fuel loadings, fuel consumption, fire behavior, tree mortality,

and plant response. Two technical aids—both with

user guides that can help in writing and explaining

prescribed fire objectives and designing fire prescriptions—are relevant for applications across the United

States and much of Canada. They are the Fire Effects

Information System-FEIS (Fischer and others 1996)

and the First Order Fire Effects Model-FOFEM

(Reinhardt and others 1997).

FEIS—This is an easy to use, computerized knowledge management system that stores and retrieves

current information as text organized in an encyclopedic fashion. FEIS provides fire effects and related

biological, ecological, and management information in

three major categories: plant species, wildlife species,

and plant communities. The plant species category

includes for each species, information on taxonomy,

distribution and occurrence, value and use, botanical

and ecological characteristics, fire ecology, fire effects,

and references. A citation retrieval system can be

searched independently by author and keyword.

Although the system was originally developed to meet

prescribed fire needs, it is now recognized as a valuable aid for obtaining information about species ecology for any application. It can be accessed through a

U.S. Forest Service Web site:

http://www.fs.fed.us/database/feis
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 203

Chapter 9: Ecological Principles, Shifting Fire Regimes and Management Considerations Brown

FOFEM—This system was developed to predict the

direct consequences of fire, that is, first order fire

effects. FOFEM computes duff and woody fuel consumption, mineral soil exposure, fire-caused tree mortality, and smoke production for many forest and

rangeland ecosystems. An update is scheduled to add

soil heating effects. FOFEM contains a fire effects

calculator to predict effects of fire from the burning

conditions and a prescribed fire planner to compute

the burn conditions necessary to achieve a desired

effect. Users may enter their own fuel data or use

default values derived from fuel models provided for

natural and activity fuels by many forest cover types.

The model is implemented in a computer program

available for use on a PC or Forest Service computer.

To obtain a current version of the FOFEM software,

contact the authors at the Intermountain Fire Sciences Laboratory, (406) 329-4800, or PO Box 8089,

Missoula, MT 59807.

Research Needs ________________

The goals of maintaining sustainability of all ecosystem components and processes and conserving biodiversity present new challenges to land management

organizations. Knowledge of how ecosystems function

and what they provide is essential to making informed

environmental decisions. The following broadly stated

research needs indicate the knowledge required for

managing fire effects on flora and fuel that will contribute to maintaining sustainable ecosystems.

Characteristics of Fire Regimes

• What is the historical range of variability in fire

regime characteristics especially fire frequency,

seasonality, and severity for fire dependent ecosystems? This should be answered for multiple

spatial scales because of the hierarchical structure of ecosystems.

• What are the limits to ecosystem patterns and

processes that signal ecosystems are beyond the

boundaries of the historical range of variability?

• To what extent has climate influenced fire regime

characteristics in the past? How might anticipated climate change alter fire regime characteristics in the future?

Effects of Fire on Ecosystem Processes

and Biodiversity

• What are the long-term effects of fire of varying

frequencies and severities on nutrient dynamics

and vegetation?

• How does fire of varying frequency, seasonality,

and severity influence individual plant species

and plant community development? The emphasis for research should be on rare species and other

vegetation components where knowledge is lacking.

• What interactions between insects and diseases

and fire characterized historical fire regimes, and

how has this affected landscape patterns? How do

these interactions change when ecosystems exceed the natural range of variability and when

various management activities are applied?

• What is the interaction of different ecosystem

scales on ecosystem processes and biodiversity?

To what extent can coarse scale analysis account

for ecosystem processes and biodiversity?

• What are the long-term effects of largely excluding

fire from ecosystems that evolved under fire

regimes?

Restoration of Ecosystems

• What approaches and methods involving wildland

fire use, prescribed fire, silviculture, and grazing

can be used to restore ecosystems to a semblance

of the historical range of vegetation composition

and structure while meeting the resource needs of

society?

• What fuel management activities can provide an

acceptable level of fire hazard and remain compatible with ecosystem goals, especially needs for

coarse woody debris?

• How can nonindigenous plant species be managed

in combination with prescribed fire and resource

utilization activities to maintain biodiversity?

Development of Ecosystem Evaluation

Methodologies

• Continue with development of simulation models

and ecosystem evaluation techniques that can

help in understanding and managing ecosystem

dynamics. Succession and landscape models are

needed that account for interaction of fire, vegetation, fuels, and climate.

• Fire effects models at small spatial and temporal

scales are needed for rigorous fire effects hypothesis testing and as building blocks for models with

larger temporal and spatial scales.

• Determine organizational approaches that allow

complex ecosystem models requiring specialized

skills and high speed computer facilities to be

accessible to all land management organizations

and units.
204 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Notes

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USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 205

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Appendices

Appendix A: Common and Scientific Names of Plant Species_____________

Common names mentioned in the text and scientific names follow the nomenclature of the U.S. Department of

Agriculture (1999) PLANTS database except for some wet grassland species, which follow the Integrated

Taxonomic Information System (http://www.itis.usda.gov/). For some species a second common name is shown

because it is commonly used.

Common name Scientific name

Trees and Shrubs

a’ali’i Dodonaea viscosa

Alaska-cedar Chamaecyparis nootkatensis

alder Alnus spp.

alligator juniper Juniperus deppeana

American beech, beech Fagus grandifolia

American elm, white elm Ulmus americana

American mountain-ash Sorbus americana

antelope bitterbrush Purshia tridentata

Arizona pine Pinus ponderosa var. arizonica

Arizona white oak Quercus arizonica

Ashe juniper Juniperus ashei

aspen, trembling aspen, quaking aspen Populus tremuloides

Atlantic white-cedar Chamaecyparis thyoides

baccharis Baccharis spp.

baldcypress Taxodium distichum

balsam fir Abies balsamea

balsam poplar Populus balsamifera

basin big sagebrush Artemisia tridentata ssp. tridentata

basswood Tilia americana

batis Batis maritima

bayberry Morella spp.

beaked hazel Corylus cornuta

bear oak Quercus ilicifolia

Bebb willow Salix bebbiana

big sagebrush Artemisia tridentata

bigleaf maple Acer macrophyllum

bigleaf sumpweed Iva frutescens

bitter cherry Prunus emarginata

bitternut hickory Carya cordiformis

black ash Fraxinus nigra

black cherry Prunus serotina

black cottonwood Populus trichocarpa

black greasewood Sarcobatus vermiculatus

black mangrove Avicennia germinans

black oak Quercus velutina

black spruce Picea mariana

black walnut Juglans nigra

blackbrush Coleogyne ramosissima

blackgum, black tupelo Nyssa sylvatica

blackjack oak Quercus marilandica
240 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Appendices

blue huckleberry Vaccinium membranaceum

blue oak Quercus douglasii

blue spruce Picea pungens

bluejack oak Quercus incana

bog blueberry Vaccinium uliginosum

bog labrador tea Ledum groenlandicum

boxelder, Manitoba maple Acer negundo

Brazilian pepper Schinus terebinthifolus

buckeye Aesculus

buckwheat tree Cliftonia monophylla

bur oak Quercus macrocarpa

bursage spp. Ambrosia spp.

butternut Juglans cinerea

buttonwood, button mangrove Conocarpus erectus

cabbage palmetto Sabal palmetto

California black oak Quercus kelloggii

California red fir Abies magnifica

California sagebrush Artemisia californica

canyon live oak Quercus chrysolepis

Caribbean pine, Honduras pine Pinus caribaea

Carolina ash, pop-ash Fraxinus caroliniana

cenzia, purple sage Leucophyllum frutescens

chamise Adenostoma fasciculatum

Chapman oak Quercus chapmanii

chestnut oak Quercus prinus

Chinese tallow Sapium sebiferum

chokecherry Prunus virginiana

coast Douglas-fir Pseudotsuga menziesii var. menziesii

coast live oak Quercus agrifolia

coastalplain staggerbush Lyonia fruticosa

cocoplum Chrysobalanus icaco

common persimmon Diospyros virginiana

creeping barberry Mahonia repens

creosotebush Larrea tridentata

cypress Taxodium spp.

