United States

Department

of Agriculture

Forest Service

Rocky Mountain

Research Station

General Technical

Report RMRS-GTR-202

November 2007

Fire Ecology and Management of the 

Major Ecosystems of Southern Utah

Sharon M. Hood and Melanie Miller, Editors
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Hood, Sharon M.; Miller, Melanie, editors. 2007. Fire ecology and management of the major 

ecosystems of southern Utah. Gen. Tech. Rep. RMRS-GTR-202. Fort Collins, CO: U.S. 

Department of Agriculture, Forest Service, Rocky Mountain Research Station. 110 p.

Abstract

This document provides managers with a literature synthesis of the historical conditions, 

current conditions, fire regime condition classes (FRCC), and recommended treatments for the 

major ecosystems in southern Utah. Sections are by ecosystems and include: 1) coniferous 

forests (ponderosa pine, mixed conifer, and Engelmann spruce-subalpine fir), 2) aspen, 3) 

pinyon-juniper, 4) big and black sagebrush, and 5) desert shrubs (creosotebush, blackbrush, 

and interior chaparral). Southern Utah is at the ecological crossroads for much of the western United States. It contains steep environmental gradients and a broad range of fuels and 

fire regimes associated with vegetation types representative of the Rocky Mountains, the 

Great Basin, Northern Arizona and New Mexico, and the Mohave Desert. The Southern Utah 

Demonstration Area consists of contiguous state and federal lands within the administrative 

boundaries of the Bureau of Land Management (BLM), Fishlake and Dixie National Forests, 

National Park Sevice, and State of Utah, roughly encompassing the southern 15 percent of 

Utah (3.24 million ha). The vegetation types described are similar in species composition, 

stand structure, and ecologic function, including fire regime to vegetation types found on 

hundreds of millions of hectares in the 11 western states. 

Key Words: Fire regime condition class, disturbance, fire ecology, fuel treatment 

Acknowledgments

The Joint Fire Science Program (#99-1-3-29) and the Fire Modeling Institute of the Rocky 

Mountain Research Station provided funding for this work. Kevin Ryan first developed the 

idea of this literature synthesis. The maps were created by Don Long and Heather Kreilick. 

We thank Matthew Rollins and Steve Barrett for their reviews of the fire regime condition 

classes sections. The reviews provided by Curt Deuser, Sherel Goodrich, Robert Keane, 

Ward McCaughey, Walt Mueggler, Doug Page, Jolie Pollet, Louis Provencher, Tom Roberts, 

Bruce Roundy, Steve Sutherland, Wayne Shepperd, and Bruce Welch greatly improved the 

publication. 
Authors

Sharon M. Hood is a Forester with the Rocky Mountain Research Station at the Fire Sciences Laboratory in Missoula, 

Montana. She completed a B.S. in forestry from Mississippi 

State University and a M.S. in forestry from Virginia Polytechnic 

and State University. 

Melanie Miller is a Fire Ecologist with the Bureau of Land 

Management, Fire and Aviation Directorate (Boise), National 

Interagency Fire Center. She is stationed at the Rocky Mountain 

Research Station Fire Sciences Laboratory, Missoula, Montana. 

She completed a B.S. Honours Degree in physical geography 

at the University of Calgary in 1972, and a M.S. in forest fire 

science at the University of Montana in 1976.

Donald Long is a Fire Ecologist currently working on many of 

the technical aspects of the vegetation mapping and modeling 

effort for the LANDFIRE project at the Fire Sciences Lab of the 

Rocky Mountain Research Station in Missoula, Montana. He 

has worked on a variety of research projects concerned with 

ecosystem dynamics, fuel and vegetation inventory and mapping, and fire behavior and effects. He earned a B.S. degree 

in forest science from the University of Montana in 1981 and a 

M.S. in forest resources at the University of Idaho in 1998.

Kevin C. Ryan is a Program Manager with the Rocky Mountain 

Research Station at the Missoula Fire Sciences Laboratory in 

Missoula, Montana. He completed a B.S. in forest biology at the 

Colorado State University in 1973 and a M.S. in forest ecology 

in 1976, also from Colorado State University. He received a 

PhD. in forest ecology from the University of Montana in 1993. 

Kevin began his career in fire research at the Rocky Mountain 

Research Station in Fort Collins, Colorado, in 1976. Between 

1976 and 1979 he was a Research Forester with the Pacific 

Northwest Research Station in Seattle, Washington. He has 

been conducting and managing fire research projects at the 

Missoula Fire Sciences Laboratory since 1979.

Mike Battaglia is an Ecologist with the Rocky Mountain 

Research Station in Fort Collins, Colorado. He completed a 

B.S. in biology at the University of South Carolina, an M.S. in 

forestry at Virginia Tech, and a Ph.D in silviculture/fire ecology 

at Colorado State University.

Wayne D. Shepperd retired as a Research Silviculturist, at 

the U.S. Forest Service, Rocky Mountain Research Station 

in Fort Collins, Colorado, in January 2007. He holds a B.S. in 

outdoor recreation, and a M.S. and Ph.D. in silviculture from 

Colorado State University and is a recognized expert on the 

ecology, growth, and management of Rocky Mountain forests. 

His research included fire, fuels, vegetation, and management 

studies throughout the west.

Dale Bartos is an Aspen Ecologist with the Rocky Mountain 

Research Station at the Forestry Sciences Laboratory in Logan, 

Utah. He completed a B.S. in general science, a M.S. in botany 

from Fort Hays State University, and a Ph.D. in range science 

at Colorado State University.

Robin J. Tausch is a Research Range Ecologist for the Rocky 

Mountain Research Station in Reno, Nevada. Dr. Tausch has 

a B.S. in wildlife management from Humboldt State University, 

a M.S. in range management from the University of Nevada, 

Reno, and a Ph.D. in range ecology from Utah State University. 

His 37 years of research experience in the Great Basin has 

focused on the ecology and paleoecology of pinyon/juniper 

woodlands and the associated sagebrush ecosystems of which 

they are a part.

Stanley G. Kitchen is a Research Botanist for the Rocky 

Mountain Research Station at the Shrub Sciences Laboratory 

in Provo, Utah. He holds a B.S. from Utah State University and 

a M.S. from Brigham Young University. His research interests 

include fire ecology, disturbance and restoration ecology of 

shrublands, seed biology, and plant materials development.

E. Durant McArthur is Program Manager for the Rocky 

Mountain Research Station, Grassland, Shrubland, and Desert 

Ecosystem Research Program and is located at the Shrub Sciences Laboratory, Provo, Utah. A Forest Service scientist since 

1972, much of his research experience has been in shrubland 

ecosystem genetics and ecology. He has a B.S., M.S., and 

Ph.D. in genetics, botany, and biology from the University of 

Utah and postdoctoral experience at the University of Leeds, 

United Kingdom.

Matt Brooks is a Research Botanist with the U.S. Geological 

Survey, Western Ecological Research Center. He completed 

a B.S. in biology at U.C. Irvine, an M.A. in biology at C.S.U. 

Fresno, and a Ph.D. in biology at U.C. Riverside. His research 

focus is on the ecology and management of fire and non-native 

invasive plants in western North America.

Todd Esque is a Research Ecologist with the U.S. Geological Survey, Western Ecological Research Center, Las Vegas 

Field Station. His primary research focus is plant and animal 

responses to disturbances such as fire, invasive species, and 

urbanization.

Tim Duck is the Fuels Program Manager for the Arizona Strip 

District of the Bureau of Land Management in St. George, Utah. 

He completed his B.S. in ecology and evolutionary biology at 

the University of Arizona, and has worked as a Wildlife Biologist, 

Ecologist, Wildland Firefighter, and Prescribed Fire Specialist 

for the Bureau of Land Management since 1980.

i

Editors

Sharon M. Hood is a Forester with the Rocky Mountain Research Station at the Fire Sciences Laboratory in Missoula, Montana. 

She completed a B.S. in forestry from Mississippi State University and a M.S. in forestry from Virginia Polytechnic and State 

University. 

Melanie Miller is a Fire Ecologist with the Bureau of Land Management, Fire and Aviation Directorate (Boise), National 

Interagency Fire Center. She is stationed at the Rocky Mountain Research Station Fire Sciences Laboratory, Missoula, Montana. 

She completed a B.S. Honours Degree in physical geography at the University of Calgary in 1972, and a M.S. in forest fire 

science at the University of Montana in 1976. 

Fire Ecology and Management of the 

Major Ecosystems of Southern Utah

Sharon M. Hood and Melanie Miller, Editors

iii

chapter 1 – Introduction.......................................................................................... 1

Background ..........................................................................................................................1

Project Objectives.................................................................................................................1

Study Area............................................................................................................................2

Fire Regime Conditions Classes ..........................................................................................2

Management Implications.....................................................................................................3

References ...........................................................................................................................5

chapter 2 – Ponderosa pine, mixed conifer, and spruce-fir Forests .. 7

Introduction...........................................................................................................................7

Ponderosa Pine Forests .................................................................................................... 7

Historical Conditions.............................................................................................................7 

Current Conditions...................................................................................................9

Current Disturbances.............................................................................................12

 Fire .................................................................................................................12

 Insects and Disease.......................................................................................12

 Ungulates .......................................................................................................12

 Exotic Weeds .................................................................................................13

Fire Regime Condition Classes..........................................................................................13

Fire Regime Condition Class 1 (FRCC 1) .............................................................13

Fire Regime Condition Class 2 (FRCC 2) .............................................................13

Fire Regime Condition Class 3 (FRCC 3) .............................................................13

Recommended Treatments ................................................................................................14

Mixed Conifer Forests...................................................................................................... 16

Historical Conditions...........................................................................................................16

Current Conditions.................................................................................................18

Current Disturbances.............................................................................................20

 Fire .................................................................................................................20

 Insects and Disease.......................................................................................20

 Ungulates .......................................................................................................21

 Selective Harvesting and Fire Exclusion........................................................21

Fire Regime Condition Classes..........................................................................................21

Fire Regime Condition Class 1 (FRCC 1) .............................................................21

Fire Regime Condition Class 2 (FRCC 2) .............................................................21

Fire Regime Condition Class 3 (FRCC 3) .............................................................21

Recommended Treatments ................................................................................................22

Engelmann Spruce and Subalpine Fir Forests.............................................................. 23

Historical Conditions...........................................................................................................24

Current Conditions.................................................................................................25

Current Disturbances.............................................................................................26

 Ungulates .......................................................................................................27

 Fire .................................................................................................................27

 Insects and Disease.......................................................................................27

Fire Regime Condition Classes..........................................................................................28

Fire Regime Condition Class 1 (FRCC 1) .............................................................28

Fire Regime Condition Classes 2 and 3 (FRCC 2 and FRCC 3)...........................29

Contents

v
Recommended Treatments ................................................................................................29

Spruce-fir Treatments at the Landscape Scale .....................................................30

Spruce-fir Treatments at the Stand Scale..............................................................31

Summary ............................................................................................................................31

References .........................................................................................................................32

chapter 3 – Aspen....................................................................................................... 39

Introduction.........................................................................................................................39

Historical Conditions...........................................................................................................41

Current Conditions..............................................................................................................41

Ungulates ..............................................................................................................42

Insects and Disease ..............................................................................................43

Fire Exclusion........................................................................................................43

Ecosystem Impacts Caused by Aspen Loss..........................................................43

Fire Regime Condition Classes..........................................................................................46

Fire Regime Condition Class 1 (FRCC 1) .............................................................46

Fire Regime Condition Class 2 (FRCC 2) .............................................................46

Fire Regime Condition Class 3 (FRCC 3) .............................................................48

Recommended Treatments ................................................................................................48

Doing Nothing........................................................................................................50

Commercial Harvest..............................................................................................50

Fire ........................................................................................................................50

Mechanical Root Stimulation.................................................................................51

Removal of Vegetative Competition ......................................................................51

Protection of Regeneration from Herbivory ...........................................................51

Regenerating from Seed .......................................................................................52

References .........................................................................................................................52

Chapter 4 – Pinyon/Juniper WoodlandS............................................................ 57

Introduction.........................................................................................................................57

Historical Conditions...........................................................................................................58

Current Conditions..............................................................................................................59

Fire Regimes ......................................................................................................................63

Fire Regime Condition Classes..........................................................................................64

Recommended Treatments ................................................................................................65

Tree Removal ........................................................................................................65

Prescribed Fire ......................................................................................................66

Mechanical Thinning..............................................................................................66

Seeding .................................................................................................................67

Herbicides..............................................................................................................67

Summary ............................................................................................................................67

References .........................................................................................................................68

Chapter 5 – Big and Black Sagebrush Landscapes..................................... 73

Introduction.........................................................................................................................73

Sagebrush Biology and Reproduction...................................................................74

Geographic Distribution.........................................................................................75

Historical Conditions...........................................................................................................75

Ecological Distribution and Associated Species ....................................................76

Fire Effects ............................................................................................................79

Big and Black Sagebrush Fire Regimes................................................................80

Euro-American Settlement ....................................................................................83

vi
Current Conditions..............................................................................................................84

Ungulate Impacts...................................................................................................84

Exotic Weed Impacts.............................................................................................84

Fire Regime Condition Classes..........................................................................................84

Fire Regime Condition Class 1 (FRCC 1) .............................................................85

Fire Regime Condition Class 2 (FRCC 2) .............................................................85

Fire Regime Condition Class 3 (FRCC 3) .............................................................85

Fire Regime Condition Class Assessment ............................................................86

Recommended Treatments ................................................................................................87

Prescribed Fire ......................................................................................................88

Herbicides..............................................................................................................88

Mechanical Treatments .........................................................................................89

Restoration Plantings ............................................................................................89

References .........................................................................................................................90 

Chapter 6 – Creosotebush, Blackbrush, and Interior

 Chaparral Shrublands ........................................................................ 97

Introduction.........................................................................................................................97

Creosotebush Scrub ........................................................................................................ 97

Introduction.........................................................................................................................97

Historical Conditions...........................................................................................................99

Current Conditions............................................................................................................100

Fire Regime Condition Classes........................................................................................101

Fire Regime Condition Class 1 (FRCC 1) ...........................................................101

Fire Regime Condition Class 2 (FRCC 2) ...........................................................101

Fire Regime Condition Class 3 (FRCC 3) ...........................................................101

Recommended Treatments ..............................................................................................101

Blackbrush...................................................................................................................... 102

Introduction.......................................................................................................................102

Historical Conditions.........................................................................................................103

Current Conditions............................................................................................................104

Fire Regime Condition Classes........................................................................................104 

Fire Regime Condition Class 1 (FRCC 1) ...........................................................104

Fire Regime Condition Class 2 (FRCC 2) ...........................................................105

Fire Regime Condition Class 3 (FRCC 3) ...........................................................105

Recommended Treatments ..............................................................................................105

Interior Chaparral ........................................................................................................... 106

Introduction.......................................................................................................................106

Historical Conditions.........................................................................................................107

Current Conditions............................................................................................................107

Fire Regime Condition Classes........................................................................................108

Fire Regime Condition Class 1 (FRCC 1) ...........................................................108

Fire Regime Condition Class 2 (FRCC 2) ...........................................................108

Fire Regime Condition Class 3 (FRCC 3) ...........................................................108

Recommended Treatments ..............................................................................................108

Acknowledgments ............................................................................................................109

References .......................................................................................................................109

vii

USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 1

Background

Many areas throughout the United States are facing the 

triple threat of increasing fire severity, residential growth 

in areas prone to wildland fire, and suppression costs 

and losses. In addition, substantial changes are occurring in the way we plan and implement management on 

federal lands relative to use of wildland fire, prescribed 

fire, and mechanical fuel management. Past emphasis 

in fire management has been on wildfire suppression 

and prescribed fire in support of other resources such as 

hazard reduction and site preparation in harvested areas 

and wildlife habitat improvement. Federal financial support has only recently supported the large-scale use of 

prescribed burning and mechanical fuels treatments to 

reduce unnatural fuel accumulations in non-wilderness 

areas.

The Southern Utah Fuel Management Demonstration 

Project was an effort to develop, evaluate, and compare 

methods to incorporate wildland fuels management into 

landscape scale land use planning processes for approximately 5.3 million ha (13 million acres) of southern 

Utah and 0.8 million ha (2 million acres) of northern 

Arizona (Ryan and Long 2004). This area was chosen 

because it is at an ecological crossroads for much of the 

western United States. It contains steep environmental 

gradients and a broad range of fuels and fire regimes 

associated with vegetation types representative of the 

Rocky Mountains, the Great Basin, and the Mohave 

Desert. The project developed GIS data layers for fuels, 

vegetation, and terrain that provided the inputs necessary 

to conduct fire behavior, fire effects, and succession 

modeling analyses. Several fuel treatments were also 

implemented at a variety of scales in the project area 

to demonstrate the use of fire behavior models for fuel 

treatment planning (Long and others 2003; Mathews 

2003; Stratton 2004).

Chapter 1

Introduction

Sharon M. Hood, Donald Long, Melanie Miller, and Kevin C. Ryan

Project Objectives

This literature synthesis provides an ecological context for 

the larger Southern Utah Fuel Management Demonstration 

Project. The synthesis reviews the pertinent literature to address 1) historical conditions, 2) current conditions, 3) fire 

regime condition classes, and 4) recommended treatments 

for each of the major ecosystems found in the Southern 

Utah study area. Sections are by ecosystems and include: 1) 

coniferous forests (ponderosa pine, mixed conifer, and Engelmann spruce-subalpine fir), 2) aspen, 3) pinyon-juniper, 4) 

big and black sagebrush, and 5) desert shrubs (creosotebush, 

blackbrush, and interior chaparral).

Historical conditions are described in terms of characteristic species composition, structure, size class, age 

distribution, and fuel complexes that existed in presettlement times. In addition, disturbance frequency, 

size, and severity of historical disturbance regimes are 

discussed. Authors also address the ways disturbance 

historically affected community characteristics and 

position on landscape.

Current conditions of each focus ecosystem are then 

compared to historical. Current fuel condition descriptions may include additional information on amount 

of downed wood, amount of live and dead shrub and 

herbaceous fuels, annual production, range of annual 

variation caused by weather, and production relative 

to dominant life form. Where appropriate, authors 

also address changes in hydrologic function, such as 

streamflow, water yield, sediment production, and the 

potential effects of fire. Current disturbances are also 

described, including fire and fire suppression, insects, 

disease, ungulates (domestic and wildlife), exotic plants, 

patterns and frequencies of disturbances, as well as the 

resilience of each ecosystem to disturbance.

Each chapter includes a Fire Regime Condition Class 

(FRCC) (Schmidt and others 2002) assessment of the 
2 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 1—Introduction

ecosystem with descriptions of each condition class’s 

characteristics within that particular ecosystem and an 

estimate of the area currently categorized as FRCC 1, 2, 

or 3. Authors address the key components that are at risk 

of being lost if a wildfire were to occur under current 

conditions. Moreover, authors recommend treatments 

for each ecosystem by condition class and describe 

treatments in terms of timing and season, the scale of 

the treatment, and pre- and post-treatment management 

considerations.

Study Area

The Southern Utah Demonstration Area roughly encompasses the southern 15 percent of Utah and consists 

of contiguous state and federal lands within the administrative boundaries of the Bureau of Land Management 

(BLM), Fishlake and Dixie National Forests, National 

Park Service, and state of Utah (fig. 1). Dominant vegetation types found in the demonstration area include 

various associations of pinyon-juniper, ponderosa pine, 

sagebrush-grass, aspen, spruce-fir, interior chaparral, 

and desert shrubs (fig. 2). These vegetation types are 

similar in species composition, stand structure, and 

ecologic function to vegetation types found on hundreds 

of millions of hectares in western United States.

Fire Regime Conditions Classes

A fire regime condition class (FRCC) is a classification 

of the amount of departure from the historical natural 

fire regime (Hann and Bunnell 2001). Coarse-scale 

FRCC classes were first defined and mapped by Hardy 

and others (2001) and Schmidt and others (2002). They 

are a metric for reporting the number of hectares in need 

of hazardous fuel reduction and evaluating the efficacy 

of wildland fuel treatment projects (Rollins and others 

2006). This departure results in changes to one or more 

of the following ecological components: vegetation 

characteristics (species composition, structural stages, 

stand age, canopy closure, and mosaic pattern); fuel 

composition; fire frequency, severity, and pattern; and 

other associated disturbances (for example, insect and 

diseased mortality, grazing, and drought) (Hann and 

others 2004).

FRCC stratifies three condition classes by five natural 

historical fire regimes. A natural fire regime is a general classification of the role fire would play across a Figure 2—Major vegetation groups in southern Utah 

project area.

Figure 1—Southern Utah Fuel Management Demonstration 

Area location.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 3

Chapter 1—Introduction

landscape in the absence of modern human mechanical 

intervention, but includes the influence of aboriginal 

burning. The five natural historical fire regimes are 

classified based on average number of years between 

fires (fire frequency) combined with the severity (amount 

of replacement) of the fire on the dominant overstory 

vegetation (table 1) (Hann and others 2004).

The three conditions classes are based on low (FRCC 1), 

moderate (FRCC 2), and high (FRCC 3) departure from 

the central tendency of the natural historical fire regime 

(table 2) (Hann and Bunnell 2001; Hardy and others 2001; 

Schmidt and others 2002). Low departure (FRCC 1) 

describes fire regimes and successional status operating 

within the historical range of variability. FRCC 2 and 

FRCC 3 characterize conditions outside the historical 

range (Rollins and others 2006). Characteristic vegetation and fuel conditions are those that occurred within the 

natural fire regime. Uncharacteristic conditions are those 

that did not occur within the natural fire regime, such 

as invasive species, “high graded” forest composition 

and structure, or repeated annual grazing that maintains 

grassy fuels across relatively large areas at levels that 

will not carry a surface fire. Determination of amount 

of departure is based on comparing a composite measure 

of fire regime attributes (vegetation characteristics; fuel 

composition; fire frequency, severity, and pattern) to the 

central tendency of the natural (historical) fire regime. 

The amount of departure is then classified to determine 

the fire regime condition class (table 3) (Hann and others 2004). Additional FRCC information can be found 

at http://www.frcc.gov/ (Hann and others 2003).

Management Implications

This literature synthesis provides land managers and 

planners in southern Utah and surrounding areas with the 

current state of knowledge of the dominant ecosystems in 

Table 1—Natural (historical) fire regime classes from Hann and Bunnell (2001) for modeling landscape dynamics in the lower 

48 States. Historical Range of Variability (HRV)–the variability of regional or landscape composition, structure, and 

disturbances during a period of time of several cycles of the common disturbance intervals and similar environmental 

gradients prior to extensive agricultural or industrial development.

Fire Frequency

regime (Fire return

class interval, in years) Severity Modeling assumptions

I Frequent (0 to 35) Low Open forest or savannah structures maintained by frequent fire; 

also includes frequent mixed severity fires that create a mosaic 

of different age post-fire open forest, early to mid-seral forest 

structural stages, and shrub or herb dominated patches (generally < 40 ha [100 acres]).

II Frequent (0 to 35) Stand replacement Shrub or grasslands maintained or cycled by frequent fire; fires 

kill non-sprouting shrubs such as sagebrush, which typically 

regenerate and become dominant within 10 to 15 years; fires 

remove tops of sprouting shrubs such as mesquite and chaparral, which typically resprout and dominate within 5 years; fires 

typically kill most tree regeneration such as juniper, pinyon 

pine, ponderosa pine, Douglas-fir, or lodgepole pine.

III Less frequent (35 to 100) Mixed Mosaic of different age post-fire open forest, early to mid-seral 

forest structural stages, and shrub or herb dominated patches 

(generally < 40 ha [100 acres]) maintained or cycled by infrequent fire.

IV Less frequent (35 to 100) Stand replacement Large patches (generally > 40 ha [100 acres]) of similar age 

post-fire shrub or herb dominated structures, or early to midseral forest cycled by infrequent fire.

V Infrequent (> 100) Stand replacement Large patches (generally > 40 ha [100 acres]) of similar age

 post-fire shrub or herb dominated structures, or early to mid to

 late seral forest cycled by infrequent fire.
4 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 1—Introduction

Table 2—Condition classes from Hann and Bunnell (2001) for modeling landscape dynamics and departure from historical 

natural range of variability for the lower 48 States. Historical Range of Variability (HRV)–the variability of regional or 

landscape composition, structure, and disturbances, during a period of time of several cycles of the common disturbance intervals and similar environmental gradients prior to extensive agricultural or industrial development.

Condition Departure

class from HRV Description

Class 1 None, minimal, low Vegetation composition, structure, and fuels are similar to those of the historic regime 

and do not pre-dispose the system to risk of loss of key ecosystem components. Wildland fires are characteristic of the historical fire regime behavior, severity, and patterns. 

Disturbance agents, native species habitats, and hydrologic functions are within the 

historical range of variability. Smoke production potential is low in volume.

Class 2 Moderate Vegetation composition, structure, and fuels have moderate departure from the historic 

regime and predispose the system to risk of loss of key ecosystem components. 

Wildland fires are moderately uncharacteristic compared to the historical fire regime 

behaviors, severity, and patterns. Disturbance agents, native species habitats, and 

hydrologic functions are outside the historical range of variability. Smoke production 

potential has increased moderately in volume and duration.

Class 3 High Vegetation composition, structure, and fuels have high departure from the historic 

regime and predispose the system to high risk of loss of key ecosystem components. 

Wildland fires are highly uncharacteristic compared to the historical fire regime behaviors, 

severity, and patterns. Disturbance agents, native species habitats, and hydrologic 

functions are substantially outside the historical range of variability. Smoke production 

potential has increased with risks of high volume production of long duration.

Table 3—A simplified description of the fire regime condition classes and associated potential risks (Hann and others 2003).

Fire regime

condition class Description Potential risks

Condition Class 1 Within the natural Fire behavior, effects, and other associated disturbances are 

(historical) range of similar to those that occurred prior to fire exclusion (suppression)

variability of vegetation and other types of management that do not mimic the natural fire

characteristics; fuel regime and associated vegetation and fuel characteristics.

composition; fire Composition and structure of vegetation and fuels are similar to the natural

frequency, severity and (historical) regime.

pattern; and other Risk of loss of key ecosystem components (for example, native species, 

associated disturbances. large trees, and soil) is low.

Condition Class 2 Moderate departure from Fire behavior, effects, and other associated disturbances are moderately

the natural (historical) departed (more or less severe).

regime of vegetation Composition and structure of vegetation and fuel are moderately altered.

characteristics; fuel Uncharacteristic conditions range from low to moderate.

composition; fire Risk of loss of key ecosystem components is moderate.

frequency, severity and

pattern; and other

associated disturbances.

Condition Class 3 High departure from the Fire behavior, effects, and other associated disturbances are highly

natural (historical) regime departed (more or less severe).

of vegetation characteristics; Composition and structure of vegetation and fuel are highly altered.

fuel composition; fire Uncharacteristic conditions range from moderate to high.

frequency, severity and Risk of loss of key ecosystem components is high.

pattern; and other

associated disturbances.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 5

Chapter 1—Introduction

the area. The review of historical and current conditions 

highlights how post-European settlement has changed the 

southern Utah landscape and the problems facing land 

managers today. The FRCC sections are intended as a 

general overview of ecosystems conditions. They should 

not be used to determine actual fire regime conditions 

in a given area. The Interagency Fire Regime Condition Class Guidebook and other tools are designed for 

assigning specific fire regime condition classes (Hann 

and others 2004, www.frcc.gov).

The recommended treatment sections include the commonly used methods to treat areas for fuel accumulation, 

exotic weed control, and other objectives. New methods, 

such as mastication, are always emerging to provide 

managers with more treatment alternatives. Research 

of these treatments will be necessary to improve our 

understanding of their effects on the ecosystems. With an 

understanding of the ecosystems and their responses to 

treatments and accurate spatial data on fuels, vegetation, 

fire regimes, and values, we can develop collaborative 

strategies for managing fuels in southern Utah on a 

landscape basis.

References

Hann, W. J.; Bunnell, D. L. 2001. Fire and land management planning and implementation across multiple scales. International 

Journal of Wildland Fire. 10: 389-403.

Hann, W.; Havlina, D.; Shlisky, A. 2003. Interagency and The 

Nature Conservancy fire regime condition class website. U.S. 

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

of the Interior, The Nature Conservancy, and Systems for Environmental Management [Online]. Available: http://www.frcc.

gov/ [2007 February 20].

Hann, W.; Shlisky, A.; Havlina, D.; Schon, K.; Barrett, S.; DeMeo, 

T.; Pohl, K.; Menakis, J. P.; Hamilton, D.; Jones, J.; Levesque, M. 

2004. Interagency Fire Regime Condition Class Guidebook [Online.] Available: http://www.frcc.gov/ [2006 May 12]. 120 p.

Hardy, C. C.; Schmidt, K. M.; Menakis, J. P.; Sampson, R. N. 2001. 

Spatial data for national fire planning and fuel management. 

International Journal of Wildland Fire. 10: 353-372.

Long, Donald; Ryan, Kevin; Stratton, Rick; Mathews, Ed; Scott, 

Joe; Mislivets, Maureen; Miller, Melanie; Hood, Sharon. 2003. 

Modeling the effects of fuel treatments for the Southern Utah Fuel 

Management Project. In: Omi, Philip N.; Joyce, Linda A., tech. 

eds. Fire, fuel treatments, and ecological restoration: conference 

proceedings; 2002 16-18 April; Fort Collins, CO. Proc. RMRSP-29. Fort Collins, CO: U.S. Department of Agriculture, Forest 

Service, Rocky Mountain Research Station: 387-395.

Mathews, E. 2003. Use of FVS/FFE in the Duck Creek UT fuel 

management demonstration area. In: Proceedings of the second 

international wildland fire ecology and fire management congress 

and fifth symposium on fire and forest meteorology. American 

Meteorological Society: Orlando, FL. [Online]. Available: http://

www.firelab.org/index.php?option=com_content&task=blogcate

gory&id=96&Itemid=99 [2007, February 20]. Abstract.

Rollins, M. G.; Keane, R. E.; Zhu, Z.; Menakis, J. P. 2006. An 

overview of the LANDFIRE prototype project. In: Rollins, M. G.; 

Frame, C., tech. eds. The LANDFIRE prototype project: nationally 

consistent and locally relevant geospatial data for wildland fire 

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

Rocky Mountain Research Station, Fort Collins, CO: 5-43.

Ryan, Kevin C.; Long, Donald G. 2004. The Southern Utah Fuels 

Management Demonstration Project: Project# 99-1-3-29. Final Report 

to the Joint Fire Science Program. On file at the Fire Sciences Lab, 

Rocky Mountain Research Station, Missoula, MT. 433 p.

Schmidt, K. M.; Menakis, J. P.; Hardy, C. C.; Hann, W. J.; Bunnell, 

D. L. 2002. Development of coarse-scale spatial data for wildland 

fire and fuel management. Gen. Tech. Rep. RMRS-GTR-87. U.S. 

Department of Agriculture, Forest Service, Rocky Mountain 

Research Station, Fort Collins, Colorado. CD+41 p.

Stratton, R. D. 2004. Assessing the effectiveness of landscape fuel 

treatments on fire growth and behavior. Journal of Forestry. 

102: 32-40.

USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 7

Introduction

Before European settlement of the interior west of 

the United States, coniferous forests of this region were 

influenced by many disturbance regimes, primarily 

fires, insects, diseases, and herbivory, which maintained 

a diversity of successional stages and vegetative types 

across landscapes. Activities after settlement, such as 

fire suppression, grazing, and logging significantly 

altered these disturbance regimes. As a result, forest 

structure and species composition have departed from 

historical conditions on many landscapes and this has 

led to increased forest densities, forest type conversions, 

and greater contiguity of many western forests. These 

forests are now more susceptible to large-scale insect 

infestations, disease outbreaks, and severe wildland 

fires than in the past, possibly endangering overall forest ecosystem health. The purpose of this paper is to 

address the historical and current conditions of coniferous forests of southern Utah to aid in the development 

of treatments to restore the ecological composition, 

structure, and function of these ecosystems.

The distribution of coniferous forests in southern Utah 

is mainly influenced by both climate and disturbance 

regime. Climatic factors, such as temperature and 

precipitation, determine where certain forest types can 

grow. Ponderosa pine (Pinus ponderosa) forests are the 

lowest elevation coniferous forest type in southern Utah 

occurring just above the warmer, drier Colorado pinyon-Utah juniper (Pinus edulis-Juniperus osteosperma) 

woodlands (Youngblood and Mauk 1985). As elevation 

increases, mixed conifer forests consisting of ponderosa 

pine, white fir (Abies concolor), Douglas-fir (Pseudotsuga menziesii), blue spruce (Picea pungens), limber pine 

(Pinus flexilis), Engelmann spruce (Picea engelmannii), 

and subalpine fir (Abies lasiocarpa) occur. Species 

dominance in these mixed conifer forests is mainly 

determined by disturbance history and microclimate. 

Chapter 2

Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

Michael A. Battaglia and Wayne D. Shepperd

In the cool, moist, higher elevations above 3,048 m 

(10,000 ft), Engelmann spruce-subalpine fir forests 

dominate when there is minimal disturbance.

Ponderosa Pine Forests

Historical Conditions

Historically, ponderosa pine was found on warm, dry 

sites on plateaus and mountains of central and southern 

Utah at elevations ranging from 1,981 to 2,743 m (6,500 to 

9,000 ft) (Madany and West 1980; Youngblood and 

Mauk 1985). Ponderosa pine forests bordered the shrub 

and woodland communities on its lower elevation range 

and mixed with Douglas-fir, white fir, blue spruce, and 

aspen at higher elevations (Powell 1879; Stein 1988a). 

Historical accounts of ponderosa pine acreage in southern 

Utah at the time of settlement are limited. The Fishlake 

National Forest estimates that ponderosa pine forests 

once occupied 9 percent, or 54,632 ha (135,000 acres), 

of the 640,212 ha (1,582,000 acres) analyzed within its 

boundaries (USDA Forest Service, Fishlake National 

Forest 1999).

Historical accounts of ponderosa pine forests for 

southern Utah and the southwestern United States 

indicate that these forests were open with a large 

diversity of grasses and flowers, with scattered pockets 

of shrubs (Alter 1942; Cooper 1960; Ogle and DuMond 

1997; Powell 1879). Several of these historical descriptions commented on the abundance of thick, lush, 

and high grass that would provide excellent forage for 

livestock.

The large diversity and amount of grasses in ponderosa 

pine forests was due in part to the structure of the forest. 

Pre-settlement ponderosa pine forests were irregularly 

spaced, uneven-aged stands with trees growing together 

in small even-aged groups and grassy meadows between 
8 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

the groups (Covington and Moore 1994a; Dutton 1882; 

Mast and others 1999; Schubert 1974). Although some 

groups could be overstocked (Schubert 1974), estimations for tree density and basal area in pre-settlement 

ponderosa pine forests are low – approximately 99 to 

148 trees/ha (40 to 60 trees/acre) with basal areas of 

11.5 m2/ha (50 ft2/acre) (Covington and Moore 1994b; 

Rasmussen 1941) – indicating the great size of the mature 

ponderosa pine trees. For example, a 1911 survey on 

the Manti-LaSal National Forest indicated that mature 

ponderosa pine trees grew as large as 1.2 to 1.5 m (4 to 

5 ft) in diameter, although it was more common to see 

trees that were 1.1 m (3.5 ft) in diameter and many of 

these large trees showed evidence of fire scars (Peterson 

1971). Other descriptions noted ponderosa pine trees over 

30.5 m (100 ft) tall and 30.5 to 46 cm (12 to 18 inches) 

in diameter (Ogle and DuMond 1997).

The prevalence of large diameter ponderosa pine trees 

and the open structure of the forests are attributed to 

fire and ponderosa pine’s insulative bark, which allows 

it to survive low intensity, surface fires. Frequent, low 

intensity fires created a diversity of vegetative structures 

by creating gaps, thinning seedlings, releasing nutrients, 

encouraging light-intolerant pine germination, and reducing the invasion of fire-intolerant, shade tolerant species. 

In pre-settlement forests, frequent fires spread easily 

through fine fuels such as grasses and needles. During 

dry spells, these fine fuels would allow surface fires 

to spread over large areas. The surface fires would kill 

seedlings, saplings, and shrubs and consume the large 

fuels such as branches and logs. With frequent surface 

fires, most woody fuels were consumed and large fuel 

loadings rarely accumulated; therefore, severe fires 

were rare because of low fuel volumes (Bradley and 

others 1992). Estimates for fuel loads in pre-settlement 

ponderosa pine forests are limited. However, Covington 

and Moore (1994b) estimate that forest floor and woody 

fuel loadings on North Kaibab ponderosa pine forests 

were less than 2.24 Mg/ha (1 ton/acre). 

Fuels loads and climate were the driving force for 

pre-settlement fires in ponderosa pine forests. Fires 

occurred primarily in dry years following wet years 

(Fulé and others 2000; Swetnam and Betancourt 1990) 

during the early growing season (Heyerdahl and others 

2005; Heyerdahl and others 2006) suggesting that wet 

pre-fire years allow fine fuel production and buildup that 

facilitate burning. Mean fire return intervals (MFRI) in 

ponderosa pine forests of southern Utah vary according 

to site topography and elevation (Heyerdahl and others 

2005; Heyerdahl and others 2006). For example, in 

three canyons on the Paunsaugunt plateau of the Dixie 

National Forest, composite MFRI ranged from 15 to 

18 years (Stein 1988b). On the Old Woman Plateau of 

the Fishlake National Forest, the MFRI was 27 years 

with a range of 6 to 62 years (Heyerdahl and others 

2005). Buchanan and Tolman (1983) reported MFRI 

of 4 to 7 years for Bryce Canyon National Park with 

a fire occurring at least once a decade within a different area of the forest. They suggest that the fires were 

small in extent because few of the fire-scarred trees in 

the same immediate vicinity were burned in the same 

year. In Zion National Park, MFRI ranged from 2.7 to 

25 years on the Horse Pasture Plateau, but within the 

park on the isolated Church Mesa surrounded by a barren expanse, MFRI was significantly longer, averaging 

69 years with a range of 56 to 79 years (Madany and 

West 1980; Madany and West 1983). Since fire could 

not start somewhere else on the landscape and spread 

to the isolated mesa, lightning strikes that hit the mesa 

were the only source of ignition, resulting in a longer 

fire return interval. MFRI in a ponderosa pine/Gambel 

oak forest on the North Rim of the Grand Canyon in 

Arizona was 3 to 4 years with a range of 1 to 11 years. 

The frequency decreased slightly to 6.8 years with a 

range of 2 to 24 years when calculations were made 

for fires that scarred 25 percent or more of the sample 

size within the same year, indicating a larger scale fire 

(Fulé and others 2000).

In general, most pre-settlement fires in ponderosa pine 

forests in the interior west and Rocky Mountains were 

of low to moderate severity surface fires (Barrett 1981; 

Brown and others 1999; Brown and Sieg 1999; Cooper 

1960; Covington and others 1997; Fulé and others 1997; 

Heyerdahl and others 1994; Swetnam and Baisan 1996). 

Crown fires were rare in open, lightly stocked stands 

because crowns did not overlap to allow running crown 

fires and ponderosa pine self-prune their branches, 

which keeps the foliage separate from the surface fuels. 

However, on moist sites that did not burn as frequently, 

more surface fuel probably accumulated and ingrowth 

of Douglas-fir and white fir served as ladder fuels, 

thereby increasing the chance for stand replacement 

and mixed-severity fires (Bradley and others 1992).

Although frequent periodic fire was the major driving 

force that influenced ponderosa pine forest structure, 

other disturbances also contributed to the thinning of 

forests and accumulation of surface fuels for the fires. 

Mountain pine beetle (Dendroctonus ponderosae), western pine beetle (Dendroctonus brevicomis), roundheaded 

pine beetle (Dendroctonus adjunctus), and pine engraver 

(Ips pini) all probably attacked stressed ponderosa pine 

trees in southern Utah. Endemic populations most likely 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 9

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

reproduced in stressed or weakened trees, as they do 

today, and killed a few trees per acre. Epidemics most 

likely occurred when tree densities became greater than 

27.5 m2/ha (120 ft2/acre) of basal area (Schmid and others 

1994), which would only occur in the absence of fire. 

Outbreaks of roundheaded pine beetles were probably 

sporadic and short-lived (Negron and others 2000). On 

the Kaibab plateau in northern Arizona, Blackman (1931) 

estimated that pre-settlement outbreaks of mountain pine 

beetle occurred approximately every 20 years. Even if 

the ponderosa pine trees were able to resist an insect 

attack, the tree would still be tree more susceptible to 

fire due to the exposed resin on the bark around the 

attacked areas (Bradley and others 1992).

Diseases, such as Armillaria root disease (Armillaria 

mellea) and dwarf mistletoe (Arceuthobium spp.), also 

existed in pre-settlement ponderosa pine forests, but it 

is unknown to what extent and severity. Woolsey (1911) 

reported that 1 to 2 percent of ponderosa pine surveyed in 

Arizona and New Mexico were infected with dwarf mistletoe. Fire often determined the distribution and intensity of 

dwarf mistletoe infection in coniferous forests (Alexander 

and Hawksworth 1976). Dwarf mistletoe infections were 

probably kept to a minimum by frequent fires because 

severe infections lead to high accumulations of dead trees, 

highly flammable witches’ brooms, and other surface fuels 

that would have burned more severely (Parmeter 1978). If 

stands were heavily infected, a fire would be more severe 

and kill the entire stand, thereby removing the infection 

source. However, a partial burn that left scattered infected 

trees could actually lead to rapid infection of regeneration 

(Alexander and Hawksworth 1976). Armillaria root disease 

was probably more common since it resides in overmature 

trees (Schubert 1974). The diseased trees often occur in 

groups of trees and are susceptible to wind breakage, which 

adds to the surface fuel loads. With frequent fires, however, 

the accumulation of such large fuels was minimal, but could 

probably result in severe fires in affected areas.

Herbivory in pre-settlement ponderosa pine forests 

was probably kept to a minimum since herbivore populations were quite low in southern Utah and ponderosa 

pine is not utilized for browse by most wild ungulates. 