cyrilla, swamp cyrilla Cyrilla racemiflora

dahoon Ilex cassine

digger pine, California foothills pine Pinus sabiniana

Douglas-fir Pseudotsuga menziesii

dwarf chinkapin oak Quercus prinoides

dwarf huckleberry Gaylussacia dumosa

eastern baccharis, groundsel-tree Baccharis halimifolia

eastern cottonwood Populus deltoides

eastern hemlock Tsuga canadensis

eastern redcedar Juniperus virginiana

eastern white pine Pinus strobus

Engelmann spruce Picea engelmannii

eucalyptus Eucalyptus spp.

fetterbush Lyonia lucida

flowering dogwood Cornus florida

forage kochia Kochia prostrata

fourwing saltbush Atriplex canescens

Fraser fir Abies fraseri

gallberry, inkberry Ilex glabra

Gambel oak Quercus gambelii

Common name Scientific name
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 241

Appendices

giant sequoia Sequoiadendron giganteum

gooseberry, currant Ribes spp.

grand fir Abies grandis

gray birch Betula populifolia

green ash Fraxinus pennsylvanica

greenbriar Smilax glauca, Smilax spp.

ground blueberry Vaccinium myrsinites

hickory Carya spp.

hoaryleaf ceanothus Ceanothus crassifolius

honey mesquite Prosopis glandulosa

horsebrush Tetradymia spp.

incense-cedar Calocedrus decurrens

interior live oak Quercus wislizenii

interior ponderosa pine Pinus ponderosa var. scopulorum

jack pine Pinus banksiana

Jeffrey pine Pinus jeffreyi

Joshua tree Yucca brevifolia

juniper Juniperus spp.

koa Acacia koa

large gallberry Ilex coriacea

laurel oak Quercus laurifolia

leatherleaf Chamaedaphne calyculata

live oak Quercus virginiana

loblolly pine Pinus taeda

loblolly-bay Gordonia lasianthus

lodgepole pine Pinus contorta

longleaf pine Pinus palustris

lyonia Lyonia spp.

manzanita Arctostaphylos spp.

melaleuca Melaleuca quinquenervia

mesquite Prosopis spp.

mockernut hickory Carya tomentosa

Mormon tea Ephedra torreyana

mountain alder Alnus incana

mountain big sagebrush Artemisia tridentata ssp. vaseyana

mountain hemlock Tsuga mertensiana

mountain-laurel Kalmia latifolia

myrsine Myrsine quianensis

myrtle oak Quercus myrtlifolia

myrtle, wax myrtle, southern bayberry Morella cerifera

noble fir Abies procera

northern pin oak Quercus ellipsoidalis

northern red oak Quercus rubra

northern white-cedar, e. white-cedar Thuja occidentalis

oneseed juniper Juniperus monosperma

oak Quercus spp.

Oregon white oak Quercus garryana

Pacific madrone Arbutus menziesii

Pacific ponderosa pine Pinus ponderosa var. ponderosa

Pacific silver fir Abies amabilis

paloverde spp. Cercidium spp.

paper birch Betula papyrifera

persimmon Diospyros spp.

piedmont staggerbush Lyonia mariana

pignut hickory Carya glabra

Common name Scientific name
242 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Appendices

pin cherry, fire cherry Prunus pensylvanica

pin oak Quercus palustris

pinyon pine See singleleaf pinyon, true pinyon

pitch pine Pinus rigida

pond cypress Taxodium ascendens

pond pine Pinus serotina

ponderosa pine Pinus ponderosa

poplar Populus spp.

post oak Quercus stellata

rabbitbrush Chrysothamnus spp.

raspberry, blackberry Rubus spp.

red alder Alnus rubra

red bay Persea borbonia

red elderberry Sambucus racemosa ssp. pubens

red mangrove Rhizophora mangle

red maple Acer rubrum

red pine Pinus resinosa

red raspberry Rubus idaeus

red spruce Picea rubens

redberry juniper Juniperus erythrocarpa

redstem ceanothus Ceanothus sanguineus

redwood Sequoia sempervirens

rhododendron Rhododendron spp.

Rocky Mountain Douglas-fir Pseudotsuga menziesii var. glauca

Rocky Mountain juniper Juniperus scopulorum

Rocky Mountain lodgepole pine Pinus contorta var. latifolia

Rocky Mountain maple Acer glabrum

rosemary Ceratiola ericoides

rusty staggerbush Lyonia ferruginea

sagebrush Artemisia spp.

salmonberry Rubus spectabilis

sand live oak Quercus virginiana var. maritima

sand pine Pinus clausa

sand post oak Quercus stellata var. margaretta

sand shinnery oak Quercus havardii

Saskatoon serviceberry Amelanchier alnifolia

saw palmetto Serenoa repens

scarlet oak Quercus coccinea

scrub oak Quercus dumosa

scrub palmetto Sabal etonia

shadscale Atriplex confertifolia

sheep-laurel Kalmia angustifolia

shore pine Pinus contorta var. contorta

shortleaf pine Pinus echinata

silver maple Acer saccharinum

silver sagebrush Artemisia cana

silverbell Halesia Ellis

singleleaf pinyon Pinus monophylla

Sitka spruce Picea sitchensis

slash pine Pinus elliottii

snakeweed Gutierrezia spp.

snowbrush ceanothus Ceanothus velutinus

sourwood Oxydendrum arboreum

southern magnolia Magnolia grandiflora

Common name Scientific name
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 243

Appendices

southern red oak, cherrybark oak Quercus falcata

speckled alder Alnus rugosa

spiny hopsage Grayia spinosa

spruce pine Pinus glabra

subalpine fir Abies lasiocarpa

sugar maple Acer saccharum

sugar pine Pinus lambertiana

sugarberry Celtis laevigata

swamp bay Persea palustris

swamp chestnut oak Quercus michauxii

swamp tupelo Nyssa biflora

sweetbay Magnolia virginiana

sweetgum Liquidambar styraciflua

sweetpepperbush, poor man’s soap Clethra alnifolia

sycamore Platanus occidentalis

Table Mountain pine Pinus pungens

tamarack Larix laricina

tanoak Lithocarpus densiflora

tarbush Flourensia cernua

Texas persimmon Diospyros texana

thimbleberry Rubus parviflorus

thinleaf alder Alnus incana ssp. tenuifolia

threetip sagebrush Artemisia tripartita

toyon Heteromeles arbutifolia

trefoil Lotus spp.

true pinyon, Colorado pinyon Pinus edulis

turkey oak Quercus laevis

Utah juniper Juniperus osteosperma

varnish leaf Dodonea virginiana

Virginia pine Pinus virginiana

water oak Quercus nigra

water tupelo Nyssa aquatica

western hemlock Tsuga heterophylla

western juniper Juniperus occidentalis

western larch Larix occidentalis

western redcedar Thuja plicata

western white pine Pinus monticola

white ash Fraxinus americana

white bully Sideroxylon salicifolium

white bursage Ambrosia dumosa

white fir Abies concolor

white mangrove Laguncularia racemosa

white oak Quercus alba

white sage Salvia apiana

white spirea Spiraea betulifolia

white spruce Picea glauca

whitebark pine Pinus albicaulis

willow Salix spp.

willow oak Quercus phellos

winterfat Krascheninnikovia lanata

Wyoming big sagebrush Artemisia tridentata ssp. wyomingensis

yaupon Ilex vomitoria

yellow birch Betula alleghaniensis

Common name Scientific name
244 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Appendices

yellow buckeye Aesculus octandra

yellow paloverde Cercidium microphyllum

yellow-poplar Liriodendron tulipifera

Grasses and Forbs

alligatorweed Alternanthera philoxeroides

American white waterlily Nymphaea odorata

annual fleabane Erigeron annuus

annual ryegrass, Italian ryegrass Lolium perenne spp. multiflorum

arrowhead Sagittaria spp.

arrowleaf balsamroot Balsamorhiza sagittata

basin wildrye Leymus cinereus

beakrush, Tracy’s beaksedge Rhynchospora tracyi

big bluestem Andropogon gerardii var. gerardii

black grama Bouteloua eriopoda

blackeyed Susan Rudbeckia hirta

blue grama Bouteloua gracilis

bluebunch wheatgrass Pseudoroegneria spicata

blue-eyed grass Sisyrinchium campestre

bluestem spp. Schizachyrium spp.

bottlebrush squirreltail Elymus elymoides

broadleaf arnica Arnica latifolia

brome Bromus spp.

broomsedge, broomsedge bluestem Andropogon virginicus

buffalograss Buchloe dactyloides

bulltongue arrowhead Sagittaria lancifolia

Burma reed Neyrundia reynaudiana

bush muhly Muhlenbergia porteri

cane, switch cane Arundinaria gigantea

cattail Typha spp.

chairmaker’s bullrush Schoenoplectus americanus

chalky bluestem Andropogon capillipes

cheatgrass Bromus tectorum

cogongrass Imperata cylindrica

Columbian bluestem, bush beardgrass Schizachyrium condensatum

common camas Camassia quamash

cordgrass Spartina spp.

curly-mesquite Hilaria belangeri

Curtis’ dropseed Sporobolus curtissii

cutgrass Zizaniopsis spp.

deathcamas Zigadenus venenosus

dropseed, sacaton Sporobolus spp.

fescue Festuca spp.

fireweed Epilobium angustifolium

fountain grass Pennisetum sataceum

gayfeather Liatris spp.

glacier lily Erythronium grandiflorum

golden brodiaea Brodiaea ixioides

green arrow arum Peltandra virginica

gulf cordgrass Spartina spartinae

hairawn muhly Muhlenbergia capillaris

heartleaf arnica Arnica cordifolia

Common name Scientific name
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 245

Appendices

hydrocotyle, marshpennywort Hydrocotyle spp.

Idaho fescue Festuca idahoensis

Indian paintbrush Castilleja spp.

Indian ricegrass Oryzopsis hymenoides

Indiangrass Sorghastrum nutans

inland saltgrass Distichlis spicata

Lehmann lovegrass Eragrostis lehmanniana

little bluestem Andropogon scoparius

little bluestem, creeping bluestem Schizachyrium scoparium

lupine Lupinus spp.

maidencane Panicum hemitomon

milkvetch Astragalus spp.

molasses grass Melinis minutiflora

muhly spp. Muhlenbergia spp.

natal redtop, rose natalgrass Melinis repens

needlegrass rush, black rush Juncus roemerianus

panicum Panicum spp.

pasque flower Anemone pratens

pickerelweed Pontederia cordata

pickleweed Salicornia spp.

pili grass Heteropogon contortus

pinegrass Calamagrostis rubescens

pineland threeawn, wiregrass Aristida stricta

pitcherplant Sarracenia purpurea

pond lily Nuphar spp.

prairie cordgrass Spartina pectinata

prairie violet Viola pedatifida

red brome Bromus rubens

reed Phragmites spp.

rough fescue Festuca scabrella

rush Juncus spp.

saltgrass Distichlis spp.

saltmeadow cordgrass Spartina patens

saltmeadow rush Juncus gerardii

sand cordgrass Spartina bakeri

sand dropseed Sporobolus cryptandrus

Sandberg bluegrass Poa secunda

sawgrass Cladium spp.

seaside tansy Borrichia spp.

sheathed cottonsedge Eriophorum vaginatum

showy aster Aster conspicuus

showy partridgepea Cassia fasciculata

sideoats grama Bouteloua curtipendula

sky blue aster Aster azureus

slender bluestem Schizachyrium tenerum

smooth cordgrass Spartina alterniflora

spikerush Eleocharis spp.

star-flowered Solomon’s seal Smilacina stellata

strawberry Fragaria spp.

switchgrass Panicum virgatum

thatching grass Hyparrhenia rufa

Thurber’s needlegrass Achnatherum thurberiana

Common name Scientific name
246 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Appendices

tobosa Pleuraphis mutica

trumpet pitcherplant, yellow pitcherplant Sarracenia flava

twinflower Linnaea borealis

waterlily Nymphaea spp.

western wheatgrass Pascopyrum smithii

western yarrow Achillea millefolium

wheatgrass Agropyron spp.

wild columbine Aquilegia canadensis

wild hollyhock Iliamna rivularis

wild sarsaparilla Aralia nudicaulis

wildrice Zizania spp.

wiregrass Aristida spp.

Mosses, Ferns, Cactus, and Lichens

bracken fern Pteridium aquilinum

cholla, prickly pear Opuntia spp.

cup lichen Cladonia spp.

dicranum Dicranum spp.

knight’s plume moss Ptilium crista-castrensis

mountain fern moss Hylocomium splendens

Old world or small-leaf climbing fern Lygodium microphyllum

prickly pear cactus Opuntia humifusa

saguaro Carnegia gigantea

Schreber’s moss Pleurozium schreberi

sphagnum Sphagnum spp.

Common name Scientific name
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 247

Appendices

FIRESUM–FIRE SUcession Model is a stand level,

individual tree ecosystem process model developed for western conifers, especially ponderosa

pine and whitebark pine, to simulate effects of

different fire regimes on tree composition, stand

structure, and fuel loadings (Keane and others

1990, 1989).

FVS–Forest Vegetation Simulation is a stand level

mensurational model (Wykoff and others 1982).

FFE–Fire and Fuels Extension to the FVS is a stand

level model for simulating surface fuel loadings,

tree characteristics, expected fire behavior, and

expected tree mortality (Beukema and others

1999). This model is particularly useful for growth,

mortality, falldown, and decay of conifer trees.

Fire-BGC–The Fire Biogeochemical Mechanistic process model can be used to investigate stand- and

landscape-level changes in ecosystem processes

and characteristics in fire-dominated environments (Keane and others 1996). It is useful for

evaluating effects of climate change.

SIMPPLLE–Simulating Patterns and Processes at

Landscape Scales is a rule-based model designed

for Northern Rocky Mountain forest types (Chew

1997). It starts at a coarse scale and adds fine scale

only as needed to produce acceptable performance.

VDDT—The Vegetation Dynamics Development Tool

is a nonspatial, deterministic model where successional pathways connecting successional stages

are used to explore community dynamics

(Beukuma and Kurz 1995). VDDT only simulates

one vegetation type at a time but is useful and

efficient for simulating disturbance and succession on mid-scale to fine-scale landscapes. This

model can be readily used by managers and is

undergoing development for national applications.

CRBSUM–The Columbia River Basin Succession

Model was used to simulate landscape changes

for the Interior Columbia Basin Ecosystem Management Project (Keane and others 1996). It

incorporates disturbance as a stochastic process

and models succession for individual landscape

pixels. CRBSUM 2 was created from CRBSUM to

improve the simulation of fire processes over

time and space (Keane and Long 1998).

LANDSUM–the Landscape Succession Model was

derived from CRBSUM to operate on a polygon

level rather than pixels. This allows it to be used

at finer scales of resolution (Keane 1999).

FIREPAT–Fire Pattern Succession Model attempts

to more realistically model fire by simulating fire

ignition and size to compute number of pixels

disturbed by fire (Keane and Long 1998). It

operates at a coarse scale and models succession

similar to CRBSUM.

INTELAND–Integrated Terrestrial Landscape Model

simulates natural processes in boreal forest ecosystems (Gauthier and others 1998). It was designed as a GIS-based model to help define natural system baselines for disturbance regimes,

vegetation dynamics, wildlife species composition, and landscape diversity.

Appendix B: Succession Simulation Models __________________________
248 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Appendices

The terminology here was derived from the following

references: fuels and fire behavior from Agee (1993),

Brown and others (1982), and Ryan and Noste (1985);

fire occurrence from Agee (1993), Johnson (1992), and

Romme (1980); and plant reproduction from Allaby

(1992), Helms (1998), and Sutton and Tinus (1983).