Kay (1995) argues that elk (Cervus canadensis) and 

deer (Odocoileus hemionus) in the 1800s were rare in 

the western United States, including Utah, due to the 

efficiency of Native American hunters and the predation 

of the ungulates by wolves (Canis lupus), coyotes (Canis 

latrans), and other carnivores. Photographs taken in 

1872 by early surveyors showing multi-aged regenerating aspen stands suggest low population levels of elk 

and deer (Kay 1995). In contrast, bighorn sheep (Ovis 

canadensis) existed in all the mountain ranges of Utah 

(Dalton and Spillet 1974) and utilized the open ponderosa

pine forests (Smith and others 1999) due to the high 

visibility they provided (Wakelyn 1987). Most likely, 

herbivores influenced ponderosa pine forest structure 

by eating the grasses that carried fire, which prompted 

Native Americans to light fires to enhance the habitat 

for hunting purposes.

Disease, insects, and fires worked in concert to shape 

ponderosa pine forest community characteristics. These 

disturbances helped create heterogeneous landscape 

structure by creating gaps in the forest canopy, maintaining various age classes, and reducing forest density (Lundquist 1995a; Lundquist 1995b; Lundquist and Negron 

2000). Reduction in forest density decreased competition 

for water and nutrients and subsequently increased tree 

vigor, which lowered susceptibility to insect and disease 

attack (Christiansen and others 1987; Kegley and others 

1997; Larsson and others 1983; Wargo and Harrington 

1991). Frequent, low-intensity fires reduced encroachment of ponderosa pine into meadows and influenced 

species composition by reducing the invasion of shadetolerant and fire-intolerant species (Weaver 1967; Wright 

1978). The frequent fires prevented surface and ladder 

fuel buildup, released nutrients (Covington and Sackett 

1984; Covington and Sackett 1992), and encouraged 

germination (Bailey and Covington 2002). Frequent 

fires also allowed ponderosa pine forests to reach high 

elevations by reducing the competition with other tree 

species that were less fire resistant, such as Douglas-fir 

and white fir.

Current Conditions

The settlement of southern Utah has drastically altered 

the ponderosa pine ecosystem. The combined effects of 

fire suppression, logging, and grazing have altered the 

extent, location, and structure of ponderosa pine forests. 

Ponderosa pine still borders the Colorado pinyon-Utah 

juniper woodland (Youngblood and Mauk 1985), but 

forest inventories have shown a significant decrease 

in its quantity and extent (Stein 1988a; USDA Forest 

Service, Fishlake National Forest 1999). Ponderosa pine 

forests have gained some acreage from riparian zones, 

aspen, sagebrush, and mountain brush, but have lost 

significant acreage to Douglas-fir and white fir invasion (USDA Forest Service, Fishlake National Forest 

1999; Heyerdahl and others 2005; Heyerdahl and others 

2006). In the entire state of Utah, ponderosa pine now 

covers 240,560 ha (594,436 acres), or 3.7 percent, of the 

forested land (O’Brien 1999). Comparison of historical 

versus current acreage is only available for the Fishlake 
10 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

National Forest where ponderosa pine once occupied 

54,632 ha (135,000 acres). Today, ponderosa pine 

occupies 16,716 ha (41,307 acres), which is a 69 percent 

decrease in coverage (USDA Forest Service, Fishlake 

National Forest 1998).

The current stocking of the ponderosa pine understory 

varies from open to dense thickets (fig. 1). Many ponderosa 

pine stands in southern Utah now have shrub- dominated understories, most likely a result of continued fire suppression 

and grazing practices (fig. 2) (Bradley and others 1992). 

The most common shrubs in the understory include curlleaf 

mountain-mahogany (Cercocarpus ledifolius), greenleaf 

manzanita (Arctostaphylos patula), black sagebrush 

(Artemisia nova), Gambel oak (Quercus gambelii), mountain snowberry (Symphoricarpos oreophilus), bitterbush 

(Purshia tridentata) and common juniper (Juniperus 

communis). Mountain muhly (Muhlenbergia montana), 

a graminoid, is also a common habitat type in southern 

Utah (Bradley and others 1992; Youngblood and Mauk 

1985). Forbs do not contribute much to the understories 

of current southern Utah ponderosa pine forests (Bradley 

and others 1992).

Figure 1—Ponderosa pine (Pinus 

ponderosa) with an understory 

dominated by muttongrass 

(Poa fendleriana) and rubber 

rabbitbrush (Chrysothamnus 

nauseosus).

Figure 2—Well-developed 

understory of greenleaf manzanita 

(Arctostaphylos patula), Gambel 

oak (Quercus gambelii), and 

Rocky Mountain juniper (Juniperus 

scopulorum) beneath an open 

canopy of ponderosa pine (Pinus 

ponderosa).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 11

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

The increase in shrub density is dramatic in some areas. 

For example, Fulé and others (2002a) studied changes in 

a ponderosa pine/Gambel oak habitat on the north rim of 

the Grand Canyon and reported that Gambel oak density 

has substantially increased since settlement. Gambel oak 

densities in this area pre-settlement were approximately 

1 to 6 percent of the total density, whereas current density 

of Gambel oak is 20 to 70 percent of total plot density. 

On these same plots, total tree density (ponderosa pine 

included) has increased 155 to 486 percent. This increase 

in Gambel oak populations is a direct result of fire suppression. After fire, Gambel oak can resprout, but oak 

densities were historically kept low due to frequent fires. 

With fire suppression, Gambel oak was able to grow into 

sapling and pole thickets (Fulé and others 2002a).

The overstory structure of ponderosa pine forests has 

been drastically altered compared to pre-settlement forests (Heyerdahl and others 2006). Reynolds and others 

(in press) compared current southwest-wide data on 

ponderosa pine forests from Arizona (Conner and others 

1990), New Mexico (Van Hooser and others 1993), and 

northern Kaibab to Woolsey’s (1911) inventory of forests 

in 1910. Current tree densities averaged 329 trees/ha 

(133 trees/acre) compared to 89 trees/ha (36 trees/acre) 

for forests in 1910. Basal areas in 1910 averaged around 

18.8 m2/ha (82 ft2/acre), while current basal areas average around 13.3 m2

/ha (58 ft2/acre) (Ogle and DuMond 

1997; Woolsey 1911). The main difference, however, 

is the size of the trees. Due to heavy logging, current 

forests lack large old growth trees (Ogle and DuMond 

1997), but instead have many seedlings, saplings, and 

small saw-log sized trees.

The current age distribution of ponderosa pine forests 

has also changed from pre-settlement forests, especially 

at higher elevations. The age structure in lower elevation 

ponderosa pine forests follows an uneven-aged distribution with numerous seedlings, saplings, and pole-sized 

ponderosa pines (Stein 1988a), often in a clumped spatial 

pattern. Although this distribution has not changed much 

from pre-settlement forests, the magnitude in the density 

of young trees is much higher. Furthermore, the mortality of older trees has increased due to the competition 

for nutrients and water from the dense post-settlement 

trees and the stagnated nutrient cycling in the absence 

of fire (Covington and Moore 1994b). The younger age 

classes of ponderosa pine are missing in higher elevation ponderosa pine forests due to inability to regenerate 

under more shade-tolerant species such as Douglas-fir 

and white fir, which have invaded around the widely 

scattered mature ponderosa pine (Stein 1988a; Mast 

and Wolf 2004).

Surface fuel loading measurement estimates for southern Utah ponderosa pine forests are scarce. Estimates of 

total dead fuel loading for several ponderosa pine forests 

across the southwest range from 44.8 to 69.5 Mg/ha 

(20.0 to 31 tons/acre) (Bastian 2001a; Covington and 

Moore 1994b; Sackett 1979). Sackett’s (1979) survey of 

62 ponderosa pine stands across the southwest United 

States indicated that 58 percent of all the dead surface 

fuel was less than 2.54 cm (1 inch) in diameter. This 

highly flammable fuel, combined with the increased 

density of live shrubs and small trees that provide ladder fuels to carry fires from the surface into the trees 

(Cooper 1960; Covington and others 1994b; Madany and 

West 1980), have increased the chance of crown fire.

Herbaceous production in ponderosa pine forests of 

southern Utah is currently quite low and these stands 

lack the graminoid undergrowth that characterizes 

high quality range (Youngblood and Mauk 1985). A 

simulation study of the North Kaibab estimates that 

in pre-settlement ponderosa pine forests, herbage production averaged 0.67 Mg/ha (600 lbs/acre). Current 

estimates place the production at around 0.10 Mg/ha 

(100 lbs/acre) (Covington and Moore 1994b). Decreases 

in herbaceous production are a result of heavy grazing 

and an increase in both tree and shrub density, which 

increases competition. In addition, grazing reduces fine 

fuels to carry surface fire, exacerbating the problem of 

no fires, which results in stagnated nutrient cycling and 

ingrowth of more shade-tolerant tree species (Belsky 

and Blumenthal 1997).

Forests play an important role in supplying water to the 

majority of southern Utah communities. Descriptions of 

water yield from pre-settlement ponderosa pine forests 

are unknown. However, a simulation study of the North 

Kaibab estimates that in pre-settlement ponderosa pine 

forests, stream flows were around 17.5 cm (6.9 inches) 

and have decreased by 4.6 cm (1.8 inches), a 26 percent 

reduction post-settlement (Covington and Moore 1994b). 

Since ponderosa pine typically grows on drier sites 

than other conifers (for example, Douglas-fir, white fir, 

Engelmann spruce), the majority of water runoff comes 

from higher elevations where snow accumulates in the 

winter and melts in the spring. Annual water yield for 

the different drainage areas within southern Utah ranges 

between 2.5 to 5 cm (1 to 2 inches) of water, or 7.5 to 

13.5 percent of the total precipitation (Hemphill 1998). 

The use of overstory removal treatments to increase 

water yield from ponderosa pine forests is short-lived 

due to the increase in understory herbaceous and shrub 

cover (Bojorquez-Tapia and others 1990).
12 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

Wildfires can greatly affect the hydrology of a forest 

for several years post-fire depending on the severity. 

Removal of forest litter in the understory and overstory 

vegetation increases runoff, peak discharge, soil erosion, 

sedimentation, and loss of soil nutrients (Baker 1990; 

Campbell and others 1977; Dunford 1954; Rich 1962). 

Water runoff was eight times greater on a severely burned 

(majority of trees killed by fire) watershed than on an 

unburned watershed the year preceding the wildfire. 

Higher incidences of runoff events increased for 

several years in the moderately and severely burned 

watersheds due to the removal of litter cover and 

hydrophobic soil. A year after the wildfire, more than 

1.4 Mg/ha (1,254 lbs/acre) of sediment was lost in 

runoff from the severely burned watershed, although 

within another year, amount of sediment production 

returned to that of the unburned watershed (Campbell 

and others 1977).

Current Disturbances

Fire—The exclusion of fire since the late 1800s and 

early 1900s has greatly reduced periodic fires in southern 

Utah (Buchanan and Tolman 1983; Heyerdahl and others 

2005; Heyerdahl and others 2006; Madany and West 

1980; Madany and West 1983; Stein 1988b). Increased 

densities of small, young trees, build-up of surface 

fuels, and large areas of contiguous forests increase the 

likelihood of large-scale, severe fires. Lightning-caused 

wildfires in the southwest are getting larger over time, 

with some reaching tens of thousands of hectares (and 

getting bigger), in contrast to the 40 to 400 ha (100 to 

1,000 acres) surface fires of pre-settlement times 

(Heyerdahl and others 2006; Swetnam 1990). Since 

1972, in the entire Fishlake National Forest, 70 wildfires 

have burned with an average size of 650 ha (1,607 acres), 

but with some fires up to 7,440 ha (18,385 acres) in size 

(Fishlake National Forest GIS data). Estimates for all 

federal lands in the entire state of Utah from 1986 to 

1996 indicate that there were 8,335 fires with an average 

size of 50.5 ha (125 acres), but some reached 28,781 ha 

(71,120 acres) (Schmidt and others 2002).

Insects and Disease—Tree vigor has also declined as 

a result of fire suppression and the subsequent increase 

in forest density and competition. This decline in tree 

vigor increases the potential for more insect infestations. 

Forest inventory reports for the Dixie, Fishlake, and 

Manti-La Sal National Forests indicate that between 73 to 

93 percent of ponderosa pine trees within these forests 

are at moderate to high risk of attack by bark beetles 

(O’Brien and Brown 1998; O’Brien and Woudenberg 

1998; O’Brien and Waters 1998). Since 2002, infestations 

of mountain pine beetle have increased in acreage in these 

forests (Matthews and others 2005). The North Kaibab 

Plateau in Arizona has experienced several outbreaks of 

mountain pine beetle in the past century (Blackman 1931; 

Parker and Stevens 1979; Wilson and Tkacz 1995). The 

roundheaded pine beetle was at epidemic levels in 1995 

in the Dixie National Forest (Negron and others 2000). 

Outbreaks of the roundheaded pine beetle have caused 

considerable mortality across the southwest (Lucht and 

others 1974; Massey and others 1977), reducing basal 

area and total numbers of ponderosa pine trees by up to 

50 percent (Stevens and Flake 1974) and increasing large 

diameter woody fuel loadings up to eight times of what 

was already on the ground (Negron 2002). Increased 

surface fuel loads as a result of bark beetle epidemics 

increases the potential for fire hazard and the probability 

for higher-intensity fires for several years (Schmid and 

Amman 1992).

Lower tree vigor has also made ponderosa pine trees 

susceptible to several diseases. Over 20 percent of 

ponderosa pine trees in Utah are infected with dwarf 

mistletoe (Matthews and others 2005). Dwarf mistletoe increases the chance of a surface fire to transition 

into a crown fire due to the flammable witches broom 

and lower crowns (Harrington and Hawksworth 1990). 

Armillaria root disease is also infecting and killing 

mature and immature ponderosa pine trees in Utah 

(Forest Health Protection 2000) creating additional 

surface fuel loads.

Ungulates—Livestock grazing is one post-settlement 

disturbance that has indirectly contributed to increased 

ponderosa pine and shrub densities. Utah was severely 

overgrazed in the late 1800s and throughout the 20th

century (Ogle and DuMond 1997). Several large ranches 

were established in Utah and Arizona as early as 1863 

(Altschul and Fairley 1989), and cattle and sheep have 

heavily grazed the Paunsaugunt Plateau since 1866 

(Rathburn 1971). Guidelines for range allotments were 

non-existent in the 1800s, but by 1920, the Forest Service 

started to enforce them (Stein 1988a).

Overgrazing and trampling by livestock changed 

understory species composition, increased the amount of 

bare ground, decreased water storage, increased runoff, 

compacted soil, and increased erosion (T. Alexander 

1987; Belsky and Blumenthal 1997). Heavy grazing by 

livestock promoted the establishment of tree seedlings 

due to the reduction of herbaceous ground cover and an 

increase in bare soil (Madany and West 1983). Grazing 

also decreased the competition of grasses with shrubs 

and increased the density and extent of shrubs such as 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 13

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

Gambel oak, bigtooth maple, Utah serviceberry, and 

greenleaf manzanita (Mitchell 1984). With the loss 

of fine fuels (grass) for frequent fire, ponderosa pine 

was able to expand into ecotonal communities such as 

sagebrush and mountain brush.

Increased numbers of wild ungulates, such as deer, 

elk, and moose (Alces alces), have reduced herbaceous 

vegetation and aspen suckering in many stands. Presettlement populations of elk and deer were very low 

(Kay 1995; Rawley 1985), but once Native Americans 

populations were removed and wolves, mountain lions 

(Felis concolor), coyotes, and other predators were extirpated, ungulate populations exploded (Rasmussen 1941). 

Deer populations on the Kaibab Plateau had reached 

100,000 before declining in 1924 due to starvation and 

the creation of government hunting programs. Elk were 

transplanted into Utah from 1912 to 1915 (Rawley and 

Rawley 1967), and moose have recently been transplanted 

into south-central Utah (Kay and Bartos 2000). Current populations in southern Utah are approximately 

116,000 deer, 23,000 elk, and 65 moose (Utah Division 

of Wildlife 2005a,b,c).

Exotic Weeds—Overgrazing has also lead to the 

invasion of exotic species such as cheatgrass (Bromus 

tectorum). Cheatgrass originated from Eurasia where it 

coevolved with heavy grazing and has adapted well to 

heavy grazing regimes in the west where native species 

are at a disadvantage (Stebbins 1981). Cheatgrass grows 

in dense stands and cures by mid-June, about two to four 

weeks earlier than native grasses (Devine 1993). Once 

cheatgrass is dry, it can carry a fast moving fire and 

cause more frequent, intense and early-season wildfires. 

Cheatgrass has already increased fire frequency in lower 

elevation ecosystems such as sagebrush (Kitchen and 

McArthur, this volume). Cheatgrass and other exotics, 

such as mullein (Verbascum thapsus), toadflax (Linaria 

dalmatica), and thistle (Circium pulchellum), are often 

found in severely burned areas (Keeley 2003; Phillips 

and Crisp 2001; Sackett and Haase 1998). Cheatgrass has 

been observed in understories throughout the ponderosa 

pine range, from California (Keeley 2003) to Colorado 

(Fornwalt and others 2003).

Fire Regime Condition Classes

Fire Regime Condition Class 1 (FRCC 1)

Ponderosa pine forests functioning within the historic 

range of variability contain trees of all sizes and ages. 

Stands usually occur in even-aged groups or clumps. 

Periodic surface fires have pruned the branches of large 

trees well above the ground, creating an open appearance 

with long sighting distances under the forest canopy. 

Such fires also have prevented large accumulations 

of surface fuels and kept the density of smaller trees 

and woody understory species low, creating a diverse 

understory of grasses and forbs. Periodic fires have created openings of up to several acres as a result of the 

torching of clumps or groups of trees. Such openings 

increase the spatial diversity of the forest and lessen the 

occurrence of landscape-wide stand replacement crown 

fires by serving as fuel breaks. Frequent surface fires do 

not consume the soil’s organic layer that helps stabilize 

the soil surface and prevent excessive erosion. Smoke 

production is low in volume and short in duration, but 

regional landscapes would likely have been more smoky 

than today.

Fire Regime Condition Class 2 (FRCC 2)

Ponderosa pine forests existing under moderately 

altered fire regimes are denser and contain higher numbers of young trees in the understory than a historical 

stand. Large pines are still a component of the forest. 

Smaller, shade tolerant conifer associates are present at 

some places in the understory and as occasional larger 

trees. Smaller openings have been invaded by ponderosa 

pine, which creates a more contiguous forest canopy. 

Small pines also have branches closer to the ground. 

Understory species abundance and diversity is less than 

that in FRCC 1.

A wildfire occurring in a current FRCC 2 ponderosa 

pine forest is a mixture of severity and intensity, larger 

in size than historically observed, and difficult to suppress. Smoke production is probably moderate in volume 

and duration. Understory plants and small trees burn 

and provide a pathway for the fire to reach overstory 

foliage, which leads to large areas of torched trees. Soil 

productivity is more severely impacted (in other words, 

reduced) as a result of litter and upper duff consumption. Areas of bare ground are subject to some erosion, 

and sedimentation of streams decreases water quality. 

Biodiversity of herbaceous plants increases for the first 

few years post-fire, but with an increased probability 

of exotic species. Regeneration of ponderosa pine is 

possible in areas with surviving trees.

Fire Regime Condition Class 3 (FRCC 3)

Ponderosa pine forests are considerably altered from 

those in FRCC 1, either through repeated harvests or 

through vegetative succession as a result of altered fire 
14 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

regimes. Where harvests have occurred, most large 

pre-settlement trees have been removed. If post-harvest 

regeneration was successful, dense, evenly-spaced forests 

of younger pines now exist. If pine regeneration was not 

successful, oak shrubs or other conifers replaced the 

logged pine. This created either mixed species forests 

or open pine forest with dense shrub understories. In 

either case, understory grass and forb production has 

declined, along with populations of animals and birds 

that depended upon the diversity of FRCC 1 ponderosa 

pine forests.

A wildfire occurring in a current FRCC 3 ponderosa 

pine forest is extremely damaging to the ecological 

integrity of an area. The wildfire is high in severity 

and intensity, covers large acreages, and is extremely 

difficult and costly to suppress. Smoke production is 

extremely high in volume and duration. Understory 

plants, small trees, and even large trees burn, resulting in 

a completely altered landscape. Nutrients are volatized, 

soil microbes killed, and soil organic matter consumed 

where accumulations of woody fuels burned. Soil is more 

likely to become hydrophobic and increased erosion and 

sedimentation into streams is likely to impact regional 

water quality. Exotic species invasion is likely in areas 

of high fire severity. Regeneration of ponderosa pine is 

limited to areas that are not completely burned.

Ponderosa pine forests in southern Utah are somewhat 

of an anomaly compared to those in other regions. Inventory data indicate that the average ponderosa pine forest 

in southern Utah is relatively poorly stocked, containing mostly young, small diameter trees (O’Brien 1999; 

O’Brien and Brown 1998; O’Brien and Waters 1998; 

O’Brien and Woudenberg 1998). In terms of stocking, 

ponderosa pine forests would initially appear to be in 

FRCC 1. However, the lack of large old growth trees 

(O’Brien 1999; O’Brien and Brown 1998; O’Brien 

and Waters 1998; O’Brien and Woudenberg 1998) 

and 120+ years of fire exclusion would suggest that 

the majority of southern Utah ponderosa pine forests 

are not within the historical range of variation and 

are likely in FRCC 2 or 3.

We suspect that repeated timber harvests, heavy grazing, and wildfire suppression have significantly altered 

the species composition and fuel (surface and canopy) 

loadings of many ponderosa pine forests in southern 

Utah. Expansion of oak, pinyon, and juniper into the 

understories of lower elevation ponderosa pine and the 

increased presence of Douglas-fir and true fir in pine 

forests at higher elevations (Mast and Wolf 2004) are 

likely to have created conditions that will require 

restoration treatments before fire is reintroduced. 

A 69 percent post-settlement decrease in ponderosa 

pine on the Fishlake National Forest (USDA Forest 

Service, Fishlake National Forest 1999) indicates that 

type conversion from ponderosa pine to other forest 

types is occurring.

The increased presence of these associated forest 

species compared to pre-settlement conditions has 

resulted in an increased risk of stand replacement wildfire, even though the stocking of ponderosa pine is low. 

The lack of large thick-barked, fire-pruned ponderosa 

pines only increases the chances that all pines will be 

eliminated from the forest if fires should occur. Shade 

tolerant conifer seedlings such as Douglas-fir and true 

fir, and oak in the understory, provide live fuel ladders 

that will allow fire to easily reach into the crowns of 

ponderosa pine.

Recommended Treatments

In contrast to strict thinning from below where all 

small trees are removed, we advocate the use of an 

uneven-aged approach in maintaining open ponderosa 

pine stands that are at less risk to crown fire. We 

recommend converting the understocked ponderosa 

pine stands common to southern Utah to irregularly 

structured uneven-aged stands by reducing or removing shade tolerant conifers and oak and re-introducing 

frequent prescribed surface fires. Development of stand 

structures should focus on creating forests with basal 

areas and densities that occurred historically, which 

includes a component of large-diameter trees that can 

withstand wildfires. Historic ponderosa pine forest 

structure in the Grand Canyon and the Kaibab Plateau 

in nearby Arizona averaged 80 to 160 trees/ha (32 to 

65 trees/acre) with basal areas ranging between 5.9 to 

20.5 m2/ha (25.7 to 89 ft2/acres) (Fulé and others 2002a, 

2002b, 2006) and quadratic mean diameters between 

47 to 52.7 cm (18.5 to 20.7 inches) (Fulé and others 

2002b). The estimates for historic southern Utah ponderosa pine forests had 0 to 297 trees greater than 20 cm 

dbh per ha (0 to 120 trees/acre) (Heyerdahl and others 

2006).

Control of stocking or density under uneven-aged 

management can be achieved using either the Stand 

Density Index (SDI) or BDQ method. SDI utilizes 

upper and lower diameter limits, but assigns stocking 

evenly across all diameter classes except for a reduction in relative density for the smallest diameter class 

(see Long and Daniel 1990; Long 1995). As in mixed 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 15

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

conifer forests, ponderosa pine forests should contain 

30 percent maximum SDI stocking or less and contain 

trees of all sizes and ages (Long 1995). Several papers 

provide detailed examples for calculating SDI to regulate 

stocking in ponderosa pine forests (Long and Daniel 

1990; Long 1995; Shepperd 2006). Shepperd (2006) 

describes a methodology (SDI-Flex) by which an infinite 

variety of stand configurations can be maintained.

BDQ relies on basal area, diameter distribution, and 

a “Q” ratio. Under the BDQ system, managers must 

first select the upper and lower target diameters that 

define the range of tree sizes to be managed. Second, a 

“Q” factor must be chosen. The “Q” factor is the ratio 

between the number of trees per acre in one diameter 

class and those in the next smaller class (Alexander and 

Edminster 1987). Lastly, the residual basal area that 

the growing stock is reduced to following each cutting 

cycle must be chosen. Examples of calculating desired 

stocking using the BDQ method are presented in Guldin 

(1996) and in Alexander and Edminster (1987). It is 

important to note that for the success of either method, 

managers must be diligent in monitoring regeneration 

and levels of growing stock across all diameter classes 

during each cutting cycle. Marking and thinning of such 

forests should be done to create a grouped or clumped 

appearance and be irregular in both spatial and vertical structure. Silvicultural prescriptions must include a 

flexible time table for future thinnings and prescribed 

burn treatments.

In southern Utah, much of the above-mentioned activity will need to be done in the future, rather than at 

present. Regional inventory data (O’Brien 1999; O’Brien 

and Brown 1998; O’Brien and Waters 1998; O’Brien 

and Woudenberg 1998) indicate many ponderosa pine 

forests are poorly stocked and will require time to grow 

into desirable ponderosa pine stocking conditions. Initial 

manipulation may still be required to reduce stocking of 

other species such as white fir or Douglas-fir. Both of 

these species would reduce available light and resources 

for successful ponderosa pine establishment and growth. 

Furthermore, white fir and Douglas-fir would increase 

the risk for a stand replacing wildfire.

In areas where white fir and Douglas-fir are present 

in ponderosa pine stands, they should be removed in 

the first uneven-aged entry to set the successional stage 

back to a purer ponderosa pine stand. A prescribed burn 

to kill fir seedlings should follow thinning. Similarly, 

where aspen occurs in conjunction with ponderosa pine, 

the aspen should be encouraged to come in under the 

pine. If aspen is a desired component and is already 

present in the stand, SDI stocking or BDQ guidelines 

may need to be lowered to create a more open stand to 

allow aspen to thrive. Normally, marking using a group 

selection technique should be adequate to maintain 

small clones of aspen interspersed among ponderosa 

pine. Periodic prescribed burns and/or harvest entries 

can be used to maintain the presence of aspen in these 

landscapes. Where ponderosa pine grows in conjunction 

with Gambel oak or other hardwoods species, periodic 

prescribed fire is needed to keep those shrub species 

in check.

Maintaining ponderosa pine forests as open, irregularly 

spaced forests has the additional benefit of increasing 

forage production for domestic livestock and wildlife. 

Because allotment stocking rates have not been adjusted 

in many years, the gradual loss of grasses and forbs due 

to shrub and shade-tolerant conifer ingrowth has actually 

reduced herbaceous production compared to a century 

ago.

A primary goal in restoration treatments in ponderosa 

pine stands should be to increase the presence of large 

trees through silvicultural practices. Repeated timber 

harvests in many areas of southern Utah selected only 

the larger trees and have reduced the average tree size 

considerably from that of the past. In terms of unevenaged management, the largest pine stumps found on 

a site can serve as upper diameter limit targets when 

developing stocking guidelines.

If managers desire to use the BDQ or SDI-Flex 

(Shepperd and others 2006; Shepperd 2007) approach 

in achieving stocking goals for fuels reduction and forest restoration, diameter distributions should be chosen 

that do not result in an over abundance of young trees 

in the forest. For instance, increasing the Q factor (or 

flex factor) will increase the density of smaller trees, 

and decreasing the Q factor will lower the density of 

smaller trees (Shepperd and Battaglia 2002; Shepperd 

and others 2006; Shepperd 2007). Typically, if 5-cm 

(2-inch) diameter classes are used, a Q of 1.2 or less 

is desirable in ponderosa pine forests and no more 

than 1.3 in mixed conifer. Higher values will result in 

young trees that will have to be removed in subsequent 

entries.

Modeling and experience has shown that using unevenaged management techniques with periodic entries on 

sites in both ponderosa pine and mixed conifer forests 

similar to those in southern Utah can be accomplished 

every 30 years and be economically sustainable (Skog 

and others 2006). A commercial timber harvest could 

be used to remove some medium and large diameter 

material, taking care to leave a sufficient number of large 

trees to create conditions similar to historic stocking and 
16 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

basal area levels. However, biomass or other markets are 

needed for smaller diameter materials to ensure they 

are also removed from the site to reduce competition 

and surface fuel buildup. As with all restoration cuts, 

finding markets for the small diameter wood biomass 

in southern Utah to offset costs is as critical an issue 

as anywhere else in the west. Much of what needs to be 

done cannot be accomplished without markets for small 

diameter material.

Mixed Conifer Forests

Historical Conditions

Pre-settlement mixed conifer forests contained ponderosa pine, Douglas-fir, white fir, aspen, blue spruce, 

Engelmann spruce, and subalpine fir. Distribution of 

these species depended upon disturbance history, aspect, 

elevation, and available moisture (Bradley and others 

1992). Powell (1879) observed a mixture of Douglas-fir, 

white fir, ponderosa pine, and blue spruce growing from 

2,133 to 2,743 m (7,000 to 9,000 ft). Lower elevations 

were dominated by ponderosa pine with Douglas-fir 

as the second most common tree. Ponderosa pine was 

prevalent on drier, southern aspects at middle and upper 

elevations up to 2,591 m (8,500 ft). At elevations above 

2,591 m (8,500 ft), limber pine, Douglas-fir, and especially 

white fir performed best (Buchanan and Harper 1981). 

Detailed acreage estimates of pre-settlement mixed conifer 

forests for southern Utah are not available.

Lang and Stewart (1910) described the mixed conifer 

forests of the Kaibab Plateau as open and subjected to 

multiple wildfires that created partially denuded landscapes. A reconstruction of a mixed conifer forest in 

Little Park (>2,650 m [8,700 ft]) on the North Rim of 

the Grand Canyon provides a small picture of the species composition, age structure, and forest structure of a 

mixed coniferous forest in 1880 (Fulé and others 2003; 

Fulé and others 2002c). Tree density in this area averaged 40 trees/ha (98 trees/acre) with an average basal 

area of 17.7 m2/ha (77 ft2/acre). Ponderosa pine, white 

fir, and aspen each made up 24 to 27 percent of the tree 

density, followed by Douglas-fir (19 percent), and small 

amounts of Engelmann spruce (5 percent) and subalpine 

fir (1 percent). Although aspen made up 25 percent of 

the tree density, the species only contributed 4 percent 

to the basal area, indicating high numbers of sprouts. In 

contrast, ponderosa pine, white fir, and Douglas-fir each 

made up 31 to 32 percent of the basal area, suggesting 

the presence of larger, older trees (Fulé and others 2003; 

Fulé and others 2002c).

At another lower elevation site (2,400 to 2,500 m 

[7,962 to 8,300 ft]) within the Grand Canyon, Swamp 

Ridge, Fulé and others (2002a) reconstructed the 1880 

mixed conifer forest and found no Engelmann spruce 

or subalpine fir. Average tree densities were similar to 

Little Park, but with higher average basal area (28.5 m

2

/ha 

[124 ft2/acre]). Ponderosa pine was dominant in this stand 

making up 75 percent of the basal area and 53 percent 

of the tree density. White fir (13 percent) had twice the 

tree density of Douglas-fir (6 percent), but its basal area 

was only 20 percent higher, indicating an ingrowth of 

white fir seedlings. As with the Little Park area, aspen 

contributed about 27 percent of the tree density, but only 

4 percent of the basal area. Again, this suggests there 

were a high number of sprouts (Fulé and others 2002a). 

Of interest is the wide range of basal areas (15 to 54 m

2

/ha 

[66 to 235 ft2/acre]) found on the Swamp Ridge plots. 

This suggests both open and dense stands were present 

in pre-settlement mixed conifer forests as a result of 

mixed severity fires (Fulé and others 2003).

The diverse stand structure of pre-settlement mixed 

conifer forests contributed to a variety of understory 

conditions ranging from dry, open, grassy understories 

to moist, closed canopy understories with a diverse 

mixture of plant lifeforms. Although no quantitative 

pre-settlement descriptions of shrub and herbaceous 

conditions for mixed conifer forests were found in the 

literature, some conclusions can be drawn from Buchanan 

and Harper’s (1981) comparison of a 1959 and 1980 

botanical survey of Bryce Canyon National Park. This 

study reported a 30 to 44 percent decrease in average 

understory coverage in four mixed conifer community 

types and a 21 percent to 24 percent decline in understory 

diversity in three mixed conifer community types due 

to an increase in forest overstory density. Community 

types with white fir showed the largest decline in shrub 

coverage. Based on these data, we suggest that shrubs 

(especially sprouters) in pre-settlement mixed conifer 

forests, which had even lower overstory densities than 

those observed in 1959 and were subjected to fire, were 

more prevalent, and more grasses and forbs were present 

due to a greater availability of light.

Based on the varied densities and composition of presettlement mixed conifer forests, surface fuels loads 

in these stands were probably quite variable. The fuel 

complex of pre-settlement mixed conifer forests in Bryce 

Canyon National Park has been described as a timber 

overstory with a grass or herbaceous surface stratum 

(Anderson 1982; Jenkins and others 1998; Roberts and 

others 1993). Although there are no detailed inventories 

of pre-settlement surface fuel loads, Dieterich (1983) 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 17

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

estimates that surface fuel loads were probably less than 

one-fourth to one-third of present-day loadings.

Mean fire return intervals (MFRI) reported in the 

literature for mixed conifer forests of the southern Utah 

region range from 2 to 129 years depending upon location, 

species composition, and methodology used (Buchanan 

and Tolman 1983; Chappell and others 1997; Fulé and 

others 2003; Heyerdahl and others 2005, Heyerdahl 

and others 2006; Jenkins and others 1998; Stein 1988b; 

Touchan and others 1996; White and Vankat 1993; Wolf 

and Mast 1998). In general, MFRI lengthened with elevation and composition of shade-tolerant species (Fulé and 

others 2003; Heyerdahl and others 2005; Heyerdahl and 

others 2006; Wolf and Mast 1998). For example, Wolf 

and Mast (1998) reported lower elevation mixed conifer 

forests had MFRI ranging from 5 to 7.25 years, while 

higher elevation forests with a spruce component had 

longer return intervals of 10 to 19 years, with the lower 

value including all sampled trees and the higher value 

calculated for fires that scarred 25 percent or more of 

the sample size within the same year. Stands sampled 

on the Fishlake National Forest showed similar patterns 

with lower elevation mixed conifer forests having MFRI 

ranging from 15 to 30 years and higher elevation mixed 

conifer forests with an Engelmann spruce-subalpine 

fir component ranging from 45- to 57-year mean fire 

interval (Chappell and others 1997; Heyerdahl and others 

2005).

In general, most pre-settlement fires in mixed conifer 

forests were low to moderate intensity surface fires 

(Buchanan and Tolman 1983; Dieterich 1983; Fulé and 

others 2003; Heyerdahl and others 1994; Swetnam and 

Brown 1992), especially on southern and western aspects 

(Fulé and others 2003). In areas of high tree density or 

areas with high insect and disease-caused mortality, 

patchy crown fires and high intensity fires were possible 

(Bradley and others 1992). At lower elevations, fires 

burned during the dormant and early growing season, 

while higher elevations burned during the dormant 

and late growing season (Heyerdahl and others 2005; 

Heyerdahl and others 2006).

Several types of insects influence the structure and 

composition of mixed coniferous forests in southern 

Utah because they are host specific and their mode 

of attack differs (Swetnam and Lynch 1989; Swetnam 

and Lynch 1993). For instance, mountain pine beetles 

(Dendroctonus ponderosae), western pine beetles 

(Dendroctonus brevicomis), pine engravers (Ips pini), 

and roundheaded pine beetles (Dendroctonus adjunctus) 

all attack ponderosa pine, while Douglas-fir beetles 

(Dendroctonus pseudotsugae) only attack Douglas-fir. 

The Douglas-fir tussock moth (Orgyia pseudotsugata) 

and fir engravers (Scolytus ventralis) prefer both 

white fir and Douglas-fir. Western spruce budworm 

(Choristoneura occidentalis) attacks Douglas-fir, white 

fir, subalpine fir, and Engelmann spruce, but usually 

ignores ponderosa pine. Spruce beetle (Dendroctonus 

rufipennis) attacks only Engelmann spruce.

Although there are few descriptions of insect or 

disease attacks on pre-settlement mixed conifer forests 

in southern Utah, other regions of the western United 

States have experienced evidence of episodic outbreaks 

of Douglas-fir tussock moth (Wickman and Swetnam 

1997) and western spruce budworm (Swetnam and 

Lynch 1989; Swetnam and Lynch 1993; Veblen and 

others 1994). In the Blue Mountains of Oregon, Wickman and Swetnam (1997) reported Douglas-fir tussock 

moth outbreaks occurred over large areas and were 

synchronous. Western spruce budworm outbreaks in 

southern Colorado and northern New Mexico were 

shown to occur at irregular intervals over the last 300 

years with intervals between outbreaks ranging from 

14 to 58 years with an average duration of 12.9 years 

(Swetnam and Lynch 1989; Swetnam and Lynch 1993). 

Mountain pine beetle epidemics reported in the previous 

ponderosa pine section of this chapter probably have 

attacked the ponderosa pines found in mixed conifer 

forests. Spruce beetle outbreaks in mixed conifer forests 

would have been limited to the upper elevations where 

more Engelmann spruce was present.

Diseases, such as Armillaria root disease (Armillaria 

mellea), annosus root disease (Heterobasidion annosum), 

and dwarf mistletoe (Arceuthobium spp.), existed in presettlement mixed conifer forests, but it is unknown to 

what extent. Outbreaks in pre-settlement mixed conifer 

forests were probably not as detrimental as they are 

today for several reasons. First, frequent fires would 

have maintained low to moderate basal areas. These 

fires decreased competition and increased tree vigor, 

lowering the susceptibility to insect and disease attacks. 

Additionally, insects and diseases are host specific and 

the pre-settlement mixed conifer forests maintained a 

mixture of tree species. This combination would allow 

tree species not susceptible to a certain insect or disease 

attack to regenerate while the susceptible tree species 

declined.

Disturbances in mixed conifer forests affect community characteristics in several ways. Species composition 

and tree density are somewhat regulated by fire due to 

the different fire resistance characteristics of each species. Mature ponderosa pine and Douglas-fir have thick 

bark that is very resistant to fire. However, seedlings and 
18 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

saplings of Douglas-fir are vulnerable to surface fires, 

while young ponderosa pine can maintain a presence 

on sites with fire intervals as short as 6 years if fire 

severity is low (Bradley and others 1992). Young white 

fir trees are also vulnerable to fire, but as they mature, 

the bark becomes thicker and more resistant, although 

its low branching habit increases its susceptibility to 

crown fires. Fire resistance in young and mature blue 

spruce, subalpine fir, and Engelmann spruce is very 

low and these species are often killed by fire. Aspens 

are also easily killed by fire, but can revegetate a site 

quickly with new sprouts produced through root suckering (Bartos, this volume).

Fire frequency and severity are the major factors that 

determined the species composition of pre-settlement 

mixed coniferous forests. Most areas were probably 

dominated by ponderosa pine and Douglas-fir in the 

overstory because of the frequent fire regime in the presettlement era. In areas where fire intervals were longer, 

white fir was able to develop thick enough bark to resist 

low intensity surface fires and contribute to post-fire 

regeneration. At higher elevations, where fires were not 

as frequent, a greater proportion of spruce and fir were 

present. Aspens were prevalent in pre-settlement mixed 

conifer forests under both frequent and mixed mode 

fire regimes and served as natural firebreaks. In areas 

with longer fire intervals, aspen populations were much 

lower, but could increase in response to fire, serving as 

a nurse crop for conifers.

Current Conditions

Species composition, forest density, structure, and 

disturbance regimes have been altered in many mixed 

conifer forests of southern Utah since settlement. Interruption of natural fire regimes has allowed succession 

to move these forests toward more shade-tolerant species. As a result, ponderosa pine is no longer dominant 

in mixed conifer forests and aspen populations have 

declined dramatically (fig. 3). Ponderosa pine has lost 

acreage to both Douglas-fir and white fir, and in turn, 

Douglas-fir has lost acreage to white fir (fig. 4) (USDA 

Forest Service, Fishlake National Forest 1998).

For the entire state of Utah, Douglas-fir covers 

456,647 ha (1,128,400 acres) and white fir covers 

163,374 ha (403,707 acres) (O’Brien 1999). Within 

southern Utah, the Douglas-fir cover type comprises 

about 5 percent and white fir makes up 2.5 percent of 

the forested land (O’Brien and Brown 1998; O’Brien 

and Woudenberg 1998; O’Brien and Waters 1998).

Mixed conifer stands in southern Utah are dominated by a variety of tall and low shrubs (Bradley and 

others 1992; Youngblood and Mauk 1985). The most 

common shrubs found in the Douglas-fir, white fir, 

and blue spruce habitat types of southern Utah include 

ninebark (Physocarpus malvaceus), curlleaf mountain 

mahogany (Cercocarpus ledifolius), greenleaf manzanita 

(Arctostaphylos patula), mountain mahogany (Cercocarpus montanus), Gambel oak (Quercus gambelii), 

Figure 3—Colorado blue 

spruce (Picea pungens) gaining 

dominance in a stand of trembling 

aspen (Populus tremuloides) 

with an understory dominated 

by Kentucky bluegrass (Poa 

pratensis) and mountain snowberry 

(Symphoricarpos oreophilus).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 19

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

Oregon grape (Berberis repens), mountain snowberry 

(Symphoricarpos oreophilus), mountain maple (Acer 

glabrum), and common juniper (Juniperus communis). 

The majority of these shrubs resprout after fire at various degrees, with the exception of common juniper and 

curlleaf mountain mahogany (Bradley and others 1992). 

The majority of herbaceous stratums in the habitat types 

of southern Utah are depauperate, although some of 

the white fir and blue spruce areas do support a small 

amount of graminoids and forbs (Buchanan and Harper 

1981; Youngblood and Mauk 1985).

The most dramatic change in mixed conifer forests 

is the increase in basal area, tree density, and species 

composition shift toward white fir at lower elevations 

and Engelmann spruce and subalpine fir at the higher 

elevations (Bastian 2001b; Fulé and others 2003; Fulé 

and others 2002a; Heyerdahl and others 2005; Heyerdahl and others 2006; Jenkins and others 1998; Mast 

and Wolf 2004). For instance, in one area of Bryce 

Canyon, the largest and oldest trees (200 to 250 years 

old) are ponderosa pine, Douglas-fir, and white fir, 

but the regeneration for the past 100 years is mostly 

white fir and Douglas-fir (Jenkins and others 1998). 