 Fuels

Fuel comprises living and dead vegetation that can

be ignited. It is often classified as dead or alive and as

natural or activity fuels. Fuel components refer to

such items as downed dead woody material by various

size classes, litter, duff, herbaceous vegetation, live

foliage, live shrub stems and so forth.

Kinds of Fuel

dead fuels: Fallen dead vegetation such as downed

woody material, litter, duff, and organic soils and

dead upright vegetation such as cured grasses,

forbs, and dead attached shrub stemwood.

live fuels: Living plants. Especially important components include tree crowns, shrubs, grasses,

forbs, and ferns.

natural fuels: These result from plant growth and

death, loss of foliage, branch breakage, and tree

blowdown.

activity fuels: These fuels result from human activity such as logging, thinning, chaining, and herbicide use. It usually refers to residues from

cutting operations.

down, dead woody fuels: Dead twigs, branches,

stems, and boles of trees and shrubs that have

fallen and lie on or near the ground (Brown and

others 1982). Wood includes sound and rotten

components.

litter: The top layer of the forest floor (01 soil horizon),

including freshly fallen leaves, needles, fine twigs,

bark flakes, fruits, matted dead grass and a

variety of miscellaneous vegetative parts that

are little altered by decomposition. Litter also

accumulates beneath rangeland shrubs. Some

surface feather moss and lichens are considered

to be litter because their moisture response is

similar to dead fine fuel. In grasslands, litter is

the accumulated dead herbaceous material usually referred to as thatch.

duff: Partially decomposed organic matter lying beneath the litter layer and above the mineral soil.

It includes the fermentation and humus layers of

the the forest floor (02 soil horizon).

Appendix C: Glossary_____________________________________________

organic soils: The deep layers of organic matter that

frequently develop in poorly drained areas such

as bogs, swamps, and marshes.

Fuel Properties

loading: The weight per unit area of fuel often expressed in tons per acre or tonnes per hectare.

Dead woody fuel loadings are commonly described

for small material in diameter classes of 0 to 0.25,

0.25 to 1, and 1 to 3 inches and for large material

greater than 3 inches.

fuel continuity: A qualitative description of the distribution of fuel both horizontally and vertically.

Continuous fuels readily support fire spread.

The larger the fuel discontinuity, the greater the

fire intensity required for fire spread.

total fuel: The amount of biomass that potentially

could burn.

available fuel: The amount of biomass that will burn

under a given set of conditions. Moisture content

and fuel size are the primary determinants of

availability. Arrangement and compactness of

fuel may also determine availability.

fuel moisture content: This is expressed as a percent or fraction of oven dry weight of fuel. It is the

most important fuel property controlling flammability. In living plants it is physiologically

bound. Its daily fluctuations vary considerably

by species but are usually above 80 to 100 percent. As plants mature, moisture content decreases. When herbaceous plants cure, their

moisture content responds as dead fuel moisture

content, which fluctuates according to changes in

temperature, humidity, and precipitation.

Fire Behavior

type of fire: Refers to the fuels that are primarily

supporting the fire namely surface fires, ground

fires, and crown fires.

surface fires: These fires burn in litter and other live

and dead fuels at or near the surface of the

ground mostly by flaming combustion.

ground fires: These burn in the organic material

below the litter layer mostly by smoldering combustion. Fires in duff, peat, dead moss and lichens, and punky wood are typically ground

fires.

crown fires: These burn in the crowns of trees and

shrubs usually ignited by a surface fire. They are
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 249

Appendices

common in coniferous forests and chaparral type

shrublands.

fireline intensity: Also called Byram’s intensity,

this is the rate of energy release per unit length

of the fire front expressed as BTU per foot of

fireline per second or as kilowatts per meter of

fireline. This expression is commonly used to

describe the power of wildland fires.

flame length: The length of flames in the propagating

fire front measured along the slant of the flame

from the midpoint of its base to its tip. It is

mathematically related to fireline intensity and

tree crown scorch height.

total heat release: The heat released by combustion

during burnout of all fuels in BTU per square foot

or kilocalories per square meter.

fire duration: The length of time that combustion

occurs at a given point. It relates closely to

downward heating and fire effects below the fuel

surface as well as heating of tree boles above the

surface.

ground char: A qualitative measure of a fire’s heat

pulse downward into the soil. It is determined by

visually judging the extent of fuel consumption,

charring, and changes in soil texture (Ryan and

Noste 1985). It is also referred to as burn severity and depth of burn, which is a quantitative

expression of depth of forest floor consumed by

fire. It is largely determined by fire duration and

characteristics of the soil including the forest

floor.

fire severity: A qualitative measure of the immediate

effects of fire on the ecosystem. It relates to the

extent of mortality and survival of plant and

animal life both aboveground and belowground

and to loss of organic matter. It is determined by

heat released aboveground and belowground.

Ryan and Noste (1985) describe a method for

rating fire severity based on flame length and

depth of burn.

Fire Occurrence

The definitions here were based on a review of fire

history terminology at a fire history workshop (Romme

1980) and phraseology by Agee (1993).

fire cycle: A fire return interval calculated using a

negative exponential distribution, applied using

current age-class structure on the landscape. It

is the average stand age of a forest characterized

using the negative exponential distribution.

fire frequency: A general term referring to the recurrence of fire in a given area over time. It is

sometimes stated as number of fires per unit

time in a designated area. It is also used to refer

to the probability of an element burning per unit

time (Johnson 1992).

fire rotation: The length of time necessary for an

area equal in size to the study area to burn and

is equal to the fire cycle.

mean fire-return interval: The arithmetic average

of all fire intervals in a given area over a given

time period.

Plant Reproduction

axil: The upper side of the point where a leaf meets a

stem, or a branch meets another branch or the

main stem of a plant.

bulb: An underground storage organ that bears roots

on its lower surface and fleshy leaves above. It

provides a means of reproduction in perennials.

burl: A mass of woody tissue or wartlike structure

fomed on stem, branch, or root; has numerous

buds, which rarely develop further.

caudex: A largely underground woody stem base of an

otherwise herbaceous perennial forb that produces leaves and flowering stems.

corm: An underground storage organ bearing adventitious roots and scale leaves. It may function as an

organ of vegetative reproduction in perennials.

epicormic branch: A shoot arising spontaneously

from an adventitious or dormant bud on the stem

or branch of a woody plant often following exposure to increased light levels or fire.

lignotuber: A woody organ of food storage and regeneration forming a swelling of stem material,

more or less at ground level, that originates from

the axils of cotyledons or early seedling leaves.

rhizome: A creeping stem, not a root, growing beneath the surface consisting of a series of nodes

with roots commonly produced from the nodes

and producing buds in the leaf axils.

root collar: Loosely, the point along the main stemroot axis at which the primary vascular anatomy

changes from that of a stem to that of a root, usually

applied to trees. Transition point between stem

and root. It may be clearly or vaguely apparent.

root crown: A mass of woody tissue from which roots

and stems originate, usually applied to shrubs

and herbaceous plants; can be considered as the

point at which root and stem meet.
250 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Appendices

Index

A

a’ali’i 173, 239

Abies amabilis 241

Abies balsamea 239

Abies concolor 243

Abies fraseri 240

Abies grandis 241

Abies lasiocarpa 243

Abies magnifica 240

Abies procera 241

Acacia koa 241

Acer glabrum 242

Acer macrophyllum 239

Acer negundo 240

Acer rubrum 242

Acer saccharinum 242

Acer saccharum 243

Achillea millefolium 246

Achnatherum thurberiana 245

Adenostoma fasciculatum 240

Aesculus 240, 244

Aesculus octandra 244

Agropyron spp. 246

Alaska-cedar 12, 116, 239

alder 14, 17, 26, 35, 45, 48, 92, 98, 108, 115, 239, 241,

242, 243

alligator juniper 12, 134, 239

alligatorweed 89, 244

Alnus incana 241

Alnus incana ssp. tenuifolia 243

Alnus rubra 242

Alnus rugosa 243

Alnus spp. 239

Alternanthera philoxeroides 244

Ambrosia deltoidea 243

Ambrosia dumosa 143, 243

Ambrosia spp. 240

Amelanchier alnifolia 242

American beech 13, 83, 239

American elm 13, 54, 70, 72, 76, 84, 239

American mountain-ash 118, 239

American white waterlily 171, 244

Andropogon capillipes 244

Andropogon gerardii var. gerardii 244

Andropogon scoparius 245

Andropogon virginicus 244

Anemone pratens 245

annual fleabane 94, 244

annual ryegrass 202, 244

antelope bitterbrush 17, 239

Aquilegia canadensis 246

Aralia nudicaulis 246

Arbutus menziesii 241

Arctostaphylos spp. 241

Aristida spp. 246

Aristida stricta 245

Arizona pine 98, 121, 122, 239

Arizona white oak 129, 239

Arnica cordifolia 244

Arnica latifolia 244

arrowhead 85, 89, 244

arrowleaf balsamroot 20, 244

Artemisia californica 240

Artemisia cana 242

Artemisia spp. 242

Artemisia tridentata 143, 239, 241, 243

Artemisia tridentata ssp. tridentata 239

Artemisia tridentata ssp. vaseyana 241

Artemisia tridentata ssp. wyomingensis 243

Artemisia tripartita 243

Arundinaria gigantea 244

Ashe juniper 122, 131, 134, 238, 239

aspen 13, 18, 19, 20, 35, 36, 37, 38, 39, 40, 41, 42,

43, 44, 48, 49, 83, 84, 99, 110, 111, 113, 116, 117,

120, 151, 194, 201, 221, 223, 224, 228, 229, 231,

237, 238, 239

Aster azureus 245

Aster conspicuus 245

Astragalus spp. 245

Atlantic white-cedar 55, 81, 84, 85, 86, 88, 89, 94, 96,

239

Atriplex canescens 240

Atriplex confertifolia 143, 242

Avicennia germinans 239

B

baccharis 85, 90, 91, 143, 239, 240

Baccharis halimifolia 240

Baccharis spp. 239

baldcypress 90, 91, 167, 239

balsam fir 11, 35, 36, 37, 38, 41, 42, 43, 44, 45, 46,

47, 48, 49, 84, 86, 194, 239

balsam poplar 26, 35, 239

Balsamorhiza sagittata 244

basin big sagebrush 146, 239

basin wildrye 23, 244

basswood 13, 55, 76, 78, 82, 239

batis 172, 244

Batis maritima 244

bayberry 89, 239, 241

beaked hazel 20, 41, 239

beakrush 171, 172, 244

bear oak 54, 75, 79, 80, 81, 239

Bebb willow 26, 239

beech 44, 49, 55, 74, 76, 78, 82, 84, see American

beech

Betula alleghaniensis 243

Betula papyrifera 241

Betula populifolia 241

big bluestem 29, 85, 92, 93, 244

big sagebrush 142, 146, 153, 154, 190, 195, 239, 241,

243

bigleaf maple 13, 17, 108, 115, 239
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 251

Appendices

bigleaf sumpweed 85, 239

bitter cherry 108, 239

bitternut hickory 13, 239

black ash 45, 54, 239

black cherry 55, 239

black cottonwood 13, 239

black grama 141, 157, 244

black greasewood 139, 143, 155, 239

black mangrove 172, 239

black oak 12, 13, 16, 54, 73, 79, 97, 98, 122, 131,

239, 240

black rush 172, see needlegrass rush

black spruce 12, 25, 26, 27, 36, 38, 41, 42, 43, 44,

45, 46, 47, 48, 49, 50, 119, 239

black tupelo 239

black walnut 76, 239

blackeyed Susan 94, 244

blackberry 245

blackbrush 122, 138, 139, 140, 143, 146, 154, 155, 239

blackgum 13, 76, 80, 86, 94, 239

blackjack oak 12, 54, 67, 72, 79, 80, 239

blue grama 150, 244

blue huckleberry 18, 19, 240

blue oak 12, 97, 98, 122, 240

blue spruce 12, 99, 240

blue-eyed grass 94, 244

bluebunch wheatgrass 22, 152, 244

bluejack oak 67, 240

bluestem spp. 54, 55, 72, 95, 244

bog blueberry 18, 240

bog labrador tea 38, 43, 240

Borrichia spp. 245

bottlebrush squirreltail 22, 244

Bouteloua curtipendula 245

Bouteloua eriopoda 244

Bouteloua gracilis 244

boxelder 240

bracken fern 20, 21, 49, 246

Brazilian pepper 166, 240

broadleaf arnica 29, 244

Brodiaea ixioides 244

brome 133, 139, 154, 244, 245

Bromus rubens 245

Bromus spp. 244

Bromus tectorum 244

broomsedge 64, 71, 170, 244

broomsedge bluestem see broomsedge

Buchloe dactyloides 244

buckeye 76, 240, 244

buckwheat tree 70, 71, 240

buffalograss 123, 130, 150, 244

bulltongue arrowhead 85, 89, 244

Bumelia salicifolia 243

bur oak 10, 35, 54, 240

Burma reed 166, 244

bursage 122, 138, 139, 141, 143, 148, 156, 240

bush beardgrass see Columbian bluestem

bush muhly 156, 244

butternut 76, 240

button mangrove see buttonwood

buttonwood 172, 240

C

cabbage palmetto 71, 161, 162, 169, 240

Calamagrostis rubescens 245

California black oak 13, 97, 98, 122, 131, 240

California foothills pine see digger pine

California red fir 106, 107, 109, 110, 240

California sagebrush 154, 240

Calocedrus decurrens 241

Camassia quamash 244

cane 81, 240, 244

canyon live oak 13, 98, 108, 122, 131, 240

Caribbean pine 161, 162, 163, 165, 240

Carnegia gigantea 246

Carolina ash 70, 168, 240

Carya cordiformis 239

Carya glabra 241

Carya spp. 241

Carya tomentosa 241

Cassia fasciculata 245

Castilleja spp. 245

cattail 85, 88, 172, 244

Ceanothus crassifolius 241

Ceanothus sanguineus 242

Ceanothus velutinus 242

Celtis laevigata 243

cenzia 240

Ceratiola ericoides 242

Cercidium microphyllum 143, 244

Cercidium spp. 241

chairmaker’s bullrush 90, 91, 244

chalky bluestem 71, 244

Chamaecyparis nootkatensis 239

Chamaecyparis thyoides 239

Chamaedaphne calyculata 241

chamise 17, 27, 28, 33, 149, 159, 240

Chapman oak 95, 240

Chasmanthium sessiliflorum 245

cheatgrass 122, 136, 137, 146, 152, 153, 154, 155, 180,

195, 201, 244

cherrybark oak see southern red oak

chestnut oak 54, 73, 76, 79, 86, 240, 243

Chinese tallow 89, 90, 91, 92, 202, 244

chokecherry 18, 118, 240

cholla 141, 246

Chrysobalanus icaco 240

Chrysothamnus nauseosus 242

Chrysothamnus paniculatus 240

Cladium spp. 245

Cladonia spp. 246

Clethra alnifolia 242

Cliftonia monophylla 240

coast Douglas-fir 105, 106, 108, 113, 115, 240

coast live oak 98, 122, 240

coastalplain staggerbush 164, 240

cocoplum 164, 240

cogongrass 68, 166, 244

Coleogyne ramosissima 239

Colorado pinyon see true pinyon

Columbian bluestem 170, 244

common camas 20, 244
252 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Appendices