In another area of Bryce Canyon National Park, white 

fir overstory density is 33 trees/ha (81.7 trees/acre) 

compared to a ponderosa pine overstory density of 9 

trees/ha (22.8 trees/acre). More striking is the regeneration layer where white fir seedlings density is 649 

trees/ha (1,604 trees/acre) and ponderosa pine seedling density is 15 trees/ha (36.8 trees/acre) (Bastian 

2001b). At Swamp Ridge (2,400 to 2,500 m [7,962 to 

8,300 ft]) on the Northern Rim of the Grand Canyon, 

there has been a 283 percent increase in tree density 

and a 45 percent increase in basal area (Fulé and others 

2002a). Although ponderosa pine still dominates basal 

area, it only makes up 17 percent of the tree density. 

The majority of pre-settlement aged trees are ponderosa 

pine, with a substantial increase in white fir and aspen 

regeneration after fire regimes were disrupted (Fulé 

and others 2002a). At a higher elevation site within the 

same area, average basal area is 38.8 m

2

/ha (169 ft2/acre) 

with a tree density of 143 trees/ha (353 trees/acre) (Fulé 

and others 2002a). Although white fir density has not 

increased since pre-settlement in this stand, there has 

been a substantial increase in the density of Engelmann 

spruce and subalpine fir at the expense of ponderosa 

pine and Douglas-fir. Furthermore, Engelmann spruce 

and subalpine fir now contribute 19 percent to the total 

basal area in contrast to the 3 percent they contributed 

in the pre-settlement stand (Fulé and others 2002a).

Fuel loadings estimates for southern Utah mixed 

conifer forests are limited. In Bryce Canyon National 

Park, current total surface fuel loading estimates are 

around 69.5 to 71.7 Mg/ha (31 to 32 tons/acre), a 200 

percent increase since 1900 (Bastian 2001b; Roberts 

and others 1993; Jenkins and others 1998). At higher 

elevations, mixed conifer forests average total surface 

fuel loading is 118.8 Mg/ha (53 tons/acre), the majority 

Figure 4—Mixed conifer forest 

with white fir (Abies concolor) and 

ponderosa pine (Pinus ponderosa). 

The most dominant understory 

species is greenleaf manzanita 

(Arctostaphylos patula).
20 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

(87 percent) of which is 1000-hour rotten and sound 

fuels (Fulé and others 2002c). However, Sackett’s 1979 

survey of 16 southwestern mixed conifer stands indicated 

total surface fuel loadings averaged 98.6 Mg/ha (44 

tons/acre), half of which was from 1000-hour fuels.

Although no formal forest hydrology studies exist 

in southern Utah mixed conifer forests, results can be 

extrapolated from research in mixed conifer forests in 

Arizona (Rich and Thompson 1974). Water yields were 

inversely related to the amount of forested area of the 

watershed. Clearcutting increased water yields in proportion to the percent of the forested area clearcut. While 

most of the yield was accounted for in the Arizona study 

by reductions in evapotranspiration, snow interception 

also likely plays a role in mixed conifer ecosystems 

(Troendle and others 1988). Current knowledge of the 

relationship between water yield and forest structure 

holds that water yield in mixed conifer forests is a 

function of the basal area/leaf area within the watershed (Shepperd and others 1992). Because snow packs 

do accumulate in mixed conifer forests, we can expect 

them to be hydrologically similar to spruce-fir forests 

(see following spruce-fir section). Peak discharge rates 

should not be appreciably altered by removal of vegetation, but duration of flow lengthens. Therefore, any 

removal or increase in forest vegetation (through fire, 

harvest, insect attack, or succession) would potentially 

translate into an increase in water yield.

Wildfires can greatly affect sediment production in 

mixed conifer forests. A 60-acre wildfire in a mixed 

conifer forest occurred on the South Fork watershed of 

Workman Creek in Arizona, destroying 74 percent of 

the basal area (Rich and Thompson 1974). Average postwildfire sediment production in this area was 50.6 m

3

/ha 

(726 ft3/acre), compared with 0.009 to 0.98 m3/ha (0.14 

to 14 ft3/acre) in unburned mixed conifer forest within 

the same watershed.

Current Disturbances

Fire—Since settlement, fire-free intervals have increased in mixed conifer forests similar to ponderosa 

pine forests. Shade tolerant species now dominate in 

many areas of the mixed conifer forest and the continuity 

of forest vegetation has increased in many landscapes. 

Instead of a mixed mode of surface fire and patchy 

crown fire, mixed conifer forests may now burn in large 

landscape crown fires (Fulé and others 2004). Loss of 

pure aspen stands, which served as natural fire breaks 

within mixed conifer landscapes, through succession, fire 

suppression, and overbrowsing has further exacerbated 

the situation.

Insects and Disease—Occurrences of western spruce 

budworm, mountain pine beetle, Douglas-fir beetle, and 

Douglas-fir tussock moth have likely increased with 

increasing density and continuity of southern Utah mixed 

conifer forests. Forest inventory reports for the Dixie, 

Fishlake, and Manti-LaSal National Forests indicate 

that between 61 and 74 percent of all Douglas-fir trees 

within the forests are at moderate to high risk of attack 

by bark beetles (O’Brien and Brown 1998; O’Brien 

and Woudenberg 1998; O’Brien and Waters 1998). In 

fact, outbreaks of Douglas-fir beetle have substantially 

increased on each these National Forests since 2000 

(Forest Health Protection 2000; Matthews and others 

2005). Douglas-fir beetle epidemics seem to be more 

synchronous on a larger scale than in pre-settlement 

forests. Epidemics of this beetle have occurred in the 

Front Range of Colorado approximately every 15 to 

35 years in the 20th century (Schmid and Mata 1996). 

Douglas-fir beetle epidemics typically occur following an 

extensive windthrow event or fire (Furniss and Carolin 

1977). In addition, Douglas-fir epidemics in Colorado 

and Wyoming have taken place following western spruce 

budworm epidemics (Schmid and Mata 1996).

Western spruce budworms defoliate Douglas-fir, white 

fir, subalpine fir, and to a lesser degree, Engelmann 

spruce. Fire suppression and past management practices 

have created multi-storied, dense, and continuous forests 

that provide abundant food sources for all larval stages 

of western spruce budworm and reduce larval dispersal 

loss (Carlson and others 1985). The changes in forest 

structure have shifted the spatial and temporal pattern 

of budworm outbreaks. Although frequency of outbreaks 

has not changed since the pre-settlement era (20 to 

33 year intervals), recent outbreaks of western spruce 

budworm have become more extensive (Swetnam and 

Lynch 1989), more severe, and synchronous (Swetnam 

and Lynch 1993). Activity of western spruce budworm 

typically increases with periods of high moisture and 

decreases in drier periods (Swetnam and Lynch 1993). 

Outbreaks can cause substantial mortality and create 

increased surface fuel loading in a short period. The 

increased activity during wetter periods, in combination with greater fine fuel production during the same 

time, could increase the fire hazard in subsequent drier 

years. Currently, western spruce budworms are attacking 

trees on the Dixie, Fishlake, and Manti-LaSal National 

Forests. The outbreak has increased in acreage since 

1999 with up to 6,014 ha (14,861 acres) impacting the 

Dixie National Forest in 2004 (Forest Health Protection 

2000; Matthews and others 2005).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 21

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

The increased densities in susceptible host species 

has likely increased the incidence of Armillaria root 

disease, annosus root disease, and dwarf mistletoe in 

southern Utah mixed conifer forests as it has in other 

areas of the western United States (Hagle and Gohenn 

1988). This increase in infection has lead to increased 

susceptibility to insect attack and greater surface fuel 

loadings due to mortality of susceptible trees.

Ungulates—The increased density of wild and 

domestic ungulates has significantly reduced the 

herbaceous vegetation in mixed conifer stands. Aspen 

regeneration in many areas of the Rocky Mountains has 

been severely reduced due to herbivory by ungulates 

(Hart and Hart 2001; Kay 2001a; Kay 2001b; Romme 

and others 1995), although few studies have addressed 

this issue in Utah. Because aspen stands are natural 

firebreaks, the reduction in aspen stand coverage has 

increased the probability of larger fires in mixed conifer 

landscapes. Furthermore, the lack of fine fuels in the 

understory to carry surface fires has lengthened the fire 

frequency and has allowed greater quantities of larger 

diameter woody fuel accumulation. 

Selective Harvesting and Fire Exclusion—Selective 

harvesting and fire exclusion has caused dense, multistoried Douglas-fir and white fir to largely replace the 

ponderosa pine component in mixed conifer stands. As 

a result, mixed conifer forests are now very susceptible 

to western spruce budworm, root disease, bark beetles, 

dwarf mistletoe, and stand-replacing fires (Swetnam 

and Lynch 1989). The higher densities and contiguity of 

forests has led to large regional insect outbreaks that are 

more severe than in the past. Larger outbreaks will result 

in continued changes in forest structure, composition, 

and function, including creation of openings, depletion 

of large diameter trees, and an increase in fire hazard 

due to greater surface fuel accumulations (Wilson and 

Tkacz 1995). With continued fire exclusion in mixed 

conifer forests, surface and ladder fuels will continue to 

coalesce with crowns of overstory trees. This change in 

vertical fuel structure will further increase the probability 

of severe stand replacement crown fires.

Fire Regime Condition Classes

Under pre-settlement fire regimes, stocking of mixed 

conifer landscapes ranged from components containing shade-intolerant, early-successional species, such 

as aspen and pine, to those containing shade-tolerant, 

late-successional species, such as spruce and fir. Mixed 

severity fire regimes operating at a variety of scales 

were constantly affecting successional pathways creating 

landscapes of infinite diversity and character.

Fire Regime Condition Class 1 (FRCC 1)

The majority of FRCC 1 mixed conifer landscapes 

contains early successional species such as aspen, 

Douglas-fir, and ponderosa pine. They are more diverse 

in spatial structure, containing openings and only patches 

of mature, late successional species. Vertical structure is 

also more diverse. Some patches are very open and late 

successional species are represented predominately in 

seedling and sapling size class. Understory vegetation 

is more abundant in these patches. These landscapes 

have experienced recent fire within the period of mean 

fire return interval. Surface fuel loads are patchy due 

to the nature of mixed severity fire regime. Heavier 

woody fuels are present in areas that experienced crown 

fire, while lighter woody fuel loadings are present in 

areas experiencing surface fire. Continuity of surface 

and ladder fuels is diverse due to the patchiness of fireinduced mortality.

Fire Regime Condition Class 2 (FRCC 2)

Mixed conifer forests in FRCC 2 have overstories 

dominated by Douglas-fir and ponderosa pine, but the 

regeneration layer and midcanopy is dominated by 

white-fir, Engelmann spruce, and subalpine fir. Aspen 

patches are maturing and smaller in extent due to invasion by conifers. Understory herbaceous composition 

is not as diverse as in FRCC 1. These forests have been 

impacted by fire suppression and exclusion, logging, 

and grazing, but not to the extent as observed in FRCC 

3. Surface fuels and ladder fuels are more contiguous 

throughout the landscapes with fewer openings present. 

Surface fuel loadings are dependent on management 

activities. Areas that have not been harvested contain 

heavier surface fuel loadings due to large diameter snags 

that have fallen and accumulated on the forest floor. 

Many of these large diameter snags are not present in 

stands that were managed since settlement due to fuel 

management activities during harvest.

Fire Regime Condition Class 3 (FRCC 3)

We suggest the majority of mixed conifer forests are 

currently in FRCC 3. Since settlement, the impacts of fire 

suppression and exclusion, logging, and heavy grazing 

have brought extensive changes to mixed conifer forests 

in southern Utah. These activities have converted forests 

that once were dominated by a mixture of ponderosa 
22 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

pine, aspen, and Douglas-fir trees into forests dominated 

mostly by white-fir, Engelmann spruce, and subalpine 

fir. Prolific regeneration of these shade-tolerant species 

has created abundant ladder fuels, which allow surface 

fires to travel into the crown fostering crown fires. Mixed 

conifer forests are higher in basal area and density than 

pre-settlement forests, which have lead to higher susceptibility to insect and disease attack and a substantial 

increase in surface fuel loads. Natural fuel breaks once 

provided by aspen are now decreasing in size due to 

aging aspen groves and the lack of aspen regeneration. 

Fire regimes have departed from historical frequencies 

by up to three or four intervals (Heyerdahl and others 

2005; Heyerdahl and others 2006). The combination of 

all these factors may lead to large landscape crown fires 

instead of mixed mode fires (Fulé and others 2003; Fulé 

and others 2004).

Recommended Treatments

The historical fire regimes of mixed conifer forests 

are more complex than that of nearly pure ponderosa 

pine forests or spruce-fir forests. Because mixed conifer 

forests are made up of a number of species, the mix and 

distribution of each these species across landscapes will 

determine what approach is needed to restore them to a 

proper functioning condition. The goal in most mixedconifer forests should be to return stands back to earlier 

stages of succession and increase the age and spatial 

diversity within most landscapes. Our goal should be to 

remove the more shade-tolerant species in the understory 

and leave species in the overstory, such as ponderosa pine 

and Douglas-fir, which can withstand periodic fires. As 

with spruce-fir, some complete removal of portions of the 

landscape may be needed to provide natural firebreaks 

and reintroduce early successional species. If an aspen 

component is present, the goal may be to increase aspen 

regeneration to rejuvenate the aspen stands. Techniques 

outlined in the aspen chapter of this document should 

be used to accomplish those goals (Bartos, this volume). Such fuel breaks can be accomplished by cutting 

or prescribed burning to create openings in the forest. 

Aspen may re-occupy those areas, creating effective fuel 

breaks without creating permanent openings. Because 

Douglas-fir and ponderosa pine are not as susceptible to 

windthrow as Engelmann spruce, subalpine fir, or white 

fir, these species should be left whenever a thinning or 

biomass reduction operation occurs within stands. This 

also helps set succession back to an earlier stage by 

allowing Douglas-fir and ponderosa pine the light and 

growing conditions they need to regenerate. Development of stand structures should focus on creating forests 

with basal areas and densities that occurred historically, 

which includes a component of large-diameter trees 

that can withstand wildfires. Historic mixed conifer 

forests in the Grand Canyon and the Kaibab Plateau 

in nearby Arizona averaged 190 to 265 trees/ha (77 to 

107 trees/acre) with basal areas ranging between 17.6 to 

28.5 m2/ha (77 to 124 ft2/acre) (Fulé and others 2003, 

2004, 2006). Ponderosa pine dominated the basal area 

in each stand (Fulé and others 2004, 2006) or shared 

dominance with Douglas-fir and white fir (Fulé and 

others 2003). In historic southern Utah forests, trees 

greater than 20 cm (8 inches) ranged from 0 to 396 per ha 

(0 to 160 per acre) (Heyerdahl and others 2006).

Once a harvest, non-commercial thinning, or mechanical mastication treatment has removed the undesirable 

trees, a prescribed burn can be used to clean up surface 

fuels and ensure that few white fir or subalpine fir seedlings survive. Existing fir seedlings will soon overwhelm 

the understory and quickly grow into the overstory if 

this is not done. Experience in southwestern Colorado on 

the San Juan National Forest has shown that due to their 

rapid growth in partially shaded conditions, understory 

subalpine fir saplings can dominate the canopy within 

25 years (data on file, RMRS, Wayne Shepperd).

After treatment, growing stock should be less than or 

equal to 30 percent of maximum stand SDI when calculated on a diameter class basis. The most shade-tolerant 

species should be removed first, leaving the less shadetolerant, most fire resistant species, such as ponderosa 

pine and Douglas-fir. Treatments should not be uniform 

across the landscape, but should range from complete 

removal of all trees in some patches, thinning in others, 

and leaving intact forested patches to emulate the effects 

of mixed-severity fire regimes that created the natural 

spatial diversity of mixed conifer forests. The 30 percent 

maximum SDI stated above should therefore be an average of stocking across the entire landscape. The tools 

and techniques for developing treatment prescriptions 

discussed earlier in the ponderosa pine forests section 

of this chapter also can be used in mixed conifer forests. 

However, the diverse nature of mixed conifer forests 

may require that the prescriptions be developed at the 

landscape scale rather than the stand scale. Any mechanical removal of trees will result in some scarification of 

the forest floor that will provide ideal seedbeds for the 

recruitment of new seedlings. Subsequent thinning or 

prescribed burning will periodically be needed to keep 

such regeneration from restocking and overstocking the 

forest. It is important to keep in mind that silvicultural 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 23

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

prescriptions must include a flexible time table of future 

thinnings and prescribed burn treatments.

Biomass reduction thinnings do not necessarily have to 

be evenly spaced. The same goals can be accomplished 

through irregular spacing and grouping of trees, provided 

that the groups are not contiguous to larger trees in a 

manner that they would provide live fuel ladders. Using 

an irregular spacing allows biomass fuel reduction goals 

to be met while retaining wildlife habitat attributes, 

such as hiding cover, and the juxtaposition of various 

vegetation structural stages classes within the forest 

canopy (Reynolds and others 1992), thereby leading to 

conditions more similar to pre-settlement.

The authors recognize that not all the acreage in 

southern Utah will be accessible for mechanical treatment. Prescribed burning is the most feasible method 

of creating diversity within many landscapes. In addition to prescribed under-burning, prescribed crown fire 

and wildland fire use can be used in late seral, mixed 

conifer, and some spruce-fir forests where aspen is 

present. Burning landscapes with a mixed severity fire 

that includes stand replacement patches can achieve the 

diversity that was present historically. When repeated 

through time, such burns can essentially recreate the 

vegetation mosaics that once existed within southern 

Utah mixed conifer landscapes. Recent wildfires in 

southern Utah that have resulted from attempts at such 

burns should not prevent their future use. Such setbacks 

are to be expected as we learn how to safely re-introduce 

fire into these complex systems.

Using prescribed under-burning to reduce surface fuel 

loadings will likely create a conflict between achieving 

the fuel reduction goals, maintaining healthy forests, and 

achieving a sound silvicultural treatment. If the prescribed 

burns occur under dry enough conditions such that the surface fuels are completely consumed in the understory, the 

fires are likely to be so severe that they will harm the roots 

of the living trees and plants. This has been demonstrated 

in ponderosa pine (Sackett and Haase 1998) and can be 

expected to be true for other shallow-rooted species such as 

Douglas-fir, aspen, and spruce. Avoidance of root damage 

is best accomplished by burning in late spring when soil 

moisture is at its highest and larger diameter woody fuels are 

not completely consumed because of high moisture content 

(Shepperd 2004). Even though the larger diameter woody 

fuels and some live fuels are not completely consumed, 

such burns can apparently still have a beneficial effect. 

A portion of the recent Hayman fire in Colorado in June 

2002 was slowed and essentially stopped by the Polhemus 

prescribed burn that occurred in Fall 2001 (Martinson and 

others 2003). Although there were ample live crown fuels 

left following the earlier prescribed burn, the reduction of 

surface fuels was such that it stopped the Hayman crown 

fire within the treated area. While we recognize that this 

would not occur in all instances, even an incomplete prescribed burn has some beneficial effect on preventing or 

slowing the spread of crown fires. We believe that combining mechanical treatments where possible with prescribed 

burning will allow us to better choose the portions of the 

landscape where we can expect prescribed crown fires to 

be contained.

In summary, any silvicultural technique that will result 

in a diversity of mixed conifer successional stages within 

southern Utah landscapes will probably be beneficial in 

reducing the risk of large landscape stand replacing crown 

fires in mixed conifer forests. Removing portions of forests or trees from landscapes will increase the amount of 

forage that is produced on those landscapes for wildlife 

and livestock. This is especially important in southern 

Utah where livestock numbers have not been reduced as 

landscapes have been filled in with trees and understory 

production has decreased. Reduction of forest biomass 

from all forests within southern Utah is also likely to have 

a beneficial effect on the water balance within those landscapes (Shepperd and others 1992). We believe that good 

wildfire management is also good ecosystem management. 

It would be ideal to treat every acre to reduce the risk of 

wildfires operating outside the historical range of variation; 

however, the reality is that not every acre will be treated. 

Instead, we should prioritize treatments in some areas and 

allow wildfires to burn in others.

Engelmann Spruce and

Subalpine Fir Forests

Mixed Engelmann spruce and subalpine fir forests 

comprise the upper extent of forest vegetation in southern Utah, occupying the coldest and wettest sites in the 

altitudinal continuum of ecologic conditions in the area. 

Precipitation regimes in these forests are dominated by 

snow, which can occupy these sites for 6 to 8 months of 

the year. Spruce-fir forests can exist on-site for extremely 

long periods, on average 500 to 600 years (R. Alexander 

1987), with reports of even longer periods (Brown and 

others 1995). Harsh climates and short growing seasons 

result in infrequent, but large-scale disturbances including fire, insect attacks, wind, and avalanches, which 

historically interacted to create coarse-scaled mosaics 

of different aged patches on the landscape (Baker and 

Veblen 1990).
24 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

Historical Conditions

Historical descriptions of spruce-fir forest structure 

and density are limited. One study on the North Rim of 

the Grand Canyon suggests that compared to contemporary forests, historic spruce-fir forests in the southwest 

were significantly less dense (16 to 24 percent) and had 

lower basal area (36 to 46 percent) with densities about 

150 trees/ha (60.7 trees/acre) and basal areas only 10 m

2

/ha 

(43.5 ft2/acre) (Fulé and others 2003). Reconstructed 

ca. 1860 spruce-fir forests on the Fishlake and Dixie 

National Forest show increases in density similar to the 

Grand Canyon site (Heyerdahl and others 2006). Most 

of the recruitment has occurred in the past 100 years 

suggesting these trees are smaller in diameter and height 

than pre-settlement forests (Heyerdahl and others 2006). 

A 1911 survey on the Manti-LaSal National Forest indicated that spruce reached 24.4 m (80 ft) tall and 61 cm 

(24 inches) in diameter, but the average size was 18 m 

(60 ft) tall and 46 cm (18 inches) in diameter (Ogle and 

DuMond 1997). Spruce-fir forest structure contained a 

variety of age classes and successional stages in varying 

patch sizes (Ogle and DuMond 1997).

Fuel structure under spruce-fir forests often promotes 

stand replacement fires. Surface fuel loads were probably 

much higher than those found at lower elevation montane 

forests due to slower decomposition rates (Uchytil 1991). 

Needles are small and fine and form a compact fuel bed 

in which fire spreads slowly and fuel beds accumulate 

under the narrow-crowned trees (Uchytil 1991). Large 

diameter woody debris and snags are also prevalent in 

spruce bark beetle outbreak areas. Mielke (1950) reported 

that 84 percent of spruce bark beetle killed trees were 

still standing after 25 years, mostly in large diameter 

classes. Relatively few trees fell within the first 10 years 

after beetle attack, but trees did fall after being infected 

with basal and root rots (Fomes pinicola).

Spruce-fir forests are predominately subject to longinterval stand replacement fire regimes rather than the 

more frequent low-intensity surface and mixed-mode 

fires occurring in lower elevation southern Utah coniferous forests. Mean fire return intervals (MFRI) in 

spruce-fir do appear to vary by region. Arno (1980) 

estimates MFRI of 50 to 130 years in the Northern 

Rockies. Veblen and others (1994) reported a MFRI of 

202 to 241 years in northwestern Colorado, while Peet 

(1981) reported a MFRI of 200 to 400 years for the 

Front Range of Colorado. In northern Utah, the overall 

MFRI was 41.3 years, however, this average is somewhat 

deceptive, as no fires occurred between 1700 and 1855 

(pre-settlement), followed by a 9-year fire interval from 

1856 to 1909 (settlement), and then a 79-year interval 

from 1910 to 1988 (suppression period). White and Vankat 

(1993) reported a MFRI of 70 to 250 years on the North 

Rim of the Grand Canyon. Lang and Stewart (1910) 

reported that the North Rim contained “vast denuded 

areas, charred stubs and fallen trunks.” The old fires 

extended over large areas at higher altitudes covering 

several square miles on either side of Big Park, with 

numerous smaller irregular areas over the remainder of 

the park (Lang and Stewart 1910: 18 to 19 cited in Fulé 

and others 2003).

Fires were predominately of three types: 1) lightningignited fires that consumed individual trees or small 

patches of forests (the most common type), 2) crown 

fires that killed most overstory trees as well as saplings 

and seedlings, (Arno 1980; Baker and Veblen 1990), or 

3) patchy fires that burned as surface fires for a short 

distance and then burned the overstory trees for a short 

distance resulting in a coarse-grained mosaic of dead 

trees or open areas with alternating patches of surviving 

trees (Baker and Veblen 1990).

Several morphologic and ecologic factors of Engelmann spruce and the true firs contribute to the longinterval stand replacement fire regimes. Spruce and fir 

species have thin bark that is not well insulated from fire 

damage, which allows the trees to easily be killed by low 

intensity fire. The species have long, dense crowns that 

typically reach close to the ground. This creates abundant 

live fuel ladders that allow surface fires to climb into 

the upper forest canopy. This is especially prevalent in 

older stands and where spruce bark beetle outbreaks 

have initiated a regeneration response (due to openings 

in the canopy) creating multi-storied structures. The 

accumulation of snags and heavy surface fuels resulting 

from beetle outbreaks persist for a long period of time 

(Brown and others 1998), and further contributes to a 

stand replacement fire regime.

Long fire return intervals in spruce-fir forests are 

primarily a result of their occurrence at high elevations 

that receive more precipitation and cooler temperatures 

throughout the growing season. In these snow dominated 

precipitation regimes, it is difficult for fire to burn in a 

normal year. Following snowmelt in the spring, there is 

abundant soil moisture throughout the growing season 

and cool temperatures keep understory fuels moist and 

hard to ignite. Therefore, fires are more likely to occur 

in the fall, late summer, or after an unusually dry winter 

instead of in the spring.

Windthrow is second to fire as the landscape-wide 

disturbance affecting spruce-fir forests. The amount 

and degree of damage varies depending upon the wind 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 25

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

event, stand structure, species composition, topography, 

and stand’s fire history (Kulakowski and Veblen 2002; 

Veblen and others 2001). For instance, older spruce 

trees are shallow-rooted and prone to blowdown (R. 

Alexander 1987); however, young post-fire stands are 

less affected due to shorter trees, less canopy gaps, and 

often a greater component of aspen (Kulakowski and 

Veblen 2002). With severe blowdowns, huge surface 

fuel loadings can result that increase the severity of 

any fire occurring in the area for years after the event. 

In addition, outbreaks of spruce bark beetle are often 

triggered by blowdowns (Schmid and Frye 1977).

Spruce bark beetles play a major role in the development and maintenance of high elevation coniferous 

forests in Utah. The beetles, in association with fire, 

help maintain a variety of successional stages and age 

classes of spruce-fir forests across the landscape. Endemic populations live in windthrown trees and probably 

have little effect on forest dynamics. In areas with an 

extensive windthrow event, however, ample food supplies allow rapid buildup of beetle populations (Massey 

and Wygant 1954; Wygant 1958). These epidemic levels 

of spruce bark beetle populations are reported to have 

occurred on average every 116 years (since 1700) in presettlement spruce-fir forests of central Colorado (Baker 

and Veblen 1990; Veblen and others 1994). Epidemic 

populations attack mature, larger diameter (>46 cm 

[18 inches]) Engelmann spruce trees (Schmid and Frye 

1977) and as a result, average tree diameter, height, age, 

and densities of stands are reduced, dominance in basal 

area shifts from Engelmann spruce to subalpine fir, 

and accelerated growth of residual trees are observed 

(Schmid and Frye 1977; Veblen and others 1991).

Ungulates were likely to have had minimal influence 

on subalpine forests prior to European settlement. The 

presence of large predators probably kept deer and elk 

populations at much lower levels than today and the lack 

of domestic livestock further insured minimal impact.

Disturbances existing prior to settlement affected 

community characteristics of subalpine forests and their 

position on the landscape in several ways. High elevation 

fires set back succession from forest to meadow and 

maintained existing meadows. Some fires in forest communities near timberline created herb or shrub dominated 

seral communities, which are slow to regenerate back 

to spruce (Huckaby 1991). Fires occurring at the lower 

elevation interface with mixed conifer forests allowed 

those sites to regenerate to aspen and Douglas-fir, thus 

maintaining a wider range for those forests than occurs 

today. Fires that occurred in spruce-fir beetle-killed or 

windthrow areas were likely to have been extremely 

severe, insuring that those areas remained open for long 

periods of time.

Current Conditions

Spruce-fir forests have expanded into the mixed coniferous forests, as well as into high elevation meadows 

of southern Utah (Fulé and others 2003; Heyerdahl and 

others 2005; Heyerdahl and others 2006; USDA Forest Service, Fishlake National Forest 1999; White and 

Vankat 1993). Expansion into the lower elevations is a 

result of succession from aspen forests to mixed conifer forests due to fire suppression and because aspens 

provide suitable habitat for the establishment of shadetolerant conifers (Shepperd and Jones 1985). Expansion 

into higher meadows over the past 100 years is largely 

due to grazing by domestic animals or wildlife, which 

scarified seedbeds and reduced competition between 

tree seedlings and the herbaceous community (Allen 

1989; Moir and Huckaby 1994). In addition, if climatic 

warming is occurring, it would increase the length and 

warmth of the growing season, possibly improving 

seedling survival (Moir and Huckaby 1994; Moir and 

others 1999). Furthermore, fire suppression has allowed 

seedlings to establish on the edges of meadows and reduce the extremely high soil moisture making it easier 

for additional seedlings to establish in the center of the 

meadows.

The structure of spruce-fir forests in southern Utah is 

predominately uneven-aged (fig. 5) (Hanley and others 

1975; Mielke 1950; O’Brien and Brown 1998; O’Brien 

and Waters 1998; O’Brien and Woudenberg 1998; Pfister 

1972). Engelmann spruce is the major species, followed 

by subalpine fir and aspen (Bradley and others 1992; 

Fulé and others 2003). Pure Engelmann spruce stands 

and spruce-fir forests (where spruce and subalpine fir 

are codominant) consist of all ages, although the majority of these trees are 51 to 150 years old (O’Brien and 

Brown 1998; O’Brien and Waters 1998; O’Brien and 

Woudenberg 1998). In addition, there are some Engelmann spruce trees 151 to 250 years old in the Dixie 

National Forest (O’Brien and Brown 1998). Surveys on 

the Markagunt and Aquarius plateaus in southern Utah 

in the early 1970s revealed that subalpine fir stocking 

was uneven-aged, but not all-aged. A 50- to 70-year old 

prolific regeneration component existed in the understory, which corresponded with a spruce beetle outbreak 

in the 1930s. Subalpine fir was also present in the 70- to 

130-year old age class indicating its ability to maintain 

itself under the Engelmann spruce canopy for long 
26 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

periods. The predominance of subalpine fir seedlings 

in the understory of southern Utah forests (Hanley and 

others 1975) is indicative of its ability to successfully 

reproduce on duff-covered seedbeds, but spruce’s average longevity over subalpine fir keeps it dominant in the 

overstory. Fulé and others (2003) found 73 percent of 

their plots on the North Rim had fire-initiated groups 

of trees dominated by Engelmann spruce dating back 

to 1788.

Diameter distributions for Engelmann spruce and subalpine fir in southern Utah are not available, but can be 

found for the entire state of Utah (O’Brien 1999). On a 

landscape level, the diameter distribution of Engelmann 

spruce and spruce-fir forests exhibit an uneven-aged 

structure with a large numbers of trees less than 12 cm 

(5 inches) in diameter.

Fuels/biomass data collected by Fulé and others 

(2002c) from North Rim spruce-fir forests report duff 

depths of 2.3 cm (0.9 inches) and litter depths of 0.8 cm 

(0.3 inches) with downed wood loadings of 141 Mg/ha 

(63 tons/acre), the majority of which was split between 

1000-hour rotten and sound fuels. Shrubs and small trees 

predominate in southern Utah spruce-fir forests. Abies 

lasiocarpa/Ribes montigenum habitat type occupies 

approximately half of the spruce-fir in Utah (Pfister 

1972). Annual herbaceous production ranged from 

1,201 kg/ha (1,072 lbs/acre) to 2,541 kg/ha (2,267 lbs/

acre) in similar communities in central Utah with forbs 

contributing to over 80 percent of the production (Ralphs 

and Pfister 1992).

Due to the long history of water yield research in 

subalpine forests (Leaf 1975; Troendle and others 

1988), more is known about the hydrologic function of 

spruce-fir forests than any other western forest type. 

Streamflow and water yield in these snow-dominated 

precipitation regimes is driven by the density of the 

forest, which principally affects interception of snow 

by the canopy and subsequent sublimation back to the 

atmosphere (Troendle and others 1988). Therefore, any 

disturbance that removes trees from the forest will affect 

the hydrologic water balance of the system. Potential 

effects due to fire have been studied in a subalpine 

dominated forest in the Shoshone National Forest in 

Wyoming after a fire (Troendle and Bevenger 1996). 

Since most high elevation watersheds have peak annual 

discharge at snowmelt, they found peak annual discharge 

unchanged by the fire. However, water yields increased 

25 percent on the burned watershed. Suspended sediment production in terms of concentration was two times 

greater in the burned watershed and suspended load was 

four times that of the unburned watershed. The highest 

concentration occurred during an individual summer 

storm event. These findings indicate that large-scale 

crown fires in southern Utah spruce-fir forests could 

significantly alter the water balance of the system until 

sufficient vegetation recovery occurs to reduce sediment 

transport.

Current Disturbances

Post-settlement spruce-fir forests are still subjected to 

the same disturbances observed in pre-settlement times, 

mainly stand-replacing fires, blowdowns, and insect attacks. However, new disturbances such as grazing, logging, increased fire ignitions, and fire suppression have 

emerged in these high-elevation coniferous forests.

Figure 5—Uneven-aged Engelmann spruce (Picea 

engelmannii) and subalpine fir (Abies lasiocarpa) stand 

with an understory dominated by heartleaf arnica (Arnica 

cordifolia) and tall bluebell (Mertensia arizonica).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 27

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

Ungulates—Between the time of settlement and the 

early 1900s, settler activities impacted spruce-fir forests. These activities continue to influence the disturbance regime of present times. First, the introduction 

of livestock grazing in the high elevation meadows 

had several impacts on the ecosystem. During early 

settlement, overgrazing led to loss of topsoil that created unproductive areas because no vegetation could 

reestablish. This led to several summer-time floods and 

sheet erosion (T.Alexander 1987; Keck 1972; Ogle and 

Dumond 1997). Although overgrazing reduced fine fuel 

loads, the presence of sheepherders and loggers actually 

increased the fire frequency in some spruce-fir forests 

(Bird 1964; Roberts 1968; Wadleigh and Jenkins 1996) 

until the Forest Service implemented fire suppression 

after 1910. Early logging of spruce-fir forests was limited to accessible areas and only the best species were 

removed leaving the subalpine fir (Ogle and Dumond 

1997). Today, we are left with large acreages of young, 

small diameter subalpine fir and Engelmann spruce.

Spruce-fir forests continue to be heavily use as summer range in southern Utah. The increased acreage 

and density of spruce-fir discussed earlier translates to 

a decrease in overall forage available on allotments in 

spruce-fir forests. Furthermore, species composition of 

herbaceous plant communities continues to change as a 

result of preferential selection of specific species by the 

livestock. Continuation of traditional livestock grazing 

rates in these allotments only puts more pressure on a 

dwindling grazing resource and reduces potential for 

fire ignition.

Fire—Although fire intervals for spruce-fir forests are 

typically several hundred years, a portion of spruce-fir 

forest acreage would probably have burned in years of 

severe drought if not for fire suppression efforts. These 

fire suppression effects have only stalled the inevitable 

large stand-replacing fire. Large stand-replacing fires 

during drought years are normal for spruce-fir forests. 

One notable fire is the Outlet fire, which occurred within 

Grand Canyon National Park in May 2000. The extreme 

drought and occurrence of numerous other fires in subalpine forests throughout the western United States in 

recent summers are strong evidence that the potential 

exists for more large scale fires to occur in southern 

Utah.

Insects and Disease—Insect activity has also dramatically increased in southern Utah in recent years 

(fig. 6). Spruce bark beetle populations have been at 

epidemic levels since 1991 on the Dixie National Forest 

Figure 6—Engelmann spruce (Picea engelmannii) at Cedar Breaks National Monument, UT, killed by 

recent spruce beetle (Dendroctonus rufipennis) activity. 
28 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

and since 1989 (Knapp and others 1991) on the MantiLaSal National Forest (Knapp and others 1989). Large 

areas of the Dixie, Fishlake, and Manti-LaSal National 

Forests have experienced severe disturbances caused by 

spruce bark beetle. Most spruce trees greater than 15 to 

20 cm (6 to 8 inches) were killed during these outbreaks 

(Dymerski and others 2001; Forest Health Protection 

2000; Matthews and others 2005). For example, 73 

percent of spruce trees greater than 12 cm (5 inches) 

in diameter on the Manti-LaSal National Forest were 

killed over a period of about 5 to 7 years – a 47 percent 

decline in the spruce component (Dymerski and others 

2001; Samman and Logan 2000). From 2000 to 2004, 

spruce bark beetles have killed over 366,000 trees on 

over 40,469 ha (100,000 acres) of southern Utah National 

Forests (Forest Health Protection 2000; Matthews and 

others 2005). Hazard ratings indicate that 62 percent 

of spruce-fir forest types in the Manti-LaSal National 

Forest, 45 percent in the Dixie National Forest, and 42 

percent in the Fishlake National Forest are at moderate to high risk of attack by bark beetles (O’Brien and 

Brown 1998; O’Brien and Woudenberg 1998; O’Brien 

and Waters 1998).

Mortality of subalpine fir trees in the National Forests in southern Utah is also on the increase (Matthews 

and others 2005). From 2000 to 2004, western spruce 

budworm had defoliated over 3,287 ha (8,000 acres) 

on the Fishlake National Forest and over 24,281 ha 

(60,000 acres) on the Dixie National Forest (Forest 

Health Protection 2000; Matthews and others 2005). 

Defoliation on the Manti-LaSal National Forest remains low with approximately 243 to 485.6 ha (600 to 

1,200 acres) per year since 2002 (Matthews and others 

2005). Decline and mortality of subalpine fir resulting 

from a complex of secondary biotic agents have affected 

over 24,281 ha (60,000 acres) in the National Forests 

of southern Utah from 2000 to 2004 (Forest Health 

Protection 2000; Matthews and others 2005). These 

secondary agents include twig beetles, woodborers, 

engraver beetles, secondary bark beetles, root diseases, 

cytospora canker, and fir broom rust (Forest Health 

Protection 2000).

Fire Regime Condition Classes

The long fire return interval in spruce-fir forests and 

the wide range of possible forest structure conditions 

that may naturally develop between stand replacing 

fires pose somewhat of a dilemma in assigning FRCC 

classes to these forests. Because most existing forest 

conditions fall within the past range of natural variability, 

we believe most spruce-fir forests could be classified 

as FRCC 1. What may move these forests into FRCC 2 

or 3 is the distribution of forest structural stages across 

the landscape. We believe the continuity of fuels at the 

landscape scale is what differentiates the FRCC class 

in spruce-fir forests. Fire exclusion and subsequent 

ingrowth into gaps have resulted in more contiguous 

surface and crown fuels at the landscape scale. Although 

these surface and crown fuel accumulations are not out 

of the historical range on a stand level scale, at landscape 

scales, fires would likely be larger than what occurred 

in the past. Therefore, we believe that FRCC 2 and 3 

landscapes would be those that are currently completely 

forested as opposed to being partially forested 200 to 

300 years ago. A hundred or more years of fire exclusion in spruce-fir forests since settlement has also likely 

altered the distribution of age classes over the landscape, 

which has moved a greater portion of forests toward 

fuels conditions associated with older forests that are 

described below.

Fire Regime Condition Class 1 (FRCC 1)

Spruce-fir forests classified as FRCC 1 span a variety 

of successional stages and age classes distributed across 

the landscape. These forests include young post-fire 

forests, forests after insect epidemics, middle-aged and 

mature multi-storied forests, and overmature old growth 

forests. Each stage is a result of past disturbance history 

and is dispersed on the entire landscape intermixed with 

open meadows.

Young stands, less than 100 to 150 years old, established after a stand replacement fire. Such forests are 

clumpy with canopy gaps and openings between stands. 

Low surface fuel accumulations and a rich and diverse 

understory exist in the openings between the clumps. 

Heavy surface fuel loadings occur within the clumps 

and consist of partially decomposed and burned, fallen 

snags from the previous stand. However, these surface 

fuels are not contiguous.

As the canopy in these young forests begins to close, 

clumps begin to merge and expand. The outer edges of 

the clumps have younger fully crowned trees, but the 

inside of the clumps (approximately one ha [2.5 acres] in 

size) have relatively young trees with different diameter 

distributions (due to density). These forests do not have 

an appreciable buildup of surface fuels because mortality 

is low at this stage in their life cycle.

As the spruce matures, the stand is susceptible to disturbances such as windthrow, spruce beetle attacks, and 

disease. These disturbances help maintain the canopy 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 29

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

gaps within the forest and create multi-storied stands. 

Subalpine fir also becomes an important component 

of the forest species composition as the spruce forest 

matures. Subalpine fir increases the presence of ladder 

fuels and increases surface fuel accumulation due to its 

short life span.

Overmature and old growth spruce-fir forests are 

subjected to repeated disturbances (fire, insects, and 

disease) that result in partial mortality and heavy surface 

fuel accumulation. These forests typically have fewer 

canopy gaps and contain a significant component of 

younger fir and numerous standing snags.

If the mature, overmature, or old growth spruce-fir 

forests have been recently attacked by spruce beetle, 

the majority of large diameter trees are dead. The accumulation of snags and heavy surface fuels resulting 

from beetle outbreaks persist for a long period of time 

(Brown and others 1998) and further contributes to a 

stand replacement fire regime. The dominant species 

of these forests changes from spruce to fir and creates 

abundant live fuel ladders that allow surface fires to 

climb into the upper forest canopy.