common persimmon 67, 72, 73, 240

Conocarpus erectus 240

cordgrass 55, 84, 85, 88, 90, 91, 172, 244, 245

Cornus florida 240

Corylus cornuta 239

creeping barberry 18, 240

creeping bluestem 71, see little bluestem

creosotebush 122, 138, 139, 140, 141, 143, 148, 155,

156, 240

cup lichen 95, 246

curly-mesquite 130, 141, 244

currant 118, see goosberry

Curtis’ dropseed 71, 244

cutgrass 85, 244

cypress 55, 56, 59, 70, 76, 81, 84, 89, 94, 122, 162, 163,

164, 165, 166, 167, 168, 169, 170, 172, 240, 242

cyrilla 70, 81, 94, 240

Cyrilla racemiflora 240

D

dahoon 71, 164, 240

deathcamas 20, 244

dicranum 46, 246

Dicranum spp. 246

digger pine 11, 97, 98, 122, 240

Diospyros spp. 241

Diospyros texana 243

Diospyros virginiana 240

Distichlis spicata 245

Distichlis spp. 245

Dodonaea viscosa 239

Dodonea virginiana 243

Douglas-fir 11, 15, 25, 35, 97, 98, 99, 100, 102, 105,

106, 107, 108, 109, 110, 111, 113, 114, 115, 116, 117,

118, 151, 187, 188, 193, 195, 199, 240, 242

dropseed 23, 71, 244, 245

dwarf chinkapin oak 81, 240

dwarf huckleberry 71, 240

E

eastern baccharis 85, 90, 240

eastern cottonwood 10, 13, 35, 76, 240

eastern hemlock 76, 240

eastern redcedar 76, 85, 92, 240

eastern white-cedar see northern white-cedar

eastern white pine 35, 44, 74, 75, 77, 79, 80, 240

Eleocharis spp. 245

Elymus elymoides 244

Engelmann spruce 12, 35, 99, 111, 112, 118, 240

Ephedra torreyana 241

Epilobium angustifolium 244

Eragrostis lehmanniana 245

Erigeron annuus 244

Eriophorum vaginatum 245

Erythronium grandiflorum 244

eucalyptus 17, 171, 240

Eucalyptus spp. 240

F

Fagus grandifolia 239

fescue 22, 122, 139, 152, 244, 245

Festuca idahoensis 245

Festuca scabrella 245

Festuca spp. 244

fetterbush 59, 95, 240

fire cherry see pin cherry

fireweed 20, 21, 33, 244

Flourensia cernua 243

flowering dogwood 73, 240

forage kochia 152, 244

fountain grass 170, 171, 244

fourwing saltbush 240

Fragaria spp. 245

Fraser fir 86, 240

Fraxinus americana 243

Fraxinus caroliniana 240

Fraxinus nigra 239

Fraxinus pennsylvanica 241

G

gallberry 58, 63, 64, 67, 71, 80, 81, 95, 164, 240, 241

Gambel oak 18, 20, 28, 129, 240

gayfeather 20, 244

Gaylussacia dumosa 240

giant sequoia 12, 106, 186, 241

glacier lily 20, 244

golden brodiaea 29, 244

gooseberry 27, 241

Gordonia lasianthus 241

grand fir 11, 98, 100, 108, 115, 116, 118, 241

gray birch 80, 83, 84, 241

Grayia spinosa 243

greasewood 122, 138, 139, 146, 155

green arrow arum 85, 89, 244

green ash 54, 76, 84, 241

greenbriar 241

ground blueberry 71, 241

groundsel-tree see eastern baccharis

gulf cordgrass 85, 88, 172, 244

Gutierrezia spp. 242

H

hairawn muhly 172, 244

Halesia Ellis 242

heartleaf arnica 20, 33, 244

Heteromeles arbutifolia 243

Heteropogon contortus 245

hickory 6, 13, 14, 16, 53, 54, 56, 57, 61, 62, 64, 65,

68, 72, 73, 74, 75, 76, 239, 241

Hilaria belangeri 244

hoaryleaf ceanothus 27, 241

Honduras pine see Caribbean pine

honey mesquite 14, 130, 135, 136, 194, 241

horsebrush 18, 154, 241

hydrocotyle 89, 245
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 253

Appendices

Hydrocotyle spp. 245

Hylocomium splendens 246

Hyparrhenia rufa 245

I

Idaho fescue 22, 152, 245

Ilex cassine 240

Ilex coriacea 241

Ilex glabra 240

Ilex vomitoria 243

Iliamna rivularis 246

Imperata cylindrica 244

incense-cedar 12, 100, 241

Indian paintbrush 20, 245

Indian ricegrass 22, 245

Indiangrass 85, 92, 130, 245

inkberry see gallberry

inland saltgrass 85, 172, 245

interior live oak 122, 241

interior ponderosa pine 98, 121, 122, 241

Italian ryegrass see annual ryegrass

Iva frutescens 239

J

jack pine 5, 11, 25, 35, 36, 37, 38, 39, 41, 42, 44,

128, 194, 241

Jeffrey pine 11, 97, 98, 100, 101, 241

Joshua tree 139, 141, 142, 143, 241

Juglans cinerea 240

Juglans nigra 239

Juncus gerardii 245

Juncus roemerianus 245

Juncus spp. 245

juniper 12, 121, 122, 128, 129, 130, 131, 132, 133,

134, 139, 141, 142, 143, 144, 157, 158, 193, 199,

239, 241, 242, 243

Juniperus ashei 239

Juniperus deppeana 239

Juniperus erythrocarpa 242

Juniperus monosperma 241

Juniperus occidentalis 243

Juniperus osteosperma 243

Juniperus scopulorum 242

Juniperus spp. 143, 241

Juniperus virginiana 240

K

Kalmia angustifolia 242

Kalmia latifolia 241

knight’s plume moss 43, 246

koa 173, 241

Kochia prostrata 244

Krascheninnikovia lanata 143, 243

L

Laguncularia racemosa 243

large gallberry 81, 241

Larix laricina 243

Larix occidentalis 243

Larrea tridentata 240

laurel oak 72, 241

leatherleaf 80, 96, 241

Ledum groenlandicum 240

Lehmann lovegrass 245

Leucophyllum frutescens 240

Leymus cinereus 244

Liatris spp. 244

Linnaea borealis 246

Liquidambar styraciflua 243

Liriodendron tulipifera 244

Lithocarpus densiflora 243

little bluestem 29, 67, 71, 85, 92, 245

live oak 13, 95, 98, 108, 122, 129, 131, 132, 161, 162,

163, 166, 169, 240, 241, 242

loblolly pine 11, 15, 16, 28, 54, 63, 67, 68, 69, 70, 71,

72, 73, 76, 241

loblolly-bay 70, 81, 168, 241

lodgepole pine 5, 10, 11, 24, 25, 27, 35, 99, 106, 107,

108, 109, 110, 111, 113, 114, 116, 117, 118, 120, 151,

193, 200, 241, 242

Lolium perenne spp. multiflorum 244

longleaf pine 10, 11, 29, 53, 54, 56, 57, 59, 60, 62,

63, 65, 66, 67, 68, 69, 70, 71, 72, 165, 191, 193,

200, 241

Lotus spp. 243

lupine 20, 21, 245

Lupinus spp. 245

Lygodium microphyllum 246

lyonia 70, 94, 241

Lyonia ferruginea 242

Lyonia lucida 240

Lyonia mariana 241

Lyonia spp. 241

M

Magnolia grandiflora 242

Magnolia virginiana 243

Mahonia repens 240

maidencane 85, 89, 172, 245

mandrone

Manitoba maple 35, 240, see boxelder

manzanita 149, 241

marshpennywort see hydrocotyle

melaleuca 68, 162, 166, 168, 169, 170, 172, 241

Melaleuca quinquenervia 241

Melinis minutiflora 245

Melinis repens 245

mesquite 14, 122, 130, 131, 132, 134, 135, 136, 138,

141, 142, 143, 145, 148, 149, 156, 157, 194, 241,

244

milkvetch 21, 139, 245

mockernut hickory 14, 72, 73, 241

molasses grass 170, 245

Morella cerifera 241

Morella spp. 239

Mormon tea 139, 241

mountain alder 241 see thinleaf alder

mountain big sagebrush 146, 190, 195, 241

mountain fern moss 27, 43, 46, 246
254 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Appendices