Fire Regime Condition Classes 2 and 3 

(FRCC 2 and FRCC 3)

Spruce-fir forests classified as FRCC 2 or FRCC 3 

consist of a landscape with contiguous forest stocking, 

a lack of meadows, and a lack of successional and age 

class diversity. These forests have species composition 

and fuel accumulation patterns that are not consistent 

with patterns observed after a natural disturbance. The 

degree of departure of these factors determines the 

FRCC classification. For example, the expansion of 

spruce-fir forests into meadows or the upper elevations 

of mixed conifer forests due to fire exclusion could be 

classified as FRCC 2. Multi-watershed contiguous forests 

that are mature and contain no canopy gaps or other natural 

features that could interrupt the continuity of fuels over an 

extremely large area might be classified as FRCC 3.

Past harvesting techniques impact FRCC characteristics on a stand level. If they were implemented on a 

large portion of the landscape, the FRCC classification 

could shift toward 2 or 3. Past harvests often removed 

the larger diameter spruce trees, which left the smaller 

spruce and fir trees behind and effectively changed 

the age structure and species composition of the forest. Similar structural changes occur after a spruce 

beetle infestation except there is a major difference in 

the post-disturbance fuel complex since all of the dead 

trees remain on-site. After a spruce beetle infestation, 

large diameter snags are present in the stand, eventually 

fall, and then contribute to the nutrient pool, surface 

fuel loading, and eventually to high intensity wildfire. 

Salvage logging removes the large diameter snags and effectively reduces large diameter surface fuel loadings and 

potential nutrient inputs. In any case, harvested stands 

would have higher loadings of fine fuels if activity fuels 

were not removed or treated. Modern whole-tree harvest 

methods alter fuels differently. Since all unmerchantable trees and tops are removed, surface fuels are not 

as heavy as in past harvests. Early logging of spruce-fir 

forests in Southern Utah was limited to accessible areas 

(Ogle and Dumond 1997), so fuels were not altered on 

all portions of landscapes.

Recommended Treatments

Although most existing spruce-fir forests in southern 

Utah are likely to be in FRCC 1, or possibly FRCC 2, losing 

key ecological components that define spruce-fir forests 

is still possible if extremely large patches of these forests 

are burned completely in a single stand replacement fire 

event. Spruce and fir are generally shade tolerant species 

that establish best in partial shade. They do not establish 

well or quickly in the full sunlight existing after a stand 

replacing fire. Because Engelmann spruce seeds require 

mineral soil to germinate, subalpine fir seedlings usually 

outnumber spruce seedlings in the understory of spruce-fir 

stands (Alexander and others 1984).

Successional pathways following fire can vary 

dramatically (Stahelin 1943). Since aspen acts as a nurse 

crop for young spruce and fir seedlings, if an aspen 

component is present, spruce-fir can recover in 100 to 

150 years providing a residual seed source is present 

(Shepperd and Jones 1985). If aspen is missing and the 

burn is severe, a meadow opening is created that will 

take much longer to recover to spruce-fir. In high elevation forests, it takes over 40 years for spruce trees to 

reach dbh (1.37 m [4.5 ft tall]) under open conditions. 

This growing period increases on south and west facing 

slopes and decreases on north and east aspects where 

water is not a limiting factor (R. Alexander 1987). Spruce 

and fir seedlings are dependent on wind dispersal, so if 

fires are large in scale, it takes a long time for seedlings 

to establish and create a new forest. Effective seeding 

distance is four to five times tree height (R. Alexander 

1987). Very severe fires that create large openings will 

result in a type conversion to meadow or aspen for hundreds of years. Moderate fires will have some spruce 

and fir survival, which will increase the reestablishment 

rate of spruce-fir into burned openings.
30 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

As mentioned above, past logging in spruce-fir forests 

will likely affect fuels and fire behavior. Increasing the 

spacing of trees and allowing more light to reach the 

understory will further affect fire behavior by increasing understory vegetation production and wind speeds 

to promote a warmer, drier microclimate. However, a 

counteracting effect of logging to decrease crown fuels 

would be the increased accumulation of woody fuels in 

the understory, which would increase the likelihood of 

high severity fires. Initially, any disturbance resulting 

in the partial removal of trees would likely decrease 

fire spread rate. However, the tendency of spruce-fir to 

develop a multi-storied structure following disturbance 

would eventually negate this effect.

A possible management goal in spruce-fir landscapes 

should be to allow stand replacement fires to occur while 

reducing the risk of entire landscapes succumbing to 

such fires. Techniques that could be used to accomplish 

this goal include thinning to reduce stocking of younger 

forests, complete cleanup of slash with whole tree harvests or mechanical harvesting, or using techniques 

such as block or strip cutting that add fuel breaks in the 

forested landscape. All of these techniques would break 

up the continuity of fuels and increase the chance that 

a portion of the forest would remain after a fire.

Use of prescribed fire in spruce-fir forests is less 

likely to be beneficial in this ecosystem because of 

the sensitivity of the species to damage by fire and the 

heavy accumulation of woody fuels and duff layers. 

Even cool, slow burning fires would damage the roots 

and bark of the trees (Bradley and others 1992) and 

would likely get into the crowns. Shelterwood systems, 

if used in some portions of the landscapes, would break 

up the age class structure, create gaps between crowns, 

and possibly lessen the likelihood that stand replacing 

fires would sweep over large areas. However, use of 

mechanical treatment is limited to areas that are roaded 

and accessible. Many spruce-fir forests are on steep 

slopes in inoperable terrain, or are in wilderness areas 

where management options are quite limited.

Spruce-fir Treatments at the 

Landscape Scale

Because spruce and fir are species that evolved 

with stand replacement fire regimes, underthinning 

or reduction of density is not recommended to protect 

individual stands from catastrophic replacement. Almost 

any spruce-fir forest is capable of burning. Even young 

forests can burn because their crowns generally reach 

down to the ground. Breaking up landscapes of spruce-fir 

into different successional stages is one way to increase 

the chances of spruce-fir survival at large scales. This 

would be akin to practicing uneven-aged management 

at a large scale, similar to what occurred naturally in 

the past. This would allow fires to burn through some 

areas and not others resulting in spruce-fir stands interspersed with pure aspen stands and openings such as 

alpine meadows or shrublands.

Openings created by harvesting large areas in sprucefir forests will take decades to regenerate if they are 

greater than four to five times tree height wide (the 

effective seed distance) (R. Alexander 1987). Openings 

this large could be used as fuel breaks since it would 

take a long time to regenerate. Because of the heavy 

surface fuels under many spruce-fir stands, ensuring that 

openings be maintained in a meadow-like appearance 

might require broadcast burning recently cut units to 

reduce surface fuels and eliminate residual small trees. 

Ample evidence of the utility of this technique is evident 

from older, large spruce-fir cut blocks created in the 

central Rockies in the late 1950s to early 1960s. Some 

of these areas are still inadequately restocked 40 years 

after clearcutting and broadcast burning.

The creation of such openings should consider the 

landscape form so that a natural appearance is maintained 

and management activities do not appear to be geometric, 

irregular, or artificial in appearance. Areas selected for 

spruce-fir forest removal to create openings should be 

where former openings likely once existed. Examples 

include sides of valleys, bottom areas, and other areas 

that were likely to be unstocked at sometime in the past. 

One way to examine the landscape for spatial distribution 

patterns might be to use a technique similar to that of 

Moir and Huckaby (1994). They determined age cohorts 

on the Fraser Experimental Forest in central Colorado 

by identifying where past disturbances have occurred. 

Mangers could utilize such patterns in planning where 

to create fuel breaks in landscapes.

In some stands, the fir component of spruce-fir forests 

in southern Utah will affect fuel loadings and susceptibility to crown fire over time. Subalpine fir is much 

shorter lived than spruce and is likely to cycle in and 

out of spruce-fir stands through time (Aplet 1988; data 

on file RMRS). Assuming similar dynamics occur in 

southern Utah, spruce-fir forests are likely to have 

increased surface fuel loadings when the fir dies out 

and creates forests that are more susceptible to stand 

replacement fires. A similar relationship exists where 

aspen is a component of spruce-fir forests. As aspen 

dies out, stands accumulate large buildups of surface 

fuels. This creates areas that are more susceptible 

to stand replacement fire. However, most higher 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 31

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

elevation spruce-fir forests are more likely to contain 

higher portions of spruce and be stocked with very old 

trees (Brown and others 1995), this indicates that fire 

is a rare occurrence at or near timberline. Removing 

the fir component from mixed stands is unlikely to 

reduce long-term fire risk. Unless all fir seed sources 

are removed, fir will quickly reestablish and contribute 

to a live fuel problem in the future.

Spruce-fir Treatments at the Stand Scale

We believe target stand structures are inappropriate in spruce-fir because FRCC is really applicable 

at the scale of landscapes instead of stands. However, 

management activities that add age classes and spatial 

diversity to the landscape would ultimately maintain 

FRCC 1 stands and benefit general ecosystem sustainability. Any management activity should leave a similar 

surface fuel complex that would naturally have occurred 

in a given stand configuration. The complete cleanup 

of slash mentioned above for fuel break creation would 

not apply here since these areas are part of the general 

forested matrix in which an almost infinite variety of 

fuel loadings could naturally occur. If the decision is 

made to treat fuels at the stand scale for reasons other 

than restoration (for example, protection of infrastructure), there are still limitations to what can be done in 

spruce-fir forests.

Thinning from below in mature spruce-fir forests, 

similar to what is done under even-aged shelterwood or 

seed tree regeneration in other forest types, is of limited 

usefulness in lessening the risk of catastrophic wildfire. 

The amount of material that can be safely removed 

without increasing risk of windthrow (R. Alexander 

1987) is such that it is doubtful sufficient space could be 

maintained between crowns to appreciably lessen crown 

fire risk. At the same time, logging would likely increase 

the amount of downed woody fuel on the forest floor. 

Careful cleanup of logging slash, as well as monitoring 

of the overstory to remove subsequent windthrow, would 

be essential to keep surface fuel loadings at reasonable 

levels; however, these types of treatments are very costly. 

Since such techniques are normally used to regenerate 

spruce-fir forests, ample regeneration could be expected 

within a few years. The resulting increase in live fuel 

loadings would then require precommercial thinning to 

reduce risk of crown fire, further increasing the cost.

Uneven-aged management that maintains overall health 

and vigor of spruce-fir forests by periodically reducing 

stocking levels through time seems to be a reasonable 

way to meet a variety of ecosystem objectives for sprucefir forests. Although nothing is likely to prevent crown 

fire, uneven-aged management can prevent excessive 

mortality and reduce both surface and crown fuels from 

levels that currently exist in many stands.

Summary

Pre-settlement coniferous forests of southern Utah were 

influenced by various disturbance regimes, including 

fire, insects, diseases, and herbivory, which maintained 

a heterogeneous landscape structure consisting of a 

diversity of successional stages and vegetative types. 

The combined effects of fire suppression, logging, and 

grazing have altered the extent, location, and structure 

of these forests. Identifying pre-settlement disturbance 

regimes and forest structures can serve as a reference 

to return the landscape to a more ecological sustainable 

system.

It is imperative that managers understand that each 

forest type in southern Utah has coevolved under different disturbance regimes that shaped their structure 

and community dynamics. Subsequently, hazardous fuel 

reduction and restoration treatments should differ for 

each coniferous forest type. Treatment area prioritization 

should focus on the ponderosa pine and mixed conifer 

forests types where fire regimes and vegetation attributes 

have been significantly altered from their historical range 

of variability. These areas will require moderate to high 

levels of mechanical restoration treatments before fire 

can be reintroduced to restore the historical fire regime. 

In contrast, fire regimes in spruce-fir forests have not 

been drastically altered from their historical range of 

variability and the risk of losing key ecological components is low to moderate. Therefore, hazardous fuel 

reduction and restoration treatments in the spruce-fir 

forests should be of low priority.

In ponderosa pine forests, treatments should focus 

on converting to uneven-aged management, reducing 

or removing shade tolerant conifers and oak, and reintroducing frequent prescribed surface fires. Mixed 

conifer forest treatments should focus on reducing the 

amount of shade tolerant species and leaving more 

fire-resistant tree species such as ponderosa pine and 

Douglas-fir. Mixed conifer forests should maintain 

a diversity of vegetation successional stages and age 

classes juxtaposed to each other. Reintroduction of 

prescribed fire should not be limited to under-burning 

since historical fire regimes included some crown fire in 

mixed conifer ecosystems. Spruce-fir forest treatments 

should focus on maintaining a landscape of different age 

structures, successional stages, and fuel breaks to lessen 
32 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 2—Ponderosa Pine, Mixed Conifer, and Spruce-fir Forests

the risk of entire landscapes burning in one event. Fire 

suppression in spruce-fir forests should be tempered to 

allow some forests to burn.

To establish and continue a realistic restoration strategy, 

we must maintain the fire disturbance cycle, whether 

with prescribed burning or wildfire. The end result of 

any treatment is to create a heterogeneous landscape 

structure consisting of a diversity of successional stages 

and vegetative types that if subjected to wildfire, would 

retain key ecosystem components.

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USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 39

Chapter 3

Aspen

Dale L. Bartos

Introduction

Quaking aspen (Populus tremuloides Michx.) is 

the most widely distributed broadleaf tree in North 

American (Little 1971; Sargent 1890). Aspen forests 

occur from Labrador on the east coast to Alaska in the 

north to Mexico in the south. In its eastern range, aspen 

is relatively continuously distributed. In the western 

United States, however, it occurs on the more suitable 

sites on mountains and high plateaus (Jones 1985). On 

xeric sites, it is found primarily in riparian zones. Most 

western aspen occur on public lands and exist as pure 

clones, successional to conifer, or as small-scattered 

decadent groves. At least 75 percent of aspen in the 

western United States occurs in two states, Utah and 

Colorado. Almost 25 percent of Utah’s forests are aspen 

(Mueggler 1988).

Aspen is portrayed as an excellent indicator of 

ecological integrity as well as landscape health (Kay 

1991a,b; Woodley and Theberge 1992; Woodley 1993; 

Woodley and others 1993). Aspen ecosystems in the 

western United States yield numerous products and 

benefits, including forage for livestock, habitat for 

wildlife, water for downstream users, esthetics, sites for 

recreational opportunities, wood fiber, and landscape 

diversity (Hessl 2002). Loss and potential loss of aspen 

from these western landscapes is attributed primarily to 

the reduction of wildfire, long-term overuse by ungulates 

(both wildlife and domestic livestock), and successional 

processes. During 2006, considerable attention was 

given to aspen die-off. This die-off is described as the 

rapid death of entire clones, including the root system, 

with no potential for aspen survival (http://www.fs.fed.

us/rmrs/research/highlights/aspen-restoration.php).

Aspen is an early successional, shade intolerant species with a short life span—usually less than 200 years 

in the western United States (Perala 1991). Aspen exist 

primarily as one of three types: 1) stable, 2) successional 

to conifers or seral, and 3) decadent and falling apart 

(Bartos 2001). Unlike other western forest trees, aspen 

are unique because they reproduce almost exclusively 

by suckering from the parent root system. Generally, a 

disturbance or die-back is necessary to stimulate regeneration of the aspen stands. It is speculated that these 

self-regenerating stands have existed for thousands of 

years. Unlike other tree species, if aspen stands are lost 

from the landscape they usually will not return through 

the natural process of seed dispersal.

Most aspen stands will eventually be replaced by 

conifers, sagebrush (Artemisia spp.), or possibly other 

shrub communities if the current conditions that have 

prevailed for the past 100 to 150 years (for example, 

lack of fire, excessive wildlife, and livestock browsing) continue. Current unpublished Forest Inventory 

and Analysis (FIA) data for the state of Utah show that 

there has been approximately a 60 percent decrease in 

aspen-dominated public lands since European settlement 

(Bartos and Campbell 1998a). This decrease is generally uniform across the entire state. When aspen lands 

convert to conifer dominated landscapes, substantially 

less water is available for streamflow, undergrowth 

biomass production is greatly reduced, and there is a 

marked decline in the diversity of plants and animals.

Development of useful management recommendations concerning the aspen, aspen/conifer, and aspen/

sagebrush ecosystems requires a good understanding 

of the structure and function of these systems. Aspen 

literature for the western United States has been summarized through the early 1980s (DeByle and Winokur 

1985). The published proceedings of a more recent 

aspen symposium held in June 2000 was a supplement 

to this earlier work (Shepperd and others 2001). Shepperd and others (2006) recently synthesized the existing 

information on the ecology and management of aspen 

in the Sierra Nevada of California.

The purpose of this paper is to discuss the historical and current conditions of aspen, with an emphasis 

on aspen in Utah. This paper also identifies clones at 
40 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 3—Aspen

risk, provides a method of treatment prioritization, and 

offers management recommendations of aspen stands. 

Landscapes that need priority management are mixedconifer/aspen, particularly where subalpine fir (Abies 

lasiocarpa (Hook.) Nutt.) dominates, and aspen/sagebrush transition zones. Recommendations to regenerate 

aspen include the need for immediate, large-scale (200 

to 400+ ha [500 to 1000+ acres]) treatments that mimic 

natural mosaic patterns.

An aspen stand typically contains numerous genetically 

identical stems (ramets) that propagated vegetatively 

from seedlings at some time in the past when conditions were favorable. Aspen stems originate from root 

suckers, some of which still may be interconnected 

via a shared root system. Shepperd and Smith (1993) 

reported that aspen stems establish independent root 

systems by approximately 25 years of age, with few 

stems still connected to the original root system. Barnes 

(1975) speculated that these self-regenerating clones 

have existed for thousands of years, being perpetuated 

over time by disturbance – primarily fire. Some clones 

occupy large areas. One such clone, referred to as Pando 

(Latin for I spread), occurs on the Fishlake National 

Forest in central Utah (Mitton and Grant 1996; Grant 

1993). The authors claim that this 43 ha (106 acres) clone 

is the world’s largest living organism. Kemperman and 

Barnes (1976) originally identified the Pando clone and 

speculated that clones as large as 81 ha (200 acres) may 

exist.

Successful sexual reproduction of western aspen is 

extremely rare (Mitton and Grant 1996). Jelinski and 

Cheliak (1992) describe a “window of opportunity” that 

may allow seedling establishment at intervals of 200 

to 400 years. Abnormally cool, moist years may allow 

sexual reproduction of aspen seedlings (Elliott and Baker 

2004). There are some documented cases of successful aspen seedling establishment in the west. Romme 

and others (2005) documented the establishment and 

persistence of sexually produced aspen seedlings after 

the 1988 fires in Yellowstone National Park. Elliott and 

Baker (2004) hypothesized that expanding aspen clones 

at tree line in the San Juan Mountains of Colorado were 

established by seed. However, unlike other western 

tree species, aspen seedling establishment is rare and 

undependable (Barnes 1966). Therefore, once aspen is 

lost from the landscape, it will seldom establish from 

seed in the Intermountain West. Vegetative reproduction by suckers generally requires a disturbance (fig. 1) 

or die-back that alters the hormonal balance of a clone 

(Bancroft 1989; Schier and others 1985). Frey and others 

(2003) recently reviewed the process of aspen sucker 

regeneration. Bancroft (1989) and Bartos (2001) discuss 

what happens when there is a disruption in the flow of 

two hormones (cytokinin and auxin) within an aspen 

tree. Basically, when the tree is killed or stressed, the 

flow of sucker-suppressing auxin from the crown down 

to the root system is disrupted, which allows a second 

hormone, cytokinin, to stimulate suckering.

Figure 1—One-year old aspen 

(Populus tremuloides) sprouts after 

the Sanford Fire, UT.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 41

Chapter 3—Aspen

To abide by the words of Aldo Leopold: “...if we are 

serious about restoring or maintaining ecosystem health 

and ecological integrity, then we must first know what 

the land was like to begin with” (quote in Covington 

and Moore 1994). For millennia, wildfire dominated 

the natural functioning of the aspen ecosystem (Baker 

1925; Parfit 1996; Pyne 1995). Baker (1918) and Schier 

(1976) describe western aspen as a fire dependent species. Jones (1985) stated “...fire is responsible for the 

abundance of aspen in the west and for the even-aged 

structure of so many stands.” Conversely, reduction of fire 

in the Interior West is probably one of the main causes 

of the decline of aspen from so much of our landscapes 

because it permits the encroachment of shade tolerant 

conifers that reproduce from seed (Bartos 2001; Schier 

1975).

Historical Conditions

Fire has been a major component of the vegetation 

landscape prior to humans arriving on the scene, and 

later became a critical tool for man. Pyne (1995) suggests 

that fire and humans co-evolved. The native inhabitants 

of the North American continent have long been considered natural conservationists living in harmony with 

the environment. However, it is now becoming clear that 

the American Indians had a profound influence on the 

landscape, particularly with respect to the use of fire. 

Ongoing fire history studies in southern Utah indicate 

that during the approximately 400 years prior to settlement, fire-free intervals varied from 20 to 60 years. 

These fire free intervals were longer with increases in 

elevation (Chappell 1997). When European man arrived 

in the western United States around 1850, they found a 

mosaic of vegetation profoundly influenced by Native 

Americans over a period of 12,000+ years (Kay 1997). 

These “pristine” landscapes were in a continual state 

of flux as a result of both natural and human caused 

burning.

Early settlers further impacted the land through removal of fine fuels by grazing domestic livestock, thus 

limiting the spread of fire. In the early part of the 20th 

century, land managers instigated a vigorous campaign 

of fire control. These two practices (grazing and fire 

control) have had a profound effect on the vegetation we 

now see on western landscapes. As a result of this lack 

of fire, aspen stands on the landscape have been greatly 

reduced. Because western aspen reproduces primarily by 

suckering, the elimination of fire combined with excessive browsing of the asexual reproduction by ungulates 

has caused many aspen dominated sites to convert to 

other vegetation types. Barnes (1975) speculates that 

many of the existing clones have been present since 

the last ice age, while Knight (1994) states some clones 

may have persisted since Pleistocene times. Therefore, 

clones that have persisted for thousands, if not hundreds 

of thousands, of years are in danger of being eliminated 

in as little as a couple of centuries.

Brown and Simmerman (1986) detailed the difficulty 

of burning pure aspen, while DeByle and others (1987) 

describe the aspen type as “asbestos” in nature and 

discuss its use as natural firebreaks. Earlier, Fechner 

and Barrows (1976) discussed the use of pure aspen 

as firebreaks. They reported that fire ignition rates in 

quaking aspen stands are less than half those for all other 

cover types in Colorado. Crown fires often drop to the 

ground when they reach pure aspen stands and therefore, 

fire spreads only a short distance into the stands. As 

early as 1925, Baker (1925) cited grazing as the primary 

reason many aspen stands no longer burn. He stated 

that intense grazing pressure reduces the undergrowth 

in aspen clones such that fire spread is inhibited. Fires 

before grazing became heavy in the mountains of Utah 

were as frequent as 7 to 10 years, but stopped about the 

time grazing began (Baker 1925). DeByle and others 

(1987) gave four possible reasons why fires are less frequent in aspen today than historically: 1) effective direct 

control of wildfires, 2) grazing, 3) removal of American 

Indians from historic range, and 4) succession. This 

phenomenon of aspen no longer burning needs to be 

further evaluated, especially now that there is so much 

concern with respect to the wildland/urban interface.

Current Conditions

Aspen is a disturbance-dependent species that flourished when these western lands burned periodically. 

With suppression of this natural force, many of these 

lands have converted, or are in the process of converting, 

to other vegetation types, such as conifer (for example, 

subalpine fir) or sagebrush. Three of the most critical 

products being lost from the aspen system as a result of 

this conversion are water, undergrowth vegetation, and 

biodiversity.

DeByle and Winokur (1985) reported that there were 

more than 2.8 million ha (7 million acres) of aspen 

in the Interior West and, Mueggler and Campbell 

(1986) reported over 648,000 ha (1.6 million acres) of 

aspen-dominated forests in Utah. The Rocky Mountain Research Station’s FIA Project has compiled data 
42 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 3—Aspen

representing current and historical acreage of aspen in 

the Interior West. The historical data result from summing all acres that currently contain at least one aspen, 

either living or dead. This assumes that this acreage is, 

or once was, occupied by aspen. For the state of Utah, 

FIA data show that there has been almost a 60 percent 

decrease in aspen dominated lands since the arrival of 

European man (table 1). This loss is fairly uniform across 

the entire state as seen from the data for individual forests. Similar trends are seen elsewhere throughout the 

western United States. Analysis of FIA plots in Idaho, 

Colorado, and Wyoming show regional aspen cover loss 

(Rogers 2002). Rogers and others (1998) report that 31 

percent of the FIA plots in Colorado aspen forest types 

are “unstable” and are possibly in transition to other 

types. However, Manier and Laven (2002) and Kaye 

and others (2003) found no evidence of aspen decline 

in some areas of Colorado. Lachowski and others (1996) 

and Wirth and others (1996) used remote sensing and 

geographic information systems (GIS) to evaluate the 

loss of aspen in the Gravelly Mountains in southwestern 

Montana. They found that aspen decreased by approximately 47 percent over a 45-year period (1947 to 1992) 

and attributed this decrease to succession to conifer 

species. Brown (1995), in a review article, states that 

the deterioration of aspen in Oregon and Washington 

appears similar to the decline of aspen in other parts 

of the west. In the South Warner mountains in northern 

California, Di Orio and others (2005) found a 24 percent 

decline in aspen cover between 1946 and 1994 and a 

decrease in the mean size of aspen stands.

Ungulates

Ungulates have a major impact on aspen regeneration 

(Bartos and Mueggler 1981; Bartos and others 1994; 

Kay 1985, 1990, 1995; Kay and others 1994; McCain 

and others 2003; U.S. Department of Agriculture 1994). 

Excessive browsing on aspen suckers by wildlife (elk 

[Cervis canadensis], deer [Odocoileus hemionus], or 

moose [Alces alces]) or by livestock (cattle or sheep) can 

suppress or eliminate regeneration in aspen stands. Excessive livestock use of young aspen suckers can threaten 

stand maintenance in climax aspen communities, especially if the use is in August or September (McCartney 

1993). Early in the past century, Sampson (1919) noted 

that cattle were not a detriment to the establishment of 

aspen suckers “unless the range is stocked with cattle 

beyond its natural carrying capacity.” He further noted 

that sheep would eliminate all aspen reproduction where 

sites were grazed 3 consecutive years. Smith and others 

(1972) concluded that proper livestock management is 

essential to regeneration of aspen following removal 

of older stands. A recent model-based assessment of 

aspen and elk herbivory in Rocky Mountain National 

Park suggests that a reduction of elk populations up 

to 80 percent is necessary to allow regeneration of 90 

percent of the park’s declining aspen stands (Weisberg 

and Coughenour 2003).

The Pando clone (Kemperman and Barnes 1976) mentioned earlier is a good example of the problems associated with the decline of aspen. Managers observed that 

this single-story clone was not regenerating successfully. 

A cutting treatment was imposed on a portion of the 

Table 1—Current and historical acres of aspen found in Utah (Unpublished data provided 

by the USDA Forest Service, Rocky Mountain Research Station’s Forest Inventory and Analysis Project.).

Current aspen Historical aspen Decline

National Forest hectares (acres) hectares (acres) (percent)

Ashley 41,036 (101,358) 130580 (322,532) 69

Uinta 70,645 (174,492) 115527 (285,351) 39

Wasatch-Cache 52,071 (128,615) 151351 (373,837) 66

Dixie 61,965 (153,053) 177213 (437,715) 65

Fishlake 57,469 (141,948) 127014 (313,724) 55

Manti-LaSal 64,318 (158,866) 136845 (338,008) 53

Southern Utah 183,752 (453,867) 441072 (1,089,447) 58

Northern Utah 163,751 (404,465) 397457 (981,720) 59

Total 347,503 (858,332) 838529 (2,071,167) 59
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 43

Chapter 3—Aspen

clone in the early 1990s. The initial treated area produced 

suckers; however, none of them survived. Subsequently, 

an additional portion was cut and fenced with a 2-m 

(7-ft) high woven wire fence to exclude large animals. 

This resulted in successful regeneration in the cut and 

fenced area that should be stand-replacing. Apparently, 

deer and cattle browsing effectively suppressed sucker 

reproduction in the unprotected portion of the stand. 

The Coconino National Forest in Arizona installed two 

large elk exclosures after a wildfire to facilitate aspen 

regeneration. Elk browsed 85 percent of aspen shoots 

in high severity burned areas outside the exclosures and 

34 percent of the shoots on the intermediate burned 

areas (Bailey and Whitman 2002). This suggests elk 

pressure is greatest in areas where recent high severity 

burns produce large numbers of aspen shoots.

This potentially severe ungulate influence on successful aspen regeneration adds a critical component 

to management considerations in certain parts of the 

west (Shepperd and Fairweather 1994). Burned clones 

subjected to repeated browsing only hastens their demise 

(Bartos and others 1994; Kay 1990). Therefore, treatments to induce suckering must not be initiated before 

relief from excessive browsing is obtained (U.S. Department of Agriculture 1994). Kay (1997) speculates that 

the historic abundance of aspen in the Intermountain 

area suggests that these ecosystems developed with 

relatively low levels of ungulate pressure. Larsen and 

Ripple (2003, 2005) found that aspen recruitment into 

the overstory was significantly higher on the National 

Forests bordering the northern portion of Yellowstone 

National Park than inside the park where elk pressure is 

greater. They also found that overstory aspen recruitment 

inside the park was only occurring in areas protected 

from ungulate browsing, such as fenced exclosures and 

scree slopes. This further supports the need to control 

or modify browsing intensity, where necessary, if these 

western aspen communities are to survive by vegetative 

reproduction.

Insects and Disease

Quaking aspen plays host to a wide array of both 

insects and diseases. Hinds (1985) states that many diseases attack aspen, however, very few kill or seriously 

injure living trees. For example, leaf spot (Marssonina 

populi) is a common leaf disease of aspen in the west. 

If this disease prevails for 2 or more years, it can cause 

twig and branch mortality, but rarely kills mature trees 

(Harniss and Nelson 1984). Many of the prevalent 

diseases in aspen affect both regeneration and mature 

trees (Hinds 1985). In both cases, trees damaged and 

weakened by animals are easy targets for introduction 

of disease. Disease plays a part in thinning excessive 

regeneration of aspen in areas where stand replacing 

episodes have occurred. Hart and Hart (2001) working in western Wyoming found that larger aspen stems 

(>15 cm [5.9 inches]) had less mortality due to disease/

animal wounds than smaller stems. This observation 

substantiates the fact that disease tends to reduce aspen 

regeneration and smaller trees. Aspen die-off, mentioned 

earlier, is occurring from northern Arizona to Canada 

and currently, is most prevalent in southern Utah and 

western Colorado. Some speculate that this die-off 

phenomenon is a result of prolonged drought conditions 

coupled with either opportunistic diseases and/or insects. 

More detailed studies are needed to define this problem 

in order to make management recommendations that 

address the issue (Aspen Summit 2006).

More than 30 insect species are listed by Furniss 

and Carolin (1977) as utilizing aspen. Some of these 

species cause minimal damage, some just attack and 

impact weakened or dying trees, and some may severely 

damage or kill otherwise healthy trees. Insect impacts 

coupled with other factors, such as ungulate use, can 

have major impacts on regeneration. Jones (1985) gives 

a thorough discussion of insects and their impact on 

western aspen.

Fire Exclusion

The 7- to 10-year fire return interval suggested by 

Baker (1925) indicates these lands were burning often 

prior to European settlement. Because of the disruption 

in the fire regime and natural succession, we are finding aspen that are considerably older than the normal 

pathological rotation age of 70 to 80 years (Personal 

communication, John Guyon, Pathologist, Forest Health 

Protection, Ogden, UT). This aging of aspen on the landscape gives rise to the potential for more catastrophic 

disease and insect events. Additionally, this excessive 

mortality will add to the fuel loads.

Ecosystem Impacts 

Caused by Aspen Loss

Loss of aspen from western landscapes translates to a 

loss of water, forage, and biodiversity, as well as other 

benefits (Bartos and Campbell 1998b). Harper and others (1981) suggest that water yields can be expected to 

decrease 5 percent as conifers replace aspen. Gifford 

and others (1984) concluded from modeling studies that 

a net water loss of 18.6 cm (7.3 inches) occurs when 

spruce (Picea spp.) replaces aspen, and a loss of 7.2 cm 
44 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 3—Aspen

(2.8 inches) of water when subalpine fir forests replace 

aspen. Therefore, aspen-to-conifer succession appears 

to have the potential to markedly reduce water yields in 

the western United States (Bartos and Campbell 1998b). 

This loss of water means that it is not available to produce undergrowth vegetation, recharge soil profiles, or 

increase streamflow. In dry climates, such as the Great 

Basin, this loss of water is substantial and should be of 

great concern to the public.

Aspen forests traditionally have been considered prime 

grazing lands in the Intermountain West. Forty to 70 percent of the undergrowth production found in association 

with major aspen community types in the Intermountain 

Region consists of palatable forage (fig. 2) (Mueggler 

1988). Mueggler found that the most productive aspen 

communities produced as much as 3,200 kg/ha (2,900 

lbs/acre) air-dry undergrowth material and averaged 

over 1,350 kg/ha (1,200 lbs/acre). Mueggler (1985) found 

when conifers comprised as little as 15 percent of the 

total tree basal area on the Wasatch Plateau of Utah, 

undergrowth production was reduced approximately 

50 percent (fig. 3a). Mueggler (1988) reported an even 

greater reduction on the Fishlake and Dixie National 

Forests of southern Utah, with undergrowth production 

reduced two-thirds when conifers comprised as little as 

15 percent of the total tree basal area (fig. 3b). Similar 

reductions in undergrowth forage production occur in 

other forest types as conifer basal area increases (Jameson 

1967). Once conifer invasion approaches 50 percent of 

the total tree basal area in aspen stands, undergrowth 

production is likely to be only a small fraction of what 

it once was on these formerly prime grazing lands.

When we consider that more than half of our aspen 

dominated stands in the Interior west have converted 

to conifer during the past 150 years, we obviously can 

conclude we have experienced a large loss of forage 

production. If this dramatic decrease in vegetation is not 

considered when determining stocking rates for grazing 

allotments, an overgrazing problem is compounded. For 

example, if a 50 percent decline of aspen stands on an 

allotment occurred without a corresponding adjustment 

of allowable stocking, the remaining forage producing 

areas will be subjected to much heavier utilization 

than previously. This excessive pressure will adversely 

impact aspen regeneration as well as the other forage 

resources.

Figure 2—Aspen (Populus tremuloides) with productive herbaceous understory, 

including cowparsnip (Heracleum lanatum), snowberry (Symphoricarpos oreophilus), 

and numerous other forbs and grasses. In addition, subalpine fire (Abies lasiocarpa) 

and other conifers are establishing in the understory. 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 45

Chapter 3—Aspen

Figure 3—Effect of conifer invasion on annual undergrowth production: A. The Wasatch Plateau 

of central Utah. B. The Fishlake and Dixie National Forests of southern Utah. Re-analysis of data 

reported in Mueggler (1985).
46 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 3—Aspen

Aspen communities have exceedingly high 

biodiversity–second only to riparian areas on western ranges (Kay 1997). When aspen lands become 

dominated by either conifer or sagebrush, marked 

changes in both flora and fauna occur and biodiversity 

is compromised. Not only is there a loss of forage 

production, but there is also a substantial decrease 

in plant species richness associated with the loss of 

aspen.

Winternitz (1980) reported that the density and diversity of birds was greater in aspen than conifer stands, 

and McGraw/Bergstrom (1986) found that mature stands 

of aspen had the most bird species compared to younger 

stands or those being invaded by conifers. Bird species 

diversity also increases with the size of aspen stands 

(Johns 1993). Linder (1995) found that bird community 

structure is positively correlated with the size of aspen 

patches.

Other aspects of biodiversity loss are not so readily 

apparent. For example lichens, bats, and snails are also 

affected by the decline of aspen. McCune (1997) indicates 

that a loss of aspen would adversely affect the epiphytic 

lichen communities that depend on mature to old aspen. 

Vonhof (1996) found in Southern British Columbia that 

both big brown bats (Eptesicus fuscus) and silver-haired 

bats (Lasionycteris noctivagans) prefer trembling aspen 

for roost sites. In the Cypress Hills, Saskatchewan, all 

little brown bats (Myotis lucifugus) roost in aspen trees 

(Kalcounis and Hecker 1996). Crampton and Barclay 

(1996) monitored two species of bats (little brown bats 

and silver-haired bats) in aspen stands in Lac La Biche, 

Alberta, and found that the bats preferred newly dead 

aspen with heart rot and trees with low leaf cover. Beetle 

(1997) determined that 21 species of land snails occurred 

in aspen stands in Yellowstone National Park and that 

these stands had a greater abundance and variety of 

snails than coniferous forests on more acidic soils.

Fire Regime Condition Classes

Fire Regime Condition Class 1 (FRCC 1)

Fire Regime Condition Class 1 (FRCC 1) indicates 

that the fire regimes are within a historical range and 

there is little risk of losing aspen from the ecosystem. 

This class includes the stable and seral stands that 

are reproducing and have little conifer encroachment. 

Mueggler (1989) speculated that as much as a third of 

the aspen in the west could be classified as climax or 

stable. Smith and Smith (2005) reported 16 percent of 

the aspen stands on the Uncompahgre Plateau, CO, 

were stable. Therefore, assuming a 60 percent decline 

in the historical range of aspen, only a portion of aspen 

dominated land (10 to 15 percent) would be considered 

functioning normally, in other words, regenerating naturally without major interventions. Stable clones are not 

subject to succession, primarily because there is a lack 

of conifer seed source and most natural regeneration is 

not consumed by ungulates. This leaves only a few aspen 

in Utah that are able to survive and perpetuate without 

naturally occurring wildfire or other disturbance. These 

are approximate figures for Utah and may be more or 

less depending on various situations. Caution should be 

used when extrapolating to aspen outside of Utah.

FRCC 1 aspen stands contain no conifers, are multistoried with numerous age classes, and usually have a 

fairy ring or skirt of regeneration around the outside of 

the clones, which makes them hard to see through or into 

(fig. 4) (Bartos 2001). These stands may also have present 

a large shrub component of species such as chokecherry 

(Prunus virginiana L.) or snowberry (Symphoricarpos 

albus (L.) Blake). Aspen stands in FRCC 1 regenerate 

without many external forces, such as disease or abiotic 

factors, acting upon them. The fire return interval in 

these stands is characterized by either frequent, very 

low intensity fires that cause some overstory mortality 

and subsequent regeneration, or more infrequent higher 

intensity fires every 100 to 200 years that result in new 

even-aged stands (Brown and Simmerman 1986). These 

high intensity fires in pure stands of aspen result when 

there are a lot of fine fuels, and/or a considerable shrub 

component coupled with extremely dry conditions. The 

low intensity fire return interval could be as frequent 

as 7 to 10 years (Baker 1925; DeByle and others 1987). 

In such cases, FRCC 1 seral stands should be able to 

regenerate enough sprouts to successfully restock the 

stand after fire. Baker (1925) states that 6,000 aspen 

sprouts per hectare (2,500 sprouts/acre) by the end of 

year 3 are successful post-fire stocking levels. However, 

ungulate pressure must also be low enough to prevent 

excessive browsing of the new aspen shoots.

Fire Regime Condition Class 2 (FRCC 2)

Fire Regime Condition Class 2 (FRCC 2) indicates 

that the fire regimes have been altered and there is a 

moderate risk of losing key ecosystem components. In 

Utah, approximately 25 to 30 percent of the aspen are 

in FRCC 2. Seral FRCC 2 stands have a large conifer 

component that over time will convert the stand to a 

mixed conifer type if no disturbance occurs (fig. 5). 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 47

Chapter 3—Aspen

Figure 4—Healthy aspen (Populus tremuloides) stand with a younger age class at the 

outer edge. 

Figure 5—Aspen (Populus tremuloides) stand with a co-dominant overstory of subalpine fir 

(Abies lasiocarpa) and Engelmann spruce (Picea engelmannii). Dominant grass and shrub 

species are slender wheatgrass (Elymus trachycaulus) and gooseberry currant (Ribes 

montigenum).
48 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 3—Aspen

Stable FRCC 2 stands have limited sprouting and some 

of the mature stems are dead and dying. Sprouting may 

be limited because of low light levels in the case of 

conifer encroachment, or if ungulate browsing pressure 

is heavy, the clone may be sprouting but the sprouts are 

summarily eaten.

Fire Regime Condition Class 3 (FRCC 3)

Fire Regime Condition Class 3 (FRCC 3) contains all 

aspen that are not included in the previous two classes. 

Fire return intervals have been drastically altered in 

FRCC 3 stands and there is a high risk of losing components of the ecosystem. Approximately 60 percent 

of the aspen component in Utah is classified as FRCC 

3. Because of the lack of burning or other disturbance, 

these once aspen dominated lands are being replaced 

primarily by conifers. In FRCC 3 stands, conifer canopy 

cover exceeds aspen canopy cover and they are now 

considered mixed-conifer rather than aspen cover types. 

Without treatment or disturbance, these decadent aspen 

stands will soon be completely replaced by conifers and 

lost from the landscape (fig. 6). As these aspen stands 

age and are replaced by conifers, the aspen begin to die 

(around 120 years) and fall to the ground and increase 

fuel loadings. Higher fuel loadings can cause fires to 

burn more severely, creating significant soil heating. 

Most aspen suckers regenerate within 6 cm (2.4 inches) 

of the surface and virtually all regenerate within the 

upper 12 cm (4.7 inches) of soil (Perala 1991). Therefore, 

high severity fires could reduce suckering potential by 

killing the roots, and subsequent stocking levels may 

not be high enough to successfully regenerate the stand. 

Aging aspen stands also may not be able to regenerate 

successfully due to a reduced capacity of the root system 

to sprout after disturbance. Browsing by both wild and 

domestic ungulates further complicates the situation. In 

areas where ungulate pressure is intense, disturbance 

events are not successful because all suckers are browsed 

and some aspen clones are ultimately lost. Excessive 

utilization needs to be addressed before fire or other 

treatments are returned to the system (U.S. Department 

of Agriculture 1994). If one goal of management is to 

restore aspen, then animal use needs to be monitored, 

evaluated, and adjusted. Otherwise, animals utilizing 

the aspen regeneration can slow or defeat restoration 

efforts.

Recommended Treatments

Restoring aspen to a level near its natural range of 

variability requires that land managers take an aggressive management approach in the very near future. Land 

managers in the western United States are currently very 

interested in actively treating their deteriorating aspen 

stands. Five major risk factors have been identified 

Figure 6—Former aspen 

(Populus tremuloides) stand now 

dominated by ponderosa pine 

(Pinus ponderosa), Douglas-fir 

(Pseudotsuga menziesii), and 

white fir (Abies concolor).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 49

Chapter 3—Aspen

to assist managers in evaluating their aspen resource 

(Bartos and Campbell 1998a). These risk factors are: (1) 

conifer cover >25 percent, including reproduction in the 

understory; (2) sagebrush cover >10 percent; (3) aspen 

canopy cover <40 percent; (4) dominant aspen trees >100 

years of age; (5) aspen regeneration 1.5 to 4.5 m (5 to 

15 ft) tall and <1,200 stems per ha (<500 stems/acre).

A prioritized key (table 2) was developed by Campbell 

and Bartos (2001) that incorporates the above risk factors 

and is applicable to most aspen stands found in southern 

Utah. In part, this key was developed with knowledge 

and expertise of aspen systems that occur in this region. 

The stands at greatest risk and highest priority are those 

having conifer canopy cover that exceeds aspen canopy 

cover. These are currently mixed-conifer rather than 

aspen cover types. However, with proper treatments, the 

aspen cover type can usually be restored and sustained. 

Mueggler (1989) developed a general management decision model for maintaining aspen stands in the west 

(fig. 7). From work in Ontario, Peterson and Peterson 

(1992) suggested that stands targeted for treatment need 

at least 40 parent aspen stems per hectare (16 stems/acre) 

to produce the minimal acceptable stocking and about 

125 parent aspen stems per hectare (50 stems/acre) to 

fully stock a stand. However, Shepperd (2004) reported 

successful regeneration of an aspen clone in Arizona 

with only two remaining parent stems when fencing 

was used to protect suckers.

Numerous techniques are available for restoring these 

decadent aspen clones or late successional clones dominated by conifers. These include prescribed burning, 

clear cutting, cutting and burning, fencing, spraying, 

ripping, and chaining (Bartos and Harniss 1990; Bartos 

and Lester 1984). Such treatments should give rise to 

abundant aspen suckers if the three essential elements 

to the aspen regeneration triangle are met: 1) hormonal 

stimulation, 2) proper environment, and 3) protection of 

new suckers (Shepperd 2001, 2004). Shepperd (2001) lists 

seven aspen regeneration alternatives: 1) doing nothing, 

2) commercial harvest, 3) prescribed fire, 4) mechanical root stimulation, 5) removal of vegetative competition, 6) protection of regeneration from herbivory, and 

7) regenerating from seed.

Table 2—Key to the risk factors used to prioritize areas with aspen for restoration and conservation actions in the Intermountain West. Assumption: Aspen are present with a density of at least 50 

mature trees per hectare (20 mature trees per acre). Note: Couplet 1 refers to relative cover; 

couplets 2 to 5 use absolute cover (Campbell and Bartos 2001).

Risk Factor Priority

1. a. Conifer species comprise at least half of the canopy cover. Highest priority

b. Aspen comprises more than half of the total canopy cover. 2

2. a. Aspen canopy cover is less than 40 percent; and sagebrush High priority 

 usually a dominant understory species exceeds 15 percent cover.

b. Not as above. 3

3. a. Conifer cover (including overstory and understory) exceeds Moderate to high priority

 25 percent. 

b. Conifer cover is less than 25 percent. 4

4. a. Aspen regeneration (1.5 to 4.5 m [5 to 15 feet tall]) is less Moderate priority

 than 1,200 stems/ha (500 stems/acre).

b. Aspen regeneration exceeds 1,200 stems/ha (500 stems/acre). 5

5. a. Any two of the following three risk factors are represented: Low to moderate priority

 1 – Aspen canopy cover is less than 40 percent.

 2 – Dominant aspen trees are greater than 100 years old.

 3 – Sagebrush cover exceeds 10 percent.

b. Two of the three risk factors in 5a are not represented. 6

6. a. One of the three risk factors in 5a is represented. Low priority

b. None of the risk factors above are represented. Candidate for properly 

functioning condition
50 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 3—Aspen

Figure 7—General management decision model for maintaining aspen (Populus 

tremuloides) stands in the western United States. Adapted from Mueggler (1989).

Doing Nothing

This action is suitable for stands in FRCC 1. Management intervention is likely unnecessary if the stand 

is “showing little sign of decline, disease, or distress 

from competition, contains multiple age classes, or is 

successfully suckering” (Shepperd 2001).

Commercial Harvest

Where aspen markets exist, clearfelling is the preferred 

harvest method to regenerate a declining stand. Partial 

harvests result in fewer suckers and expose the residual 

mature stems to breakage, windthrow, and sunscald 

(Shepperd 2001). Dormant season logging promotes more 

suckering than spring or summer harvests (Perala 1991; 

Frey and others 2003). Ohms (2003) found clearfelling 

aspen clones produced significantly more suckers and 

increased sucker height compared to the control clones. 

Treatment response was even more successful in the areas 

of the clearcuts where browsing was excluded. The successful regeneration of the aspen stand after clearfelling 

is dependent upon the size of the area treated, unless 

browsing is controlled (Mueggler and Bartos 1977). If 

ungulate use is not controlled, treated areas should be 

greater than 5 ha (12 acres).

Fire

In climax aspen stands, suckering following the 

death of decadent stems is usually sufficient to maintain the stands. However, in many areas where aspen 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 51

Chapter 3—Aspen

is successional to conifers, fire can promote sufficient 

suckering to maintain aspen stands and remove encroaching conifers. Working in central Utah, Baker (1925) 

stated that a 50-year fire return interval would keep 

most aspen stands free of conifers. A comprehensive 

fire guide for treating aspen in the western United States 

was developed by Brown and Simmerman (1986). In this 

guide, they evaluated fuels and flammability for five 

specific types: aspen/shrub, aspen/tall forb, aspen/low 

forb, mixed/shrub, and mixed/forb. Various factors were 

evaluated and a probability for a successful burn was 

given. This information should be useful to the resource 

manager for developing prescribed fire burn plans.

Bradley and others (1992) prepared a detailed report 

that discusses fire ecology of forests and woodlands in 

Utah. This paper identifies various fire groups, some 

of which include an aspen component. Habitat types 

developed by Mauk and Henderson (1984), Youngblood 

and Mauk (1985), and some of the community types 

discussed by Padgett and others (1989) and Mueggler 

(1989) are assigned to 13 “Fire Groups” (Bradley and 

others 1992). Five of these fire groups contain aspen 

and would be relevant in planning any prescribed burn. 

These five specific fire groups are: (a) 7-community 

types where aspen appears to be climax or a long-term 

seral dominant; (b) 3-Ponderosa Pine Habitat Type 

(Pinus ponderosa P. & C. Lawson); (c) 5-Douglas-Fir 

Habitat Type (Pseudotsuga menziesii (Mirbel) Franco); 

(d) 6-White Fir and Blue Spruce Habitat Types (Abies 

concolor (Gord. & Glend.) Lindl. Ex Hildebr. and Picea 

pungens Engelm.); and (e) 10-Subalpine Fir Habitat 

Types. Items b through e contain aspen that are successional to various coniferous species indicated in the 

habitat type.

Fire that kills most of the aspen overstory produces 

more shoots than less intense fires that result in lower 

overstory mortality. Keyser and others (2005) reported 

high severity fires (>75 percent aspen mortality) in aspen stands produced more shoots, killed encroaching 

conifers, and had greater shoot height growth than low 

severity fires (25 to 75 percent aspen mortality) or unburned aspen stands. Aspen sprout aboveground biomass 

was 10 times greater on high severity sites (100 percent 

aspen mortality) than intermediate severity burned sites 

(~50 percent aspen mortality) after a wildfire in Arizona 

(Bailey and Whitman 2002).

Mechanical Root Stimulation

Severing aspen roots from parent stems can stimulate 

suckering. Shepperd (1996) found that tipping over mature aspen stems resulted in significantly more sprouts 

than stems that had been chainsaw felled. Ripping 

around the perimeter of a decadent aspen clone can 

also stimulate suckering. Variations of ripping around 

aspen clones have proved successful on the Coconino 

National Forest in Arizona (Shepperd 2004). This method 

can expand the size of the clone and introduce new age 

classes without cutting existing mature stems.

Removal of Vegetative Competition

Removing competing vegetation increases light to the 

understory (Shepperd 2001). Vegetation removal can 

be done alone or in conjunction with other treatments 

to increase aspen regeneration. In northern California, 

conifers up to 66 cm (26 inches) diameter at breast height 

(DBH) were removed around existing aspen stems to 

encourage suckering (Jones and others 2005). Aspen stem 

density was significantly higher in the treated stands 

4 years post-treatment when compared to controls. In 

a separate study to re-invigorate the clone, all conifers 

were removed and an elk-proof fence was installed 

around a declining aspen clone consisting of only two 

remaining live aspen stems (Shepperd 2004). Four years 

post-treatment, the stand had produced numerous sprouts 

and the treatment was considered successful.

Protection of Regeneration from Herbivory

In some areas, confounding factors exist that will 

challenge the manager to be successful. Treatment to 

induce aspen suckering is not enough in areas where 

there is extensive ungulate pressure (both domestic and 

wildlife) because uncontrolled ungulate use could cause 

some aspen treatments to be unsuccessful. Such actions 

to induce suckering must not be initiated before relief 

from excessive browsing is obtained (U.S. Department 

of Agriculture 1994; Brown 1995). Treatments to stimulate aspen regeneration that are less than 2 ha (5 acres) 

may concentrate deer use and cause excessive browsing 

(Mueggler and Bartos 1977). In Colorado, harvesting 

several adjacent large (6 to 8 ha [15 to 20 acres]) units 

at one time resulted in successful aspen regeneration, 

even with large ungulate populations (Crouch 1983). 

Additional protection, such as temporary fencing, may 

be required to permit the growth of the aspen suckers 

beyond the reach of browsing animals if ungulate populations are high. Excessive browsing of sucker reproduction is perhaps the most detrimental influence on the 

successful regeneration of stands burned or otherwise 

treated to simulate suckering, as well as in stable aspen 

communities which usually produce ample suckers for 

regeneration as the aspen overstory gradually thins with 
52 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 3—Aspen

age. Obtaining ample suckers to replace aspen stands is 

usually not a problem. Allowing these suckers to grow 

beyond the reach of ungulates is the issue (Shepperd 

and Fairweather 1994; Rolf 2001).

Various barriers to protect aspen regeneration under 

these conditions were evaluated by Kota (2005) in the 

Black Hills of South Dakota. He found that a barrier of 

hinged trees was useful in protecting aspen regeneration. Hinging is accomplished by felling the tree (live or 

dead) and keeping the bole on the stump, thus creating a 

barrier around patches of aspen regeneration. Standard 

fencing practices can also be used but are sometimes 

cost prohibitive, especially if large areas need to be protected. Kay and Bartos (2000) surveyed all known aspen 

exclosures on the Dixie and Fishlake National Forests in 

Utah in 1995 and 1996. The exclosures were originally 

established between 1930 and 1970 and were designed 

to exclude either all ungulates (livestock and wildlife) 

or just livestock. Aspen in the total exclusion exclosures 

successfully regenerated without other treatments and 

developed multi-sized, multi-aged stems. Aspen in the 

livestock-exclusion exclosures either failed to produce 

new stems greater than 2 m (6.5 ft) tall, or regenerated 

at lower densities than the total-exclusion plots. Outside the exclosures, aspen either failed to regenerate 

successfully or regenerated at lower densities than the 

livestock-exclusion plots. Aspen regenerated successfully 

on the livestock excluded plots and outside the exclosures 

only when mule deer populations were low.

Regenerating from Seed

The feasibility of seeding aspen in areas of abundant 

soil moisture is currently being tested (Shepperd 2001). 

If successful, this technique could be used to restore 

aspen where it has been lost from the system.

References

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Station, Salt Lake City, Utah.

Bailey, J. K.; Whitman, T. G. 2002. Interactions among fire, aspen, 

and elk affect insect diversity: reversal of a community response. 

Ecology. 83: 1701-1712.

Baker, F. S. 1918. Aspen reproduction in relation to management. 

Journal of Forestry. 16: 389-398.

Baker, F. S. 1925. Aspen in the central rocky mountain region. 

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Bancroft, B. 1989. Response of aspen suckering to pre-harvest 

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Barnes, B. V. 1966. The clonal growth habit of American aspens. 

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Barnes, B. V. 1975. Phenotypic variation of trembling aspen in 

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Bartos, D.L.; Campbell, R.B. 1998a. Decline of quaking aspen in the 

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USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 57

Chapter 4

Pinyon/Juniper Woodlands

Robin J. Tausch and Sharon Hood

Introduction

Pinyon-juniper woodlands occur in 10 states and cover 

large areas in many of them. These woodlands can be 

dominated by several species of pinyon pine (Pinus 

spp. L.) and juniper (Juniperus spp. L.) (Lanner 1975; 

Mitchell and Roberts 1999; West 1999a). A considerable 

amount of information is available on the expansion of 

the woodlands that has occurred over large parts of the 

geographic ranges of the tree species involved (Miller and 

Tausch 2001). In southern Utah, the woodlands contain 

Utah juniper (Juniperus osteosperma), singleleaf pinyon 

(Pinus monophylla), and two-needled pinyon (Pinus 

edulis). Singleleaf pinyon is present in the western, and 

two-needled, or Colorado pinyon, in the eastern portions 

of the region. Each species can occur alone, or in a mix 

of one of the pinyon species with Utah juniper. Overall, information on woodland ecology for the southern 

Utah study area is limited. For this reason, the available 

literature for the woodlands in general, but particularly 

for the Great Basin, will be summarized and possible 

implications for southern Utah indicated.

Wherever they are found, pinyon and juniper woodlands are a landscape scale phenomenon (fig. 1). These 

trees have large ecological amplitudes and are capable 

of invading into, and dominating, a wide range of communities (Miller and Tausch 2001; West and others 

1978a,b; West and others 1998). Their range can extend 

Figure 1—Pinyon-juniper landscape with an old chaining treatment. 
58 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 4—Pinyon/Juniper Woodlands

from the upper edge of salt desert shrub communities at 

the lowest elevations to the lower fringes of subalpine 

communities at the higher elevations (West and others 

1998). The dynamics of these woodlands and their associated vegetation processes need to be understood at 

landscape scales. In the Great Basin, most of the communities where the trees have established were dominated by sagebrush (Artemisia spp.) (West and others 

1978b). The woodland tree species are found associated 

with a range of sagebrush species and subspecies, and 

each taxon can represent a range of plant communities. 

Sagebrush presence usually extends in elevation both 

below and above the range of potential tree distribution, 

and their associated ecosystems are the matrix within 

which the trees occur. As a result, the dynamics of the 

woodlands are linked with, and in many ways dependent 

on, the dynamics of the shrub-dominated ecosystems 

involved.

Where they co-occur, sagebrush and woodland communities can have different states or levels of co-dominance within the overall successional dynamics of the 

sagebrush/woodland ecosystem complex of a particular 

landscape area. It is not possible to really understand 

or manage pinyon and juniper at these landscape scales 

without understanding the entire topography, soils, and 

vegetation complex for each landscape area of interest. 

Because this ecosystem complex is dynamic and highly 

variable across the landscape, identification of community type is determined from the species composition of 

the associated, usually sagebrush-dominated, communities (West and others 1978b; 1998). At any woodland 

location on the landscape, its successional status and 

associated ecosystems are the result of complex interactions of topography, soils, environmental conditions, 

past patterns of disturbance, and how successional processes have operated through time. In some locations in 

southern Utah, Gambel oak (Quercus gambellii) can be 

a part of the community (Thompson 1999; West 1999b) 

and can influence community dynamics.

Historical Conditions

How the patterns of disturbance were spatially distributed across the landscape and the subsequent successional 

changes through time following those disturbances were 

much different prior to Euro-American settlement than 

afterward. Prior to settlement, the primary disturbance 

was fire (Gruell 1999; Miller and Tausch 2001). The 

pattern and behavior of fire was closely related to the 

unique interactions of topography, soils, environmental 

conditions, and vegetation composition present at that 

time on each landscape area of interest. These complex 

community types contained an equally complex mix of 

fuel types and levels that determined fire pattern and 

behavior. Across the Great Basin region, trees were 

present on less than one-third of the area they currently 

occupy. For areas where trees were present, their preEuro-American settlement densities were on average 

about one-fourth to one-tenth or less of the density present at the beginning of the twenty-first century (Bauer 

2006; Miller and others, in press). Similar densities 

appear to have been present on the eastern Colorado 

Plateau portion (Floyd and others 2000; Romme and 

others 2003) of the southern Utah study area.

Vegetation cover prior to pre-European settlement 

varied widely, depending on local conditions, from less 

than 20 percent to over 80 percent on the most productive sites (West and others 1998). The majority of sites 

were below 50 percent total cover (Miller and Tausch 

2001). Similar averages and variability in cover appear 

to have been present in the southern Utah area. Total 

vegetation cover appears to have always been relatively 

the same for similar sites, whether they were sagebrushdominated or tree-dominated. Total biomass, however, 

varied from about seven to nearly 20 times greater when 

a site was tree-dominated versus sagebrush-dominated 

(Tausch and Tueller 1990).

The mix of sagebrush- and tree-dominated sites over 

the pre-Euro-American settlement landscape and the 

distribution of size and age classes within tree-dominated 

sites depended on the interactions between disturbance 

patterns and post-disturbance successional development. 

The primary control on these differences appears to be 

landscape variation in the pattern and frequency of fire. 

The heterogeneity of the landscape, combined with variation in successional processes, associated heterogeneous 

mix of community types, and associated fuel types and 

fire regimes, resulted in the maintenance of vegetation 

that varied widely across the landscape.

Pre-settlement old growth woodlands were commonly 

found on relatively fire safe sites with limited surface 

fuel loads (fig. 2) (Burkhardt and Tisdale 1969; Burwell 

1998, 1999; Holmes and others 1986; Miller and Rose 

1995, 1999; West and others 1998). The high level of 

landscape and associated vegetation heterogeneity 

present prior to European settlement resulted in a high 

degree of edge between sagebrush and tree-dominated 

communities (Tausch and Nowak 1999). These heterogeneous conditions often represented optimum habitat for 

many species of wildlife (Miller and Tausch 2001).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 59

Chapter 4—Pinyon/Juniper Woodlands

Then, as now, larger fires tended to occur during 

periods of drought (Betancourt and others 1993; Swetnam and Betancourt 1998). Insects, diseases, and native ungulates appear to have played a widespread but 

relatively minor role. Information is more limited for 

the Colorado Plateau than the Great Basin, but it indicates fire may have been less frequent in many areas 

compared to the Great Basin (Floyd and others 2000; 

Romme and others 2003). Overall, there was a dynamic 

balance between disturbance and succession resulting 

in a complex shifting distribution of the woodland and 

sagebrush dominance throughout the landscape.

It is the interaction between topography, vegetation, 

and fire that influenced both the patterns of disturbance 

and the kinds of communities that were found on a particular position on the landscape at a particular point 

in time. The deeper soils in the canyon bottoms and 

swales are generally more productive, particularly for 

the herbaceous species. These locations appear to have 

had the highest fire frequencies (Bauer 2006; Burwell 

1998; Gruell 1999; Swetnam and Basian 1996). As soils 

become shallower, generally as the topography becomes 

steeper, the abundance of perennial herbaceous vegetation is limited to years with above average precipitation. On these locations, fires appear to have been less 

frequent, increasing the probability of dominance by 

trees. The most fire-safe sites, generally on the steepest 

slopes or shallowest soils, were generally the locations of 

woodlands that were often several centuries old (Miller 

and others 1999; Waichler and others 2001). These sites 

also have generally low levels of productivity of perennial herbaceous vegetation. A few pre-settlement aged 

woodlands appear to be present from nothing more than 

the off-chance of not having burned for over 200 years 

(Miller and others, in press).

Current Conditions

Euro-American settlement activities have caused major 

changes to the composition of vegetation within the Great 

Basin (Miller and Tausch 2001; Rowland and Wisdom 

2005). The rapid woodland expansion observed during 

the late 1800s and early 1900s resulted from a combination of conditions (Miller and Tausch 2001; Miller 

and others, in press): (1) heavy livestock grazing that 

removed the herbaceous vegetation (fine fuels), (2) the 

associated reduction in the presence of fire (Heyerdahl 

and others 2001; Savage and Swetnam 1990; Swetnam 

and Betancourt 1998), and (3) wet conditions that created 

an ideal situation for tree establishment (Antevs 1938). 

The resulting expansion and increasing dominance of 

the trees in the Great Basin has continued to the present 

(Burkhardt and Tisdale 1976; Cottam and Stewart 1940; 

Miller and Rose 1995, 1999; Miller and others, in press; 

Tausch and others 1981).

Livestock grazing, particularly in the late 1800s and 

early 1900s (Young and Evans 1989), generally had the 

largest impact on the vegetation composition. Grazing 

Figure 2—The presence of 

very old Utah juniper (Juniperus 

osteosperma) suggests that this 

rocky site would rarely, if ever, 

develop a grassy understory 

capable of carrying a surface fire. 
60 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 4—Pinyon/Juniper Woodlands

reduced the herbaceous vegetation cover, which resulted 

in a reduction in fire frequency (Burkhardt and Tisdale 

1969, 1976; Campbell 1954; Ellison 1960; Miller and 

Rose 1999; Miller and Tausch 2001). The reduction 

of herbaceous species by grazing also promoted an 

increase in shrub cover. The shrubs acted a nurse crop 

and promoted tree seedling establishment (Burwell 1998, 

1999; Chambers and Vander Wall 1999; Chambers and 

others 1999; Chambers 2001; Cottam and Stewart 1940; 

Eddleman 1987; Madany and West 1983; Miller and Rose 

1995, 1999). With the reduction in fire frequency, the 

new tree seedlings were able to survive and the areas of 

woodlands expanded. As with pre-settlement woodlands, 

total vegetation cover of expansion woodlands remains 

relatively similar to the shrub cover that preceded tree 

dominance (Tausch and Tueller 1990; Tausch and West 

1995). Therefore, when the shrub layer was absent, the 

establishment of the trees was more limited (Erdman 

1970; Everett and Ward 1984).

Much of the woodland expansion has been into the more 

productive sites (for example, canyon bottoms and swales). In 

the absence of fire, the trees are well adapted and competitive 

in these more productive locations. Prior to tree expansion 

these areas represented some of the more diverse and productive sagebrush ecosystems in the region and currently 

support, or will support, some of the highest levels of 

tree dominance and fuel loads. Pre-Euro-American 

settlement woodlands have had up to a 10-fold increase 

in tree densities during this period (Bauer 2006; Miller 

and others, in press). Density increases may be less on 

the eastern Colorado Plateau portion (Floyd and others 

2000; Romme and others 2003) of the southern Utah 

study area. As the area of tree dominance continues to 

increase in the Great Basin, the heterogeneous sagebrushdominated ecosystems are being replaced by homogenous 

woodlands (fig. 3) (Miller and Tausch 2001; Milne and 

others 1996; Tausch 1999a).

Before about 1870, woodlands occurred on less than 

10 percent of their currently occupied area in the northwest Great Basin (Miller and others 1999) and on less 

than 30 percent in the central and southern Great Basin 

(Creque and others 1999; Miller and others, in press; 

Miller and Tausch 2001; O’Brien and Woodenberg 1999; 

Rogers 1982; Tausch and others 1981). Little information 

is available on the pre-settlement woodlands of the 

Colorado Plateau. Expansion woodlands now cover an 

average of three to four times the pre-Euro-American 

settlement area (Chambers and others 2000a, b; Miller 

and others, in press). This woodland expansion has 

proceeded at a nearly continuous rate across the Great 

Basin over the last 100 years (Chambers and others 

2000a, b; Miller and others, in press) and possibly 

equals or exceeds previous woodland expansions of the 

Holocene (Miller and Wigand 1994). Consequently, sagebrush communities will continue to decline as this tree 

dominance continues to increase (Despain and Mosley 

Figure 3—Initial establishment of Utah juniper (Juniperus osteosperma) in a stand of big 

sagebrush (Artemisia tridentata).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 61

Chapter 4—Pinyon/Juniper Woodlands

1990; Miller and others 1994; Miller and Tausch 2001; 

Suring and others 2005; Tress and Klopatek 1987; West 

1984).

Because of the generally slow growth of the trees, it 

has taken all of the approximately 100 years since EuroAmerican settlement for a doubling of the fuel loads to 

take place. Trees in the extensive areas of woodlands that 

have established over the last century are now rapidly 

maturing, and as they do, the fuel loads are increasing 

at an accelerated rate on these sites. On the majority of 

these areas, the density needed for trees to dominate 

is now in place (Miller and others, in press). While it 

took fuel loads in the expansion woodlands the past 100 

years to double (Chambers and others 2000a, b), they 

will double again in the next 40 to 50 years (Miller and 

Tausch 2001). The expansion of tree distribution into 

new sagebrush areas is continuing (Betancourt 1987; 

Knapp and Soule 1998; Miller and others 2000; Miller 

and Tausch 2001; Suring and others 2005; West and 

Van Pelt 1986), and with it continues the increase in 

the level and continuity of tree-dominated fuel loads. 

Similar patterns appear to exist in southern Utah.

The rate of the transition from sagebrush ecosystem to 

tree-dominated woodland is variable and depends on the 

site potential, sagebrush species and subspecies present, 

and rate of tree establishment (Miller and Tausch 2001; 

Miller and others, in press). In general, a minimum of 

60 to 90 years is required for trees to dominate a site 

(Barney and Frischknecht 1974; Huber and others 1999; 

Miller and others 1999; Miller and Rose 1995; Miller 

and Tausch 2001; Tausch and Tueller 1990; Tausch and 

West 1995). The decline in sagebrush biomass is not 

proportional to the increase in tree biomass (fig. 4). When 

the trees have reached about one-half their potential 

biomass, sagebrush biomass has declined to about onethird, sometimes one-fourth, of its former level (fig. 5) 

(Miller and Tausch 2001; Tausch and West 1995). The 

pattern of the decline is relatively consistent across the 

sagebrush species and subspecies, although the rate 

involved is not (Miller and others, in press; Miller and 

Tausch 2001). This expansion may be facilitated by the 

increasing CO2 in the atmosphere (Johnson and others 

1993). Similar changes in the landscape level patterns are 

present in the woodland changes of southern Utah.

As the dominance of the trees continues to increase 

beyond Phase I or Phase II (fig. 5), not only will fuel loads 

double from current levels over the next few decades, 

but the continuity of those fuels across the landscape 

will rapidly increase. Because of the young age of the 

trees, the ongoing increases in fuel loads are primarily in the highly flammable fine fuels. Once the trees 

dominate a site, these fine fuels can reach 9,000 kg/ha 

(10 tons/acre) on more productive sites (Chambers and 

others 2000a, b). Overall, woodland sites in the Great 

Basin vary widely, but probably have average fuel levels of about two-thirds of those sampled on the more 

Figure 4—Increasing dominance and water use by Utah juniper (Juniperus osteosperma) 

are the likely cause of the die-back of big sagebrush (Artemisia tridentata) on this site. 
62 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 4—Pinyon/Juniper Woodlands

Figure 5—Comparison of both the total tree leaf biomass (closed boxes) and total 

understory leaf biomass (open boxes) over time as indexed by the range in leaf 

biomass weighted average age of pinyon for 14 plots in southwestern Utah (Tausch 

and West 1995; Miller and Tausch 2001). X-axis = (sum of [tree age * leaf biomass] 

over all trees) / total stand leaf biomass. Phase I is the early tree establishment 

phase, Phase II is the period of tree growth and increasing tree dominance, and 

Phase III is tree dominance in expansion woodland sites.

productive central Nevada sites. The increasing crown 

closure of post-settlement woodlands is increasing the 

occurrence of crown fire (Miller and Tausch 2001; 

Tausch 1999a,b; West 1999a,b). Similar patterns appear 

to be taking place in southern Utah.

Post-fire vegetation response depends on the composition of the shrub-dominated community and the level 

of tree dominance (Barney and Frischknecht 1974; 

Dhaemers 2006; Erdman 1970; Everett and Ward 1984; 

Pickett 1976; West and VanPelt 1986). As the trees 

dominate a site, there is a decrease in the herbaceous 

species (Dhaemers 2006; Erdman 1970; Koniak and 

Everett 1982; West and others 1978a; West and others 

1998), an increase in soil erosion (Wilcox and Breshers 

1994), changes in soil fertility (Rau 2005), losses in forage 

production, and changes in wildlife habitat (Miller and 

Tausch 2001). The more dominant the trees are at the 

time of disturbance, the more the plant species composition of the communities that follow the disturbance can 

change. The intense crown fires more frequently occurring on tree-dominated sites further reduces understory 

plant species survival (Tausch 1999a).

Exotic species are changing the outcome of post-fire 

response (D’Antonio and Chambers 2006). The higher 

the level of tree dominance, the higher the probability that 

a crown fire will leave an open site. These open sites are 

increasingly being dominated by exotic plant species, many 

of which are annuals, such as cheatgrass (Bromus tectorum) 

(Chambers and others 2007; Roundy and others, in press). 

The presence of cheatgrass can cause increases in fire size 

and frequency (D’Antonio and Chambers 2006; Swetnam 

and others 1999; Tausch 1999b; Whisenant 1990; Young 

and Evans 1973) and homogeneity of those communities 

across the landscape (Young 1991; Young and Evans 1973). 

More recently, exotic perennials have begun establishing 

in these areas (D’Antonio and Chambers 2006). Once this 

conversion occurs, any return to the original sagebrush 

ecosystem, or even eventually to woodland, is often no 

longer possible without extreme restoration efforts (Miller 

and Tausch 2001). Our ability to assess the susceptibility 

of the communities associated with different sagebrush 

species and subspecies to conversion to cheatgrass is improving (Chambers and others 2007; Roundy and others, 

in press).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 63

Chapter 4—Pinyon/Juniper Woodlands

The problems associated with the exotic grasses are 

increasing through time in proportion to the level of CO2

in the atmosphere. Three ecotypes of cheatgrass from 

three elevations in the Great Basin have been investigated 

for the effects of increasing levels of atmospheric CO2. 

Four levels of CO2 were used, ranging from pre-settlement to an estimate of 2020 levels. From pre-industrial 

levels to the estimated 2020 level, the productivity of 

the upper elevation ecotype doubles, the mid-elevation 

increases 2.5 times, and the low-elevation triples (Ziska 

and others 2005). Flammability also increases (Blank 

and others 2006). Cheatgrass will have an increasingly 

negative impact over time on any woodland site where 

it becomes established following fire.

Prior to European settlement, ecosystems in the Great 

Basin and southern Utah were, on a landscape basis, 

more resilient to disturbance by fire. Fire regimes 

differed between the different sagebrush species and 

subspecies-dominated ecosystems and their associated 

communities, but were relatively consistent within each 

community. With tree expansion and increasing dominance there has been an increase in homogeneity and 

a loss of resiliency. Even old growth woodland areas 

that were relatively protected from fire prior to settlement are increasingly susceptible to fire damage from 

the adjacent tree-dominated areas (Miller and Tausch 

2001) that increasingly dominate fire behavior (Hann 

and others 2004). Developing effective restoration procedures requires more study on the ecology, structure, 

and long-term dynamics of the woodlands and their 

interaction with the associated sagebrush ecosystems 

(Chambers 2005; Dhaemers 2006; Miller and others, 

in press).

Fire Regimes

Five fire regime classes have been defined for assessing landscape dynamics for historic or past fire 

patterns and frequency (table 1) (Hann and others 

2004; Hann and Strohm 2003; Romme and others 2003; 

Schmidt and others 2002; Waichler and others 2001). 

Class I was very rare in the southern Utah area prior 

to Euro-American settlement and is not covered here. 

Prior to settlement, a heterogeneous mix of fire regime 

classes II through V often existed within relatively 

Table 1—Natural (historical) fire regime classes from Hann and others (2004) as interpreted by the authors for this analysis.

Fire Frequency

regime (mean fire return

class interval, in years) Severity Community structure description

I 0 to 35 Surface mixed Open woodland or savannah structures maintained by frequent fire; also

 frequent includes frequent mixed severity fires that create a mosaic of different age 

post-fire open woodland, early to mid seral woodland structural stages, 

and shrub or perennial grass dominated patches. 

II 0 to 35 Replacement Shrub or shrub/perennial grass maintained or cycled by frequent fire: fires

 frequent kill non-sprouting shrubs, such as sagebrush, which typically regenerate 

and become dominant within 10 to 20 years; fires remove tops of sprouting shrubs such as rabbitbrush, which typically resprout and can dominate 

after several years; fires typically kill most tree regeneration. 

III 35 to 200 Mixed surface Mosaic of different age post-fire woodland, early to mid-seral (Phase I and

 infrequent Phase II, fig. 6) woodland structural stages, and shrub or shrub/perennial 

grass dominated patches maintained or cycled by infrequent fire. 

IV 35 to 200 Replacement Large patches of post-fire shrub or shrub/perennial grass dominated

 less frequent structures, or early to sometimes late seral (Phase I to Phase III, fig. 6) 

woodland cycled by infrequent replacement fire. 

V > 200 All types May have large patches of similar post-fire shrub or shrub/perennial grass

 rare dominated structures, or early to usually late seral (Phase I to usually 

Phase III, fig. 6) woodland cycled by rare replacement fire. In systems with 

little fire or only localized torching effects of lightning strikes effects the 

composition and structure may be complex. 
64 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 4—Pinyon/Juniper Woodlands

small areas of the landscape. Because of the vegetation 

heterogeneity that existed prior to settlement, the abundance and distribution of the various types were first 

controlled by topographic heterogeneity and secondly, 

were both determined and controlled by the vegetation 

heterogeneity. Some separation of, or differences in, fire 

regime probably existed between the sagebrush species 

and subspecies present prior to settlement reflecting 

differences in their specific site conditions. Even with 

the abundance and widespread distribution of areas 

with the fire regimes classes II, III, and IV, large areas 

representing fire regime V, often old growth woodlands, 

were still present and widely distributed within, and 

often surrounded by the other, fire regimes.

Since Euro-American settlement, increasing homogeneity of the vegetation has resulted in increased fuel 

loads and continuity. Areas that were in fire regime 

V are now in fire regime IV, or sometimes even III. 

This trend will continue as the surrounding vegetation 

changes. The vegetation heterogeneity that resulted from 

differences between sagebrush species and subspecies is 

generally disappearing (Miller and Tausch 2001). Many 

areas are increasingly at risk of fire and increased fire 

size and frequency (Hann and Strohm 2003; Swetnam 

and others 1999). Despite the increasing appearance of 

homogeneity with woodland expansion, the vegetation 

response following a crown fire can still be driven as 

much by the differences between the sites identified by 

the original sagebrush community as by the level of tree 

dominance.

Fire Regime Condition Classes

Three Fire Regime Condition Classes (FRCC) have 

been defined for assessing the departure of current vegetation communities from historical vegetation structure 

and fire patterns (chapter 1, table 2, this volume) (Hann 

and Strohm 2003). Overall, the appearance of a particular 

woodland site and its associated area of the landscape 

determine the effective FRCC. This is because the context of the surrounding landscape, particularly where it 

is represented by expansion woodlands, can drive fire 

behavior and severity independent of the conditions of 

a particular old growth woodland site (Hann and others 2004). Effectively restoring a mix of sagebrush and 

woodland dominance at the landscape level also requires 

the restoration of the former landscape heterogeneity 

that makes a dynamic stability with fire possible.

Prior to settlement, most of the sagebrush/woodland 

areas of the Great Basin, and apparently southern Utah 

as well, were in FRCC 1. Woodlands and sagebrush 

ecosystems were in a dynamic balance from areas that 

burned two to three times a century, to areas that burned 

about once a century, to areas where fire occurred at 

intervals greater than 200 years. Areas that burned more 

frequently were sagebrush and bunchgrass dominated 

at the time of the fire, although some trees may have 

established after a previous fire. Because of the ongoing 

changes, most of the Great Basin is at least in FRCC 

2. Many sites may already be in FRCC 3 or are rapidly 

approaching that condition.

For many woodlands, FRCC sometimes does not depend so much on what an individual pinyon-juniper stand 

looks like, but on the probable pre-settlement community 

composition and the current landscape context within 

which it is located. For example, an area that was old 

growth woodland prior to settlement could be in FRCC 1 

when surrounded by sagebrush-dominated communities. 

The surrounding areas remained in sagebrush because 

those communities supported more frequent fire of lower 

intensity. Under these conditions fires, usually did not 

crown into the adjacent old growth woodlands and they 

appear to have remained relatively fire safe (fig. 6). Most 

of the expansion woodlands, however, are occurring in 

areas that were usually sagebrush-dominated prior to 

settlement, changing the pattern and behavior of fire 

compared to what occurred prior to tree dominance. 

These tree-dominated expansion woodlands often have 

continuous canopy cover, which can support high intensity crown fires under high wind conditions. These 

adjacent, expansion woodland sites can now drive fire 

behavior (Hann and others 2004). As a result, the old 

growth woodland stand becomes FRCC 3, even where 

little change to the woodlands has occurred. However, 

many of these pre-Euro-American settlement woodlands have also experienced increases in tree density, 

sometimes up to 10 times or more over the last century 

(Bauer 2006; Miller and others, in press), which is also 

directly changing the Fire Regime Condition Class of 

these sites. Combined, these vegetation changes are 

resulting in changes in the size, intensity, and frequency 

of fires for all parts of the landscape. This includes the 

increasing risk for conversion to cheatgrass.

For the Great Basin as a whole, little of the current 

woodland area is in FRCC 1. These are the pre-settlement 

woodlands that have not seen a significant increase in 

tree density and may represent less than 10 percent of 

today’s total (Bauer 2006; Miller and Tausch 2001). Future loss of some amount of current sagebrush-dominated 

ecosystems to tree encroachment is still possible (Suring 

and others 2005). The situation in southern Utah appears 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 65

Chapter 4—Pinyon/Juniper Woodlands

to be similar. Depending on location and community, 

current vegetation attributes put one-half to two-thirds 

of the current woodland/sagebrush ecosystem complex 

in the Great Basin in FRCC 2. Most of these woodlands 

that are currently in FRCC 2 will be rapidly transitioning 

to FRCC 3 over the next 40 to 50 years. This has serious 

implications for habitat for multiple species (Wisdom 

and others 2005). Similar changes appear to be occurring in southern Utah with similar consequences.

Recommended Treatments

The goals of treating woodlands include fuel load 

reductions, restoration of sagebrush communities, 

increasing the heterogeneity of the landscape and associated disturbance processes, improving watershed 

protection, enhancing wildlife habitat, and increasing forage production (Miller and Tausch 2001). The 

locations of the treatment sites or patches should be 

based on topographic features and areas that tended 

to have a higher fire frequency and thus, historically 

were more likely dominated by sagebrush communities. 

These are areas with deeper soils and higher herbaceous 

vegetation productivity that can carry fire. Retaining 

pre-settlement woodland sites requires as much or more 

effort to restore the surrounding communities as it does 

to restore the pre-settlement site.

Many treatment procedures have been attempted, but 

they have often been unsuccessful over the long term 

because of the lack of information about treatments 

(Chambers 2005; Tausch and Tueller 1977, 1995). A 

focus on landscape scales, rather than on just individual 

project scales, can improve treatment effectiveness 

(Hann and Bunnell 2001). Central to this has been the 

general lack of recognition of the variability of the communities that the trees are capable of dominating, and 

the range of disturbance histories represented by the 

previous communities. Because there is even less direct 

information available for the southern Utah woodlands, 

the distribution and extent of similar conditions and 

patterns of change in the area need to be determined 

on a site-specific basis.

Tree Removal

Tree removal is the primary management option 

for restoring areas affected by the ongoing woodland 

expansion. However, additional treatments have been 

proposed, many of them using new techniques. First 

and second order effects, and the success and longevity 

of the outcomes of any treatment, are highly specific 

to the site and the method used, how the treatment is 

used and its timing. For a detailed description of common treatments in pinyon-juniper communities, refer to 

the restoration chapter in Monsen and Stevens (1999) 

and Monsen and others (2004). However, some general 

Figure 6—An old growth Utah 

juniper (Juniperus osteosperma) 

growing with pinyon pine (Pinus

spp.). The tree likely reached 

this age because of inadequate 

surface fuels to carry high 

intensity fire and stand density 

was too low to support crown fires. 

The increased tree density is 

increasing the risk of lethal fire. 
66 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 4—Pinyon/Juniper Woodlands

guidelines are becoming apparent as the results from 

past and current studies improve our understanding of 

how these treatments interact with the vegetation dynamics in the woodland zone (Chambers 2005; Miller and 

Tausch 2001; Monsen and others 2004). Applied in the 

right way, at the right place, and at the right time with 

the proper follow up, if needed, any of the existing or 

proposed treatments can have positive outcomes.

Prescribed Fire

Prescribed fire may be used to remove trees and restore 

sagebrush communities before tree dominance is so high 

it reduces surface fuels to a low enough level that they 

cannot carry fire. Once tree dominance is at the high 

levels of late Phase II or Phase III (fig. 5) for an extended 

period of time, susceptibility to the establishment of 

exotics such as cheatgrass increases. Once these levels 

of dominance are reached, some form of mechanical 

treatment followed by seeding is necessary to reduce 

the level of tree dominance (Chambers 2005). This allows recovery of sagebrush and herbaceous vegetation 

before the use of prescribed fire can more fully restore 

a sagebrush ecosystem.

Prescribed fire in pinyon-juniper has been used to 

control the establishment of trees, increase forb productivity, increase habitat diversity, control invasion 

of other conifers, alter herbivore distribution, enhance 

forage palatability and nutritive quality, and prepare 

sites for reseeding (Bunting 1990). While prescribed 

fire can be beneficial, many limitations exist. Vegetation response following fire depends on the composition 

of the shrub community on a site and the level of tree 

dominance (Barney and Frischknecht 1974; Dhaemers 

2006; Erdman 1970; Everett and Ward 1984; Monsen 

and others 2004; Pickett 1976; West and VanPelt 1986). 

As trees increasingly dominate a site, the associated 

sagebrush ecosystems are greatly reduced (Chambers 

2005; Erdman 1970; Koniak and Everett 1982; West and 

others 1998). This reduction in fine fuels often makes it 

difficult for a fire to carry through a mid-successional 

stand. If fire does occur, increasing tree dominance 

increases the recovery time of herbaceous plants and 

increases the potential for invasion of exotic plants and 

erosion (Bunting 1990; D’Antonio and Chambers 2006). 