mountain hemlock 98, 111, 113, 115, 116, 241

mountain-laurel 80, 86, 241

Muhlenbergia 143, 245

Muhlenbergia capillaris 244

Muhlenbergia porteri 244

muhly 245

myrsine 164, 241

Myrsine quianensis 241

myrtle 58, 67, 71, 81, 92, 94, 95, 164, 241

myrtle oak 241

N

natal redtop 170, 245

needlegrass rush 85, 245

Neyrundia reynaudiana 244

noble fir 241

northern pin oak 38, 241

northern red oak 13, 17, 54, 76, 83, 241

northern white-cedar 241

Nuphar spp. 245

Nymphaea odorata 244

Nymphaea spp. 246

Nyssa aquatica 243

Nyssa biflora 243

Nyssa sylvatica 239

O

oak 6, 10, 12, 13, 16, 17, 18, 20, 27, 28, 35, 36, 37,

38, 53, 54, 55, 56, 57, 61, 62, 64, 65, 67, 68, 70, 71,

72, 73, 74, 75, 76, 77, 79, 80, 81, 82, 83, 84, 85, 86, 89,

92, 95, 97, 98, 105, 106, 108, 121, 122, 129, 130, 131,

132, 134, 142, 157, 161, 162, 163, 166, 169, 185, 199,

239, 240, 241, 242, 243

Old world or small-leaf climbing fern 246

oneseed juniper 12, 131, 241

Opuntia humifusa 246

Opuntia spp. 143, 246

Oregon white oak 13, 97, 98, 105, 106, 241

Oryzopsis hymenoides 245

Oxydendrum arboreum 242

P

Pacific madrone 14, 105, 241

Pacific ponderosa pine 98, 241

Pacific silver fir 11, 115, 116, 241

paloverde spp. 122, 138, 141, 143, 148, 156, 157, 241

panicum 72, 89, 245

Panicum hemitomon 245

Panicum spp. 245

Panicum virgatum 245

paper birch 14, 17, 26, 35, 36, 38, 41, 42, 43, 44, 45,

49, 83, 84, 118, 241

Pascopyrum smithii 246

pasque flower 94, 245

Peltandra virginica 244

Pennisetum sataceum 244

Persea borbonia 242

Persea palustris 243

persimmon 14, 67, 72, 73, 142, 240, 241, 243,

Phragmites spp. 245

Picea engelmannii 240

Picea glauca 243

Picea mariana 239

Picea pungens 240

Picea rubens 242

Picea sitchensis 242

pickerelweed 85, 89, 245

pickleweed 85, 245

piedmont staggerbush 80, 241

pignut hickory 14, 73, 241

pili grass 173, 245

pin cherry 96, 242

pin oak 38, 241, 242

pinegrass 22, 29, 245

pineland threeawn 245

Pinus albicaulis 243

Pinus banksiana 241

Pinus caribaea 240

Pinus clausa 242

Pinus contorta 241, 242

Pinus contorta var. contorta 242

Pinus contorta var. latifolia 242

Pinus echinata 242

Pinus edulis 243

Pinus elliottii 242

Pinus glabra 243

Pinus jeffreyi 241

Pinus lambertiana 243

Pinus monophylla 242

Pinus monticola 243

Pinus palustris 241

Pinus ponderosa 239, 241, 242

Pinus ponderosa var. arizonica 239

Pinus ponderosa var. ponderosa 241

Pinus ponderosa var. scopulorum 241

Pinus pungens 243

Pinus resinosa 242

Pinus rigida 242

Pinus sabiniana 240

Pinus serotina 242

Pinus strobus 240

Pinus taeda 241

Pinus virginiana 243

pinyon pine 11, 24, 130, 157, 242

pitch pine 11, 17, 19, 24, 54, 73, 75, 76, 78, 79, 80,

81, 86, 87, 242

pitcherplant 56, 68, 69, 71, 245

Platanus occidentalis 243

Pleuraphis mutica 246

Pleurozium schreberi 246

Poa secunda 245

pond cypress 81, 242

pond-lily 85, 245

pond pine 11, 17, 24, 54, 56, 59, 67, 70, 72, 75, 76,

77, 78, 81, 84, 89, 94, 96, 165, 242

ponderosa pine 4, 10, 11, 15, 24, 25, 27, 30, 35, 97,

98, 100, 101, 102, 104, 105, 106, 107, 112, 113, 118,

121, 122, 124, 125, 126, 127, 128, 130, 151, 185, 187,

191, 192, 193, 194, 200, 241, 242, 247

Pontederia cordata 245
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 255

Appendices

poor man’s soap see sweetpepperbush

pop-ash see Carolina ash

poplar 14, 26, 35, 45, 73, 76, 82, 239, 242, 244

Populus balsamifera 239

Populus deltoides 240

Populus spp. 242

Populus tremuloides 239

Populus trichocarpa 239

post oak 13, 53, 54, 67, 72, 73, 79, 242

prairie cordgrass 85, 245

prairie violet 94, 245

prickly pear cactus 95, 130, see cholla

Prosopis glandulosa 143, 241

Prosopis spp. 241

Prunus emarginata 239

Prunus pensylvanica 242

Prunus serotina 239

Prunus virginiana 240

Pseudoroegneria spicata 244

Pseudotsuga menziesii 240, 242

Pseudotsuga menziesii var. glauca 242

Pseudotsuga menziesii var. menziesii 240

Pteridium aquilinum 246

Ptilium crista-castrensis 246

purple sage see cenzia

Purshia tridentata 239

Q

quaking aspen see aspen

Quercus agrifolia 240

Quercus alba 243

Quercus arizonica 239

Quercus bicolor 243

Quercus chapmanii 240

Quercus chrysolepis 240

Quercus coccinea 242

Quercus douglasii 240

Quercus dumosa 242

Quercus ellipsoidalis 241

Quercus falcata 243

Quercus gambelii 240

Quercus garryana 241

Quercus havardii 242

Quercus ilicifolia 239

Quercus incana 240

Quercus kelloggii 240

Quercus laevis 243

Quercus laurifolia 241

Quercus macrocarpa 240

Quercus marilandica 239

Quercus michauxii 243

Quercus myrtlifolia 241

Quercus nigra 243

Quercus palustris 242

Quercus phellos 243

Quercus prinoides 240

Quercus prinus 240

Quercus rubra 241

Quercus spp. 241

Quercus stellata 242

Quercus stellata var. margaretta 242

Quercus velutina 239

Quercus virginiana 241, 242

Quercus wislizenii 241

R

rabbitbrush 17, 154, 242

raspberry 20, 49, 242, 245

red alder 14, 98, 108, 115, 242

red bay 166, 168, 242

red brome 133, 245

red elderberry 118, 242

red mangrove 172, 242

red maple 14, 36, 37, 44, 70, 72, 73, 74, 76, 80,

81, 83, 84, 86, 91, 94, 168, 169, 242

red pine 5, 10, 11, 35, 36, 37, 38, 39, 40, 41, 83,

193, 242

red raspberry 20, 242

red spruce 36, 54, 83, 84, 86, 242

redberry juniper 131, 242

redstem ceanothus 27, 182, 242

redwood 5, 12, 97, 98, 105, 106, 107, 108, 109, 242

reed 89, 166, 244, 245

Rhizophora mangle 242

rhododendron 65, 74, 75, 242

Rhododendron spp. 242

Rhynchospora tracyi 244

Ribes spp. 241

Rocky Mountain Douglas-fir 11, 242

Rocky Mountain juniper 122, 128, 242

Rocky Mountain lodgepole pine 11, 242

Rocky Mountain maple 242

rose natalgrass see natal redtop

rosemary 89, 95, 242

rough fescue 152, 245

Rubus idaeus 242

Rubus parviflorus 243

Rubus spectabilis 242

Rubus spp. 245

Rudbeckia hirta 244

rush 84, 85, 88, 90, 172, 245

rusty staggerbush 95, 242

S

Sabal etonia 242

Sabal palmetto 240

sacaton 245 see dropseed