Bruner and Klebenow (1979) developed an index to 

predict when fire will carry through mid-successional 

pinyon and juniper based on wind speed, shrub and tree 

cover, and air temperature. Dangerous burning conditions exist when the index is greater than 130. Optimal 

prescribed burning conditions are an index between 125 

and 130. This can be modified by fuel moisture levels. 

Tree-dominated woodlands can be easier to burn than 

the mid-successional woodlands and are increasingly 

carrying large crown fires (Miller and Tausch 2001).

Mechanical Thinning

Chaining and thinning are the most commonly used 

mechanical methods to reduce tree cover. This may 

be necessary prior to prescribed burning in order to 

reduce crown fuels and stimulate understory vegetation. 

In Spanish Fork Canyon, UT, chaining increased total 

ground cover from 47 to 80 percent and forage production from < 22 kg/ha (<20 lbs/acres) to 1,120 kg/ha 

(1,000 lbs/acres) 7 years after treatment (Chadwick 

and others 1999). Similar increases were seen between 

4 and 7 years after chaining in eastern Nevada (Tausch 

and Tueller 1977, 1995) This initial increase in ground 

cover resulted in significantly less runoff and soil erosion than the control area (Farmer and others 1999). The 

size, type, and arrangement of the chain can be varied 

to accomplish different objectives and control the size 

and amount of trees removed. Stevens and Monsen 

(2004) provide basic guidelines for chaining in pinyonjuniper. Double chaining in opposite directions removes 

additional trees missed in the first pass and covers the 

seed after the area has been broadcast seeded prior to 

the second pass (Stevens 1999a). A once over chaining 

is appropriate if sufficient understory remains, trees are 

sparse and mature, and seeding is not required (Stevens 

1999b).

Chaining for tree control increases herbaceous biomass, but can be short-lived. Often after the 4- to 7-year 

increase there can be a rapid decline to pre-chaining levels 

in 25 years as a result of accelerated growth of surviving 

trees (Tausch and Tueller 1977, 1995). Although usually 

a stand alone procedure, chaining should generally be 

used only as an effective first treatment followed by a 

second treatment, such as prescribed fire, which would 

remove the surviving trees.

Thinning overstory trees with handsaws reduces tree 

cover and causes less soil disturbance than chaining 

(Loftin 1999). In a case study in New Mexico, Loftin 

(1999) reported 2.5 times greater herbaceous cover two 

growing seasons after hand felling juniper trees without 

seeding. This method can also be marketed as a fuelwood sale to offset costs. In an economic comparison 

of chaining versus thinning using chainsaws, Chadwick 

and others (1999) found thinning cost 44 percent more 

than chaining. In the same study, thinning did not create 

an effective seedbed, and subsequent forage production 

was low compared to the chaining treatment. The different responses between the two studies are most likely 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 67

Chapter 4—Pinyon/Juniper Woodlands

due to differences in pre-treatment site conditions. This 

underscores the importance of choosing appropriate 

site-specific treatments. Hand thinning can be as equally 

short-lived as chaining and should also be considered 

as either a pre-treatment procedure before prescribed 

fire, or a regularly repeated treatment.

Mastication is another increasingly popular mechanical 

thinning method. This method grinds and chips trees to 

reduce tree cover and compact fuel beds. Over 13,360 

ha (33,000 acres) have been masticated in Utah alone 

(Bruce Roundy, personal communication). Because mastication is such a new treatment, the ecological effects 

are largely unknown and warrant future research.

Seeding

Seeding may be required to prevent the establishment of 

exotic weeds if the understory is depauperate (Thompson 

and others 2006). After a tree removal treatment, seeding should occur prior to the next growing season to 

restrict the establishment of exotics (Stevens and Monsen 

2004). Fall seeding is the most ideal time to seed in the 

Intermountain West, although in southern Utah, seeding 

just prior to mid-July monsoons has also been successful (Stevens 1999b). Fixed wing aircraft, helicopters, or 

rangeland drills are normally used for seeding. Aerial 

seeding treats large areas on steeper slopes or where tree 

densities are high. Drill seeding is used in open areas 

(Thompson and others 2006). Aerial seeding followed 

by chaining after fire significantly increased seeded 

grass cover and decreased cheatgrass cover compared 

to seeding alone (Ott and others 2003).

Historically, introduced species seed mixes were used 

to control soil erosion and forage production. In recent 

years, there has been more interest in using native seed 

mixes to increase species diversity and restore ecosystems (Richards and others 1998). Successful establishment of native grasses and forbs from different seed 

mixes has been demonstrated in several recent studies 

(Ott and others 2003; Thompson 2006; Waldron and 

others 2005).

Herbicides

Herbicides to control encroaching pinyon and juniper 

trees are another alternative to reduce tree cover. Basal 

spraying of herbicides allows for highly selective application with little effect on non-target species. Tebuthiuron 

(Spike® 80W) and picloram (Tordon® 22K) are commonly used herbicides in these systems. Parker and others 

(1995) tested the two chemicals’ efficacies in controlling 

pinyon and juniper trees using basal application under 

different concentrations. Control was best for trees less 

than 1.8 m (6 ft) in height, with picloram killing over 

90 percent of the sprouts and seedlings. Tebuthiuron, a 

slower acting herbicide, killed 30 to 60 percent of the 

sprouts and seedlings after 9 months, but results were 

expected to improve over time. Mortality of trees taller 

than 1.8 m (6 ft) was between 10 and 30 percent for 

picloram and 5 and 10 percent for tebuthiuron (Parker 

and others 1995). Johnsen (1986) states that individual 

tree application is best suited for newly invaded sites 

with fewer than 500 trees/ha (200 trees/acre) under 

1.8 m (6 ft) tall. The longevity of these treatments will 

depend on the number and age class of the trees removed. 

Concentrating only on the older trees and leaving many 

of the younger trees will reduce the longevity of the 

treatment.

Broadcast application of tebuthiuron and picloram 

produce more variable results. One-seed juniper 

(Juniperus monosperma) and Rocky Mountain juniper 

(J. scopulorum) are often the most difficult to control 

(McDaniel and WhiteTrifaro 1986). Johnsen (1986) 

reported that herbicides readily killed trees on the ridges, 

but not on areas of deep soils or bottom land. Trees along 

the ridges are often old growth pockets of pinyon and 

juniper that generally should not be a target for removal. 

Areas invaded by pinyon and juniper, where herbicides 

are not as effective, are also places where fire would 

have historically limited their establishment. These are 

the areas often needing treatment. Concern over killing 

non-target species, with potentially limited mortality of 

pinyon and juniper, makes this treatment less desirable 

than individual tree application.

Summary

Management goals that deal with woodland expansion 

need to account for the landscape variability in community composition and disturbance regimes (Miller 

and Tausch 2001). Vegetation treatments should also 

focus on the source of woodland changes. In other 

words, they should focus on areas of the woodlands that 

represent expansion beyond the pre-settlement distributions. Within these areas, the focus should then be on 

woodland sites that have only recently transitioned away 

from FRCC 1. There are areas of recent tree expansion 

with vegetation attributes that indicate they may still be 

returned relatively easily to FRCC 1. These communities include the remaining sagebrush-dominated areas, 

sagebrush areas still in early Phase I (fig. 5), and the 

areas of old growth woodlands that are present within 
68 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 4—Pinyon/Juniper Woodlands

the landscape matrix. The sagebrush ecosystems and 

their associated communities, successional classes, and 

distributions need to be determined on a landscape basis 

for each management area.

There is, however, another management reality. With 

at least two-thirds of the woodland area in the Great 

Basin (and probably in southern Utah as well) representing expansion woodlands, millions of acres are 

now involved. Even under the best of conditions, only a 

minority of such a large area will be successfully treated 

before a wildfire occurs. For the remaining areas that 

are being increasingly dominated by trees that will burn 

before treatment is possible, we need to determine restoration/rehabilitation needs and possibilities following 

wildfire, particularly when cheatgrass or other exotics 

are present. Additional research is needed to help with 

the development of these procedures.

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Tausch, R.J.; Nowak, R.S. 1999. Fifty years of ecotone change 

between shrub and tree dominance in the Jack Springs pinyon 

research natural area. In: McArthur, E.D.; Ostler, K.W.; Wamboldt, 

C.L., comps. Proceedings: shrubland ecotones. RMRS-P-11. 

Ogden, UT: U.S. Department of Agriculture, Forest Service, 

Rocky Mountain Research Station: 71-77.

Tausch, R.J.; West, N.E.; Nabi, A.A. 1981. Tree age and dominance 

patterns in Great Basin pinyon-juniper woodlands. Journal of 

Range Management. 34: 259-264.

Thompson, R.M. 1999. An example of pinyon-juniper woodland 

classification in southeastern Utah. In: Monsen, S. B.; Stevens, 

R., eds. Proceedings: ecology and management of pinyon-juniper 

communities within the Interior West: 1997 September 15-18, 

Brigham Young University, Provo, UT. Proc. RMRS-P-9. Ogden, 

UT: U.S. Department of Agriculture, Forest Service, Rocky 

Mountain Research Station: 60-63.

Thompson, T.W.; Roundy, B.A.; McArthur, E.D.; Jessop, B.D.; 

Waldron, B.; Davis, J.N. 2006. Fire rehabilitation using native 

and introduced species: a landscape trial. Rangeland Ecology 

and Management. 59: 237-248.

Tress, J.A.; Klopatek, J.M. 1987. Successional changes in community structure of pinyon-juniper woodlands in North-Central 

America. In: Everett, R.L., comp. Proceedings: pinyon-juniper 

conference. Gen. Tech. Rep. INT-215. Ogden, UT: U.S. Department of Agriculture, Forest Service, Intermountain Forest and 

Range Experiment Station: 80-85.

Waichler, W.S.; Miller, R.F.; Doescher, P.S. 2001. Community 

characteristics of old-growth western juniper woodlands in the 

pumice zone of central Oregon. Journal of Range Management. 

54: 518-527.

Waldron, B.L.; Monaco, T.A.; Jensen, K.B.; Harrison, R.D.; Palazzo, 

A.J.; Kulbeth, J.D. 2005. Coexistence of native and introduced 

perennial grasses following simultaneous seeding. Agronomy 

Journal. 97: 990-996.

West, N.E. 1984. Successional patterns and productivity potentials of pinyon-juniper ecosystems. In: Developing strategies 

for rangeland management: a report. Westview Press, Boulder, 

CO: 1301-1332.

West, N.E. 1999a. Juniper-pinyon savannas and woodlands of western 

North America. In: Anderson, R.C.; Fralish J.S.; Baskin, J.M., 

eds. Savannas, barrens, and rock outcrop plant communities of 

North America. Cambridge University Press, London, United 

Kingdom: 288-308.

West N.E. 1999b. Distribution, composition, and classification of 

current juniper-pinyon woodlands and savannas across western 

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within the Interior West: 1997 September 15-18, Brigham Young 

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Department of Agriculture, Forest Service, Intermountain Forest 

and Range Experiment Station: 43-52.

West, N.E.; Tausch, R.J.; Rea, K.H.; Tueller, P.T. 1978a. Phytogeo- Phytogeographical variation within juniper-pinyon woodlands of the Great 

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sagebrush within the pinyon-juniper woodlands of the Great 

Basin. Journal of Range Management. 31: 87-92.

West, N.E.; Tausch, R.J.; Tueller, P.T. 1998. A management oriented 

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Tech. Rep. RMRS-GTR-12, Ogden, UT: U.S. Department of Agriculture, Forest Service, Rocky Mountain Research Station. 42 p.

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INT-276. Ogden, UT: U.S. Department of Agriculture, Forest 

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raptor management symposium and workshop, Scientific and 

Technical Series No. 12. Washington, D.C.: National Wildlife 

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USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 73

Chapter 5

Big and Black Sagebrush Landscapes

Stanley G. Kitchen and E. Durant McArthur

Introduction

Perhaps no plant evokes a common vision of the 

semi-arid landscapes of western North America as do 

the sagebrushes. A collective term, sagebrush is applied 

to shrubby members of the mostly herbaceous genus, 

Artemisia L. More precisely, the moniker is usually 

restricted to members of subgenus Tridentatae, a collection of some 20 woody taxa endemic to North America 

(Beetle 1960; McArthur 1979; McArthur and Plummer 

1978). As a group, the Tridentatae are distinguished from 

other members of the genus by a combination of traits 

including their woody habit, floral morphology, stem 

anatomy, plant chemistry, and chromosomal karyotype 

(McArthur 1979).

The genus Artemisia originated on the Eurasian 

landmass during the mid-Tertiary as the late-evolving 

Asteraceae rapidly diversified in response to global 

expansion of drier and cooler habitats (Beetle 1979; 

Raven and Axelrod 1974). Toward the end of this period, 

herbaceous, mesic-adapted progenitors to contemporary Tridentatae migrated across Beringia and spread 

across western North America eventually developing a 

woody habit (Beetle 1979; McArthur 1999; McArthur 

and Plummer 1978; Stebbins 1972). Opportunities for 

continued diversification were plentiful in the spatially 

and temporally diverse environment provided by the 

interaction of a complex geography with the increasingly variable climate of the Quaternary (Beetle 1979; 

McArthur and Plummer 1978). Over time, one widely 

adapted species, big sagebrush (Artemisia tridentata), 

emerged as the “most widespread and common shrub of 

western North America” (fig. 1) (McArthur and Stevens 

2004).

Figure 1—A big sagebrush 

(Artemisia tridentata) landscape.
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Chapter 5—Big and Black Sagebrush Landscapes

Scientific and management interest given to sagebrush 

in general, and big sagebrush in particular, can be measured by the considerable volume of literature generated 

primarily during the last half-century. McArthur and 

others (1979), Blaisdell and others (1982), and McArthur 

and Stevens (2004) provide useful reviews of the ecology and management of sagebrush species and ecosystems. Various papers presented in thematic symposia 

(McArthur and Welch 1986; Utah State University 1979) 

were effective in synthesizing available knowledge, 

and although dated, the published proceedings remain 

valuable reference materials. Literature summaries are 

available online by species or subspecies from the USDA 

Forest Service, Fire Effects Information System data 

base (http://www.fs.fed.us/database/feis/) (Howard 1999; 

Johnson 2000; McMurray 1986; Tirmenstein 1999a). 

Welch’s (2005) big sagebrush synthesis is the latest and 

most comprehensive review published for this species 

complex.

Our purpose is not to provide yet another general 

sagebrush review, but to explore relevant published work 

and current thought regarding structure and successional 

processes, particularly as they relate to fire in ecosystems 

dominated by big sagebrush, and to a lesser extent, black 

sagebrush (A. nova). Although we briefly consider paleo 

distribution to provide context, our temporal focus will 

be primarily the last two centuries with emphasis on 

systematic changes that resulted from Euro-American 

settlement in the mid- to late-1800s. The time periods 

before and after this major cultural/ecological shift will 

be referred to as pre- and post-settlement. The geographical area of interest is represented by a broad zone in 

central and southern Utah where the Great Basin and 

Colorado Plateau meet, referred to here as the Southern 

Utah study area, or simply, the study area. Black sagebrush is included in the discussion because of its broad 

ecological overlap with big sagebrush and because of 

its widespread distribution and importance in the study 

area. In certain cases, such as the assessment of paleo 

and historic distribution, information is lacking for clear 

distinctions among sagebrush taxa. However, we believe 

that inferences made for the group as a whole will largely 

hold true for its dominant member, big sagebrush, and 

the closely allied black sagebrush.

Sagebrush Biology and Reproduction

Big sagebrush is a medium- to long-lived (20 to 

200+ years; McArthur and Stevens 2004; Perryman 

and others 2001) aromatic evergreen shrub with one to 

several main stems (McArthur and Stevens 2004). The 

gray to black bark on older branches is shredded and 

shaggy (Beetle 1960; McArthur and others 1979). Typical persistent leaves are small, pale green to blue-green, 

narrowly wedge-shaped, with three blunt teeth on the 

broadened end (McArthur and Stevens 2004). Spring 

ephemeral leaves are larger and generally more variable 

in shape and size than persistent leaves (Miller and Shultz 

1987). Leaves and young stems are covered by a mat 

of fine hairs that provide a silvery cast. Inconspicuous, 

wind-pollinated flowers are held above foliage on fine, 

more or less erect stems. The wind-dispersed seeds 

(achenes) are small (4,000 to 6,000 seeds per g) and 

lack specialized appendages (Meyer and others 1988a; 

Welch 2005). Dispersal distance varies with topography 

and local conditions. Maximum dispersal distance for 

sagebrush seeds has been estimated at 30 m (98 ft) (Johnson and Payne 1968; Walton and others 1986), although 

the majority of seeds generally disperse less than 3 m 

(10 ft) from the mother plant (Walton and others 1986; 

Wambolt and others 1989). Seeds are short-lived and 

do not form a persistent seed bank (McDonough and 

Harniss 1974; Meyer 1990; Young and Evans 1989).

Three widely recognized subspecies of big sagebrush 

differ in a number of morphological and physiological 

traits (McArthur and Stevens 2004). Basin big sagebrush 

(ssp. tridentata) is the tallest, typically 1 to 2.5 m (3.3 to 

8.2 ft), and has an uneven-shaped crown. The crown 

shape of Wyoming big sagebrush (ssp. wyomingensis) 

is similar to that of basin big sagebrush, however, plants 

are generally less than 1 m (3.3 ft) tall. Mountain big 

sagebrush (ssp. vaseyana) is intermediate in height at 

0.6 to 1.5 m (2 to 5 ft). Seed stalks and foliage are of 

even height above the crown and give it a more flattened 

appearance than the other subspecies. Seed maturation 

and dispersal vary among subspecies and stand elevations (McArthur and Stevens 2004; Welch 2005) and 

are generally latest for basin big sagebrush (November 

to early December) and earliest for higher elevation 

populations of mountain big sagebrush (late September 

to October). Seed production is greatest for basin big 

sagebrush and least for Wyoming big sagebrush. Significant variation in palatability and nutritive content to 

wildlife and livestock has been documented and related 

to differences in the chemistry of secondary compounds 

(Welch and McArthur 1981, 1986; Welch and others 

1981, 1987; Sheehy and Winward 1981). Dormancy in 

recently dispersed seed correlates with climate (elevation) 

of the collection site (Meyer and Monsen 1991, 1992). 

Although similar in many ways, black sagebrush differs from big sagebrush in a number of traits including 

shorter stature (generally less than 60 cm [24 inches]), 

darker appearance of leaves and reproductive stems, 
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Chapter 5—Big and Black Sagebrush Landscapes

larger seed size (2,200 seeds per g), and fewer seeds 

produced (McArthur and Stevens 2004; Meyer and 

others 1988b).

Geographic Distribution

Vegetation reconstructions based upon sediment core 

pollen records and woodrat midden macrofossil assemblages reveal that sagebrush-dominated ecosystems have 

been widespread during both the warm and cold phases 

characteristic of the Quaternary (last 2 million years). 

Studies suggest that during the late Pleistocene and 

early Holocene (40,000 to 10,000 years b.p.), sagebrush 

was widespread throughout most of its modern range 

(Mensing 2001; Nowak and others 1994; Rhode and 

Madsen 1995; Thompson 1990) and sagebrush-steppe 

ecosystems likely formed an ecotone with montane 

and continental tundra (Anderson and others 1999; 

Betancourt 1990; Fall and others 1995; Vierling 1998). 

Although the specific sagebrush taxa are unknown, a 

dominant sagebrush steppe association extended eastward into what is now the Central and Southern High 

Plains (Fredlund 1995; Hall and Valastro 1995) and into 

the desert valleys and plains of the southwest (Spaulding 1990; Van Devender 1990) during this same period. 

Sagebrush largely retreated from these eastward and 

southern range extensions during the hot dry conditions 

of the mid-Holocene (8,000 to 5,000 years b.p.) to an 

area similar to its modern distribution.

Today, big sagebrush is found throughout western 

North America from southern Canada to Baja California (McArthur 1999; McArthur and Plummer 1978). 

Beetle (1960) estimated that big sagebrush-dominated 

communities occupy approximately 586,306 km2

(226,374 mi2) in 11 western states. Although this is 

considered an overestimate (McArthur and Stevens 

2004; Wright and others 1979), its widespread ecological dominance remains impressive. With an estimated 

area of dominance of 112,150 km2 (43,301 mi2) (Beetle 

1960), black sagebrush communities occupy a greater 

total area than any other member of the Tridentatae

except for big sagebrush and silver sagebrush (A. cana; 

McArthur and Stevens 2004).

Historical Conditions

A long standing controversy persists regarding the 

relationship between modern and pre-settlement distribution and condition of big sagebrush communities 

(Johnson 1986; Peterson 1995; Young and others 1979). 

One view holds that in response to livestock grazing 

practices and altered fire regimes, big sagebrush invaded 

large landscapes that were predominantly grasslands 

(Arno and Gruell 1983; Christensen and Johnson 1964; 

Cottam 1961; Cottam and Stewart 1940; Hull and Hull 

1974, Stewart 1941). In this context, big sagebrush is 

considered an indicator, or even as an agent, of grassland degradation justifying eradication in the name 

of restoration (Blaisdell and others 1982; Britton and 

Ralphs 1979). This view became entrenched in early 

range ecology dogma (Cottam and Stewart 1940; Stewart 

1941; Stoddard 1941) and has retained popularity to the 

present. The opposing view claims that, with the exception of lands converted to other uses, the distribution 

of big sagebrush landscapes is essentially unchanged 

from historic times (Hironaka 1979; Johnson 1986; 

Vale 1975; Welch 2005). This view admits to changes 

in shrub dominance in response to disturbance (for example, livestock grazing; Austin 2000), but denies the 

supposition that significant change in vegetation type 

has occurred. This view is supported by arguments that 

expansion rates for sagebrush are too slow to account for 

significant range advances in the suggested time frame 

of approximately 100 years (Welch 2005).

Early written accounts produced by trappers, explorers, 

immigrants, and settlers have been interpreted to support 

both mindsets. Young and others (1979) found support 

for grass-dominated systems in Stewart’s (1941) review 

of historical records of Utah’s rangelands. In his treatment of the same document, Welch (2005) argues that 

grassy areas were not characteristic. Vale (1975) found 

consistent references to expansive sagebrush-dominated 

landscapes from central Wyoming to western Nevada 

and central Oregon in early journal/diary descriptions 

of vegetation along major migration routes across the 

western United States. He concluded that grasslands 

were restricted to canyon and valley locations with 

favorable soil moisture conditions. His interpretations 

are strengthened by the decision to only use accounts 

of observations made before heavy use by immigrants 

and their livestock. Analysis of published series of 

photographs taken as early as the 1870s and recent retakes provide additional insight but fail to resolve the 

issue. Arno and Gruell (1983) and Kay (2003) provide 

photographic evidence and arguments in support of 

sagebrush invasion of grasslands in southwestern Montana and central Utah. In contrast, Johnson (1986), after 

examination of 1870s photos from Wyoming, northern 

Utah, and southeast Idaho, concluded that grasslands 

and shrublands have been quite stable for 115+ years in 

spite of a wide range of disturbances imposed during the 
76 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 5—Big and Black Sagebrush Landscapes

interim. Interpretation of frozen-in-time descriptions or 

photos is confounded by a lack of knowledge regarding 

disturbance history and the relative proportions of the 

various seral stages on historic landscapes (Young and 

others 1979). Thus, a definitive answer to the debate 

may never be found due to lack of reliable information 

(Johnson 1986; Young and others 1979) and because of 

the likelihood that no single answer is correct across 

the full geography of big sagebrush. However, after a 

review of the available evidence and arguments, it is our 

opinion that, allowing for defensible exceptions, changes 

in the distribution of sagebrush-dominated landscapes 

over the last 150 years came primarily in the form of 

reductions rather than expansions. The principal causes 

for these reductions are land use conversion, woodland 

expansion, and more recently, increased fire frequencies 

associated with invasive annual grasses.

Ecological Distribution and 

Associated Species

Big and black sagebrush are largely restricted to semiarid climate regimes where winter temperatures are cool 

to cold and winter-spring precipitation is sufficiently 

reliable to support spring growth (Beetle 1960; McArthur 

and Stevens 2004). Summer precipitation varies regionally, but soils at rooting depth are typically dry for much 

of the growing season. Soils on big sagebrush-dominated 

landscapes are moderately shallow to deep, well drained, 

and low in salt content (West 1979). Soil pH may vary 

from slightly acidic to moderately alkaline. Soils associated with black sagebrush tend to be drier and are 

generally more shallow or of higher percent rock than 

soils supporting big sagebrush (McArthur and Stevens 

2004; Welch 2005).

Big and black sagebrush ecosystems form a wide, 

mostly continuous band across gradients in elevation in 

the southern Utah study area with lower limits defined 

by an ecotone with salt-desert shrublands and an upper boundary somewhat restricted by dense stands of 

montane or subalpine forest (fig. 2). Imbedded at midelevations within this sagebrush-grass matrix is a broad 

zone prone to recurrent invasion by species of pinyon 

pine and juniper (fig. 3) (Miller and others 1999; Tausch 

1999; Tausch and Hood, this volume). This invasion 

belt is somewhat centered on the Wyoming-mountain 

big sagebrush transition zone (Goodrich and others 

1999). These woodland species have been more or less 

permanent occupants on fire-protected topographical 

Figure 2—Upper end of the sagebrush zone, with mountain big sagebrush (Artemisia 

tridentata spp. vaseyana), shrubs, and grass intermixed with conifers and aspen (Populus 

tremuloides).
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Chapter 5—Big and Black Sagebrush Landscapes

units since their arrival in the Holocene (Tausch 1999; 

West and others 1998). For millennia, the rate and extent 

of expansion into the sagebrush matrix and frequency 

of retreat have been regulated by climate, topography, 

and fire regime (Tausch 1999; Tausch and Nowak 2000). 

The close association of sagebrush and woodland cover 

types in this zone led Tausch and Hood (this volume) to 

suggest that they are best considered as different phases 

of a single system (West and others 1978, 1998). They 

characterize a system of pinyon-juniper woodlands superimposed over a sagebrush-grass matrix with variation 

in the relative importance of the two types as an expression of variation in successional status across complex 

topographic and disturbance landscapes. This is clearly 

a refreshing and useful way to consider these dominant 

shrubland-woodland mosaics. For a detailed consideration of sagebrush shrubland-pinyon-juniper woodland 

successional patterns and management implications, 

see Tausch and Hood (this volume). Our focus here will 

primarily be sagebrush-grassland ecological processes 

that function independent of a woodland component.

Before settlement, Wyoming big sagebrush was the 

most abundant of the three subspecies of big sagebrush 

across its geographic range (West 1979). Within the 

southern Utah study area, it is a dominant on deep 

to moderately deep soils at elevations between 1,500 

and 2,000 m (4,920 and 6,560 ft). Typical landforms 

include broad alluvial fans, low foothills, plateaus, and 

valleys receiving 170 to 350 mm (6.7 to 13.8 inches) 

annual precipitation (Goodrich and others 1999). At 

lower and upper ends of its range, it is bounded by, and 

intermixed with, elements of salt-desert shrub, black 

sagebrush, and mountain big sagebrush-pinyon-juniper 

communities (Howard 1999). The kind and abundance 

of sub-dominant shrubs and perennial grasses varies 

with soil attributes and disturbance history. Common 

shrub associates include species of ephedra (Ephedra 

spp.), rabbitbrush (Chrysothamnus spp.), snakeweed 

(Gutierrezia spp.), horsebrush (Tetradymia spp.), saltbush 

(Atriplex spp.), antelope bitterbrush (Purshia tridentata), 

and winterfat (Ceratoides lanata). Historically, perennial grasses dominated herbaceous understory. Common 

native species in the study area are Sandberg bluegrass 

(Poa secunda), bluebunch wheatgrass (Pseudoroegneria 

spicata), western wheatgrass (Pascopyron smithii), 

bottlebrush squirreltail (Elymus elymoides), Indian 

ricegrass (Stipa hymenoides), needle-and-thread (Stipa 

comata), and galleta grass (Hilaria jamesii). Perennial 

forb diversity and cover is relatively low on Wyoming 

big sagebrush sites (Bunting 1985).

Mountain big sagebrush in the study area is found 

on foothills and dry mountain slopes and ridges at 

Figure 3—Utah juniper (Juniperus osteosperma) moving into a stand of big sagebrush 

(Artemisia tridentata).
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Chapter 5—Big and Black Sagebrush Landscapes

elevations of 1,900 to 3,000 m (6,230 to 9,840 ft) in 

moderately deep loamy soils (fig. 4) (McArthur and 

Stevens 2004). Annual precipitation varies from 300 

to 700 mm (11.8 to 27.6 inches) (Goodrich and others 

1999). At higher elevations it occurs in forest openings of 

various sizes in association with quaking aspen (Populus tremuloides), Douglas-fir (Pseudotsuga menziesii), 

Englemann spruce (Picea englemannii), white fir (Abies 

concolor), and ponderosa pine (Pinus ponderosa) (Johnson 2000). At lower elevations, mountain big sagebrush 

dominates many treeless landscapes and co-exists in 

shrub — woodland mosaics with single-leaf pinyon 

(Pinus monophyla; Great Basin), two-needle pinyon 

(P. edulis; Colorado Plateau), Rocky Mountain juniper 

(Juniperus scopulorum), Utah Juniper (J. osteosperma), 

and Gambel oak (Quercus gambelii). The importance 

of co- and sub-dominant shrubs varies with topography, 

soils, and disturbance history. Common associates include mountain snowberry (Symphoricarpus oreophilus), 

common juniper (J. communis), currants (Ribes spp.), 

Oregon grape (Mahonia repens), mountain mahogany 

(Cercocarpus spp.), serviceberry (Amelanchier spp.), 

antelope bitterbrush, rabbitbrush, and green ephedra 

(E. viridis). Numerous species of perennial grasses and 

forbs combine to make a productive understory. Species 

composition and annual biomass production vary with 

site productivity potential and disturbance history.

Basin big sagebrush grows in deep, well-drained soils 

of plains, valleys, low foothills, and canyon bottoms 

at elevations of 1,500 to 2,100 m (4,920 to 6,890 ft) in 

the southern Utah study area (McArthur and Stevens 

2004). Mean annual precipitation is approximately 300 

mm (11.8 inches). It is most often found in association 

with species common to upper Wyoming big sagebrush 

and lower mountain big sagebrush although populations 

intermixed with salt tolerant black greasewood (Sarcobatus vermiculatus), shadscale (Atriplex confertifolia), 

and saltgrass (Distichlis spicata) also exist (McArthur 

and Stevens 2004).

Within the study area, black sagebrush is most abundant at elevations of 1,500 to 2,400 m (4,920 to 7,870 ft) 

(fig. 5) (McArthur and Stevens 2004) with an extended 

elevational range of 1,400 to 2,780 m (4,590 to 9,120 ft), 

thus the elevational range of black sagebrush is nearly 

equal to the combined range of the three sub species 

of big sagebrush (McMurray 1986). Black sagebrush 

segregation from big sagebrush is due to its ability to 

grow in shallow, rocky soils and mixing of the two species is generally limited to narrow ecotones. On lower 

elevations, it is common to find black sagebrush in nearly 

pure stands with only sparse herbaceous understory and 

associated shrubs. As with big sagebrush, the diversity 

and abundance of associated shrubs and understory 

species increases with elevation.

Figure 4—Dense stand of mountain big sagebrush (Artemisia tridentata spp. vaseyana).
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Chapter 5—Big and Black Sagebrush Landscapes

Fire Effects

Although extreme weather (Anderson and Inouye 

2001; Nelson and others 1989; Nelson and Tiernan 1983), 

insects (Haws and others 1990; Nelson and others 1989), 

and disease (Nelson and others 1989, 1990; Sturges and 

Nelson 1986) apparently play a significant role in big 

sagebrush population dynamics, and do so at multiple 

spatial scales, fire is believed to be the dominant disturbance force in natural populations (Wright and Bailey 

1982). Curiously, big sagebrush lacks morphological 

or physiological adaptations to survive fire or facilitate 

rapid recolonization (Welch and Criddle 2003). Plant 

stature is low to the ground and wood, bark, and foliage are highly flammable resulting in complete shoot 

mortality of burned plants (McArthur and Stevens 2004). 

Top-killed plants do not re-sprout from roots or crown 

(Blaisdell and others 1982; Britton and Ralphs 1979; 

Peterson 1995; Wright and others 1979). Seeds mature and 

disperse after the risk of fire has all but passed (Beetle 

1960; Young and Evans 1989). Seeds have no mechanism 

for long-distance dispersal (Chambers 2000; Johnson 

and Payne 1968; Walton and others 1986; Wambolt and 

others 1999; Young and Evans 1989). The soil seed bank 

is ephemeral or absent (Beetle 1960; McDonough and 

Harniss 1974; Meyer 1990, 1994; Meyer and Monsen 

1992; Young and Evans 1989). It appears paradoxical 

that the widespread landscape dominant, big sagebrush, 

is so poorly adapted to flames when fire is considered 

the “keystone disturbance” of western North American landscapes (Frost 1998; Keane and others 2002). 

Black sagebrush is no better adapted to fire than is big 

sagebrush, but dominates on sites less prone to burn. In 

contrast, many co-occurring shrubs have at least some 

ability to tolerate burning or to rapidly recolonize after 

fire (table 1). Fire adaptation by herbaceous species 

associated with big and black sagebrush varies, but 

is generally superior to that of these dominant shrubs 

(Britton and Ralphs 1979; Wright and Bailey 1982). 

Hence, a resolution to the apparent fire-big sagebrush 

paradox is not to be found solely in a species by species description of fire effects and adaptations, but in 

an examination of the manner in which fire is manifest 

on the landscape through time and space, also known 

as the fire regime.

Figure 5—Black sage (Artemisia nova) with needle-and-thread grass (Stipa comata).
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Chapter 5—Big and Black Sagebrush Landscapes

Big and Black Sagebrush Fire Regimes

Fire regime is quantified using various temporal and 

spatial parameters including frequency, seasonality, predictability, extent, and pattern (Morgan and others 2001). Fire 

regime is also expressed in terms of intensity, a measure of 

heat production per unit of time, and severity, a measure of 

fire-induced ecosystem change (Romme and others 2003; 

Ryan and Noste 1985). Fire regimes vary through time 

and across the landscape. Temporal variation is generally 

climate driven (Brown and others 2001; Grissino-Mayer 

and others 2004; Grissino-Mayer and Swetnam 2000; 

Heyerdahl and others 2002; Swetnam and Betancourt 

1998), while spatial variation is primarily a product of 

topographic variation through its effects on species composition, productivity, desiccation rates, fuel continuity, 

and wind speed (Brown and others 2001; Heyerdahl and 

others 2001; Swetnam and Baisan 1996; Taylor and Skinner 

1998). Fire characteristics vary locally in response to recent 

fire history and adjacency to fire prone landscape units 

(temporal and spatial autocorrelation; Morgan and others 

2001).

Fire frequency is the most common measure of fire 

regime. It is an expression of the mean number of years 

between fire events, or mean fire interval (MFI), for a 

defined geographic unit. Estimates and interpretations 

of MFI are dependent upon spatial scale (Baker and 

Ehle 2001); the larger the area the shorter the interval 

in which no fire occurred. Hence, the most interpretable 

estimates of fire frequency are those associated with 

relatively small geographic units (Xiaojun and Baker 

2006). Because populations of fire intervals frequently 

are not normally distributed, other measures of central tendency may be more appropriate for predicting 

fire free intervals (Grissino-Mayer 1999). However, 

differences among candidate statistics are generally 

not ecologically significant. Conversely, interval variability can be important and is often overlooked. For 

species that must regenerate from seed, such as big 

sagebrush, the length and frequency of short intervals 

is most important in determining the compatibility of 

the fire regime with species persistence (Crawford and 

others 2004). Conversely, the length and frequency of 

long intervals are also important for ecosystems prone 

to invasion by fire sensitive species. The susceptibility 

of many big sagebrush landscapes to invasion by pinyon 

pine, juniper, or other conifer species in the absence of 

fire illustrates this point (Heyerdahl and others 2006; 

Miller and others 1999; Tausch and Hood, this volume). 

Thus, big sagebrush-dominated ecosystems provide clear 

examples of how a fire-free window, defined by both 

short and long interval statistics, can be more useful in 

determining ecosystem structure than are estimates of 

central tendency alone.

Estimates of MFI for forested ecosystems are most 

often generated from two types of dendrochronological evidence. Years of low severity or surface fires are 

determined from tree ring series with datable fire scars 

(Arno and Sneck 1977). Once injured, fire scarred trees 

become more susceptible to injury from subsequent 

fires. Consequently, individual fire-recording trees may 

provide evidence of a large percentage of low severity fires that burned at one location for extended time 

periods. More complete fire chronologies are obtained 

by combining fire records from annually cross-dated 

trees growing in close proximity (Dieterich 1980). Fire 

dates from severe, stand-replacing fires can be estimated 

based upon synchronous patterns in stand establishment 

dates (Heyerdahl and others 2001).

Table 1—Fire adaptations for big sagebrush and co-occurring shrub species. Each species is rated on a scale 

of 0 to 4 for each area of adaptation where a 0 indicates no adaptation and a 4 indicates the species is 

highly adapted. See discussion and references in (Aleksoff 1999; Anderson 2004; Howard 1997, 1999, 

2003; Johnson 2000; Marshall 1995; Tirmenstein 1999a, b; Welch and Criddle 2003; Zlatnik 1999).

 Seed Seed

 Resprouting maturation dispersal Seed

Scientific name Common name capability timing distance bank

Artemisia tridentata Big sagebrush 0 0 1 1

Chrysothamnus nauseosus Rubber rabbitbrush 4 0 4 0

Purshia tridentata Antelope bitterbrush 2 2 2 2

Amelanchier alnifolia Saskatoon serviceberry 4 1 3 1

Cercocarpus montanus True mountain mahogany 3 1 3 1

Atriplex canescens Fourwing saltbush 2 0 1 3

Ephedra nevadensis Nevada ephedra 3 3 1 2

Symphoricarpus oreophilus Mountain snowberry 3 0 1 4
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 81

Chapter 5—Big and Black Sagebrush Landscapes

These methods are difficult to apply to sagebrush 

ecosystems except where fire-recording trees grow in 

isolated pockets or forest-shrubland ecotones. Such 

conditions are limited primarily to the more mesic 

mountain big sagebrush sites. Houston (1973) used fire 

scarred Douglas-fir and lodgepole pine trees growing 

at the ecotone between forest and mountain big sagebrush-grass steppe in northern Yellowstone National 

Park to estimate MFI in sagebrush steppe for that area. 

After adjusting data to reflect only pre-1900 conditions, 

he calculated mean single-tree MFI values of 32 to 70 

years for all study units and 44, 56, and 50 years for 

trees growing in the sample units most representative 

of the whole study area. He considered this an overestimate of true MFI so, using composite chronologies 

of questionable accuracy (not cross-dated), he estimated 

MFI values of 20 to 25 years. Using a similar approach, 

Arno and Gruell (1983) examined fire frequency at 

forest-mountain big sagebrush-grass steppe ecotones 

of southwest Montana. They calculated pre-1900 MFI 

values of 41, 45, and 74 years for moist, dry, and hotdry habitat types. These investigators suggested that 

these estimates were likely overestimates of MFI due 

to possible missing fire evidence and adjusted their 

estimate of forest “grassland” ecotone MFI to 35 to 40 

years. They concluded that sagebrush distribution and 

density has increased considerably in this region due to 

a reduction of fire frequency during the last 100 years. 

In a more recent study from the same region, Heyerdahl 

and others (2006) sampled fire scarred Douglas-fir 

trees from a 1,030 ha (2,544 acre) site topographically 

characterized as a mosaic of forest islands and mountain 

big sagebrush-grass elements. They estimated an average fire return interval of 37 years for the study period 

(1700 to 1860) with a range of fire-free intervals of 2 to 

84 years. This is similar to the results observed by Arno 

and Gruell (1983). They also quantified the increase in 

tree distribution and density that occurred after 1860. 

Miller and Rose (1999) estimated MFI for mountain big 

sagebrush steppe in a south central Oregon study area 

using fire scars from four isolated clusters of ponderosa 

pine trees located in the mountain big sagebrush-grass 

matrix. Composite, pre-1900 MFI ranged from 12 to 15 

years for three of the four clusters. Seven major fires 

(three or four clusters affected) occurred between 1650 

and 1880, resulting in an approximate MFI for major 

fires of 38 years. Miller and Rose concluded that, “In 

the mountain big sagebrush community, mean fire 

intervals, prior to 1871, ranged from 12 to 15 years...” 

The estimates generated by these studies are cited 

extensively in the literature and provide the basis for 

a well-developed core conventional wisdom regarding 

sagebrush and fire.

Developing estimates of big sagebrush fire frequency 

directly from fire chronologies found on proxy species 

(trees) has the advantage of temporal precision (when 

properly dated) over long time periods. Spatial ambiguity is lessened by sampling from multiple locations on 

the landscape (Brown and others 2001; Heyerdahl and 

others 2001, 2006; Morgan and others 2001; Taylor and 

Skinner 1998). There are, however, disadvantages to this 

approach. Scarred trees are often scarce and distributed 

disproportionately across the landscape leading to spatial 

gaps in the record. Also, difficult to test assumptions 

must be made regarding historic fire regime continuity 

across the shrubland-forest ecotone. Perhaps the greatest problem with this approach rests in the fact that 

no suitable proxy species exist for the great majority 

of big sagebrush habitat types. Even if fire frequency 

estimates derived from the above cited studies prove to 

be accurate, there is considerable risk of inappropriate 

extrapolation of values to other localities.

Using an alternative approach, historic fire frequency 

for big or black sagebrush-dominated communities can 

be estimated indirectly based upon post-fire succession 

rates (Welch 2005; Welch and Criddle 2003). We suggest 

that the recovery pattern of big sagebrush to pre-burn 

conditions serves as an adequate index of post-fire succession for these plant communities. Several studies 

have attempted to quantify big sagebrush recovery time 

following both wild and prescribed fires. Harniss and 

Murray (1973) determined that big sagebrush on an upper 

Snake River Plain site in Idaho required at least 30 years 

to recover to pre-burn conditions and shorter fire-free 

intervals would lead to shrub dominance by species of 

horsebrush or rabbitbrush. Humphrey (1984) examined 

community composition in eight areas of southeastern 

Idaho where time-since-burn ranged between 2 and 36 

years. His data indicated that big sagebrush (probably 

mountain big sagebrush) was still in a recovery phase 

30+ years after burning. In Wyoming big sagebrush 

steppe in southwestern Montana, Watts and Wambolt 

(1996) observed that big sagebrush canopy cover reached 

10 percent 30 years after burn treatment compared to 

the 13.5 percent for the unburned control. Wambolt 

and others (1999) observed similar delays in recovery 

for all three major subspecies after a wildfire burned 

sagebrush steppe communities north of Yellowstone 

National Park. Stand density for Wyoming, mountain, 

and basin big sagebrush was 2, 12, and 16 percent respectively, of unburned reference areas 19 years after 

the fire. In a southwestern Montana study of 13 spring 
82 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 5—Big and Black Sagebrush Landscapes

and fall prescribed burn sites (2 to 32 years post-burn), 

mountain and Wyoming big sagebrush canopy cover or 

stand density was significantly less for burned areas than 

for unburned areas in 36 of 40 comparisons (Wambolt 

and others 2001). The authors suggested that 30 years 

might be inadequate for full recovery in many cases. 