sagebrush 31, 121, 122, 130, 133, 134, 138, 139,

142, 143, 146, 148, 151, 152, 153, 154, 155, 190, 192, 195,

199, 201, 239, 240, 241, 242, 243

Sagittaria lancifolia 244

Sagittaria spp. 244

saguaro 141, 157, 246

Salicornia spp. 245

Salix bebbiana 239

Salix spp. 243

salmonberry 28, 115, 242

saltgrass 85, 88, 139, 172, 245

saltmeadow cordgrass 85, 88, 91, 245

saltmeadow rush 85, 245

Salvia apiana 243
256 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000

Appendices

Sambucus racemosa ssp. pubens 242

sand cordgrass 172, 245

sand dropseed 23, 245

sand live oak 95, 242

sand pine 11, 54, 67, 71, 84, 86, 87, 89, 90, 95, 242

sand post oak 67, 242

sand shinnery oak 129, 242

Sandberg bluegrass 152, 245

Sapium sebiferum 244

Sarcobatus vermiculatus 143, 239

Sarracenia flava 246

Sarracenia purpurea 245

Saskatoon serviceberry 242

saw palmetto 58, 65, 67, 71, 81, 95, 164, 169, 242

sawgrass 88, 171, 172, 245

scarlet oak 79, 86, 242

Schinus terebinthifolus 240

Schizachyrium condensatum 244

Schizachyrium scoparium 245

Schizachyrium spp. 244

Schizachyrium tenerum 245

Schoenoplectus americanus 244

Schreber’s moss 43, 246

scrub oak 27, 54, 75, 81, 89, 242

scrub palmetto 95, 242

seaside tansy 172, 245

Sequoia sempervirens 242

Sequoiadendron giganteum 240

Serenoa repens 242

shadscale 139, 242

sheathed cottonsedge 45, 47, 245

sheep-laurel 242

shore pine 11, 108, 115, 242

shortleaf pine 11, 14, 17, 54, 59, 63, 67, 72, 73, 79,

80, 242

showy aster 20, 21, 29, 245

showy partridgepea 27, 245

sideoats grama 29, 141, 245

Sideroxylon salicifolium 164, 243

silver maple 10, 54, 242

silver sagebrush 154, 242

silverbell 76, 242

singleleaf pinyon 128, 242

Sisyrinchium campestre 244

Sitka spruce 12, 98, 108, 115, 120, 242

sky blue aster 94, 245

slash pine 10, 11, 54, 62, 63, 67, 68, 69, 70, 71, 72,

76, 81, 94, 95, 161, 162, 165, 169, 242

slender bluestem 67, 245

small-leaf climbing fern see old world fern

Smilacina stellata 245

Smilax glauca, Smilax spp. 241

smooth cordgrass 84, 85, 90, 245

snakeweed 154, 157, 242

snowbrush ceanothus 24, 242

Sorbus americana 239

Sorghastrum nutans 245

sourwood 86, 242

southern bayberry see myrtle

southern magnolia 14, 67, 72, 76, 242

southern red oak 13, 72, 76, 243

Spartina alterniflora 245

Spartina bakeri 245

Spartina patens 245

Spartina pectinata 245

Spartina spartinae 244

Spartina spp. 244

speckled alder 35, 45, 243

sphagnum 38, 43, 46, 48, 50, 70, 94, 246

Sphagnum spp. 246

spikerush 85, 89, 171, 245

spiny hopsage 139, 243

Spiraea betulifolia 243

Sporobolus cryptandrus 245

Sporobolus curtissii 244

Sporobolus spp. 130, 245

spruce pine 243

star-flowered Solomon’s seal 20, 245

strawberry 20, 245

subalpine fir 10, 11, 43, 99, 107, 111, 112, 116, 119,

193, 243

sugar maple 14, 35, 44, 49, 55, 74, 76, 82, 83, 243

sugar pine 11, 243

sugarberry 54, 76, 243

swamp bay 70, 243

swamp chestnut oak 72, 243

swamp cyrilla see cyrilla

swamp tupelo 70, 72, 76, 81, 91, 94, 243

sweetbay 76, 80, 81, 94, 168, 243

sweetgum 14, 54, 67, 70, 73, 76, 81, 94, 243

sweetpepperbush 70, 81, 94, 243

switch cane see cane

switchgrass 85, 89, 92, 130, 245

sycamore 54, 76, 243

T

Table Mountain pine 27, 53, 54, 78, 79, 84, 86, 87, 95,

96, 243

tamarack 12, 36, 42, 45, 50, 243

tanoak 14, 17, 98, 105, 243

tarbush 122, 141, 157, 243

Taxodium ascendens 242

Taxodium distichum 239

Taxodium spp. 240

Tetradymia canescens 241

Tetradymia spp. 241

Texas persimmon 142, 243

thatching grass 170, 245

thimbleberry 18, 243

thinleaf alder 92, 243

threetip sagebrush 146, 154, 243

Thuja occidentalis 241

Thuja plicata 243

Thurber’s needlegrass 22, 245

Tilia americana 239

tobosa 122, 130, 138, 141, 142, 143, 148, 157, 246

toyon 27, 243

Tracy’s beaksedge see beakrush

trefoil 27, 243

trembling aspen see aspen

true pinyon 128, 243

trumpet pitcherplant 69, 246
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-42-vol. 2. 2000 257

Appendices

Tsuga canadensis 240

Tsuga heterophylla 243

Tsuga mertensiana 241

turkey oak 13, 17, 67, 95, 243

twinflower 20, 246

Typha spp. 244

U

Ulmus americana 239

Utah juniper 12, 128, 132, 243

V

Vaccinium membranaceum 240

Vaccinium myrsinites 241

Vaccinium uliginosum 240

varnish leaf 164, 243

Viola pedatifida 245

Virginia pine 11, 54, 63, 67, 72, 73, 75, 76, 77, 78,

79, 81, 243

W

water oak 70, 72, 243

water tupelo 76, 91, 243

waterlily 85, 171, 244, 246

wax myrtle see myrtle

western hemlock 105, 107, 108, 115, 116, 118, 120,

243

western juniper 122, 129, 130, 131, 133, 134, 243

western larch 10, 15, 25, 27, 100, 106, 107, 110, 116,

118, 243

western redcedar 113, 118, 243

western wheatgrass 22, 150, 246

western white pine 10, 108, 114, 115, 116, 118, 119,

243

western yarrow 20, 246

wheatgrass 22, 122, 139, 150, 152, 154, 244, 246

white ash 14, 81, 83, 243

white bully 164, 243

white bursage 156, 243

white elm 35 see American elm

white fir 11, 99, 100, 107, 110, 116, 243

white mangrove 172, 243

white oak 13, 54, 72, 76, 79, 81, 97, 98, 105, 106,

129, 239, 241, 243

white sage 17, 243

white spirea 18, 243

white spruce 12, 25, 36, 37, 41, 42, 43, 44, 46, 47,

48, 49, 86, 119, 243

whitebark pine 11, 99, 106, 111, 112, 113, 118, 119,

182, 189, 193, 199, 243, 247

wild columbine 20, 246

wild hollyhock 27, 31, 246

wildrice 85, 246

willow 17, 24, 26, 41, 45, 72, 89, 94, 110, 116, 118,

164, 168, 170, 172, 239, 243

willow oak 72, 243

winterfat 139, 243

wiregrass 22, 27, 29, 60, 63, 64, 65, 67, 68, 95, 164,

245, 246 see pineland threeawn

Wyoming big sagebrush 142, 146, 243

Y

yaupon 71, 243

yellow birch 55, 76, 78, 83, 84, 96, 243

yellow buckeye 76, 244

yellow paloverde 244

yellow pitcherplant see trumpet pitcherplant

yellow-poplar 14, 244

Yucca brevifolia 143, 241

Z

Zigadenus venenosus 244

Zizania spp. 246

Zizaniopsis spp. 244
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