We further anticipate that recovery periods will often 

be longer after the more intense, mid-summer wildfires. 

Twenty years after a wildfire burned a central Utah site, 

West and Yorks (2002) found that Wyoming big sagebrush 

recovery had barely started based on the low density of 

sagebrush plants. They concluded that recovery rates for 

“sagebrush semi-desert” communities are much slower 

than they are for sagebrush steppe communities. These 

studies suggest that big sagebrush requires from 20 to 

35+ years for post-fire stand recovery under favorable 

conditions and much longer intervals when conditions 

dictate a slower pace of recovery. Correspondingly, longer 

intervals are expected on xeric sites where fine fuel production under average weather conditions is inadequate 

to carry fire except under severe conditions.

There is little information regarding historic fire regimes for black sagebrush-dominated landscapes. It is 

generally believed that fire was rare on these landscapes 

due to insufficient fine fuels to carry fire (Wright and 

others 1979). Miller and Rose (1999) found evidence of 

just two fires in 300 years of record for a low sagebrush 

(Artemisia arbuscula)-western juniper (Juniperus occidentalis) community. Low sagebrush is similar to 

black sagebrush and like black sagebrush, occupies 

low productivity sites (McArthur and Stevens 2004). 

Although this information is insufficient to estimate 

MFI, it is sufficient to support the notion of long firefree intervals for sagebrush communities with low fine 

fuel production.

In cases where a significant number of big sagebrush 

seedlings establish from surviving residual seeds, stand 

recovery is relatively rapid (Bunting 1985), and the size, 

pattern, and continuity of the burn has little impact on 

recovery time. This phenomenon is most common when 

fires burn mature stands of mountain big sagebrush, but 

has also been observed with basin and Wyoming big 

sagebrush on mesic sites (Wambolt and others 1999, 

2001). Frequently, big sagebrush seedlings do not establish within 1 year post-burn, either because of a lack of 

viable seeds on the landscape, or because of the failure 

of seeds to produce viable plants (Welch and Criddle 

2003). When this occurs, big sagebrush recovery is 

dependent upon seed dispersal from unburned source 

areas and favorable weather patterns. Fire size, pattern, 

and continuity directly impact the distance that seeds 

must disperse, and thus have considerable impact on 

recovery time (Crawford and others 2004). Welch and 

Criddle (2003) provide an example that illustrates this 

effect. They measured the down wind (west to east) 

advance of mountain big sagebrush 14 years post-burn 

on a southern Idaho site and determined a mean annual 

spread rate of 13 m (43 ft). Northward spread was half that 

of the eastward spread and big sagebrush encroachment 

from the north and east burn margins was negligible. 

Based on these data, they estimated that it would take 

71 years for big sagebrush to reoccupy this 146+ ha 

(360+ acre) burn. In effect, large continuous fires without 

unburned islands result in long seed dispersal distances 

that translate into long recovery periods while recovery 

from small discontinuous fires with short seed dispersal 

distances is more rapid.

Although big sagebrush post-fire recovery time varies 

situationally, a general relationship between recovery 

time and minimum (significantly shorter than the mean) 

fire-free intervals should be expected. Therefore, estimates of MFI (or any other measure of central tendency) 

must be substantially longer than estimates of mean 

recovery time in order to capture the full variability in 

fire interval duration. Conversely, we recognize that on 

landscapes prone to conifer invasion, lengthy intervals 

would result in shrubland displacement by woodlands 

(Tausch and Hood, this volume). Taken together, these 

assumptions provide the theoretical basis for our estimates of MFI on sagebrush-dominated landscapes. We 

suggest that historic MFI values ranged from 40 to 80 

years for mountain big sagebrush and some productive 

basin and Wyoming big sagebrush communities and 

were as long as 100 to 200 years or longer for big and 

black sagebrush sites with low productivity. We offer 

broad estimates here in order to capture the range in MFI 

length we believe existed across the full ecological and 

geographical distribution of big sagebrush. A range of 

intervals lengths must be expected in conjunction with 

a single MFI value. For example, natural variability 

corresponding to a MFI of 50 years might produce 

intervals as short as 10 to 15 years and as long as 100 

to 120 years; however, most intervals would likely fall 

between 25 and 75 years. Our estimates are similar to 

pre-1900 MFI values calculated from tree-ring records 

at forest-shrubland ecotones by Heyerdahl and others 

(37 years; 2006), Houston (32 to 70 years; 1973), and 

Arno and Gruell (41 to 74 years; 1983) before data 

adjustment. Although the estimate for historic MFI of 

12 to 15 years proffered by Miller and Rose (1999) is 

unrealistic, the approximate MFI of 38 years derived 

from their data for widespread landscape fires only 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 83

Chapter 5—Big and Black Sagebrush Landscapes

approaches our estimate and is likely a more realistic 

application of their data to the mountain big sagebrush 

matrix of their study area. In their North American synthesis of fire ecology, Wright and Bailey (1982) estimate 

MFI for mountain and Wyoming big sagebrush at 50 

and 100 years – numbers that are in general agreement 

with what we suggest here.

So, what of the big sagebrush-fire paradox? It seems 

that in contrast to strategies employed by co-occurring 

shrubs (table 1), big sagebrush solved the fire problem 

by producing highly competitive, yet disposable plants. 

It does not invest resources in morphological or physiological adaptations to fire, as it never had to in its short 

evolutionary past. This was particularly true for the 2+ 

million years of the Pleistocene, during which time 

cooler climatic conditions would have rarely favored 

fire to the extent they do today. Sagebrush thrives on 

suitable landscapes as long as the fire-free intervals 

are sufficiently long to permit re-establishment of 

mature stands, and short enough to prevent displacement by forest or woodland (Miller and Tausch 2001). 

This leads us to the conclusion that past variation in 

big sagebrush distribution and dominance was to some 

degree influenced by climatically-driven changes in 

fire regime parameters. Where changes were large 

and persistent, fire regime-driven shifts among shrubsteppe, grassland, and woodland or forest ecosystems 

must have occurred. Consequently, caution is warranted 

when making comparisons of big sagebrush distribution 

or dominance from time periods experiencing distinct 

climatic patterns. Judgments made when comparing 

pre-settlement big sagebrush conditions corresponding 

to the end of the “Little Ice Age” with contemporary 

environmental conditions should be tempered by the 

context of the corresponding change in climate.

Variation in historic fire regimes might also be attributed to variation in human-caused ignitions. Fire 

was the most important tool available to aboriginal 

inhabitants for manipulating the natural environment 

(Griffin 2002; Williams 2004). It was used to promote 

the growth of desirable resource plants, enhance habitat for important animals, and drive game and insects 

(Griffin 2002). It might also have been important for 

warfare, clearing travel corridors, and providing firesafe camp sites (Williams 2004). Anecdotal and ethnographic accounts describing the use of fire within the 

big sagebrush domain are reviewed by Baker (2002), 

Griffin (2002), and Whitlock and Knox (2002). These 

authors conclude that, in spite of known patterns of fire 

use, it is difficult to find available evidence sufficient 

to attribute landscape-scale variation in pre-settlement 

fire regimes to human fire practices. Thus, we are left 

to consider the ecological consequences of historic fire 

regime variability with little ability to ascertain the human role in that variation.

Finally, we pose the question, “What does big 

sagebrush-grass climax look like in the absence of woodland invasion and fire?” Current theory suggests that, 

with time, competition from big sagebrush will eventually 

reduce perennial grasses and forbs to scattered remnants 

(Miller and Tausch 2001). Indeed, one does not need to 

look far when in sagebrush country to find dense stands 

of sagebrush with depleted herbaceous understories. 

However, the question can not be properly addressed 

without also considering the impact of domestic livestock 

on the competitive relationship between shrubs and 

perennial grasses (Austin 2000). Unfortunately, there 

are few reference areas not impacted, either historically 

or currently, by livestock grazing. The question is thus 

compounded by the additional variable and becomes, 

“What does big sagebrush-grass climax look like in the 

absence of fire, woodland invasion, and livestock grazing?” Although somewhat short term in nature, results 

from a 45-year study (Anderson and Inouye 2001) found 

that after livestock removal, a dynamic equilibrium was 

reached between shrubs and perennial grasses where 

both were well represented in the plant community. 

Other studies show significant, sustained increases in 

percent grass cover concurrent with increases in big 

sagebrush cover following livestock reduction or removal 

(Branson and Miller 1981; McLean and Tisdale 1972; 

Pearson 1965). We suggest that such a dynamic balance 

should be expected between big sagebrush and grasses 

in a variety of settings, although the actual nature of 

the balance is likely to differ substantially from place to 

place. This is not to suggest that fire is not a fundamental ecological process in sagebrush grass communities, 

nor that fire should not be used as a management tool. 

Rather, we suggest that losses of perennial grasses in 

big sagebrush-dominated communities may have more 

to do with the effects of intense, selective grazing than 

with the periodicity of fire.

Euro-American Settlement

Ecological processes associated with sagebrush/grass 

ecosystems in the southern Utah study area began to 

be altered soon after Euro-American settlement in the 

mid-1800s (Young and others 1979). First settlers were 

primarily small groups of Mormons sent by their leader, 

Brigham Young, to establish organized communities 

wherever water, timber, and forage conditions permitted. 

Numerous agricultural-based communities were started 
84 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 5—Big and Black Sagebrush Landscapes

where mountain streams met valley floors. Early settlers 

soon learned that the deep, loamy soils associated with 

tall (basin) big sagebrush were well suited for irrigated 

crops and significant acreages of this type were converted 

to cropland (Kearney and others 1914).

Originally, settlers pastured livestock anywhere good 

grass could be found close to their communities. Although few settlers owned many animals individually, 

livestock were commonly united in community herds 

(Young and others 1979). Cool-season grasses growing 

in association with big sagebrush at valley and foothill 

locations were used heavily and sometimes year-round 

when located at close proximity. The area impacted 

by grazing livestock gradually expanded as the local 

forage base depleted, herd numbers increased, and concern for potential losses to opportunistic native peoples 

decreased.

Livestock numbers, especially sheep, grew rapidly 

in the late 1870s and 1880s (Keck 1972; Murdock and 

Welsh 1971) as livestock production shifted from a subsistence to a market economy. In addition to expanding 

local herds, large numbers of sheep were trailed and 

eventually freighted by train to and from the area by 

parties with little or no interest in the grazing needs 

of local communities. Sheep herds often spent several 

months trailing between mountains in fall and summer and foothills and deserts in winter and spring. By 

1890, livestock numbers far exceeded realistic estimates 

of carrying capacity and the degradation of plant/soil 

environments was widespread. Over-grazing persisted 

for several decades resulting in widespread degradation 

of big and black sagebrush landscapes similar to that 

imposed on other plant community types.

Current Conditions

Ungulate Impacts

To some degree, pasturing of domestic livestock on 

big and black sagebrush-dominated landscapes affected, 

and continues to affect, community structure and ecological processes at all grazing intensities. Minimum 

effects include the removal of fine fuels and subsequent 

reduction in capacity to carry fire. At higher grazing 

intensities, palatable herbaceous species were weakened 

or eliminated. The ecological changes caused by these 

alterations occurred rapidly on productive landscapes 

characterized historically by relatively short fire-free intervals and a propensity for invasion by woodland or forest 

conifers. In response to a weakened, and often depleted 

herbaceous understory, big sagebrush density and canopy 

cover increased and the pace of woodland invasion was 

accelerated (Miller and Tausch 2001; Tausch 1999). The 

combination of reduced fire and accelerated woodland 

invasion resulted in loss of landscape level heterogeneity 

and a major shift in the sagebrush-woodland complex 

to increasingly more widespread and dense woodland 

dominance (Tausch and Hood, this volume). Similar 

degradation has occurred on drier landscapes, although 

woodland expansion is either absent or occurring at a 

slower pace.

Exotic Weed Impacts

Immigrating Europeans intentionally and accidentally brought the seeds of numerous plant species to 

North America. Some of these species are invasive in 

big sagebrush habitats. The winter annual, cheatgrass 

(Bromus tectorum), is particularly well adapted for big 

sagebrush habitats and has had considerable success 

invading the weakened and often depleted understories 

of basin and Wyoming big sagebrush, mid elevation 

black sagebrush, and drier mountain big sagebrush 

communities (Young 1994). During years with good 

spring moisture, cheatgrass produces large quantities 

of continuous fine fuels that cure earlier in the season 

than do native perennials and effectively lengthen the 

fire season. Where established, cheatgrass has resulted 

in shorter fire-free intervals, earlier fires, and larger, 

more continuous fires than existed historically (Peters 

and Bunting 1994; Whisenant 1990). Once established, 

cheatgrass out-competes seedlings of native perennials disrupting natural regeneration processes (Billings 

1994; Young 1994; Young and Evans 1978). Where 

the perennial herbaceous understory is depleted, the 

cheatgrass-fire cycle eventually reduces what is left 

of sagebrush-dominated communities to cheatgrassdominated annual grasslands (Pellant 1990; Young and 

Evans 1978).

Fire Regime Condition Classes

The classification of existing vegetative communities based upon the degree of departure from historic 

conditions and the risk of loss of one or more defining 

components provides a framework to guide restoration 

and management efforts. Three broad classes reflect 

increasingly greater departure from historic conditions 

implicating parallel increases in intervention needs.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 85

Chapter 5—Big and Black Sagebrush Landscapes

Fire Regime Condition Class 1 (FRCC 1)

Herbaceous species diversity and relative cover for 

healthy (FRCC 1) big or black sagebrush-grass communities in the southern Utah study area vary across 

gradients in site productivity. In general, diversity and 

relative cover decrease with increasing aridity. The 

extreme of this condition can be observed in some low 

elevation (xeric) black sagebrush communities that essentially function as stable shrublands with extensive 

bare interspaces and few scattered herbs (McArthur and 

Stevens 2004). Conversely, mesic stands of mountain 

big sagebrush have a diverse assemblage of perennial 

grasses and forbs that resist weed invasion (Anderson 

and Inouye 2001; West and Yorks 2002). Depending 

on disturbance history, percent cover of mountain big 

sagebrush varies from <5 percent, 1 to 20 years post 

burning, to 20 to 40 percent at shrub/grass equilibrium 

30 to 70 years post burn (Harniss and Murray 1973; 

Humphrey 1984; Wambolt and others 2001; Welch 2005; 

Welch and Criddle 2003). Mature sagebrush stands are 

generally multi-aged (Perryman and others 2001). On 

the landscape scale, multiple seral stages are represented 

in an ever-shifting mosaic reflecting periodic reoccurrence of fire or other disturbances (Crawford and others 

2004). Although landscape-level complexity may appear 

to decrease with the longer fire-free intervals expected 

for basin and Wyoming big sagebrush sites, microscale compositional variation may if fact increase as 

the plant community has time to fine tune responses to 

small variations in the physical environment (Anderson 

and Inouye 2001). Typically, sagebrush cover for basin 

and Wyoming big sagebrush-dominated communities 

ranges from 15 to 35 percent (Welch and Criddle 2003). 

If present in FRCC 1, cheatgrass is scattered and forms 

an insignificant portion of the herbaceous biomass. At 

mid to lower elevations, cryptobiotic crusts may be present on bare inter-shrub openings. Woodland species are 

widely scattered, if present. Fires are generally patchy. 

If grazed by livestock, light to moderate stocking rates 

and periodic rest during the growing season are needed 

to maintain the herbaceous component of this condition 

class.

Fire Regime Condition Class 2 (FRCC 2)

FRCC 2 for sagebrush-grass communities occurs 

with and without elements of woodland invasion. The 

herbaceous perennial component in this condition class 

is moderately depleted in abundance or diversity relative 

to FRCC 1. A shift in species composition reflecting 

palatability is often noticeable. The primary causes are 

chronic overstocking of livestock and periodic abusive 

grazing practices. Sagebrush cover often exceeds that 

of FRCC 1 due to competitive release by a weakened 

herbaceous understory. On drier sites, cheatgrass is generally present (some soil types excepted), but does not 

dominate except in scattered patches. Current fire-free 

intervals are often much longer than estimated historic 

intervals. However, the risks of weed invasion for this 

condition class are greater than the risks of woodland 

conversion. The probability of conversion (transition) 

from a shrub-grass community to weed-dominated 

grassland is moderate.

Tausch and Hood (this volume) give descriptions of 

big and black sagebrush landscapes susceptible to invasion by woodland trees. In FRCC 2, trees have reached 

25 to 50 percent of their potential cover for the site and 

shrub, and herbaceous cover has been reduced by up to 

75 percent. Total conversion to tree dominance (FRCC 3) 

may occur in 40 to 50 years on moderately productive 

sites (Tausch and Hood, this volume). Because cheatgrass invasion frequently occurs synchronously with 

woodland expansion, fire compounds risks. The risk 

of cheatgrass expansion with fire counters that of conversion to woodland (loss of shrubs and herbs) without 

fire. Alternative non-fire treatments or combinations of 

treatments are needed for sagebrush-grass restoration.

Fire Regime Condition Class 3 (FRCC 3)

Characteristically, perennial herbs are severely depleted in FRCC 3 big sagebrush communities of the study 

area. Weedy annuals, especially cheatgrass, dominate the 

herbaceous understory. Big sagebrush plants tend to be 

old-aged with little new recruitment. Shrub cover may 

be variable. There is a high probability that a single fire 

event will result in conversion to an annual-dominated 

community with corresponding short fire-free intervals 

and large fire size. Indeed, outside of the study area, 

extensive areas of big sagebrush plant communities 

have already experienced this type conversion (Billings 

1994; Miller and others 1999; Whisenant 1990; Young 

and Evans 1978). The current fire-free interval may be 

much longer than estimated historic intervals; however, 

the greater risks for this condition class are associated 

with shortened fire-free intervals due to the loss of 

perennial herbs and probability of invasion by weeds 

such as cheatgrass. Where susceptible to woodland 

invasion, trees may have reached densities sufficient to 

fully occupy the site (Tausch and Hood, this volume). 

Shrubs and perennial herbs are weak and scattered in 

tree openings. Litter accumulates under trees and extensive bare soil may be exposed and eroding between 
86 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 5—Big and Black Sagebrush Landscapes

trees. Long-term overgrazing of stable black sagebrush 

communities by domestic sheep can result in conversion to a FRCC 3 characterized by nearly solid stands 

of low rabbitbrush (Chrysothamnus greenii) or broom 

snakeweed (Gutierrezia sarothrae).

Fire Regime Condition Class Assessment

We assessed condition class for big and black sagebrush sites in the southern Utah study area using data 

and photographs obtained from 148 sites between 1997 

and 1999 (Davis and others 2004). Sites with boundaries 

roughly corresponding to the study area were selected 

within Utah Division of Wildlife Resources Wildlife 

Management Units 16, 19 through 25, and 27 through 

30. Study sites in which neither big nor black sagebrush 

were listed as descriptive, that appeared to have been 

recently treated (chaining or seeding), or were protected 

from grazing were excluded. A second assessment was 

made for 117 of the sites from photos and data collected 5 years after the first field visit. Data from four 

additional sites were added for this later period. The 

number of sites per sagebrush taxon was 59, 15, and 

57 for mountain, basin, and Wyoming big sagebrush, 

respectively, and 21 for black sagebrush. These data are 

the most comprehensive available and provide a sound 

basis for estimating big and black sagebrush condition 

class for the study area.

Four photographs per site/assessment period were 

used to make a qualitative assessment of condition 

class. An assessment of woodland encroachment (primarily juniper) was also made from photographs. As 

a quantitative indicator of condition for each site, we 

used the Desirable Components Index (DCI) developed 

by Davis and others (2004). The DCI is computed on a 

scale of 0 to 100 and is based on cover percentages for 

shrubs, perennial grasses, perennial forbs, and annual 

grasses and the presence of noxious weeds. Shrub values 

are adjusted based upon size class (seedlings, young, 

and mature) and vigor (normal, decadent, and dead) 

distributions. Optimal values for the DCI are obtained 

by 20+ percent shrub cover, 15+ percent perennial grass 

cover, 5+ percent perennial forb cover, no annual grasses 

or noxious weeds, shrub decadence less than 20 percent, 

and percent young shrubs greater than 10 percent. For 

more details on DCI computation see Davis and others 

(2004). We derived condition class estimates from site 

DCI scores using two scales that correspond to mesic 

and xeric habitats (table 2).

Our qualitative and quantitative estimates of site 

condition class are in general agreement (table 3). Averaged across all sites, condition class ratings associated 

with the late 1990s assessment are somewhat higher 

than those for the 2002 to 2004 assessment. This difference reflects the severity of a regional drought from 

1999 to 2004. Shrub and perennial grass mortality was 

particularly striking in 2003 at low elevations in the 

southern and eastern portions of the study area. These 

data suggest that from 10 to 20 percent of big sagebrush 

landscapes and 30 to 50 percent of black sagebrush 

landscapes in the study are currently in FRCC 1. Approximately 40 to 60 percent of the four sagebrush taxa 

are in FRCC 2. Relative area in FRCC 3 is highest for 

basin and Wyoming big sagebrush sites and lowest for 

mountain big sagebrush and black sagebrush. Moderate 

to advanced encroachment by juniper was not different 

for the three big sagebrush subspecies (40, 40, and 38 

percent for mountain, basin, and Wyoming sites, respectively). We rated encroachment as moderate to advanced 

for 29 percent of black sagebrush sites. These estimates 

of woodland encroachment may be somewhat low due 

to a possible bias against woodland-dominated sites in 

the site selection criteria.

Table 2—Scales for deriving condition class from Desirable Components 

Index (DCI) scores (Davis and others 2004).

Scale Condition class DCI score

Scale 1 – mountain big sagebrush and 1 70+

upper elevation black sagebrush 2 45-69

3 <45

Scale 2 – basin and Wyoming big sagebrush 1 55+

and lower elevation black sagebrush 2 30-54

3 <30
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 87

Chapter 5—Big and Black Sagebrush Landscapes

Recommended Treatments

A variety of treatments has been developed to modify 

and restore big sagebrush communities. The appropriateness of each is dependent upon site condition class, 

existing uses, available resources, and management goals 

(Monsen 2004). The underlying principle is to repair 

structure and ecological processes in existing vegetative 

communities (Whisenant 1999). Treatment objectives 

for FRCC 1 and many FRCC 2 sites are maintenance in 

nature and include: reduction of big sagebrush density or 

cover; removal of young pinyon or juniper trees; increase 

density, productivity, or diversity of perennial herbaceous understory; increase productivity of associated 

shrub species; and create spatial heterogeneity among 

seral stages across the landscape. Potential treatments 

that are effective in achieving those objectives include 

prescribed fire; selective herbicide application; and lowimpact mechanical treatments such as anchor chaining 

and railing. Objectives for treating lower end FRCC 2 

and FRCC 3 sites are clearly remediation focused and 

include soil stabilization, water capture and retention, 

and reconstruction of resilient shrub-perennial grass 

communities. Practices employed to achieve these 

objectives include removal of woodland trees, control of 

invasives (usually annuals but may include perennials), 

and restoration plantings of perennial herb and shrub 

elements. Successful restoration of big sagebrush-grass 

communities from FRCC 3 to FRCC 1 or 2 is generally 

an expensive multi-step process requiring combinations 

of treatments (Lancaster and others 1987) and fortuitous 

timing and quantities of precipitation after planting. 

Pre-emergent herbicides and tillage treatments have 

proven to be at least moderately effective in controlling 

invasive annuals. Bio-control methods for controlling 

cheatgrass have been investigated (Kennedy 1994; Meyer 

and others 2001); however, to date, are undeveloped or 

unproven. Prescribed fire and high impact mechanical 

treatments are employed to remove woodland trees. 

Restoration plantings require proper seedbed preparation and timely planting of appropriate seed mixes of 

adapted, compatible ecotypes (Monsen and Stevens 

2004). Appropriate long-term management practices, 

including changes in livestock use, are essential after 

treatment. In the following, we discuss the advantages 

and limitations of these treatment options.

Table 3—Big and black sagebrush condition class estimates for 152 sites in the southern Utah study 

area based upon Desirable Components Index (DCI) and qualitative assessment of photographs for each site (Davis and others 2004). Percentages are based upon 148 and 121 sites 

for the 1997 to 1999 and 2002 to 2004 assessments, respectively.

Condition class estimate

 Condition 1997 to 1999 Assessment 2002 to 2004 Assessment

Sagebrush taxon class DCI photos DCI photos

- - - - - - - - - - - - - - - - - - Percent of sites - - - - - - - - - - - - - - - - - -

Mountain big sagebrush 1 26 31 13 9

2 64 64 58 74

3 10 5 29 17

Basin big sagebrush 1 20 20 13 19

2 53 40 47 37

3 27 40 40 44

Wyoming big sagebrush 1 39 28 15 5

2 47 53 48 65

3 14 19 37 30

Black sagebrush 1 52 43 50 33

2 39 57 36 58

3 9 0 14 8

All taxa combined 1 34 31 18 11

2 53 56 50 64

3 13 13 32 25
88 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 5—Big and Black Sagebrush Landscapes

Prescribed Fire

Prescribed fire is an efficient, cost effective method 

for removing big sagebrush and woodland trees on portions of the landscape (Britton and Ralphs 1979; Bunting 

and others 1987; Wright and others 1979). Because it 

mimics natural fire, properly timed prescribed fire is 

supportive of natural ecological processes of nutrient 

cycling and plant succession. As previously discussed, 

time needed for big sagebrush recovery after burning 

can vary greatly and depends upon community composition before the burn, fire intensity (linked to season of 

burn), fire size and pattern, and weather conditions after 

the burn (Bunting and others 1987). Non-target species 

can be damaged, and increased herbaceous production 

is not always realized after burning (Britton and Ralphs 

1979; Bunting and others 1987; Welch 2005; Wright and 

Bailey 1982; Wright and others 1979). Prescribed fire, 

or wildland fire use, should be limited to stands where 

perennial grasses and forbs are sufficiently abundant to 

preclude the risk of expansion by cheatgrass or other 

fire tolerant invasives. Typical sites are mountain big 

sagebrush communities in FRCC 1 and 2. Restoration 

fires should be small or patchy, facilitating the perpetuation of a mosaic of seral stages and minimizing seed 

dispersal distances for recovering big sagebrush. Large 

patches of mature sagebrush should be left unburned as 

critical wildlife habitat. Based on a 50 year MFI, mean 

area burned per year (natural and wildfires combined) 

for mountain big sagebrush should not exceed 2 percent. 

Prescribed fire should not be considered for dry basin and 

Wyoming big sagebrush and black sagebrush stands in 

the study area due to slow recovery time and the high risk 

of conversion to weeds. Deferral from livestock grazing 

for 1 or more years before burning may be necessary to 

allow for fine fuel accumulation and curing (Whisenant 

2004; Wright and others 1979). A post-treatment rest 

from grazing of one to two growing seasons (Bunting 

and others 1987; Whisenant 2004; Wright and others 

1979) should be considered a minimum requirement 

that is not always adequate. Burning restrictions often 

result in narrow windows of opportunity for treatment 

and may require rapid mobilization. Wright and Bailey (1982), Bunting and others (1987), and Whisenant 

(2004) outline guidelines for prescribed burning of big 

sagebrush communities.

Herbicides

Herbicide treatments are used effectively as substitutes 

for fire to reduce big sagebrush cover. The volume of 

literature dedicated to the development and testing 

of various compounds for this purpose reveals the 

level of interest that existed during the mid 1900s in 

finding novel ways to control or eradicate this species 

(Crawford and others 2004; Welch 2005 and references 

therein). Herbicide selectivity and effectiveness varies 

with concentration, season of use, soil characteristics, 

and community composition (Vallentine 2004; Welch 

2005). Here we discuss the use of the two herbicides 

most frequently used to control big sagebrush. Early 

work focused on the use of 2,4-D [(2,4-D-dichlorophenoxy) acetic acid] a synthetic auxin, or plant growth 

regulator (Welch 2005). Although effective in controlling big sagebrush short-term, long-term effects on 

the plant community were difficult to predict (Watts 

and Wambolt 1996; Welch 2005). This may be due to 

its effects on non-target species, especially broadleaf 

forbs. Consequently, 2,4-D is no longer the herbicide 

of choice for big sagebrush reduction (Crawford and 

others 2004). Tebuthiuron ([N-[5-(1,1-dimethylethyl)-

1,3,4-thiadiazol-2-yl]-N,N’-dimethylurea]; Spike®), a 

photosynthesis inhibitor, is applied to the soil where 

it moves into the rooting zone with water and remains 

active for several years (McDaniel and others 2005; 

Wachocki and others 2001). It is absorbed through the 

roots and is functionally selective against big sagebrush 

at low application rates (Baxter 1998; Crawford and 

others 2004; McDaniel and others 2005; Wachocki and 

others 2001). Post-treatment increases in productivity 

for herbaceous species can be substantial (Baxter 1998; 

McDaniel and others 2005; Olson and Whitson 2002). 

Herbicide treatments, particularly the use of tebuthiuron, 

pose certain advantages over prescribed fire. Spatial precision of treatment application is greater with herbicide 

application than with prescribed fire. Longer windows 

of opportunity for treatment are available, especially 

for tebuthiuron (Baxter 1998; Marion and others 1986), 

than for prescribed fire. Damage to non-target species 

is often less with tebuthiuron than with prescribed 

fire (Baxter 1998; McDaniel and others 2005). The 

level of “thinning” and associated treatment longevity 

are effectively regulated by altering application rates 

(Crawford and others 2004; McDaniel and others 2005; 

Olson and Whitson 2002; Wachocki and others 2001), 

although these must be calibrated for soil texture and 

precipitation (Baxter 1998). There are disadvantages to 

using tebuthiuron relative to prescribed fire including 

greater per-acre cost and ineffectiveness in controlling 

woodland trees. Tebuthiuron has greatest application 

where there is a need to reduce sagebrush density or 

cover to allow existing herbaceous understory to respond 

to the competitive release. It is particularly valuable 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 89

Chapter 5—Big and Black Sagebrush Landscapes

where prescribed fire is not practical or where complete 

removal of big sagebrush is undesirable. Tebuthiuron 

may have application with FRCC 2 basin and Wyoming 

big sagebrush communities where the ecological risks 

of prescribed fire would be excessive.

A second class of herbicides is used to control invasive 

annuals, such as cheatgrass, in big sagebrush communities. Broad spectrum contact herbicides are effective, but 

collateral damage to non-target species can complicate 

their use. Soil active, pre- or post-emergent herbicides 

have proven effective in killing annual weeds in early 

post-germination stages, thus effectively depleting the 

seed bank and releasing residual perennials from weedy 

competition. Applications of sulfometuron methyl 

(Oust®) have proven effective in providing a 1- to 2-year 

window of greatly reduced competition from cheatgrass 

(Pellant and others 1999). Questions remain regarding 

residual time in the soil and impacts on established 

perennials. In recent years, considerable interest has 

been generated for imazapic (Plateau®) as a soil-active 

herbicide. Early results suggest that this herbicide is quite 

selective in its effects and that it is particularly effective 

with annual bromes (Bekedam and Pyke 2004; Porath 

and others 2003; Smith and Anderson 2003; Whitson 

2003). This kind of treatment has application on FRCC 2 

sites to encourage release of weakened herbaceous plants 

and on FRCC 3 sites to deplete the weed seed bank in 

preparation for restoration plantings.

Mechanical Treatments

A wide variety of mechanical techniques has been 

devised to eliminate invading pinyon and juniper trees 

from sagebrush grass communities and to reduce cover 

and density of big sagebrush. These treatments are generally used as precursors or sometimes simultaneously 

with restoration/reclamation plantings (Monsen 2004; 

Stevens 1999). Treatment effectiveness and management 

considerations for the mechanical control of woodland 

trees are discussed in Stevens and Monsen (2004a) and 

Tausch and Hood (this volume). Parker (1979), Mattise 

and Scholten (1994), Welch (2005), and Wiedemann 

(2005) provide brief but adequate descriptions of major 

equipment developments and their applications. Various 

plow and disk type implements kill most big sagebrush 

plants, as well as associated, species necessitating followup plantings of adapted species. Mature woodland trees 

and a high percentage of non-sprouting shrubs can be 

removed by dragging a long section of anchor chain between two crawler tractors. Actual treatment outcome is 

affected by link weight and modifications, chain length, 

relative tractor positions, treatment passes, and tractor 

speed (Stevens 1999). Low-impact treatments leave a 

majority of herbaceous species intact. A second pass of 

the chain (two-way chaining) improves juniper kill and 

is reasonably effective in burying seeds broadcast after 

the first pass as part of restoration plantings (Stevens 

1999). Chaining is a preferred technique on rough terrain up to 20 to 30 percent grade. The disk chain is an 

implement that combines design features of the anchor 

chain and disk implements. The railer and pipe harrow 

are implements that are dragged behind tractors. They 

are designed to remove mature sagebrush and leave some 

smaller plants intact. Damage to herbaceous species is 

minimal; however, the pipe harrow creates enough soil 

disturbance to facilitate seed burial and establishment of 

desired species (Welch 2005). Equipment and practices 

should be selected to minimize risks to soil erosion. 

Archeological surveys are required prior to mechanical 

treatment in order to avoid cultural site disturbance. These 

treatments are generally used for FRCC 2 and 3.

Restoration Plantings

The concept of restoration planting can be defined in either 

broad or narrow terms. Narrowly, a restoration planting is 

seen as an attempt to re-establish a native plant community that is indistinguishable, or nearly so, in composition, 

structure, and ecological process from what is perceived 

as the natural state. This view dictates a careful selection 

of source germplasms for plant propagules, usually seeds, 

which are consistent with the goal. Although a worthy 

target to aim for, actualization of this kind of restoration 

is generally difficult to achieve for big or black sagebrush 

communities. Alternatively, a broader view of restoration 

plantings includes all attempts to establish complimentary 

assemblages of plant species that structurally and functionally resemble pre-disturbance conditions in so far as the 

level of site degradation will allow. Developed cultivars of 

native and introduced species may be planted in various 

combinations deemed most likely to achieve goals of site 

stabilization and other management objectives. Intermediate approaches with varying restrictions on plant material origin are common for big sagebrush-grass plantings 

(Roundy and others 1997; Stevens and Monsen 2004b). 

Although restoration practices vary, most are designed to 

either facilitate natural repair processes or supplant them 

(Whisenant 1999). Examples of facilitative actions associated with the repair of big sagebrush-grass communities 

might include protection of residual big sagebrush islands 

to allow natural seed dispersal into surrounding treated or 

disturbed landscapes (Longland and Bateman 2002) and 

delaying livestock grazing until after perennial grass seed 

shatter to allow maximum seed production and dispersal 
90 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 5—Big and Black Sagebrush Landscapes

on soil surfaces. Although facilitative restoration can be a 

slow process, costs are relatively low, allowing treatment 

of large areas.

Most efforts at developing restoration methodologies for big sagebrush-grass communities have taken 

a more direct, essentially agronomic approach. Over 

time, equipment and methods developed for the efficient establishment of crop monocultures on uniform, 

submissive environments were adapted and modified 

for planting a wide variety of seed types on highly variable and sometimes harsh environments (Keller 1979; 

Monsen and Stevens 2004; Young and Evans 1987). 

The greatest innovations have been associated with the 

collection, cleaning, and planting of native shrubs and 

forbs (Jorgensen and Stevens 2004). General principles 

and guidelines for big and black sagebrush restoration 

plantings that have stood the test of time and experience are discussed by Monsen and Stevens (2004) and 

Stevens and Monsen (2004b). First, competition from 

weedy species must be controlled. The seedbed should 

be firm, but not overly compacted. Effects of litter or 

mulching vary by species. Larger seeds should be sown 

at a depth of 1 to 2 cm (0.4 to 0.8 inches). Seeds of small 

seeded species, such as sagebrush, must be placed at or 

near the soil surface. It is generally best to plant seeds 

of slower growing forbs and shrubs separate from those 

of grasses. The optimal time for planting is from late 

fall to early winter, allowing for maximum use of winter 

and spring soil moisture and removal of seed dormancy. 

For seed collected from wild populations, climate and 

soils of the collection site should match those of the 

treatment site. Published seed transfer zones similar to 

those produced for trees species have been developed 

for a few key species (Mahalovich and McArthur 2004). 

More are needed. Although much progress has been 

made in recent decades, seeds of many species desirable 

for restoration plantings are either not available, or are 

available only in small quantities (McArthur and others 

1987; McArthur and Young 1999; Roundy and others 

1997). Too little is known of the biology of many of 

these species to plan for their efficient use. Additional 

research is needed to ascertain relationships between 

soil water and temperature and seed germination and 

seedling growth (Roundy 1994). The potential effects 

of present and future changes in biological and physical 

environments on community stability are not well understood and are in need of thoughtful attention. Finally, 

a commitment must be forged to manage preserved and 

restored big sagebrush landscapes for long-term sustainability or the degradation – restoration cycle will become 

a permanent feature of the big sagebrush landscape.

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USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 97

Chapter 6

Creosotebush, Blackbrush, and Interior 

Chaparral Shrublands

Matthew L. Brooks, Todd C. Esque, and Tim Duck

Introduction

The vegetation, fire regime, and Fire Regime Condition Class descriptions in this paper apply broadly to the 

Mojave Desert, Colorado Plateau, and southern Great 

Basin of western North America. More detail on these 

topics, including estimated percentages within each condition class, is provided for the Mojave-Colorado Plateau 

ecotone spanning southern Nevada, the Arizona Strip 

northwest of the Grand Canyon, and southwestern Utah, 

including the area within the boundaries of the Southern 

Utah Fuel Management Demonstration Project (Hood 

and others, this volume; www.firelab.org/fep/research/

sufm/home.htm). Fire Regime Condition Classes (FRCC) 

are an interagency, standardized tool for describing the 

natural range of variation in vegetation, fuels, and fire 

regime characteristics for a particular biogeographic 

region or vegetation type. They summarize how past 

land use and land management actions (or inactions) 

may have caused the FRCC to change.

The three shrubland vegetation types that we review 

are creosotebush scrub, blackbrush, and interior chaparral, which are typically situated along an increasing 

elevation gradient where they co-occur. The interior 

chaparral vegetation type is sometimes considered to 

include both Arizona chaparral, which occurs mostly in 

Arizona and New Mexico, but also in southern Utah and 

Nevada, and mountain brush (or Petran chaparral), which 

occurs primarily farther north in Utah and Colorado. 

We focus on Arizona chaparral. Two vegetation types 

dominated by big sagebrush (Artemisia tridentata) and 

pinyon-juniper (Pinus edulis, Pinus monophylla, and 

Juniperus spp.) often intergrade between blackbrush and 

interior chaparral in the alluvial soils of broad valleys 

and foothill toe-slopes. Kitchen and McArthur (this 

volume) reviewed big sagebrush and Tausch and Hood 

(this volume) reviewed pinyon-juniper.

Creosotebush Scrub

Introduction

Low cover (5 to 30 percent) of woody shrubs of 

various heights (0.5 to 1.5 m [1.6 to 4.9 ft] characterizes Creosotebush scrub (Vasek and Barbour 1995). 

It occurs across the warm desert regions of western 

North America and is the most common plant assemblage in the Mojave Desert (fig. 1) (MacMahon 2000). 

Creosotebush scrub is typically found below 1,500 m 

(4,920 ft) on well-drained alluvial flats and slopes below 

the blackbrush zone and above the saltbush zone that 

often occur within valley basins (Vasek and Barbour 

1995). It phases into shrub-steppe in regions with high 

proportions of summer rainfall, typically encountered 

in the eastern Mojave Desert and Sonoran Desert.

Creosotebush scrub is dominated by the type-species 

creosotebush (Larrea tridentata), which has the highest cover and is the most wide-ranging plant species 

in the Mojave Desert (Rowlands and others 1982). 

It is most frequently associated with white bur-sage 

(Ambrosia dumosa), but a wide range of other plants 

can co-occur with creosotebush, including goldenhead 

(Acamptopappus spp.), saltbush (Atriplex spp.), Mormon 

tea (Ephedra spp.), goldenbush (Ericameria spp.), wild 

buckwheat (Eriogonum fasciculatum), ratany (Krameria

spp.), winterfat (Krascheninnikovia lanata), boxthorn 

(Lycium spp.), indigo bush (Psorothamnus spp.), desert 

needlegrasses and Indian ricegrass (Achnatherum spp.), 

galleta grass (Pleuraphis spp.), cholla, beavertail, and 

other cacti (Opuntia spp.), and Joshua tree, Mojave yucca, 

and others (Yucca spp.). Dominant non-native species 

include red brome (Bromus rubens), Mediterranean 

split-grass (Schismus spp.), and red-stemmed filaree 

(Erodium cicutarium).
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Chapter 6—Creosotebush, Blackbrush, and Interior Chaparral Shrublands

Within the Mojave-Colorado plateau ecotone, creosotebush scrub reaches its northeast limit in southwestern Utah near the entrance to Zion National Park 

and follows the Virgin River Valley to Lake Mead 

in Nevada. It occurs in most of this river drainage 

below 1,250 m (4,100 ft) elevation. Its distribution 

broadens below the Hurricane Fault in Utah, spreading 

northward as far as Browse along Interstate 15 and 

then following the base of the red cliffs throughout 

the Dixie Valley north of Washington, St. George, 

and Gunlock, Utah. There is a large stand south of 

St. George in the Blake Lambing Grounds and other 

stands are located throughout the Virgin River Gorge. 

Below the Gorge, the distribution spreads from the foot 

of the Beaver Dam Mountains and around Utah Hill, 

up the Beaver Dam Wash northward to Jackson Wash 

almost meeting with the creosotebush scrub community west of the town of Gunlock. From Littlefield, 

Arizona, there is continuous creosotebush scrub all 

the way to Lake Mead and Las Vegas, Nevada, with 

only minor interruptions in very rocky areas.

Creosotebush and associated plants provide much of 

the microhabitat diversity and vertical habitat structure in 

the Mojave Desert. Branches provide perches and nesting opportunities for songbirds, while desert tortoises 

Figure 1—Creosotebush (Larrea tridentata) scrub in the Mojave Desert. 

(Gopherus agassizii) and a variety of rodents incorporate the root structure into their burrows. Shrubs also 

provide concealment and escape cover from predators 

and shade from the sun.

Many Mojave Desert shrubs are not fire-tolerant 

because their drought-adaptive features (thin bark, 

shallow root system, small leaves) and high dead-to-live 

woody material ratio make them vulnerable to fire. Some 

species, including creosotebush, can resprout after burning (fig. 2); however, survival rates decline significantly 

if more than 10 percent of their aboveground biomass is 

consumed by fire (Brooks and Minnich 2006). Native 

plants are generally slow to re-establish after fire 

and recurrent fire may prevent their re-establishment 

(Brooks and Minnich 2006; Brown and Minnich 1986; 

O’Leary and Minnich 1981). The loss of native plants 

can be followed by increased dominance of non-native 

annual grasses. The post-fire vegetation has typically 

lower species diversity and plant structural diversity 

than the native community, which can negatively affect 

the desert tortoise (Brooks and Esque 2002; Esque and 

others 2003) and other desert wildlife. Repeated burning 

in creosotebush scrub can lead to significant decreases 

in plant species richness with each subsequent fire 

(M. Brooks, unpublished data).
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Chapter 6—Creosotebush, Blackbrush, and Interior Chaparral Shrublands

Historical Conditions

We do not know the pre-settlement fire conditions in 

warm desert plant communities such as creosotebush 

scrub because the typical analytical methods, such as 

dendrochronology and evaluating charcoal deposits in 

lakes, are not possible where the requisite trees and 

lakes are not present (Brooks and Minnich 2006; Esque 

and Schwalbe 2002). It is generally thought that fires 

in creosotebush scrub were an infrequent event in presettlement desert habitats because fine fuels from winter 

annual plants were probably sparse, only occurring in 

large amounts after exceptionally wet winters (Brooks 

and Esque 2002; Brooks and Minnich 2006; Brown and 

Minnich 1986; Esque and Schwalbe 2002; Humphrey 

1974; O’Leary and Minnich 1981; Salo 2003). Fires 

were probably more frequent in adjacent shrub-steppe 

areas where perennial grasses provided additional fuel 

continuity to carry fire. It appears that wildfire was 

not historically a dominating influence in creosotebush 

scrub landscapes, except possibly where it intergraded 

with shrub-steppe.

Creosotebush scrub fuels are comprised primarily of 

woody shrubs, but it is the fine fuels from annuals and 

perennial grasses that facilitate the ignition and spread 

of fires. Native annual plants usually break down rapidly 

during the summer after they are produced and do not 

create a long-lived fuelbed (Brooks 1999a). As a result, 

the historical annual plant fuelbed, prior to the invasion 

by non-native species, was probably transient, only lasting 

for one summer fire season after winters of exceptional 

high rainfall. Perennial grasses may have been more 

prevalent in creosotebush scrub before the introduction 

of livestock grazing, providing another source of fine 

fuels that could additionally have helped carry fire in 

the past (Brooks and Minnich 2006). However, stands of 

perennial bunchgrasses were probably patchy, reflecting 

localized soil conditions, and the discrete clumps of fine 

fuels created by individual bunchgrasses did not likely 

create a continuous fuelbed. 

Since settlement in the 1860s, creosotebush scrub has 

been used for livestock operations along the Mojave 

Desert-Colorado plateau ecotone and elsewhere 

throughout its range. Over the past two centuries, 

there has been widespread cattle and sheep use of 

this habitat. Widespread livestock grazing probably 

promoted the invasion of the annual grasses red brome, 

cheatgrass, and Mediterranean split-grass (Brooks 

and Pyke 2001), which create more continuous fine 

fuelbeds that can persist for years (Brooks 1999a). 

Although these non-native annual grasses have probably been present for over 100 years (Brooks 2000a, 

b; Esque and Schwalbe 2002; Salo 2003), wildfire in 

creosotebush scrub has only been prevalent in this 

Figure 2—Creosotebush (Larrea tridentata) resprouting 1 year after a low intensity fire. 
100 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 6—Creosotebush, Blackbrush, and Interior Chaparral Shrublands

region since the 1970s (Jerry Empey – BLM Color 

Country South Zone dispatch, personal communication). The recent increase in fire frequency was 

coincident with the end of the mid-century drought 

(Hereford and others 2006) and the beginning of a 

20-year period of high rainfall in the late 1970s (Brooks 

and Esque 2002; Brooks and Minnich 2006). 

Current Conditions

Fine fuels from non-native annual grasses currently 

represent the most important fuelbed component in 

creosotebush scrub (fig. 3) (Brooks and Minnich 2006). 

Non-native annual grasses typically comprise >50 

percent of the total annual plant biomass in creosotebush scrub (Brooks 1999b; Brooks and Berry 2006). 

Annual plants were monitored as livestock/wildlife 

forage in unburned areas near St. George, Utah, from 

1980 through 1995 and ranged from 0 to 700 kg/ha (624 

lbs/acre) (BLM and T. Esque, unpublished data). In unburned areas near Littlefield, Arizona, the aboveground 

production of annual plants was over 1,000 kg/ha (892 

lbs/acre) in 1993, and was comprised mostly of nonnative annual grasses and forbs (Esque 1994). Above 

ground production of annual plants in a recurrently 

burned area of the Pakoon Basin of the Arizona Strip 

was measured at >2,000 kg/ha (1,784 lbs/ac) in 2001, 

and was mostly comprised of Bromus spp. (T. Esque, 

unpublished data).

During the 1980s and early 1990s, fire frequency 

increased substantially within the Mojave Desert (Brooks 

and Esque 2002; Brooks and Matchett 2006), and 

many of these fires occurred in creosotebush scrub. In 

the Opal Mountain and Stoddard Mountain regions of 

California, non-native annual grasses fueled recurrent 

fires, and some areas burned as many as three times in 

10 years (M. Brooks, unpublished data). This increase 

in fire frequency is attributed to high rainfall in 1983 

and 1992, which produced prodigious amounts of fine 

fuels, especially of red brome. Rainfall was positively 

correlated with fire size between 1980 and 2004 in the 

low-elevation ecological zone dominated by creosotebush scrub (Brooks and Matchett 2006). It appears 

that fine fuels produced during years of high rainfall 

are largely required for the development of large fires 

in this zone (Brooks and Matchett 2006; Brooks and 

Minnich 2006).

During the middle 1990s through 2004, the number of 

fires declined in lower and middle elevation zones where 

creosotebush predominates (Brooks and Matchett 2006). 

However, the amount of area burned at the upper elevation creosotebush ecotones increased slightly, reflecting 

a trend toward increasingly larger fires. This trend was 

punctuated during the summer of 2005 when hundreds 

of thousands of acres burned in areas dominated by 

creosotebush. The current fire return interval in this 

region is still unknown, but it seems to be decreasing 

due to the abundance of non-native invasive grasses 

after particularly wet years.

Figure 3—Highly flammable 

understory of red brome 

(Bromus rubens) in 

creosotebush (Larrea 

tridentata) scrub.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 101

Chapter 6—Creosotebush, Blackbrush, and Interior Chaparral Shrublands

Fire Regime Condition Classes 

Fire Regime Condition Class 1 (FRCC 1)

FRCC 1 is characterized by vegetation and fire regime 

attributes within the natural range of variation. The risk 

of losing key ecosystem components, such as habitat 

diversity and cover sites for the desert tortoise, is low. 

This is the natural condition where invasive annual 

grasses are largely absent, thus precluding the fine fuel 

buildup usually required to carry fire at lower and middle 

elevations in the Mojave Desert (Brooks and Matchett 

2006; Brooks and Minnich 2006). Fire return intervals 

are extremely long, except for areas near the base of 

mountains that experience locally higher rainfall and 

fine fuel buildup from native annuals, or where creosotebush scrub intergrades with shrub-steppe. There 

is currently zero percent of the creosotebush scrub in 

FRCC 1 at the Mojave Desert-Colorado plateau ecotone 

in southern Utah, southern Nevada, and northwestern 

Arizona, because invasive, non-native annual grasses 

have infested the entire area. However, many of the more 

arid regions of the Mojave Desert further to the southwest 

in California may classify as Condition Class 1.

Fire Regime Condition Class 2 (FRCC 2)

FRCC 2 is characterized by vegetation and fire regime attributes that have been moderately altered from 

their natural range. The risk of losing key ecosystem 

components, such as habitat diversity and cover sites 

for the desert tortoise, is moderate. Patchy fires of limited extent are possible in more mesic regions during 

any year. Risk of larger, more continuous fires is high 

during 10 to 20 percent of all years when rainfall is 

particularly high. Fire risk can be exceptionally high 

when successive years of above average rainfall promote 

the accumulation of non-native annual grass fuels (in 

other words, production values in excess of 300 kg/ha 

(268 lb/acre) in 2 or more successive years. An example 

of this dynamic occurred in the Dixie Valley north of 

St. George, Utah, and on the Beaver Dam Slope north 

of Littlefield, Arizona (T. Esque, unpublished data). 

Approximately 80 percent of the creosotebush scrub at 

the Mojave Desert-Colorado Plateau ecotone is in this 

condition class.

Fire Regime Condition Class 3 (FRCC 3)

FRCC 3 is characterized by vegetation and fire regime 

attributes that have been significantly altered from their 

natural range. The risk of losing key ecosystem components, such as habitat diversity and cover sites for the 

desert tortoise, is high. Wildfires are frequent and widespread. Examples of FRCC 3 areas exist in the Pakoon 

Basin of northern Arizona, the Tule Desert of southern 

Nevada, and the Opal Mountain and Stoddard Valley 

areas of California. Return intervals in this condition 

class may be as short as 5 years. Fuel loads for areas in 

FRCC 3 may exceed 2,000 kg/ha (1,784 lbs/acre) during 

peak years (T. Esque, unpublished data). The invasive 

plant/fire regime cycle characterized by this condition 

class creates a feedback loop of decreasing habitat quality for wildlife and livestock. Native seedbanks may 

be depleted, lengthening the recovery times for native 

plants. Approximately 20 percent of the creosotebush 

scrub at the Mojave Desert-Colorado plateau ecotone 

is in this condition class.

Recommended Treatments

It is unlikely that areas can ever be restored to FRCC 1 

due to the widespread distribution of non-native annual 

grasses, but the size of their impact may be mitigated 

through appropriate land management policies. We 

recommend suppression of all wildfires (because most 

native vegetation in this community vegetation type 

responds poorly to fires) and minimization of surface 

disturbances that promote the dominance of non-native 

annual grasses. Prescribed fires should not be conducted 

except for small research burns designed to evaluate fire 

behavior, fire effects, and fire management techniques 

and treatments. We also recommend a program of early 

detection, evaluation, and eradication for new invasive 

plants before they become established. Many new plant 

species are in the process of invasion, and some pose 

potential fire threats due to their ability to produce 

large amounts of continuous fine fuels. New invaders 

of particular concern are mustards (for example, Brassica tournefortii and Hirschfeldia incana) and perennial grasses (for example, Pennisetum setaceum and 

Cenchrus ciliaris) (Brooks and Esque 2002).

In FRCC 2 areas suppress all wildfires and avoid 

conducting prescribed fires except for small research 

burns. Livestock grazing may reduce fine fuel loads 

temporarily and may be effective for managing fuels in 

small defined areas, such as at the wildland urban interface. However, regular grazing is required to maintain 

these managed fuel zones, except during years of very 

low rainfall. Regular grazing may reduce dominance 

of late seral native plants and increase the dominance 

of non-native and early seral plants that are often more 

flammable.
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Chapter 6—Creosotebush, Blackbrush, and Interior Chaparral Shrublands

We recommend studies designed for the restoration of FRCC 3 sites, which can include research on 

adapted plant materials, propagation techniques, and 

planting techniques. Suppress all wildfires and avoid 

conducting prescribed fires except for small research 

burns. Livestock grazing to reduce fuel loads may be 

counterproductive in the long run if it also hinders the 

re-establishment of late seral native plants.

Blackbrush

Introduction

The blackbrush vegetation type is characterized by 

relatively high cover (50 percent) of low statured (50cm 

[20 inches] tall), woody evergreen shrubs (fig. 4). It occurs at the bioregional transition between the Mojave and 

Great Basin deserts, from California through Nevada, 

Arizona, and Utah (Bowns 1973). Blackbrush is typically found in the elevational zone from 1,220 to 1,520 

m (4,000 to 4,985 ft) above the creosotebush zone and 

below the interior chaparral or big sagebrush/pinyonjuniper zones (Beatley 1976; Bradley and Deacon 1967; 

Randall 1972). Within the Mojave-Colorado plateau 

ecotone, blackbrush is found on dry slopes and benches 

above the river canyons of southern Utah and northern 

Arizona (Turner 1994). It is also found mid-slope on 

mountain ranges throughout this ecotone.

Blackbrush is dominated by the type-species, blackbrush (Coleogyne ramosissima), which can comprise 

90 to 95 percent of the total plant cover (Shreve 1942). 

Cover of blackbrush is highest in late seral stands on 

shallow, sandy soils with strong petrocalcic (caliche) 

horizons where it is the primary dominant plant species. 

Cover of blackbrush is lowest in deeper, silty soils, or 

at its upper or lower ecotones, where it is co-dominant 

with other native species such as creosotebush, juniper, 

desert almond (Prunus fasciculata), Anderson wolfberry 

(Lycium andersonii), Joshua tree (Yucca brevifolia), bladder sage (Salazaria mexicana), desert needle grasses, 

Indian ricegrass, and galleta grass. Dominant non-native 

species include the annuals red brome, cheatgrass, and 

red-stemmed filaree.

Blackbrush is used as winter forage by deer and bighorn 

sheep (Bowns and West 1976), and provides cover for 

nongame birds and small mammals (Brown and Smith 

2000). It also protects soil from water and wind erosion 

and promotes soil fertility (Bowns 1973).

Figure 4—Blackbrush (Coleogyne ramosissima) dominated site with thread snakeweed 

(Gutierrezia microcephala), desert peach (Prunus fasciculata), and Joshua tree (Yucca brevifolia). 
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 103

Chapter 6—Creosotebush, Blackbrush, and Interior Chaparral Shrublands

Blackbrush shrublands are one of the more flammable 

vegetation types in the desert bioregion due to the high 

proportion of fine fuels and optimal fuel bulk density. 

Blackbrush fires are usually stand-replacing, burning 

plants to the ground and killing most of them. Individual 

fires can kill upwards of 80 percent of the seeds of all 

species in the soil seedbank (Brooks and Draper 2006). 

It is commonly thought that blackbrush stands take centuries to recover (Bowns 1973; Webb and others 1987). 

However, when only a portion of a blackbrush plant is 

consumed, it may survive and resprout from the root 

crown (Bates 1984; M. Brooks, personal observation). 

Analyses of historical photographs from Joshua Tree 

National Park and southern Nevada also indicate that 

blackbrush stands can recover after as little as 50 to 75 

years (Minnich 2003; M. Brooks, unpublished data). It 

seems probable that the ability of blackbrush to resprout 

after burning varies across its wide geographic range, 

which extends from the Colorado Plateau and southern 

Great Basin on through the Mojave Desert. In general, 

the frequency of blackbrush resprouting after partial 

consumption by fire seems to be highest in more mesic 

areas at the edges of its geographic range (M. Brooks, 

personal observation).

Historical Conditions

Blackbrush is considered a poor livestock forage species and ranchers noted that during the late 1930s and 

early 1940s, wildfires increased production of livestock 

forage in blackbrush rangeland of southern Nevada and 

northwestern Arizona (Anonymous 1945). In an attempt 

to further increase forage production, ranchers and the 

Bureau of Land Management (BLM) began a program 

of prescribed burning in the 1940s, during which time 

approximately 20 percent of the 161,875 ha (400,000 

acres) of blackbrush were burned by prescribed fire or 

wildfire in southern Nevada (BLM, Las Vegas, Nevada, 

grazing district 5) (Croft 1950). Many blackbrush fires 

also occurred in northwestern Arizona during this time 

(BLM, Arizona Strip, Arizona, grazing district 2). Additional blackbrush burning likely occurred at least 

through the 1960s because a policy review during that 

time by the Range and Forestry Officer of the Bureau 

of Land Management in Nevada recommended that 

blackbrush burning be continued to increase livestock 

forage (Dimock 1960). Before 1940, fires in these regions 

were relatively uncommon (Croft 1950). The long-term 

effects of these mid-century range burns are currently 

being evaluated using repeat photography of historical photos originally taken 5 to 10 years post-fire and 

analyses of field reports, memos, and vegetation plot 

data collected during the late 1940s through the early 

1960s by range conservationists and foresters from the 

BLM and the Forest Service, Intermountain Forest and 

Range Experiment Station (M. Brooks, unpublished 

data).

Prior to European contact (pre-settlement), late seral 

blackbrush stands were probably more extensive than 

they are today. The vast expanses of blackbrush rangeland that were burned to improve livestock production 

during the mid-1900s are still dominated by early seral 

species and have been re-colonized only sporadically by 

blackbrush (M. Brooks, unpublished data). Blackbrush 

within the Desert Wildlife Range and Nevada Test Site in 

southern Nevada that has not been managed for livestock 

production since the 1930s, and likely was not burned 

for rangeland improvement, does not currently contain 

evidence of widespread historical burning (M. Brooks 

and T. Esque, personal observations). It therefore appears 

that extensive burning to remove blackbrush probably 

created many of the vegetation stands where blackbrush 

is either absent or a sub-dominant species today.

The historical fuel complex in late seral blackbrush 

stands was probably similar to that observed in relatively 

undisturbed sites today, except for the current prevalence 

of Bromus spp. and Erodium cicutarium in many stands 

(Brooks and Matchett 2003). Vegetation characteristics of 

these stands were characteristic of blackbrush in FRCC 

1 described below. Shrub cover was likely comprised 

primarily of blackbrush at 30 to 50 percent total cover, 

and interspaces were probably mostly bare, even during years of high rainfall, due to root competition from 

blackbrush. Other species, such as perennial grasses 

and early seral shrubs, probably occurred sporadically, 

as they do today, along wash stringers and on steep 

hillslopes where cover of blackbrush is typically low.

Low amounts of fine fuels in interspaces probably 

limited fire spread to only extreme fire weather conditions during which high winds, low relative humidity, 

and low fuel moisture led to high intensity stand-replacing crown fires. Natural fire return intervals appear to 

have been on the order of centuries (Webb and others 

1987). The long intervals without fire allowed late seral 

blackbrush stands to re-establish.
104 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 6—Creosotebush, Blackbrush, and Interior Chaparral Shrublands

Current Conditions

Blackbrush is considered to be one of the most flammable native plant assemblages in the Mojave Desert. 

Many large fires have occurred in this vegetation type 

since the 1980s in the Spring Mountains and Mormon 

Mountains in Nevada, the Beaver Dam Mountains in 

Utah, the Black Mountains and Virgin Mountains in 

Arizona, and at Joshua Tree National Park in California 

(Brooks and Esque 2002). Although fire is generally 

no longer advocated as a tool for range improvement, 

ignitions from lightning and accidental ignitions along 

roads have been sufficient to burn significant acreage 

of blackbrush during the past few decades. 

At Joshua Tree National Park in California, blackbrush 

was burned during the early 1990s to reduce woody fuel 

loads at the wildland-urban interface between Joshua 

Tree and the town of Yucca Valley. During the first few 

post-fire years, the landscape was dominated by native 

annual wildflowers, but by the fourth post-fire year, the 

non-native annual grasses red brome and cheatgrass

became the dominant annual plants and remained 

as such into the 2000s (Brooks and Matchett 2003; 

M. Brooks, personal observations). The appearance 

of this new flashy fuelbed resulted in a change in fire 

management at Joshua Tree. It put a stop to the use of 

fire as a management tool until prescriptions could be 

identified that would not create continuous fuelbeds of 

non-native annual grasses. At the current rate of burning 

during 1980s and 1990s, managers estimate that all the 

blackbrush at Joshua Tree National Park will burn by 

2015 to 2020 (Hank McCutchen, Chief of Resources, 

personal communication). This is a significant concern 

because the blackbrush stands located there are disjunct 

from the rest of the blackbrush range, and if all the stands 

were to burn, it is very likely that blackbrush would not 

be able to re-establish.

The fuel complex in blackbrush appears to be more 

conducive to burning now than in the past. Non-native annual grasses currently occur in most blackbrush 

stands (M. Brooks, personal observations), although 

their dominance can vary significantly among sites 

(Brooks and Matchett 2003). Post-fire landscapes are 

even more dominated by these non-native grasses, which 

raises concerns that they will promote recurrent fire and 

prevent the re-establishment of Coleogyne ramosissima. 

This link between fine fuels and fire size is supported 

by recent analyses that demonstrate that years of high 

rainfall, which lead to high production of fine fuels, 

are correlated with larger fires in the elevations where 

blackbrush occurs in the Mojave Desert (Brooks and 

Matchett 2006). This is in contrast to the conclusions 

of Minnich (2003) who states that fine fuels that respond to short pulses of rainfall have less of an effect 

on fire regimes in blackbrush than woody fuels that 

accumulate slowly over time. Both conclusions probably 

have some validity, with fine-fuels taking precedence 

at lower elevations of the blackbrush zone and woody 

fuels taking precedence at higher elevations. At lower 

elevations, spacing between blackbrush plants is often 

relatively high (~1m [3.3 ft]). This requires fine fuels to 

carry fire under most circumstances, whereas at higher 

elevations spacing between blackbrush plants is often 

relatively small (<50cm [20 inches] (M. Brooks, personal 

observation).

Reports from the mid-1900s also acknowledge the role 

that non-native annual grasses, especially red brome, 

can play in facilitating the spread of fire in blackbrush 

(Dimock 1960; Holmgren 1960; Jenson and others 

1960), although there was disagreement as to whether 

the burned landscapes were more or less susceptible 

to reburning than unburned landscapes. Jenson and 

others (1960) thought the chances of reburning were 

low because they observed low fine fuel levels in postfire landscapes. However, their observations were made 

during the mid-century drought (Hereford and others 

2006) when fine fuel loads were on the low end of their 

possible range. In contrast, Holmgren (1960), who accompanied Jenson and others on the same field visits, 

thought that the danger of accidental fire in blackbrush 

would be higher in areas that previously burned than 

in unburned areas if high winter rainfall had produced 

more red brome biomass and other fine fuels. Prior to 

the invasion of red brome and cheatgrass during the 

late 1800s to early 1900s (Brooks 2000a; Young 2000), 

fine fuel loads were likely not as great in either burned 

or unburned blackbrush stands. This resulted in fewer 

fires, possibly smaller fires, and fewer reburns.

Fire Regime Condition Classes 

Fire Regime Condition Class 1 (FRCC 1)

FRCC 1 is characterized by vegetation and fire regime 

attributes within the natural range of variation. The 

risk of losing key ecosystem components, such as high 

blackbrush cover and associated protection from soil 

erosion, is low. Mature blackbrush stands fall into this 

condition class. These stands are typically late seral 

with occasional early seral patches created by infrequent 

stand-replacing fires.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 105

Chapter 6—Creosotebush, Blackbrush, and Interior Chaparral Shrublands

The fire regime for FRCC 1 is active crown fire 

carried primarily by blackbrush and perennial grasses 

(Achnatherum spp. and Pleuraphis spp.) on deep silty 

soils. Burns are complete and stand-replacing, fire intensity is high, and fire return intervals are >100 years. 

Long fire return intervals allow for the typically slow 

process of blackbrush re-establishment. Approximately 

20 percent of the blackbrush at the Mojave DesertColorado Plateau ecotone in southern Utah, southern 

Nevada, and northwestern Arizona is currently in this 

condition class.

Fire Regime Condition Class 2 (FRCC 2)

FRCC 2 is characterized by vegetation and fire regime 

attributes that have been moderately altered from their 

natural range. The risk of losing key ecosystem components, such as high blackbrush cover and associated 

protection from soil erosion, is moderate. Blackbrush 

stands with an intermix of late seral unburned patches 

and early seral burned patches are in this condition 

class. Blackbrush stands with patches that have been 

degraded by overgrazing, prescribed burning in the 

mid-1900s, or other forms of surface disturbance also 

fall into this condition class. These disturbances reduce cover of blackbrush and increase cover of early 

seral shrubs such as rabbitbrush (Chrysothamnus spp.), 

snakeweed (Gutierrezia spp.), and wild buckwheat, early 

seral herbaceous perennials such as desert globemallow 

(Sphaeralcea ambigua) and milkvetch (Astragalus spp.), 

and non-native annual plants such as red brome, cheatgrass, and red-stemmed filaree. Burned stands without 

livestock over-grazing that are situated on deep, silty 

soils can also have a large perennial grass component 

(Achnatherum spp. and Pleuraphis spp.). 

Fires in late seral patches are active crown fires carried by blackbrush, perennial grasses (Achnatherum

spp. and Pleuraphis spp.), and non-native annual grasses 

(Bromus spp.). Burns are complete and fire intensity is 

high. Fires within early seral patches are passive crown 

fires or surface fires carried by perennial grasses and 

non-native annual grasses between the sparse cover of 

early seral shrubs. Burns are patchy and fire intensity 

is low to moderate.

Fire return intervals in FRCC 2 stands are approximately 50 to 100 years. This shorter fire return interval, and 

over-grazing pressure from livestock, helps to maintain 

dominance by early seral species and may prevent reestablishment by blackbrush. Approximately 70 percent 

of the blackbrush at the Mojave Desert-Colorado plateau 

ecotone is in this condition class.

Fire Regime Condition Class 3 (FRCC 3)

Condition Class 3 is characterized by vegetation 

and fire regime attributes that have been significantly 

altered from their natural range. The risk of losing key 

ecosystem components, such as high blackbrush cover 

and its associated protection from soil erosion, is high. 

Blackbrush stands that burned during the 1900s, and 

have reburned at least once, fall into this condition class. 

These stands are typically dominated by non-native 

annuals and early seral perennials. Recurrently burned 

stands without livestock over-grazing can also have a 

large perennial grass component (Achnatherum spp. 

and Pleuraphis spp.). 

The fire regime for this condition class is typically 

surface fire carried primarily by non-native annual 

plants. Burns are patchy, fire intensity is low, and fire 

return intervals are <50 years. Re-establishment by 

blackbrush is highly unlikely under this fire regime. 

Approximately 10 percent of the blackbrush at the 

Mojave Desert-Colorado plateau ecotone is in this condition class, but this percentage could increase if ignition 

rates increase with burgeoning human populations and 

if fine fuel continuity increases due to increased dominance of non-native annual grasses that could occur if 

atmospheric CO2 and rainfall levels increase (Brooks 

and Pyke 2001).

Recommended Treatments

In FRCC 1 areas suppress human-caused fires, but 

consider allowing lightning-caused wildfires to burn 

unless significant populations of red brome or cheatgrass 

are present, or there are other reasons for suppression. 

Prescribed fires should not be conducted except for small 

research burns designed to evaluate fire behavior, fire 

effects, and fire management techniques and treatments. 

Only apply fuels management treatments at the wildland 

urban interface to reduce fire hazard, or in wildland areas where fuel breaks are deemed necessary to achieve 

management goals. Realize that regular maintenance 

may be required to maintain these managed fuel zones 

because fuel treatments that involve replacement of late 

seral woody fuels with early seral fine fuels will reduce 

fire intensity, but may increase susceptibility to ignition 

and rates of fire spread.

In FRCC 2 areas suppress all wildfires. Prescribed 

fires should not be conducted except for small research 

burns. Minimize livestock grazing and other surface 

disturbances on early seral stands, or where early seral 
106 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 6—Creosotebush, Blackbrush, and Interior Chaparral Shrublands

and late seral stands are intermixed. Do not apply fuels 

management treatments on late seral stands, except possibly at the wildland urban interface to reduce fire hazards or in wildland areas where fuel breaks are deemed 

necessary to achieve management goals. Realize that 

regular maintenance may be required to maintain these 

managed fuel zones because fuel treatments that involve 

replacement of late seral woody fuels with early seral 

fine fuels will reduce fire intensity, but may increase 

susceptibility to ignition, fire spread rates, and fire 

frequency. Fuels treatments on early seral stands dominated by non-native annual plants may include the use of 

grass-specific herbicides. Carefully managed livestock 

grazing may reduce fine fuel loads temporarily, but may 

hinder the re-establishment of blackbrush and other late 

seral species and thus, may be counterproductive in the 

long-term.

In FRCC 3 areas suppress all wildfires. Prescribed 

fires should not be conducted except for small research 

burns. Extreme measures may be required in these 

stands. Revegetation with blackbrush and other late 

seral shrubs and perennial grasses, and exclusion of 

livestock grazing and other surface disturbances, may 

be necessary. Control of non-native annual grasses using 

herbicides or early season prescribed fire implemented 

immediately before revegetation treatments may improve 

initial establishment rates of revegetated plants.

Interior Chaparral

Introduction

Interior chaparral (Arizona chaparral) is characterized 

by moderate cover, ranging from 40 percent at dry sites 

to 80 percent at wetter sites, of moderately tall statured (1 

to 2.5 m [3.3 to 8.2 ft] tall), woody evergreen shrubs with 

dense crowns (fig. 5) (Carmichael and others 1978). It is 

best represented in the foothills and mountain slopes and 

canyons in the sub-Mogollan region of central Arizona 

(Pase and Brown 1994). Disjunct stands also occur in 

the mountains of southwestern Utah, southern Nevada, 

northwestern and southeastern Arizona, southern New 

Mexico, southwest Texas, and northern Mexico. It generally occurs between 1,000 to 2,000 m (3,280 to 6,560 

ft) in elevation above the desert grassland/shrub-steppe 

zone and below the pinyon-juniper/ponderosa pine (Pinus 

ponderosa) zone (Keeley 2000).

Within the Mojave-Colorado Plateau ecotone, interior chaparral is best represented in the Pine Valley 

Mountains, Bull Valley Mountains, and Zion National 

Park in Washington County (Utah), the Virgin Mountains 

of Mojave County (Arizona), and the Spring Mountains 

and Gold Butte regions of Clark County and various 

other mountain ranges in Lincoln County (Nevada). 

In this region, it is characterized by turbinella live oak 

Figure 5—Interior chaparral dominated by turbinella live oak (Quercus turbinella).
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 107

Chapter 6—Creosotebush, Blackbrush, and Interior Chaparral Shrublands

(Quercus turbinella), buckbrush (Ceanothus greggii), 

pointleaf manzanita (Arctostaphylos pungens), Wright 

silktassel (Garrya wrightii), and narrowleaf yerbasanta (Eriodictyon angustifolium). Understory grass 

species, such as grama (Bouteloua spp.) and three-awn 

(Aristida spp.), occur throughout the range of interior 

chaparral.

Interior chaparral is an important vegetation community for wildlife. Game animals such as javelina, deer, 

and bighorn sheep use several plant species as forage and 

others as cover. Distinct from the communities above 

and below it, interior chaparral provides a significant 

amount of wildlife habitat diversity.

Most chaparral species are adapted to a fire-prone 

system, but methods of response vary. Turbinella live 

oak is well adapted to survive fire, typically resprouting vigorously from the root crown and rhizomes in 

response to fire or other disturbance (Pase 1969), 

while buckbrush and pointleaf manzanita regenerate 

from long-lived seeds that accumulate in the soil and 

germinate prolifically following fire. Areas dominated 

by non-sprouting chaparral species, including pointleaf 

manzanita, may develop a persistent cover of herbaceous 

species following fire, especially where Bromus spp. are 

present. Interior chaparral dominated by turbinella live 

oak, have fire return intervals from 74 to 100 years. At 

least 20 years may be required before these sites can 

reburn (Cable 1975).

Historical Conditions

In some areas prior to the 1900s, interior chaparral 

was considered good livestock range, but overgrazing 

has removed much of the perennial grass component 

since then (Nichol 1937). Most people now view interior 

chaparral as a nuisance or obstacle to human activity. 

Ranchers have routinely attempted to get rid of the brush 

that competes with forage grass and makes travel difficult. Livestock grazing and fire suppression have at 

least partly caused shrub encroachment into grassland 

and woodland/forest encroachment into shrublands 

during the 1900s in the southwestern United States 

(Hastings and Turner 1965; Leopold 1924; Miller and 

Rose 1999).

Many lower elevation interior chaparral sites have been 

managed for livestock grazing since the 1880s (Pase and 

Brown 1994). Where fire was used to maintain grass 

forage, interior chaparral probably did not encroach into 

lower elevation grasslands. However, where fire was not 

used, and the removal of fine fuels by livestock grazing 

and fire suppression further decreased the frequency of 

wildfire, interior chaparral very likely did encroach into 

lower elevation grasslands. Large areas near the early 

settlements of Prescott and Globe, AZ, were reported to 

be grasslands in the 1860s and became dense stands of 

interior chaparral by 1936 (Cable 1975). Aldo Leopold 

(1924) reported a substantial increase in “brush” cover 

since the 1880s at the expense of herbaceous plant cover 

after 40 years of livestock grazing. 

Aldo Leopold made additional observations at the interior chaparral–grassland ecotone in southern Arizona, 

which sheds some light on the pre-settlement fire regime 

of this region. He noted during the early 1920s that 

there were multiple fire scars on ancient juniper stumps 

embedded in even-aged chaparral stands consisting of 

shrubs <40 years old. This suggests that the fire scars 

were created during low intensity grassland fires that 

pre-dated the current chaparral stands (Leopold 1924). 

Based on observations such as these, Leopold concluded 

that there had been no widespread fires in the chaparral–grassland ecotone in southern Arizona between the 

early 1880s and early 1920s. He further hypothesized 

that previous grassland fires at these same sites occurred 

at intervals of approximately once every 10 years before 

the advent of widespread livestock grazing. 

Higher elevation interior chaparral sites likely did not 

receive as much grazing pressure, but fire suppression, 

especially at the interface with ponderosa pine forests, 

may have resulted in forest encroachment into chaparral 

shrublands. For example, where old chaparral stands 

intergrade with woodlands or forests at higher elevations, 

chaparral species such as Pringle manzanita (Arctostaphylos pringlei) and Fendler ceanothus (Ceanothus 

fendleri) may be replaced by ponderosa pine, emory oak 

(Quercus emoryi), or Arizona oak (Quercus arizonica) 

after long fire-free intervals (Pase and Brown 1994).

Historical fire return intervals in interior chaparral 

were likely 50 to 100 years (Cable 1975). However, this 

is an average over its entire range, and local intervals 

probably vary widely.

Current Conditions

Interior chaparral presents a more complex management challenge than the forests above or the deserts 

below, facing some of the challenges of both. Like the 

ponderosa pine forests, chaparral communities are firedependant. Exclusion of fire from interior chaparral can 

lead to encroachments by woodland and forest species. 

Like the creosotebush scrub and blackbrush desert 
108 USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007

Chapter 6—Creosotebush, Blackbrush, and Interior Chaparral Shrublands

shrublands, non-native annual grasses can increase fire 

frequency to the point where even the fire-adapted interior chaparral cannot recover. Thus, interior chaparral 

requires fire, but not too much fire.

Conventional wisdom in the fire suppression community is that chaparral either does not burn, or when 

it does burn, it burns very intensely. Due to threats to 

humans and their property, as well as to other high value 

resources, full suppression remains the primary response 

to wildfires that occur under the hot, dry conditions of 

summer in the southwest.

Fire Regime Condition Classes

Fire Regime Condition Class 1 (FRCC 1)

FRCC 1 is characterized by vegetation and fire regime 

attributes within the natural range of variation. The risk 

of losing key ecosystem components, such as cover for 

wildlife, is low. These stands are typically late seral 

with occasional early seral patches created by infrequent 

stand-replacing fires. The fire regime for FRCC 1 is active crown fire carried primarily by turbinella live oak, 

buckbrush, and manzanita species. Burns are complete 

and stand-replacing, fire intensity is high, and fire return 

intervals are between 50 and 100 years. 

Long fire return intervals allow for the re-establishment of seed banks and the development of the fuel 

loads and spatial continuity necessary for fire to occur. 

Approximately 40 percent of the interior chaparral at 

the Mojave Desert-Colorado plateau ecotone in southern 

Utah, southern Nevada, and northwestern Arizona is in 

this condition class.

Fire Regime Condition Class 2 (FRCC 2)

FRCC 2 is characterized by vegetation and fire regime 

attributes that have been moderately altered from their 

natural range. The risk of losing key ecosystem components, such as cover for wildlife, is moderate. There are 

two types of chaparral within FRCC 2: 1) sites where 

the fire return intervals are greater than the natural 

range of variation (>100 years) and encroachment of 

interior chaparral by conifer woodlands has occurred (10 

percent of the Mojave Desert-Colorado plateau ecotone 

region) and 2) sites where fire return intervals, or other 

landscape-scale disturbances, are within or slightly less 

than the natural range of variation (≤50 to 100 years) 

and the post-disturbance community contains a significant intermix of late seral unburned patches and early 

seral burned patches dominated by non-native grasses 

(20 percent of the Mojave Desert-Colorado plateau 

ecotone region). These sites typically retain significant 

amounts of area with late seral interior chaparral characteristics, but the trend is toward FRCC 3.

Fire Regime Condition Class 3 (FRCC 3)

FRCC 3 is characterized by vegetation and fire regime 

attributes that have been significantly altered from their 

natural range. The risk of losing key ecosystem components, such as cover for wildlife, is high. Chaparral 

stands that have burned repeatedly and lost most of the 

shrub cover, have lost the shrub seed bank, and have 

significant amounts of Bromus spp. or other non-native 

grasses fall into this condition class. Non-native annuals and early seral perennials typically dominate these 

stands. 

The fire regime for this condition class is typically 

surface fire carried primarily by non-native annual 

plants. Burns are patchy, fire intensity is low, and fire 

return intervals are <20 years. Re-establishment by 

native interior chaparral shrub species is diminished 

under this fire regime and becomes less likely as native 

chaparral seed banks disappear. 

Approximately 30 percent of the interior chaparral at 

the Mojave Desert-Colorado plateau ecotone is in this 

condition class, but this percentage could increase if 

ignition rates increase with burgeoning human populations and if fine fuel continuity increases with increased 

dominance of non-native annual grasses that could occur 

if atmospheric CO2 and rainfall levels increase.

Recommended Treatments

In FRCC 1 areas suppress human-caused fires. Consider 

allowing lightning-caused fires to burn unless significant 

populations of Bromus spp. are present, especially near 

stands of non-sprouting chaparral species, or there are 

other reasons for suppression. Prescribed fires should not 

be conducted except for small research burns designed 

to evaluate fire behavior, fire effects, and fire management techniques and treatments. Apply fuel management 

treatments at the wildland urban interface to reduce fire 

hazard and in wildland areas where fuel breaks are deemed 

necessary to achieve management goals. Where possible, 

use indirect fire suppression tactics that provide a reasonable containment strategy to protect human life, property, 

and other valuable resources while allowing some acres to 

burn during the high-intensity summer fire season where 

deemed appropriate.
USDA Forest Service Gen. Tech. Rep. RMRS-GTR-202. 2007 109

Chapter 6—Creosotebush, Blackbrush, and Interior Chaparral Shrublands

Since this is a fire-dependent ecosystem, counter the 

effects of fire suppression in FRCC 1 areas by implementing burn treatments unless a significant threat of 

conversion to a Bromus-dominated system is present. 

The goal of these treatments would be to replicate 

normal patch size/mosaic patterns to provide diversity 

and to reduce the potential size of wildfires. Treatment 

frequency should not occur more often than 20 years at 

any site. Treatment areas should not be reseeded with 

herbaceous vegetation unless necessary to compete with 

non-native annual grasses. Treatments on sites dominated 

by seed propagators (for example, pointleaf manzanita) 

may need to be seeded with chaparral species.

Realize that regular maintenance may be required 

to maintain FRCC 1 managed fuel zones because fuel 

treatments that involve replacement of late seral woody 

fuels with early seral fine fuels will reduce fire intensity, 

but may increase susceptibility to ignition and rates of 

fire spread.

Treat FRCC 2 sites that have had fire excluded and 

are being encroached by conifer woodlands in a manner 

similar to FRCC 1, except where non-native grasses are 

a significant threat. In other words, suppress humancaused fires, but consider allowing lightning-caused fires 

to burn unless significant populations of Bromus spp. 

are present or there are other reasons for suppression. 

Prescribed fires should not be conducted except for small 

research burns. Apply fuels management treatments at 

the wildland urban interface to reduce fire hazard and 

in wildland areas where fuel breaks are deemed necessary to achieve management goals. Where possible, use 

indirect fire suppression tactics that provide a reasonable 

containment strategy to protect human life, property, 

and other valuable resources while allowing some acres 

to burn during the high-intensity summer fire season.

Treat FRCC 2 sites where disturbances have occurred, 

but the post-disturbance community has changed to one 

with a significant intermix of late seral unburned patches 

and early seral burned patches with non-native grasses, 

in a manner similar to FRCC 3. In other words, suppress 

all fires; revegetate with live oak, buckbrush, manzanita, and other late seral shrubs and perennial grasses; 

exclude livestock grazing and other surface disturbing 

activities at least until plants become established, and 

control non-native annual grasses using herbicides). 

In FRCC 3 areas suppress all wildfires. Prescribed 

fires should not be conducted except for small research 

burns. Revegetation may be necessary using turbinella 

live oak, Ceanothus spp., and manzanita species. and 

other late seral shrubs and perennial grasses, along 

with exclusion of livestock grazing and other surface 

disturbances. Control of non-native annual grasses and 

other invasive plants using herbicides immediately before 

revegetation treatments may improve initial establishment rates of revegetated plants and increase survival 

rates of regenerating plants.

We recommend research in adapted plant materials, 

propagation techniques, and planting techniques for 

restoration of these FRCC 3 sites. The grass/fire cycle 

creates a feedback loop of decreasing habitat quality 

for wildlife and livestock. Livestock grazing to reduce 

fuel loads may be counterproductive in the long run if 

it also hinders the re-establishment of late seral native 

plants.

Acknowledgments

This document reflects many of the peer-review 

comments and suggestions provided by Henry Bastian, 

Curt Deuser, Karen Phillips, Louis Provencher, and 

Tom Roberts. Financial support for the development of 

this document was provided by grant #00-2-32 to Matt 

Brooks from the Joint Fire Science Program.

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